Developments in Geotechnical Engineering, 82
Geoenvironmental Engineering
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D e v e l o p m e n t s in G e o t e c h n i c a l E n g i n e e r i n g , 82
Geoenvironmental Engineering A b d e l M.O. M o h a m e d
McGill University, Montreal, Quebec, Canada and H o g a n E. Antia
McGill University, Montreal, Quebec, Canada
1998 ELSEVIER Amsterdam m Lausanne -- New York--
Oxford -- Shannon -- Singapore -- Tokyo
ELSEVIER SCIENCE B.V. Sara Burgerhartstraat 25 P.O. Box 211, 1000 AE Amsterdam, The Netherlands
Librarv of Congress Cataloging in Publication Data:
Mohamed, Abdel~Mohsen Onsy. Geoenvlronmental engineering / Abdel
M.O. M o h a m e d and Hogan E.. Antia. p. cm. -- ( D e v e l o p m e n t s In g e o t e c h n l c a l e n g i n e e r i n g ; 82) Includes b i b l i o g r a p h i c a l r e f e r e n c e s and index. ISBN 0-444-89847-6
1. Environmental geotechnology. I I . Tit, le. III. Series. TD171.9.M65 1998 628--dc21
I.
Antia,
Hogan E . ,
1957-
.
98-6453 CIP
ISBN: 0-444-89847-6 91998 Elsevier Science B.V. All rights reserved. No part of this publication may be reproduced, stored in a retrieval system or transmitted in any form or by any means, electronic, mechanical, photocopying, recording or otherwise, without the prior written permission of the publisher, Elsevier Science B.V., Copyright & Permissions Department, P.O. Box 521, 1000 AM Amsterdam, The Netherlands. Special regulations for readers in the USA - This publication has been registered with the Copyright Clearance Center Inc. (CCC), 222 Rosewood Drive, Danvers, MA 01923. Information can be obtained from the CCC about conditions under which photocopies of parts of this publication may be made in the USA. All other copyright questions, including photocopying outside of the USA, should be referred to the publisher. No responsibility is assumed by the publisher for any injury and/or damage to persons or property as a matter of products liability, negligence or otherwise, or from any use or operation of any methods, products, instructions or ideas contained in the material herein. ~ ) The paper used in this publication meets the requirements of ANSI/NISO Z39.48-1992 (Permanence of Paper). Printed in The Netherlands
Further titles in this series: Volumes 2, 3, 5-7, 9, 10, 12, 13, 15, 16A, 22 and 26 are out of print 1. 4. 8.
G. SANGLERAT - THE PENETROMETER AND SOIL EXPLORATION R. SILVESTER - COASTAL ENGINEERING. 1 AND 2 L.N. PERSEN - ROCK DYNAMICS AND GEOPHYSICAL EXPLORATION Introduction to Stress Waves in Rocks 11. H.K. GUPTA AND B.K. RASTOGI - DAMS AND EARTHQUAKES 14. B. VOIGHT (Editor) - ROCKSLIDES AND AVALANCHES. 1 and 2 17. A.P.S. SELVADURAI - ELASTIC ANALYSIS OF SOIL-FOUNDATION INTERACTION 18. J. FEDA - STRESS IN SUBSOIL AND METHODS OF FINAL SETTLEMENT CALCULATION 19. ,&,. KI~ZDI- STABILIZED EARTH ROADS 20. E.W. BRAND AND R.P. BRENNER (Editors) - SOFT-CLAY ENGINEERING 21. A. MYSLIVE AND Z. KYSELA - THE BEARING CAPACITY OF BUILDING FOUNDATIONS 23. P. BRUUN - STABILITY OF TIDAL INLETS - Theory and Engineering 24. Z. B A Z A N T - METHODS OF FOUNDATION ENGINEERING 25. ,~. K[=ZDI - SOIL PHYSICS - Selected Topics 27. D. STEPHENSON - ROCKFILL IN HYDRAULIC ENGINEERING 28. P.E. FRIVIK, N. JANBU, R. SAETERSDAL AND L.I. FINBORUD (Editors) - GROUND FREEZING 1980 29. P. PETER - CANAL AND RIVER LEVI~ES 30. J. F E D A - MECHANICS OF PARTICULATE MATERIALS - The Principles 31. Q. Z,~RUBA AND V. MENCL - LANDSLIDES AND THEIR CONTROL Second completely revised edition 32. I.W. FARMER (Editor) - STRATA MECHANICS 33. L. HOBST AND J. ZAJiC - ANCHORING IN ROCK AND SOIL Second completely revised edition 34. G. SANGLERAT, G. OLIVARI AND B. CAMBOU - PRACTICAL PROBLEMS IN SOIL MECHANICS AND FOUNDATION ENGINEERING, 1 and 2 35. L. RI~THATI - GROUNDWATER IN CIVIL ENGINEERING 36. S.S. VYALOV - RHEOLOGICAL FUNDAMENTALS OF SOIL MECHANICS 37. P. BRUUN (Editor) - DESIGN AND CONSTRUCTION OF MOUNDS FOR BREAKWATER AND COASTAL PROTECTION 38. W.F. CHEN AND G.Y. BALADI - SOIL PLASTICITY - Theory and Implementation 39. E.T. HANRAHAN - THE GEOTECTONICS OF REAL MATERIALS: THE ~g~k METHOD 40. J. ALDORF AND K . E X N E R - MINE OPENINGS - Stability and Support 41. J.E. G I L L O T - CLAY IN ENGINEERING GEOLOGY 42. A.S. CAKMAK (Editor) - SOIL DYNAMICS AND LIQUEFACTION 43. A.S. CAKMAK (Editor) - SOIL-STRUCTURE INTERACTION 44. A.S. CAKMAK (Editor) - GROUND MOTION AND ENGINEERING SEISMOLOGY 45. A.S. CAKMAK (Editor) - STRUCTURES, UNDERGROUND STRUCTURES, DAMS, AND STOCHASTIC METHODS 46. L. RE~THATI- PROBABILISTIC SOLUTIONS IN GEOTECTONICS 47. B.M. DAS - T H E O R E T I C A L FOUNDATION ENGINEERING 48. W. DERSKI, R. IZBICKI, I. KISIEL AND Z. MROZ - ROCK AND SOIL MECHANICS 49. T. ARIMAN, M. HAMADA, A.C. SINGHAL, M.A. HAROUN AND A.S. CAKMAK (Editors) - RECENT ADVANCES IN LIFELINE EARTHQUAKE ENGINEERING 50. B.M. DAS - EARTH ANCHORS 51. K. T H I E L - ROCK MECHANICS IN HYDROENGINEERING 52. W.F. CHEN AND X.L. LIU - LIMIT ANALYSIS IN SOIL MECHANICS 53. W.F. CHEN AND E. MIZUNO - NONLINEAR ANALYSIS IN SOIL MECHANICS 54. F.H. CHEN - FOUNDATIONS ON EXPANSIVE SOILS 55. J. VERFEL - ROCK GROUTING AND DIAPHRAGM WALL CONSTRUCTION 56. B.N. WHITTAKER AND D.J. REDDISH - SUBSIDENCE - Occu.rrence, Prediction and Control 57. E. NONVEILLER - GROUTING, THEORY AND PRACTICE 58. V. KOL,~,I~AND I. NEMEC - MODELLING OF SOIL-STRUCTURE INTERACTION 59A. R.S. SINHA (Editor) - UNDERGROUND STRUCTURES - Design and Instrumentation 59B. R.S. SINHA (Editor) - UNDERGROUND STRUCTURES - Design and Construction 60. R.L. HARLAN, K.E. KOLM AND E.D. GUTENTAG - W A T E R - W E L L DESIGN AND CONSTRUCTION 61. I. K A S D A - FINITE ELEMENT TECHNIQUES IN GROUNDWATER FLOW STUDIES 62. L. FIALOVSZKY (Editor) - SURVEYING INSTRUMENTS AND THEIR OPERATION PRINCIPLES
63. 64. 65. 66. 67. 68. 69. 70. 71. 72. 73. 74. 75. 76. 77. 78. 79. 80. 81.
H. GIL - THE THEORY OF STRATA MECHANICS H.K. G U P T A - RESERVOIR-INDUCED EARTHQUAKES V.J. LUNARDINI - HEAT TRANSFER WITH FREEZING AND THAWING T.S. NAGARAI - PRINCIPLES OF TESTING SOILS, ROCKS AND CONCRETE E. JUHASOVA - SEISMIC EFFECTS ON STRUCTURES J. F E D A - CREEP OF SOILS - and Related Phenomena E. DUL,&,CSKA - SOIL SETTLEMENT EFFECTS ON BUILDINGS D. MILOVI(~ - STRESSES AND DISPLACEMENTS FOR SHALLOW FOUNDATIONS B.N. W H I T T A K E R , R.N. SINGH AND G. S U N - ROCK FRACTURE MECHANICS - Principles, Design and Applications M.A. MAHTAB AND P. GRASSO - GEOMECHANICS PRINCIPLES IN THE DESIGN OF TUNNELS AND CAVERNS IN ROCK R.N. YONG, A.M.O. MOHAMED AND B.P. WARKENTIN - PRINCIPLES OF CONTAMINANT TRANSPORT IN SOILS H. BURGER (Editor)- OPTIONS FOR TUNNELING 1993 S. H A N S B O - FOUNDATION ENGINEERING R. PUSCH - WASTE DISPOSAL IN ROCK R. PUSCH - ROCK MECHANICS ON A GEOLOGICAL BASE T. SAWARAGI - COASTAL ENGINEERING - WAVES, BEACHES, WAVE-STRUCTURE INTERACTIONS O. STEPHANSSON, L. JING AND CHIN-FU TSANG (Editors) - COUPLED THERMO- HYDROMECHANICAL PROCESSES OF FRACTURED MEDIA J. HARTLI~N AND W. WOLSKI (EDITORS) - EMBANKMENTS ON ORGANIC SOILS Y. KANAORI (EDITOR) - EARTHQUAKE PROOF DESIGN AND ACTIVE FAULTS
PREFACE
The new social and economic era calls for the integration of ecology and economy in a system of cause and effect. The central element in this shift is sustainable development. Fundamental to the achievement of sustainable development is the requirement for environmentally responsible waste management and restoration of the environment. Solutions to the complex problems confronted by waste management and environmental restoration industry are currently handled by the geoenvironmental engineering profession, consisting of geotechnical, environmental and chemical engineers, chemists, geologists, microbiologists, and soil scientists. These professionals employ a synergy of information developed in a variety of disciplines to solve these challenging problems. This book provides a comprehensive introduction to a complex interdisciplinary field. It is written to serve as a textbook for senior undergraduate and graduate students. Researchers and practicing engineers should also find it very useful. Geoenvironmental engineers need a good background in soil biology, chemistry, mechanics, mineralogy, and physics. In recognition of this need, this book summarizes relevant aspects of various soil physics, mineralogy, and chemistry as well the chemistry of pollutants. This treatment will provide sufficient background to students to enable them think about how to approach waste management and environmental restoration problems. With this in mind, the book contains a number of unique elements: (1) An appreciation of the geoenvironmental engineering practice within the context of global environmental problems, and the practice requirements to achieve environmental protection; (2) Basic elements of sustainable development and risk management strategies -- ecologically viable, economically feasible, and socially desirable. In doing so, the geoenvironmental engineering practice will be environmentally sound and politically acceptable; (3) Fundamentals of soil biology, chemistry, and mineralogy; (4) Basic elements of sources and characteristics of waste, and waste management concepts from an international viewpoint; (5) Fundamentals of soil-water-waste interactions, and their impact on fate and mobility of pollutants in subsurface environment; (6) Basic tools for subsurface assessment of polluted soils (e.g., geophysical, geotechnical, and geochemical techniques, risk assessment, geostatistics, and diffusion modelling); and (7) Fundamentals of remedial techniques (e.g., containment, electrical, physical and chemical extraction, and bioremediation). The book is organized into three parts. Part one provides the background material for complete understanding of the main elements that govern the problem of waste management and waste restoration. It contains six chapters (i.e., Chapters 1 to 6). Chapter 1 explores the global environmental problems and their impact on geoenvironmental engineering practice. Chapter 2 discusses the concept of sustainable development and how it can be achieved through proper geoenvironmental engineering practices. Chapter 3 discusses the sources and characteristics of waste. Also, the waste management problem is discussed from an international outlook. Chapter 4 details the fundamentals of soil mineralogy and chemistry. Chapter 5 introduces the basic interaction vii
viii mechanisms between soil, water, and pollutants. The impact of various interaction mechanisms on the changes of the transport parameters is emphasized. Chapter 6 discusses the fate and mobility of pollutants in the subsurface environment from a physico-chemical viewpoint. Part two examines the methods currently used for assessing subsurface pollution and evaluating its impact on human health and the environment. It contains four chapters (i.e., Chapters 7 to 10). Chapter 7 discusses the currently used geophysical, geotechnical, and geochemical techniques for subsurface assessment. Chapter 8 introduces the geostatistical concept and its application to subsurface pollution. Chapter 9 discusses the modelling aspects of pollutant transport in subsurface environment. Chapter 10 outlines the methodology for evaluating human health risk based on various exposure pathways. Part three examines various remedial options. It contains twelve chapters (i.e., Chapters 11 to 22). Chapter 11 introduces the concept of risk management and discusses the necessary basic elements of a sustainable risk management strategy. Subsurface control systems are discussed in Chapters 12 to 15. Physico-chemical extraction techniques are discussed in Chapters 16 to 18. Chapter 16 discusses vapour extraction. The use of solvents (Chapter 17) and surfactants (Chapter 18) for extracting pollutants from contaminated soils is explored. Fundamentals of electrochemical extraction processes are discussed in Chapter 19. Solidification/Stabilization treatment principles are introduced in Chapter 20. Bioremediation fundamentals are discussed in Chapter 21. Finally, Chapter 22 provides an application of the environmental impact assessment procedure to a hypothetical polluted site. Barely 60 years ago, few theories and concepts were available for explaining soil behaviour. Considerable progress has been made in developing viable tools and techniques for evaluating the specific interactions between soil constituents and pore fluid, and in describing the behaviour of soil in response to external constraints. Undoubtedly, much remains to be done, and significant milestones lie ahead. With the ultimate goal of serving society, we must continue to pay close attention to the bridge linking theory and practice. The authors are indebted to their graduate students, and their colleagues at McGill and at other universities and research units, for their valuable and timely input into the development of much of the materials contained in this book.
A.M.O. MOHAMED H.E. ANTIA 1998
CONTENTS
PREFACE .................................................................
vii
CHAPTER ONE: GEOENVIRONMENTAL E N G I N E E R I N G IN A G L O B A L ENVIRONMENT ....................................................... 1.1 INTRODUCTION ....................................................... 1.2 SCOPE OF ENVIRONMENTAL PROBLEMS ................................ 1.2.1 P o l l u t i o n o f O c e a n s and International R i v e r s . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.2.2 W a t e r Scarcity and D e g r a d a t i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.2.3 G l o b a l P o p u l a t i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.2.4 C o n t r o l o f C h e m i c a l s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.2.5 T r a n s - b o u n d a r y M o v e m e n t o f H a z a r d o u s W a s t e s . . . . . . . . . . . . . . . . . . . . . . . . . 1.2.6 A c i d R a i n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.2.7 D e f o r e s t a t i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.2.8 D e s e r t i f i c a t i o n and Soil E r o s i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.2.9 G l o b a l W a r m i n g . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.2.10 D e p l e t i o n o f the O z o n e Layer . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.2.11 D e c r e a s i n g Species o f Wildlife . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.2.12 E n v i r o n m e n t a l P o l l u t i o n in D e v e l o p i n g C o u n t r i e s . . . . . . . . . . . . . . . . . . . . . . . . 1.3 I N T E R C O N N E C T I O N OF G L O B A L E N V I R O N M E N T A L P R O B L E M S . . . . . . . . . . 1.4 GEOENVIRONMENTAL ENGINEERING ASPECTS ......................... 1.5 ACTIONS TOWARD RESTORING THE ENVIRONMENT .................... 1.5.1 C o m p r e h e n s i v e P o l i c y R e s p o n s e . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.5.2 E n v i r o n m e n t a l P o l i c y . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.5.3 E n v i r o n m e n t a l Ethics . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.6 SUMMARY AND CONCLUDING REMARKS ..............................
1 1 2 2 4 4 5 5 6 6 7 7 8 9 9 10 10 12 12 13 14 15
CHAPTER TWO: SUSTAINABLE DEVELOPMENT ............................ 2.1 INTRODUCTION ...................................................... 2.2 APPROACHES TO SUSTAINABLE DEVELOPMENT ........................ 2.2.1 E c o n o m i c s and Sustainability . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.2.2 E n v i r o n m e n t a l Sustainability . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.2.3 Social Sustainability . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.2.4 L a n d Sustainability . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.3 SUSTAINABLE DEVELOPMENT AND THE AMBIENT ENVIRONMENT ...... 2.3.1 A s s i m i l a t i v e C a p a c i t y in E n v i r o n m e n t a l M a n a g e m e n t . . . . . . . . . . . . . . . . . . . . 2.3.2 W a t e r Q u a l i t y M a n a g e m e n t and Sustainable D e v e l o p m e n t . . . . . . . . . . . . . . . . 2.4 ENVIRONMENTAL IMPACT ASSESSMENT .............................. 2.4.1 G u i d i n g Principles ............................................... 2.4.2 A t t r i b u t e s o f a Successful E I A . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.4.3 E v a l u a t i v e T o o l s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.5 ENGINEERING FOR SUSTAINABLE DEVELOPMENT ......................
17 17 18 18 20 20 21 22 22 23 24 24 25 26 26
ix
2.6
SUMMARY AND CONCLUDING REMARKS
..............................
CHAPTER THREE: SOURCES, CHARACTERISTICS, AND M A N A G E M E N T OF WASTES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1 3.2
3.3
3.4
3.5
INTRODUCTION ...................................................... SOURCES OF WASTES ................................................ 3.2.1 M u n i c i p a l Solid W a s t e . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2.2 P e s t i c i d e W a s t e s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2.3 M i n i n g W a s t e s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2.4 E l e c t r o p l a t i n g a n d M e t a l F i n i s h i n g I n d u s t r y . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2.5 M e t a l S m e l t i n g a n d R e f i n i n g Industries . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
33 33 33 34 36 36 38 38
3.2.6 3.2.7 3.2.8
40 42 44
Pulp and Paper Wastes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Petroleum Refining Wastes ......................................... P a i n t and A l l i e d Industries . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
HAZARDOUS WASTE ................................................. 3.3.1 U n i t e d States o f A m e r i c a . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.3.2 Canada ......................................................... 3.3.3 E u r o p e a n C o m m u n i t y . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.3.4 G e r m a n y . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.3.5 The Netherlands . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . HAZARDOUS WASTE MANAGEMENT .................................. 3.4.1 US Regulation Disposal Philosophy .................................. 3.4.2 Canada ......................................................... 3.4.3 European Community ............................................. SUMMARY AND CONCLUDING REMARKS ..............................
CHAPTER FOUR: SOIL SYSTEM 4.1 4.2
4.3
4.4 4.5 4.6 4.7
30
INTRODUCTION
............................................ ......................................................
44 44 47 47 48 48 48 48 55 55 58 59 59
SOIL PHASES .........................................................
59
4.2.1 Gas Phase . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.2.2 F l u i d P h a s e . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.2.3 Solid P h a s e . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . MINERAL COMPOSITION .............................................. 4.3.1 Primary and Secondary Minerals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.3.2 T r a c e E l e m e n t s in Soil M i n e r a l s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . SOIL MINERAL TRANSFORMATIONS ................................... C R Y S T A L C H E M I S T R Y OF SILICATES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . STRUCTURAL COMPONENTS OF SOIL CLAYS ........................... 4.6.1 Silica, G i b b s i t e , and Brucite Sheets . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . PROPERTIES OF LAYER SILICATES ..................................... 4.7.1 Kaolins ......................................................... 4.7.2 H y d r o u s M i c a (Illite) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.7.3 Montmorillonite .................................................. 4.7.4 V e r m i c u l i t e s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
59 61 63 64 64 66 67 67 70 71 75 75 76 76 77
4.7.5
77
Chlorites . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
xi
4.8 4.9
4.10 4.11
4.7.6 Sepiolite and P a l y g o r s k i t e . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.7.7 M i x e d - l a y e r C l a y s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.7.8 Soil Clays . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . SOIL ORGANIC MATTER .............................................. C H A R G E D E V E L O P M E N T IN S O I L S . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.9.1 C o n s t a n t Surface Charge M i n e r a l s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.9.2 C o n s t a n t Surface Potential M i n e r a l s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . SURFACE FUNCTIONAL GROUPS ...................................... SUMMARY AND CONCLUDING REMARKS ..............................
C H A P T E R FIVE: S O I L - W A T E R - P O L L U T A N T I N T E R A C T I O N . . . . . . . . . . . . . . . . . . 5.1 5.2 5.3 5.4
5.5 5.6
5.7 5.8 5.9
5.10
INTRODUCTION ...................................................... ADSORPTION MECHANISMS ........................................... ADSORPTION MEASUREMENTS ........................................ METAL CATION ADSORPTION ......................................... 5.4.1 M e t a l C a t i o n A d s o r p t i o n by Soil . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.4.2 M e t a l C a t i o n A d s o r p t i o n by Soil Constituents . . . . . . . . . . . . . . . . . . . . . . . . . . ADSORPTION EQUILIBRIUM ........................................... MOLECULAR ADSORPTION MODELS ................................... 5.6.1 Electric D o u b l e L a y e r Structure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.6.2 G o u y - C h a p m a n M o d e l . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.6.3 StemModel .................................................... 5.6.4 Other M o d e l s .................................................. ORGANIC POLLUTANT-SOIL ORGANIC MATTER INTERACTION .......... SOIL ORGANIC MATTER-SOIL MINERALS INTERACTION ................ INFLUENCE OF POLLUTANTS ON SOIL HYDRAULIC CONDUCTIVITY ..... 5.9.1 Influence o f Inorganic C h e m i c a l s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.9.2 Influence o f Organic C h e m i c a l s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . SUMMARY AND CONCLUDING REMARKS .............................
C H A P T E R SIX: FATE A N D E F F E C T S OF P O L L U T A N T S 6.1 6.2 6.3
6.4
6.5
6.6
...................... INTRODUCTION ..................................................... POLLUTANT PATHWAYS ............................................. ENVIRONMENTAL FATE ............................................. 6.3.1 Surface W a t e r C o m p a r t m e n t . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.3.2 Soil C o m p a r t m e n t . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.3.3 Sediment Compartment ........................................... BIOAVAILABILITY ................................................... 6.4.1 Availability of Heavy Metals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.4.2 A v a i l a b i l i t y o f Inorganic P h o s p h a t e s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.4.3 A v a i l a b i l i t y o f Organic C h e m i c a l s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . EFFECTS OF POLLUTANTS ........................................... 6.5.1 Uptake of Pollutants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.5.2 T y p e s o f P o l l u t a n t Effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . PARAMETER IDENTIFICATION ........................................
78 78 78 79 80 80 80 82 83 85 85 85 89 89 89 91 93 97 97 98 107 109 110 113 114
114 117 121 123 123 123 124 124 125 125 126 127 128 129 130 130 130 131
xii
6.7 6.8
6.6.1 Physico-chemical Parameters ....................................... 6.6.2 Fate P a r a m e t e r s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.6.3 Effect Parameters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.6.4 B i o l o g i c a l P a r a m e t e r s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . TIERED TEST PROGRAM ............................................. SUMMARY AND CONCLUDING REMARKS .............................
C H A P T E R S E V E N : SITE I N V E S T I G A T I O N
132 134 137 137
140 140
.................................. INTRODUCTION ..................................................... SITE INVESTIGATION APPROACH ..................................... PHASE I INVESTIGATIONS ............................................ 7.3.1 Collecting Information ............................................ 7.3.2 Field Reconnaissance . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.3.3 Development of a Conceptual Model ................................ 7.3.4 E s t a b l i s h i n g The W o r k P l a n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . P H A S E II I N V E S T I G A T I O N S . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
143 143 143 144 144 146 147 147 148
GEOPHYSICAL TECHNIQUES ......................................... 7.5.1 Ground Penetration Radar ......................................... 7.5.2 Electromagnetic ................................................ 7.5.3 Surface Resistivity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.5.4 Seismic Surveys ................................................ 7.5.5 Borehole Logging ................................................ 7.5.6 V i d e o C a m e r a s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . HYDROGEOLOGICAL INVESTIGATIONS ...............................
148 149 149
164 165 166 167
7.8
7.6.1 Drilling methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.6.2 Sampling Methods ............................................... 7.6.3 W e l l Installation T e c h n i q u e s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.6.4 M o n i t o r i n g W e l l D e s i g n C o m p o n e n t s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.6.5 Well Decontamination Procedures .................................. HYDROGEOCHEMICAL INVESTIGATION ............................... 7.7.1 Subsurface Environment .......................................... 7.7.2 Sampling Considerations .......................................... GEOCHEMICAL DATA COLLECTION ...................................
7.9 7.10
7.8.1 S o u r c e s o f Errors . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.8.2 S a m p l i n g M e t h o d s and T y p e s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . GEOCHEMICAL DATA ANALYSIS ..................................... SUMMARY AND CONCLUDING REMARKS .............................
7.1 7.2 7.3
7.4 7.5
7.6
7.7
155 160 162 164 164
171 171
171 174 179 179
181 182 182
CHAPTER EIGHT: GEOSTATISTICS ........................................
185
8.1 8.2
185 186 186 186 187 188
INTRODUCTION ..................................................... DATA ANALYSIS CONCEPTS ......................................... 8.2.1 Histogram ...................................................... 8.2.2 O g i v e . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.2.3 S u m m a r y Statistics . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.2.4 N o r m a l D i s t r i b u t i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
xiii SEMIVARIOGRAM ................................................... INTRINSIC MODELLING .............................................. STRUCTURAL ANALYSIS ............................................. 8.5.1 Stationary Models . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.5.2 N o n s t a t i o n a r y M o d e l s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.5.3 M o d e l S u p e r p o s i t i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . KRIGING ............................................................ SOLUTION METHODOLOGY .......................................... APPLICATION ....................................................... SUMMARY AND CONCLUDING REMARKS .............................
188 191 192 192 195 196 198 201 203 207
CHAPTER NINE: SUBSURFACE POLLUTANT TRANSPORT ................... 9.1 INTRODUCTION ..................................................... 9.2 MODELLING PROCESS ............................................... 9.3 T R A N S P O R T P R O C E S S E S IN S O I L S . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.3.1 Advection ...................................................... 9.3.2 D i f f u s i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.3.3 Dispersion ..................................................... 9.4 TRANSPORT EQUATION .............................................. 9.5 SOLUTE TRANSPORT MODELS ........................................ 9.5.1 Conservative Tracer .............................................. 9.5.2 R e a c t i v e C h e m i c a l Species . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.5.3 Spill o f P o l l u t a n t s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.5.4 P o l l u t a n t P l u m e . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.6 METHODS FOR CALCULATING TRANSPORT PARAMETERS ............. 9.6.1 L a b o r a t o r y M e t h o d s for H y d r a u l i c C o n d u c t i v i t y T e s t i n g . . . . . . . . . . . . . . . . . 9.6.2 L a b o r a t o r y M e t h o d s for A d s o r p t i o n C h a r a c t e r i s t i c s . . . . . . . . . . . . . . . . . . . . . 9.6.3 Estimation of Transport Parameters .................................. 9.7 SUMMARY AND CONCLUDING REMARKS .............................
209 209 209
8.3 8.4 8.5
8.6 8.7 8.8 8.9
CHAPTER TEN: RISK ASSESSMENT ........................................ 10.1 INTRODUCTION ..................................................... 10.2 BASIC ELEMENTS OF HUMAN HEALTH RISK ASSESSMENT ............. 10.3 HAZARD IDENTIFICATION ........................................... 10.4 EXPOSURE ASSESSMENT ............................................ 10.5 TOXICITY ASSESSMENT ............................................. 10.5.1 I n t r o d u c t i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.5.2 S o u r c e s o f T o x i c i t y I n f o r m a t i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.5.3 T o x i c o l o g i c a l P a r a m e t e r s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.5.4 E x p o s u r e R o u t e C o n s i d e r a t i o n s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.6 RISK CHARACTERIZATION ........................................... 10.6.1 C a l c u l a t i o n o f C a r c i n o g e n i c R i s k s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.6.2 C a l c u l a t i o n o f N o n c a r c i n o g e n i c H a z a r d s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.7 SUMMARY AND CONCLUDING REMARKS .............................
210 210 211 213 214 218 218 222 223 225 227 227 230 239 244
245
245 246 247 248 254 254 255 256 259 259 260 261 262
xiv
C H A P T E R ELEVEN: RISK M A N A G E M E N T . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.1 11.2
11.3 11.4 11.5
11.6
11.7 11.8
INTRODUCTION ..................................................... ELEMENTS OF A RISK MANAGEMENT PROGRAM ...................... 11.2.1 H a z a r d s Identification P r o g r a m . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.2.2 C o n s e q u e n c e A n a l y s i s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.2.3 R i s k M i t i g a t i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . QUANTIFIED RISK ASSESSMENT ...................................... R O L E OF R E G U L A T O R Y A G E N C I E S . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . REGULATORY APPROACHES ......................................... 11.5.1 R i s k - B a s e d M i t i g a t i o n Criteria . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.5.2 N u m e r i c a l l y - B a s e d M i t i g a t i o n Criteria . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . MITIGATION TECHNOLOGIES FOR POLLUTED SOILS ................... 11.6.1 Natural A t t e n u a t i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.6.2 C o n t a i n m e n t . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.6.3 R e m o v a l and T r e a t m e n t . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.6.4 In-Situ T r e a t m e n t . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . SELECTION OF MITIGATION OPTIONS ................................. SUMMARY AND CONCLUDING REMARKS .............................
C H A P T E R TWELVE: SUBSURFACE CONTROL SYSTEMS --- G R O U N D W A T E R E X T R A C T I O N - - - . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.1 12.2
12.3
12.4
12.5
12.6
12.7
INTRODUCTION ..................................................... WELL HYDRAULICS ................................................. 12.2.1 Steady State E q u a t i o n s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.2.2 U n s t e a d y State E q u a t i o n s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.2.3 C a p t u r e Z o n e A n a l y s i s and O p t i m i z a t i o n M o d e l l i n g . . . . . . . . . . . . . . . . . . . . GROUNDWATER MANAGEMENT BY HYDRODYNAMIC CONTROLS ...... 12.3.1 E x t r a c t i o n s y s t e m s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.3.2 Injection S y s t e m s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . HYDRAULIC BARRIER SYSTEMS ...................................... 12.4.1 Barriers U s i n g Extraction S y s t e m s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.4.2 Barriers using Injection S y s t e m . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.4.3 Barriers U s i n g B o t h Extraction and Injection S y s t e m s . . . . . . . . . . . . . . . . . . . DESIGN AND CONSTRUCTION CONSIDERATIONS ........................ 12.5.1 Site C h a r a c t e r i z a t i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.5.2 A s s e s s m e n t o f P r o p e r t y Constraints . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.5.3 R e m e d i a l S y s t e m D e s i g n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.5.4 M o n i t o r i n g o f S y s t e m P e r f o r m a n c e . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.5.5 E s t a b l i s h m e n t o f C l e a n u p Goals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . CRITICAL EVALUATION OF GROUNDWATER MANAGEMENT ........... 12.6.1 Tailing and R e b o u n d I m p a c t s on R e m e d i a t i o n . . . . . . . . . . . . . . . . . . . . . . . . . 12.6.2 C o n t r i b u t i n g Factors to Tailing and R e b o u n d . . . . . . . . . . . . . . . . . . . . . . . . . . SUMMARY AND CONCLUDING REMARKS .............................
263 263 263 265 265 266 267 268 268 269 270 276 276 277 277 278 278 279
281 281 282 283 285 288 289 289 290 290 290 296 296 298 298 299 299 301 302 302 302 303 304
XV
C H A P T E R THIRTEEN: SUBSURFACE C O N T R O L SYSTEMS --- A L T E R N A T I V E S TO G R O U N D W A T E R E X T R A C T I O N --- . . . . . . . . . . . . .
307
13.1 13.2
INTRODUCTION ..................................................... SLURRY WALLS ..................................................... 13.2.1 Structure F o r m a t i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.2.2 Factors A f f e c t i n g Structure F o r m a t i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.2.3 I n f l u e n c e o f Pollutants on Slurry W a l l H y d r a u l i c C o n d u c t i v i t y . . . . . . . . . . . . 13.2.4 Slurry T r e n c h i n g Quality Control P a r a m e t e r s . . . . . . . . . . . . . . . . . . . . . . . . . .
307 308 308 309 315 319
13.3
GROUTING .......................................................... 13.3.1 T y p e s o f Grouts . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.3.2 C o m p a t i b i l i t y o f C h e m i c a l Grouts w i t h P o l l u t a n t s . . . . . . . . . . . . . . . . . . . . . . .
321 321 323
13.3.3 G r o u t i n g M e t h o d s
324
13.4
SHEET PILING .......................................................
327
13.5
GROUND FREEZING
327
13.6
ELECTROKINETICS
13.7
13.8
...............................................
................................................. ..................................................
328
13.6.1 E l e c t r o o s m o s i s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
328
13.6.2 E l e c t r o p h o r e s i s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
328
13.6.3 Electrolysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . REACTIVE BARRIERS ................................................
328 329
13.7.1 In-Situ R e a c t i v e Z o n e s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
330
13.7.2 T y p e s o f R e a c t i v e M e d i a . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.7.3 E n g i n e e r i n g A s p e c t s o f the R e a c t i v e Barrier S y s t e m . . . . . . . . . . . . . . . . . . . .
331 332
SUMMARY AND CONCLUDING REMARKS
333
.............................
C H A P T E R FOURTEEN: COVERING SYSTEMS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
335
14.1
INTRODUCTION
335
14.2
F U N C T I O N S OF C O V E R I N G S Y S T E M S . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
14.3
T Y P E S OF C O V E R I N G S Y S T E M S . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
336
14.4
SOIL-BASED COVERING SYSTEMS
337
14.5
TOP LAYER
.....................................................
....................................
.........................................................
335
338
14.5.1 Physical P a r a m e t e r s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.5.2 E n v i r o n m e n t a l P a r a m e t e r s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.5.3 C h e m i c a l P a r a m e t e r s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
341 344 345
14.6
DRAINAGE LAYER ...................................................
347
14.7
INFILTRATION BARRIER LAYER
349
......................................
14.7.1 Physical P a r a m e t e r s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
351
14.7.2 C h e m i c a l P a r a m e t e r s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
359
14.8
BIOTIC BARRIERS
360
14.9
MINING WASTE COVERING SYSTEMS
...................................................
360 361
14.9.2 W a t e r Barrier C o n c e p t . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
364
14.9.3 O r g a n i c Barrier C o n c e p t 14.10
.................................
14.9.1 D r y Barrier C o n c e p t . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ..........................................
TYPES OF COVERING MATERIALS
....................................
365 367
14.10.1
N a t u r a l Materials . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
368
14.10.2
M o d i f i e d Soils . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
368
xvi
14.11
14.10.3 Synthetic materials . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.10.4 Waste Materials ........................................... SUMMARY AND CONCLUDING REMARKS .............................
368 368 369
CHAPTER FIFTEEN: LINING SYSTEMS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
371
15.1 15.2 15.3 15.4
INTRODUCTION ..................................................... CLASSIFICATION OF LINING SYSTEMS ................................ CLAY LINERS ....................................................... DESIRABLE PROPERTIES OF CLAY LINERS ............................ 15.4.1 L o w H y d r a u l i c C o n d u c t i v i t y . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.4.2 H i g h A t t e n u a t i o n o f C h e m i c a l s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
371 372 372 373 373 374
15.4.3 15.4.4 15.4.5 15.4.6 15.4.7
375 375 377 377 377
15.5 15.6
15.7
15.8
15.9
Low Diffusivity ................................................. Ductility . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . S l ope stability . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . A d e q u a t e Interface S t r e n g t h . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . L o n g T e r m Stability . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
15.4.8 C o n s t r u c t a b i l i t y . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . REQUIRED PROPERTIES OF CLAY LINERS ............................. F A C T O R S C O N T R O L L I N G P R O P E R T I E S OF C L A Y L I N E R S . . . . . . . . . . . . . . . . 15.6.1 C o m p o s i t i o n a l F a c t o r s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
377 377 378 378
15.6.2 E n v i r o n m e n t a l F a c t o r s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . POST CONSTRUCTION CHANGES ..................................... 15.7.1 P h y s i c a l P r o c e s s e s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.7.2 C h e m i c a l P r o c e s s e s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.7.3 B i o l o g i c a l P r o c e s s e s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . OTHER TYPES OF LINING SYSTEMS ................................... 15.8.1 Soil C e m e n t . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.8.2 Soil C a r b o n a t e s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.8.3 Soil A m o r p h o u s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.8.4 Soil S e a l a n t s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.8.5 F l e x i b l e M e m b r a n e Liners . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . SUMMARY AND CONCLUDING REMARKS .............................
381 382 382 386 394 395 395 396 397 398 399 402
CHAPTER SIXTEEN: SOIL V A P O U R EXTRACTION . . . . . . . . . . . . . . . . . . . . . . . . . .
405
16.1 16.2 16.3
INTRODUCTION ..................................................... TECHNOLOGY DESCRIPTION ......................................... SITE C O N D I T I O N S . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.3.1 V o l a t i l e O r g a n i c P o l l u t a n t D i s t r i b u t i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
405 405 407 407
16.3.2 G r o u n d w a t e r T a b l e . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.3.3 Infiltration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.3.4 L o c a t i o n o f H e t e r o g e n e i t y . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.3.5 T e m p e r a t u r e . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.3.6 A t m o s p h e r i c P r e s s u r e . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . SOIL PROPERTIES ................................................... 16.4.1 Soil H y d r a u l i c C o n d u c t i v i t y . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
407 408 408 408 409 410 410
16.4
xvii 16.4.2 Soil Organic Matter Content . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . CHEMICAL PROPERTIES ............................................. 16.5.1 Henry's Constant and Solubility . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.5.2 Adsorption Equilibria and the Octanol-Water Partition Coefficient . . . . . . . . . 16.5.3 Diffusivity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.5.4 Density and Viscosity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.6 CONTROL VARIABLES ............................................... 16.6.1 Air Withdrawal Rate . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.6.2 Well Configuration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.6.3 Well Spacing and Surface Covering . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.6.4 P u m p i n g Duration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.6.5 Inlet Air Concentration and Moisture Content . . . . . . . . . . . . . . . . . . . . . . . . . 16.7 RESPONSE VARIABLES .............................................. 16.7.1 Soil-related Variables . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.7.2 Extracted Air-related Variables . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.7.3 System-related Variables . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.8 M O D E L L I N G OF V A P O U R R E M O V A L R A T E . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.9 CASE S T U D Y . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.9.1 Temperature . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.9.2 Air F l o w Rate . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.9.3 Organic Carbon Fraction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.9.4 Soil Moisture Content . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.10 ISSUES R E L A T E D TO SVE A P P L I C A T I O N . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.10.1 Feasibility Study . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.10.2 Physical Testing . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.10.3 Final System Design . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.10.4 Monitoring . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.11 I N T E G R A T E D S Y S T E M S . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.11.1 Air Sparging . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.11.2 Bioventing . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.11.3 Thermal Enhancements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.12 S U M M A R Y A N D C O N C L U D I N G R E M A R K S . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
411 412 412 413 414 414 415 415 417 417 418 418 419 419 420 420 421 423 423 424 425 425 426 426 427 427 427 427 427 428 429 430
CHAPTER SEVENTEEN: SOLVENT EXTRACTION PROCESSES ...............
433 433 433 435
16.5
17.1 17.2 17.3
17.4
17.5 17.6
INTRODUCTION ..................................................... PROCESS DESCRIPTION . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . EXTRACTION TECHNOLOGIES ........................................ 17.3.1 Metal Mining . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.3.2 Metals Mining Methods Versus Hazardous Waste Treatment Techniques . . . . SELECTIVE SEQUENTIAL EXTRACTION ............................... 17.4.1 Description o f the Technique . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.4.2 Retention Phases . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.4.3 Mobility and Bioavailability . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . EXTRACTION VIA HYDROCHLORIC ACID . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . EXTRACTION VIA CHELATING AGENTS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
435 436 437 437 441 445 447 449
xviii
17.7 17.8
17.6.1 C o n c e p t a n d D e f i n i t i o n s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.6.2 C l a s s i f i c a t i o n a n d P r o p e r t i e s o f C h e l a t i n g A g e n t s . . . . . . . . . . . . . . . . . . . . . . 17.6.3 E D T A Stability C o m p l e x e s in S o l u t i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.6.4 R e m e d i a t i o n V i a C h e l a t i n g A g e n t s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . EXTRACTION VIA ORGANIC ACIDS ................................... SUMMARY AND CONCLUDING REMARKS .............................
C H A P T E R EIGHTEEN: SURFACTANT E X T R A C T I O N PROCESSES . . . . . . . . . . . . 18.1 18.2
449 451 454 460 463 464
INTRODUCTION ..................................................... SURFACTANT PROPERTIES .......................................... 18.2.1 D e f i n i t i o n and T y p e s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
465 465 466 466
18.2.2 18.2.3 18.2.4 18.2.5
468 470 470 471
Surface Tension . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Critical M i c e l l e C o n c e n t r a t i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Solubilization and Detergency . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Dispersion .....................................................
18.6
18.2.6 M i c e l l e F o r m a t i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ORGANIC POLLUTANT PROPERTIES ................................... 18.3.1 S o l u b i l i t y . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.3.2 H y d r o l y s i s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.3.3 V a p o u r P r e s s u r e . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.3.4 C h e m i c a l A l t e r n a t i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . SOIL SYSTEM PROPERTIES ........................................... 18.4.1 A c t i v i t y . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.4.2 P o r o s i t y . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.4.3 M o i s t u r e C o n t e n t . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.4.4 Soil O r g a n i c M a t t e r C o n t e n t . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.4.5 B u o y a n c y F o r c e s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.4.6 H y d r a u l i c C o n d u c t i v i t y . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . INTERACTION MECHANISMS OF VARIOUS COMPONENTS .............. 18.5.1 S o i l - O r g a n i c P o l l u t a n t Interaction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.5.2 P o l l u t a n t - S u r f a c t a n t Interaction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.5.3 S o i l - S u r f a c t a n t Interaction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . M I C E L L E F O R M A T I O N IN S U R F A C T A N T S O L U T I O N S . . . . . . . . . . . . . . . . . . .
471 474 474 474 476 477 477 477 480 481 482 483 485 486 486 486 488 488
18.7 18.8 18.9
MICELLE SOLUBILIZATION ........................................... S U R F A C T A N T F L O W IN P O L L U T E D S O I L . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . SUMMARY AND CONCLUDING REMARKS .............................
492 494 498
18.3
18.4
18.5
C H A P T E R NINETEEN: E L E C T R O C H E M I C A L REMEDIATION . . . . . . . . . . . . . . . .
499
19.1 19.2
INTRODUCTION ..................................................... CONDUCTION OF ELECTRICITY ....................................... 19.2.1 O h m ' s L a w . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19.2.2 M o b i l i t i e s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ELECTROCHEMICAL REACTIONS .....................................
499 499 499 500 501
19.3.1 E l e c t r o l y s i s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19.3.2 E l e c t r o d e P o t e n t i a l . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
501 503
19.3
xix
19.4
19.5
19.6
19.7
19.8
19.3.3 T h e r m o d y n a m i c s o f E l e c t r o d e P o t e n t i a l . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19.3.4 E l e c t r o c h e m i c a l R e a c t i o n P r o c e s s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . T R A N S P O R T OF E L E C T R O A C T I V E SPECIES . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19.4.1 M i g r a t i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19.4.2 D i f f u s i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19.4.3 C o n v e c t i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ELECTROKINETIC PHENOMENA ...................................... 19.5.1 E l e c t r o o s m o s i s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19.5.2 S t r e a m i n g Potential . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19.5.3 S e d i m e n t a t i o n Potential . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19.5.4 E l e c t r o p h o r e s i s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ELECTROKINETIC REMEDIATION ..................................... 19.6.1 p H V a r i a t i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19.6.2 R e d o x P o t e n t i a l . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19.6.3 S u r f a c e C h a r g e o f C l a y M i n e r a l s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19.6.4 Soil B u f f e r i n g C a p a c i t y . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19.6.5 C o n t r o l o f p H and R e d o x p o t e n t i a l . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ELECTRODIALYSIS REMEDIATION ................................... 19.7.1 E l e c t r o d i a l y s i s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
505 508 509 509 510 510 510 512 515 516 517 518 519 520 520 521 523 523 523
19.7.2 T r e a t m e n t E f f i c i e n c y . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . SUMMARY AND CONCLUDING REMARKS .............................
525 527
CHAPTER TWENTY: SOLIDIFICATION/STABILIZATION PROCESSES . . . . . . . . 20.1 20.2
20.3 20.4
20.5
20.6 20.7
20.8
20.9
INTRODUCTION ..................................................... S/S P R O C E S S E S . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.2.1 I n o r g a n i c - b a s e d P r o c e s s e s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.2.2 O r g a n i c - b a s e d P r o c e s s e s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . CEMENT TYPES AND COMPOSITION .................................. CEMENT HYDRATION ............................................... 20.4.1 K i n e t i c s o f C e m e n t H y d r a t i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.4.2 H y d r a t i o n o f P h a s e s in P o r t l a n d C e m e n t . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
529 529 529 529 530 530 533 534 534
FACTORS INFLUENCING THE SET OF PORTLAND CEMENT 20.5.1 A c c e l e r a t i n g A d m i x t u r e s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
540 540
..............
20.5.2 R e t a r d i n g A d m i x t u r e s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . SOIL-PORTLAND CEMENT INTERACTION ..............................
540 541
SOIL-LIME INTERACTION ............................................ 20.7.1 H y d r a t i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.7.2 I o n E x c h a n g e a n d F l o c c u l a t i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.7.3 P o z z o l a n i c R e a c t i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.7.4 C a r b o n a t i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . C O N T R O L L I N G F A C T O R S IN S O I L S T A B I L I Z A T I O N . . . . . . . . . . . . . . . . . . . . . . 20.8.1 S o i l - P o r t l a n d C e m e n t S t a b i l i z a t i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.8.2 S o i l - L i m e S t a b i l i z a t i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . K I N E T I C S O F R E A C T A N T S A N D P R O D U C T S IN S T A B I L I Z E D S O I L . . . . . . . . . 20.9.1 C l a y M i n e r a l s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
542 542 543 543 543 543 543 544 545 545
XX
20.9.2 C e m e n t i n g A g e n t s ( C S H and C A H )
.................................
546
20.10
20.9.3 Ettringite . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . INORGANIC WASTE-PORTLAND CEMENT INTERACTION ................
546 547
20.11
20.10.1 Cations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.10.2 A n i o n s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ORGANIC WASTE-PORTLAND CEMENT INTERACTION ..................
547 549 551
20.12
E V A L U A T I O N OF P O T E N T I A L L E A C H A B I L I T Y . . . . . . . . . . . . . . . . . . . . . . . . . .
553
20.12.1 L e a c h i n g Tests . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.12.2 L e a c h i n g M e c h a n i s m . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.12.3 L e a c h a t e T r a n s p o r t M o d e l l i n g . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
553 554 554
20.12.4 Factors A f f e c t i n g Leachability . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
556
SUMMARY AND CONCLUDING REMARKS
557
20.13
C H A P T E R T W E N T Y ONE: BIOREMEDIATION 21.1 21.2
21.3
21.4
INTRODUCTION SOIL BIOMASS
..............................
559
.....................................................
559
......................................................
559
21.2.1 Definition . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
559
21.2.2 Bacteria . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21.2.3 Fungi . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . MICROBIAL METABOLISM AND GROWTH .............................
560 561 561
21.3.1 M i c r o b i a l M e t a b o l i s m . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21.3.2 M i c r o b i a l G r o w t h . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
561 562
MICROBIAL REACTIONS
564
21.4.1
21.5
.............................
.............................................
Biomass Growth .................................................
564
21.4.2 A e r o b i c R e s p i r a t i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
566
21.4.3 A n a e r o b i c M e t a b o l i s m . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
567
MICROBIAL TRANSFORMATION
OF I N O R G A N I C C O M P O U N D S
..........
567
21.5.2 P h o s p h o r u s T r a n s f o r m a t i o n s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
569
21.5.3 Metal T r a n s f o r m a t i o n s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21.6
MICROBIAL TRANSFORMATION
............
573 573 575
21.6.3 P o l y m e r i z a t i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
577
21.6.4 M i c r o b i a l A c c u m u l a t i o n
579
BIOTRANSFORMATION 21.7.1 H y d r o l y s i s
21.8
571
OF ORGANIC COMPOUNDS
21.6.1 B i o d e g r a d a t i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21.6.2 C o m e t a b o l i c T r a n s f o r m a t i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ..........................................
21.6.5 N o n e n z y m a t i c T r a n s f o r m a t i o n 21.7
567
21.5.1 N i t r o g e n T r a n s f o r m a t i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
.....................................
REACTIONS ...................................
.....................................................
581 581 581
21.7.2 O x i d a t i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
581
21.7.3 R e d u c t i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
582
FACTORS AFFECTING MICROBIAL TRANSFORMATION 21.8.1
.................
C h e m i c a l Factors . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
583 583
21.8.2 E n v i r o n m e n t a l Factors . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
589
21.9
K I N E T I C S OF M I C R O B I A L T R A N S F O R M A T I O N
590
21.10
SUMMARY AND CONCLUDING REMARKS
.........................
.............................
591
xxi
CHAPTER TWENTY TWO: CASE STUDY .................................... H I S T O R Y OF T H E H Y P O T H E T I C A L SITE . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.2 WATER QUALITY .................................................... 22.3 SEDIMENT QUALITY ................................................. 22.4 DEVELOPMENT OF CLEANUP ALTERNATIVES ......................... 22.5 C O M P A R A T I V E A N A L Y S I S OF C L E A N U P A L T E R N A T I V E S . . . . . . . . . . . . . . . 22.5.1 P e r m a n e n t E n v i r o n m e n t a l Criteria . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.5.2 T e m p o r a r y E n v i r o n m e n t a l Criteria . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.5.3 T e c h n i c a l Criteria . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.5.4 T e c h n i c a l and E c o n o m i c Criteria . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.5.5 E c o n o m i c Criteria . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.5.6 N o n - D i s c r i m i n a t o r y Criteria . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.5.7 N o n - P e r t i n e n t Criteria . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.5.8 I m p o r t a n c e o f Rating . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.5.9 R e m e d i a l A l t e r n a t i v e Selection . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . SUMMARY AND CONCLUDING REMARKS ............................. 22.6
609
GLOSSARY
611
22.1
REFERENCES
............................................................... AND SUGGESTED
READING
.................................
INDEX ....................................................................
593 593 593 594 596 599 600 602 602 603 603 603 604 605 605
653 693
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C H A P T E R ONE
GEOENVIRONMENTAL ENGINEERING IN A G L O B A L E N V I R O N M E N T
1.1
INTRODUCTION
In providing the foundations for the advancement of human life, it is inevitable that we have to work with the land. Land is a broader concept and is defined by the Intemational Food and Agriculture Organization (FAO, 1976) as:
"An area of the earth's surface, the characteristics of which embrace all reasonably stable, or predictably cyclic, attributes of the biosphere vertically above and below this area, including those of the atmosphere, the soil and the underlying geology, the hydrogeology, the plant and animal populations and the results of past and present human activity, to the extent that these attributes exert a significant influence on present and future uses of the land by man." This definition requires full interaction between the four ecosystems, i.e., atmosphere, hydrosphere, geosphere, and biosphere, that constitute the land. The atmosphere is the envelope of gasses surrounding the earth and it is subdivided into regions on the basis of altitude (Linsley, 1987; Parker and Corbitt, 1993). Hydrosphere refers to water in various forms: oceans, lakes, streams, snowpack, glaciers, polar ice caps, and water under the ground (groundwater) (Friedman, 1987; Parker and Corbitt, 1993). Geosphere refers to the complex and variable mixture of minerals, organic matter, pore fluid, and air that make up the soil. Finally, the term biosphere refers to living organisms and their environments on the surface of the earth (Manahan, 1991). The movement of pollutants within the land has a profound effect upon its bioavailability. The rates of transfer are important and can affect the land livability. The natural processes that promote these changes are ever present and are responsible for the magnitude and direction of the exchange, be they desirable or undesirable. In approaching the creation of a new environment which is related to the land, we must consider the problems of waste management and restoration of the environment from the global environment viewpoint. Solutions to the complex problems confronted by waste management and environmental restoration industry are currently handled by the geoenvironmental engineering profession, consisting of geotechnical, environmental and chemical engineers, geologists, chemists, microbiologists and soil scientists. These professionals employ a synergy of information developed in a variety of disciplines to solve the challenging problems confronted in waste management and environmental restoration. In this context, geoenvironmental engineering can be defined as the application of earth science principles to the solution of land environmental problems. In seeking solutions to restore the degradation of the environment, we need to consider the
2
GEOENVIRONMENTAL ENGINEERING IN A GLOBAL ENVIRONMENT
interconnecting nature of the various ecosystems. This means that a cooperative and holistic global effort should be considered in developing a viable solution to global environmental problems. Therefore, it is necessary for the geoenvironmental engineer to be aware of the various problems and how they may impact on his or her approach in developing a sound remedial solution. The global environmental problems, that impact on the geoenvironmental engineering practice, and actions geared towards restoring the environment are the subjects of discussion in this chapter.
1.2
SCOPE OF E N V I R O N M E N T A L P R O B L E M S
Human activity is bringing about significant changes in the global environment at an unprecedentedly fast pace. It is not an exaggeration to say that the survival of human society itself is threatened by changes in the composition of the atmosphere, the rain which falls on the ground surface, and the amount of greenery which embraces the land. These changes are by no means recent. The expanding scope of human activity, pursuant to the industrial revolution, has steadily overweighted the global environment. The situation has been made all the more serious by the expansion of the global economy and the increase in population since World War II. Every area of our planet's life support system, i.e., air, land, water, and natural ecosystems, is showing signs of damage, as evidenced by: (1) pollution of air, land, and water, (2) water scarcity and degradation, (3) global population increase, (4) growing quantities of wastes, (5) trans-boundary movement of hazardous waste, (6) acid rain, (7) deforestation in terms of land degradation and its contribution to climatic change, (8) desertification and soil erosion, (9) global warming, (10) depletion of the ozone layer, and (11) decreasing species of wildlife. Given these developments, the world has in recent years taken an increasingly great interest in the environment, and there is widespread recognition that it should begin to act in concert in the interest of the future generation. In fact the state of the global environment is now continuously monitored under the UN Environmental Program (UNEP), established by agreement reached at the 1972 UN Conference on the Human Environment, and by the World Meteorological Organization (WMO) and the World Health Organization (WHO). In the following sections we will briefly discuss the various signs of damage enumerated above. For in-depth discussions the reader should consult specialized textbooks and literatures. 1.2.1
Pollution of Oceans and International Rivers
In addition to pollutants which flow into the seas through rivers or directly from the coast, a wide variety of pollutants are released into the seas due to the navigation of ships and their mishaps and the development of oil fields on the sea bed, etc. According to the findings of a survey conducted by the Intergovernmental Oceanographic Commission (IOC) of UNESCO and WMO, released in 1982, many cases of oil pollution were observed on the Bay of Mexico, along the routes of Japanese tankers and along the routes of tankers between the United States and Europe. Oil pollution will sometimes inflict significant damage on birds, marine mammals and other wild organisms, the tourist industry and fisheries, etc. Regarding the global effect, there is the possibility that oil spread over the sea surface would impact on the circulation of water and air in the environment. The pollution of regional seas hemmed in by a number of countries is classic international environmental pollution problem. Particularly in the North Sea, which some of the world's most
SCOPE OF ENVIRONMENTAL PROBLEMS
3
prominent industrial countries surround, there has been concem about the impacts on the ecosystem from the inflow of a wide variety of heavy metals, chemicals, oils and nutrient salts. In the summer of 1988, large numbers of fish and seals died from unknown causes. In the Baltic, the Mediterranean and other seas, pollution is also an issue. To conserve the ecosystems of those regional seas, UNEP urged the countries which share those seas to take common measures. As a result, programs for the conservation of 11 regions in the world have been formulated by more than 130 countries and 11 UN and regional organizations. Efforts (e.g., research, monitoring) are being made to reduce pollution loads. Like marine pollution, the pollution of fresh water is also important, and intemational cooperation has been rendered, primarily by coastal countries, to conserve the water quality of intemational rivers. For example, the water quality of the Rhine, which runs through Westem Europe, deteriorated due to economic development and population increase of its basin. In the early 1970s, fish came up to the water surface, and other damage was caused, due to a lack of dissolved oxygen (DO), and as there was a rise in the discharge of heavy metals and organic chemicals, there arose concem about impacts on drinking water. Given this development, Switzerland, France, West Germany at that time, Luxembourg and the Netherlands established an Intemational Commission for the Protection of the Rhine Against Pollution and adopted a Rhine Chemical Treaty and a Treaty for the Protection of the Rhine against pollution by chlorides. As a result of such international cooperation, the water quality has improved to 80% of the normal value in terms of the quality of dissolved oxygen, and the concentration has decreased for almost all heavy metals. Further efforts are still required for the purification of the Rhine.
Figure 1.1. Global water availability in selected regions (World Resources Institute and World Bank estimates); Renewable resources per capita (m3/year).
4
1.2.2
GEOENVIRONMENTAL ENGINEERING IN A GLOBAL ENVIRONMENT
Water Scarcity and Degradation
Water scarcity and degradation are growing concems for countries around the world. Global demand for fresh water is doubling every 21 years according to the FAO. As industrial, agricultural and domestic pollution threaten finite supplies, water is becoming an increasingly precious resource. Across the world today, renewable water resources available per person are roughly half what they were in 1960. This figure is expected to drop by half again by the year 2025 according to the estimates of the World Bank, shown in Figure 1.1. Clearly, if water resources are not better managed, they would present a burden on economic growth as well as a potential danger to human health and the environment.
1.2.3
Global Population
The global population is projected to increase from the current estimate of 5.5 billion to 8.5 billion over the next thirty years (UN, 1992). It is estimated that out of the 8.5 billion people, 7.1 billion will live in developing countries. As shown in Figure 1.2, the population of the developing regions has increased dramatically. From 1965 to 1995, urban population in Africa, Asia, Central America and South America increased at average annual rates of 4.7%, 3.5%, 3.4% and 3.3%, respectively (WRI, 1994). Europe, United States and Canada experienced average annual growth rates of 1.0% and 1.3%, respectively.
Figure 1.2. Global population estimates.
The relatively high rates of population increase in urban areas of developing regions require rapid infrastructure development. Unfortunately, the gross national products (GNPs) of most of these countries are not growing at a pace that is compatible with their infrastructural needs. Adequate
SCOPE OF ENVIRONMENTAL PROBLEMS
5
programs have not been implemented to control waste and protect human health. In most cities of the developing world, services associated with waste management are not provided to about 30% of the population (Bartone and Bernstein, 1992). Furthermore, less than 5% of the solid waste management budget is typically allocated to waste disposal. In contrast, industrialized countries devote 25-30% of such budgets to waste disposal. Wastes generated in industrialized countries are more toxic and greater in quantity per capita than in developing countries. 1.2.4
Control of Chemicals
In conjunction with the development of the chemical industry, tens of thousands of varieties of chemicals are industrially produced. The amount of chemicals produced is increasing at a strikingly fast pace. A wider variety of chemicals are used in various sectors, including those closely related to citizens' daily lives. Given the rise in the utilization of chemicals, there are increased opportunities for their discharge into the environment. Chemicals like PCB and other organochlorine agents are responsible for pollution on a global scale. Environmental pollution by chemicals is now a common and grave problem in the world. The increased problem of environmental pollution by chemicals is due to their massive and extensive use without a full assessment of their potential impacts on health and the environment. The evaluation of the effects of chemicals and the control of their production, distribution, use and disposal have become very important. Systems have been developed, for example, in European and North American countries and Japan whereby reporting is obligated on the safety of new chemicals before their production, import or marketing. In light of major accidents for which chemicals are responsible, procedures have been established in many countries to document responses to accidents and the amount of chemicals in stock, etc. The Organization for Economic Cooperation and Development (OECD) and other international organizations have taken measures to adopt a system for reporting on new chemicals, prevention and response to large scale accidents, assessment of the safety of chemicals, exchanges of information about the control of chemicals, researches and studies in each country, and cooperative work to collect data on environmental safety tests on existing chemicals common to each country, etc. 1.2.5
Trans-boundary Movement of Hazardous Wastes
In Europe, the trans-boundary movement of hazardous wastes became an issue in the early 1980s. In one case soil polluted with dioxin was found in northern France eight months after causing an explosion at an Italian pesticide plant ( the Seveso incident). In the late 1980s toxic waste was exported from western countries to developing countries in Africa and South America without proper disposal arrangements and there appeared many cases where it was illegally brought in or left untreated. This in turn makes the trans-boundary movement of hazardous wastes a global issue which transcends national borders. The causes of the trans-boundary movement of hazardous wastes include: (1) rising costs of disposal in the home country, (2) diminishing capacity for disposal of certain types of wastes in those countries, (3) potential future liability for any damages by wastes disposed into or onto land in the home country, (4) tightening of laws, regulations and policies concerning disposal of certain types, e.g., prescriptive disposal routes, such as incineration being required for liquids containing certain organic substances, (5) tightening of laws, regulations and
6
GEOENVIRONMENTAL ENGINEERING IN A GLOBAL ENVIRONMENT
policies governing on-site disposal operations for wastes performed by a generator on his own premises, (6) general economic growth which may result in more total generation of wastes, (7) existence of disposal facilities which may serve several countries, (8) market opportunities for materials which can be recovered, reclaimed or recycled from wastes otherwise destined for final disposal, and (9) existence of an appropriate disposal facility in a foreign country which is closer than a similar facility in home country. In light of these incidents, UNEP formulated a basic guideline on the legitimate control of hazardous waste (Cairo Guideline). The OECD and the European Community (EC) adopted various regulations and directives obligating the exporters of hazardous wastes to inform and get permission from the governments of the exporting and importing countries. The Organization of African Unity (OAU) adopted a resolution in 1988 under which a ban would be put on agreements or contracts with foreign governments or enterprises on the disposal of hazardous wastes in Africa. In addition, the 1996 Mediterranean Action Plan (MAP), held in Izmir, adopted a protocol that provides for the following: (1) commitment to abate the production of hazardous wastes, (2) ban on exports of wastes from developed to developing countries, (3) reciprocal commitment by developing countries not to accept wastes, and (4) strict control of the passage of ships transporting wastes through the territorial seas of each state. 1.2.6
Acid Rain
Acid rain refers to the precipitation phenomenon that incorporates anthropogenic acids (those acids that are the result of human activities) and other acidic materials. The deposition of acid materials into the land occurs in both wet and dry forms as rain, snow, fog, dry particles, and gases. The effect of acid rain on a particular ecosystem depends largely on the sensitivity of the ecosystem to acid deposition (Johnson and Gordon, 1987). There is also the ability of the ecosystem to neutralize the acid as well as the concentration and composition of acid reaction products and the amount of acid added to the system. In attempts to cope with the damage caused by acid rain, western nations conducted a Convention Against Long Distance Air Pollution Across National Borders in 1979. The outcome of the convention was incorporated in a protocol of the Helsinki Conference mandating that the discharge of sulfur oxides be decreased by upwards of 30% from the 1980 levels. 1.2.7
Deforestation
The forest is not only a source for supply of firewood and charcoal resources, fruits and the like but also a habitat for birds and beasts. In recent years, the importance of forests as a sink for carbon dioxide, a global warming agent, has been reconfirmed. Deforestation is defined as the change or transfer of forest land into non-afforestation purposes. A check of the availability of global forest resources by region indicates, as shown in Table 1.1, that they are abundantly available in South America (embracing broad expanses of tropical rain forests, primarily in the Amazon basins) and in the former Soviet Union where polar needle-leaved trees are extensively distributed. It was projected by the US Department of Agriculture in the 1980s that forests around the world were decreasing at an annual rate of 18-20 million hectares and that this net deforestation would remain at the same pace till the end of the 20th century.
SCOPE OF ENVIRONMENTAL PROBLEMS Table 1.1: Existence of world forest resources (FAO, 1988) Area of land Area of forests (in million hectares) (in million hectares) Africa 2964 686 N. America 2137 686 S. America 1753 900 Asia 2679 539 Europe 473 157 USSR 2227 944 Oceanic 843 156 World 13077 4069
1.2.8
7
% of forests 23 32 51 20 33 42 19 31
Desertification and Soil Erosion
Desertification is defined as the drop in soil productivity in dry and semi-dry areas. It began to draw attention around the world after 1968 drought, which continued until the 1980s. Devastation of the environment, death of many people, and the appearance of refugees were the result of that drought. In response, the first UN Conference on Desertification (UNCOD) was held in 1977. In 1984, a fact-finding survey was conducted to assess the implementation of the program adopted by the UNCOD against desertification. It concluded that: (1) the land impoverished by desertification increased at an annual rate of 6 million hectares per year, (2) the areas of net economic productivity continued to rise at an annual rate of 21 million hectares per year, (3) about 60% of rainfall farmland, and 30% of irrigated farmland were adversely affected, (4) the total area of desertified farmland accounted for 75% of the total area of production, and (5) the rural population affected by serious desertification had increased from 57 million in 1977 to 135 million in 1983. Soil erosion is closely related to desertification. It represents the phenomenon in which the top soil is lost by wind and rain, hence inflicting damage on farm production. Cultivation on the slopes, inappropriate felling and farming methods are cited as causative factors for soil erosion. With respect to soil erosion, the annual worldwide soil loss is 16 billion tons greater than newly generated soil. It is projected that the world's major grain producers (USA, the former Soviet Union, India and China) account for half of the total loss. 1.2.9
Global Warming
The constituents of the atmosphere are primarily nitrogen (N2) , oxygen ( 0 2 ) , and argon (Ar). The concentration of water vapour (H 20) is highly variable. There are many minor constituents of trace gases, such as carbon dioxide (CO2), methane (CH4), hydrogen (Ha), nitrous oxide (N20), chlorofluorocarbons (CFCs), and tropospheric ozone (O3), among others, that play important roles in radiative and biological processes. In addition to the gaseous constituents, the atmosphere contains suspended and liquid particles. An important consequence of the changes in the constituents of the atmosphere, due to human activity, is the tendency of the temperature close to the earth's surface to rise, a phenomenon
8
GEOENVIRONMENTAL ENGINEERING IN A GLOBAL ENVIRONMENT
referred to as the greenhouse effect or global warming. The rise in the temperature of the earth is analogous to the rise in the temperature in a greenhouse when the energy from the sun is trapped and cannot escape from the enclosed space. As temperature goes up, there will presumably be rises in the sea level and changes in the pattern of precipitation and the volume of evaporation, thus producing grave impacts on ecosystems and human society. It would be extremely difficult to reverse the climate change and the sea level rise once they have occurred. This suggests a need to take action without much delay, notwithstanding the uncertainties that are present. There is a mounting recognition in the world of the serious nature of global warming. As a result, the Intergovernmental Panel on Climate Change (IPCC) was established, with the cooperation of UNEP and WMO. It consists of three working groups. The first working group evaluates scientific findings on the prediction mechanism and uncertainties in climate changes. The second group studies the social and economic impacts while the third group studies policy and strategy to cope with climate change, such as legislation, technology transfer, financial, educational and economic measures.
1.2.10 Depletion of the Ozone Layer Ozone is an important constituent of the atmosphere (Prinn, 1987). It is found in trace quantities throughout the atmosphere, the largest concentration being in the lower stratosphere between the altitudes of 15 and 30 km. The ozone layer results from the dissociation, by solar radiation, of molecular oxygen in the upper atmosphere and nitrogen dioxide in the lower atmosphere. Ozone plays an important role in the formation of photochemical smog and in the purging of trace species from the lower atmosphere. Although ozone is present in small quantities, it plays a critical role in the biosphere by absorbing ultraviolet radiation, which would otherwise be transmitted to the surface of the earth. This radiation is lethal to simple unicellular organisms (algae, bacteria) and to the surface cells of higher plants and animals. It also damages the genetic material of cells and is responsible for sunburn of human skin. In addition, the incidence of skin cancer has been statistically correlated with the observed surface intensities of ultraviolet wavelengths. In recent years, there has been an increased fear that the ozone layer might be depleted by CFCs and other chemicals. CFCs have been extensively used for solvents, refrigerants, and foaming agents, among others. CFCs decompose under intense solar ultraviolet radiation and release chlorine atoms. The released chlorine atoms then decompose ozone catalytically via chain reaction, leading to the depletion of the ozone layer. With the resulting increase in ultraviolet radiation, a wide variety of adverse effects are produced on the growth of farm products, on marine ecosystems, and on human health. An increase in the incidence of skin cancer and cataract typically characterizes the human health effect of ultraviolet radiation. Under the auspices of UNEP, many countries -- including the industrialized nations of the world -- signed an agreement in 1985 in Vienna to regulate the production of CFCs. This initial agreement was followed by the Montreal Protocol in 1987, which called for a 50% reduction in the manufacture of CFCs by the end of the century. In view of the strength of the scientific evidence linking ozone depletion with the release of CFCs, the initial provisions were strengthened in 1990 through the London Amendments to the protocol: that CFCs be essentially phased out by the end of the century. Because of the long residence times of CFCs in the atmosphere, ozone depletion would
SCOPE OF ENVIRONMENTAL PROBLEMS
9
continue well into the next century, even if the protocol were fully enforced. 1.2.11 Decreasing Species of Wildlife The surface of the earth, known as the biosphere, is where humans live. Wild fauna and flora form extensive and diversified ecosystems in the physical environment of the biosphere. Wildlife generally contributes to the maintenance of the balance of the natural environment through, for example, the circulation of energy and substances. It forms the basis for mankind's biological existence and offers blessings in arts, science and culture. Many species have become extinct because of meteorological and topographical changes and/or defeat in their struggles for existence. The increase in harmful ultraviolet rays and climate change, in conjunction with rises in the concentration of greenhouse gases, are considered new factors which might adversely affect the existence of species in the future. The feeling of the international community about the conservation of species was well reflected in the Declaration on the Human Environment at the UN Conference in 1972:
"Man has special responsibility to safeguard and wisely manage the heritage of wildlife and its habitat which are now gravely imperilled by a combination of adverse factors. Nature conservation including wildlife must therefore receive importance in planning for economic development." Nonetheless, the depletion of species continues significantly in tropical regions, which feature the richest diversity of species. In particular, the depletion of wild species is spurred in tropical rain forests, which are considered as a treasure house for wild organisms. The primary cause of the drastic depletion of species in tropical regions is the rapidly ongoing depletion and deterioration of forests. Given this situation, UNEP has started studies on new legal instrument, including conservation, for the promotion of a comprehensive policy for the conservation of biological diversity. 1.2.12 Environmental Pollution in Developing Countries In developing countries, industrialization is increasingly being embraced as the means to prosperity. In the initial stage, industrialization was promoted mainly in the light industry. Emphasis was gradually shifted to the heavy and chemical industries, such as the petrochemical and metal industries. As those industries with heavy environmental loads were constructed, industrial pollution came to the full front. Also, the high population growth rate in developing countries has contributed to the deforestation of tropical forests, thus leading people who are no longer able to maintain their livelihood in the rural areas to massively migrate to cities and rapidly accelerate the urban area population. As a result, pollution is on the uprise, and living environment and health conditions are deteriorating. With regard to air pollution, the data collected by the monitoring network of UNEP and WHO in major world cities indicate that the concentration of sulfur dioxide in cities in developing countries is the same or higher than that in cities of the industrialized countries. The same may be said of suspended particulate matter in the air. In Mexico City, the level of air pollution is extremely high, to the degree that the quantity of suspended particulate matter that a person inhales per day
10
GEOENVIRONMENTAL ENGINEERING IN A GLOBAL ENVIRONMENT
corresponds to the smoking of 40 cigarettes. In many cities in the industrialized countries, there are signs of an improvement, whereas gradual aggravation appears to be the case in cities in the developing countries. With regard to water quality, problems are posed by the inadequate treatment of waste water or the permeation of pesticides into the underground water and their release into rivers. It was noted in UNEP's Environmental Conditions in the World in 1989 that all monitoring points were polluted to varying degrees. The monitoring points were located at: (1) the Ruiji river in Tanzania (with dieldrin at 3.0 lag/l), (2) the Ganca Janchito river in Colombia (with DDT at 0.3 lag/l), (3) the Gombak river in Malaysia (with dieldrin at 30.6 lag/l), and (4) all monitoring points in Indonesia (with PCBs at 0.4-6.9 lag/l).
1.3
INTERCONNECTION OF GLOBAL ENVIRONMENTAL PROBLEMS
As we have seen in the foregoing, global environmental problems have come to the front in various aspects. These global environmental problems are basically caused by the excessive loads from the quantitative expansion and qualitative changes of human activity. With the industrial revolution as a turning point, the striking expansion of human activity and the population increase have made significant demands on the global environment. The increased emission of carbon dioxide responsible for global warming and acid rain is a typical problem associated with the quantitative expansion of human activity. The deforestation of tropical forests, ongoing desertification and the decrease of biodiversities are mainly caused by loads on the land environment. The depletion of ozone layer by CFCs, pollution by chemicals and the trans-boundary movements of hazardous wastes are caused by the massive and extensive use of a wide variety of substances newly developed by humans. Each global environmental problem is also interrelated. For example, CFCs, which deplete the ozone layer, is one of the causative substances responsible for global warming. Due to climate change, the forests are weakened, and desertification as well as global warming are accelerated. The deforestation of tropical forests is the primary cause of the depletion of wild species. Marine pollution deters the absorption by the seas of carbon dioxide and accelerates global warming. Therefore, there are many aspects where one problem concerning the global environment causes, or results from, yet another problem. Ecosystems around the earth exist on a delicate balance. It will be extremely difficult to restore this balance, once it has fallen apart. Consequently, if the present generation does not come up with adequate protective measures, and the environment deteriorates on a global scale, many burdens will eventually be imposed onto the next generation. In this context, it is necessary to comprehensively step up conservation of the global environment, with a far-reaching long term perspective that transcends the generations.
1.4
GEOENVIRONMENTAL ENGINEERING ASPECTS
The preceding discussion regarding global environmental problems clearly highlights various challenges to the waste management and environmental restoration industry. As indicated previously, the solutions to these complex problems are currently handled by the geoenvironmental engineering
GEOENVIRONMENTAL ENGINEERING ASPECTS
11
profession, which is concerned with land environmental problems. The development of solutions to restore the degradation of the environment may be additionally complicated by the interconnecting nature of the various ecosystems, i.e., atmosphere, hydrosphere, geosphere, and biosphere, that constitute the land. This means that a cooperative and holistic global effort should be considered in developing a viable solution to global environmental problems. The movement of pollutants within the land has a profound effect upon its bioavailability. Once a pollutant enters one of the mobile phases, i.e., air or water, it disperses rapidly due to fluid movements. Movement within a phase is termed interface mass transfer, diffusion, or dispersion. Interface mass transfer is important to the movement of pollutants between the various phases of the ecosystem. People and other organisms that constitute the biosphere reside, to varying degrees, within the other three spheres. Several direct and indirect routes of pollutants through the land and eventually to humans exist and are shown in Figure 1.3.
~ G~&e
POLLUTANT
WATER AIR
Direct contact
9
ter
HUMAN Nutrients
SPHERE Plants & Animals
Figure 1.3. Direct and indirect routes of pollutants through the land and eventually to human.
The indirect contact modes are of a more subtle nature. The continual intake of air and water is an indirect source of many pollutants, which are residuals in the air and water phases. Pathways for the entry of pollutants in the land environment are shown in Figure 1.4. At present, we do not understand: (1) how the interactive physical, chemical and biological processes regulate the total environment, and (2) how such a system will respond to anthropogenic influences. To develop this understanding, researchers must develop a knowledge base that embraces many disciplines and accommodates many cultures. Since policy makers are demanding a firmer base for national and international decisions, such a knowledge based system is urgently needed. Therefore, what is required is the development of an initiative which considers all relevant dimensions of the driving forces of global change. Ultimately, it is a combination of scientific
12
GEOENVIRONMENTAL ENGINEERING IN A GLOBAL ENVIRONMENT
knowledge, public understanding and political will that will initiate the actions that can preserve our global environment for future generations.
e.9
v L (Atmosphere) A,R jr
E
.9
~. .N_ ",~ oration ~ I .-o o "~ -o o > < Dissolution/ ~dsorptionf < > -'~ SOIL f WATER i . o: :u :a n:. . ] ,~1 (Geosphere) (Hydrosphere) Adsorption~J ~I~ N
. ~ p
~
t-
~
,m
~,.
t~
o
-"
t
Manufacturing Transportation
I,
Desorption
Sources of Pollutants Customer Consumer
Treatment Disposal
Figure 1.4. Pathways for entry of pollutants into the land environment.
1.5
ACTIONS TOWARD RESTORING THE ENVIRONMENT
1.5.1 Comprehensive Policy Response The global environment issues are so grave that they appear to threaten the very foundation of human survival. As we have discussed previously, global environmental issues interlock in forming a group of issues which, with the joint cooperation of the international community, need to be comprehensively addressed in a broader and long term perspective. It is important that all nations take up conservation of the global environment as an important policy task and take initiatives in realizing sustainable development on a global scale. It is necessary for every government to strengthen the comprehensive approach based on pollution prevention and environmental management in administration of the environment while incorporating the perspective of environmental conservation in the various policy sectors concerned (e.g., economic policy) to realize an energy and resource-saving society. Moreover, it might be said that the key to global environmental conservation lies in the manner in which each member of the society views the environmental issues which exist around hem or her and what sort of response he or she makes. On this score, too, it is important for each government to take the initiative to reduce environmental loads and to enhance the consciousness of the people by encouraging the dissemination of information about the environment (i.e., environmental education), and to strive to establish a code of environmental ethics.
ACTIONS TOWARD RESTORING THE ENVIRONMENT
1.5.2
13
Environmental Policy The Economic Declaration of the Arche G-7 Summit reads:
"Environmental protection is integral to issues such as trade, development, energy, transport, agriculture and economic planning. Therefore, environmental considerations must be taken into account in economic decision-making. In fact, good economic policies and good environmental policies are mutually reinforcing." At the 1989 Tokyo Conference on Conservation of the Global Environment, it was stated that:
"The sustainable development calls for a review of not only the conventional framework of the worm economy, such as trade, direct investment, international financing and official development aid, but each country's domestic economic, financial and monetary policies." In addition, the 1992 Rio Declaration on Environment and Development stated that: (1)
With reference to sustainable development:
"In order to achieve sustainable development, environmental protection shall constitute an integral part of the development process and cannot be considered in isolation from it." (2)
With reference to environment protection:
"In order to protect the environment, the precautionary approach shall be widely applied by States according to their capabilities. Where there are threats of serious or irreversible damage, lack offull scientific certainty shall not be used as a reason for postponing cost-effective measures to prevent environmental degradation." (3)
With reference to the need for environmental impact assessment:
"Environmental impact assessment, as a national instrument, shall be undertaken for proposed activities that are likely to have a significant adverse impact on the environment and are subject to a decision of a competent national authority." The realization of sustainable development is not an easy task even in industrialized countries. It is particularly difficult to achieve in developing countries which are troubled by a vicious cycle of poverty and environmental disruption. As the first step toward the realization of sustainable development, it might be necessary, more than anything else, to check if the various policies, institutions and systems associated with a nation's economic management, energy use and resources management are really in line with the ideals of sustainable development. The concept of sustainable development and its relation to engineering and environmental impact assessment is
14
GEOENVIRONMENTAL ENGINEERING IN A GLOBAL ENVIRONMENT
discussed in Chapter 2. Certain indicators are required for the assessment of economic policy from a sustainable development viewpoint. The gross national product (GNP) and the gross domestic product (GDP), which are broadly used in the world, are not necessarily adequate. The Arche G-7 Summit called on the Organization for Economic Cooperation and Development (OECD) to develop a new indicator which would integrate the economy with the environment. In Japan, an indicator -- the net national welfare (NNW) -- which would take account of damage due to environmental pollution and costs for preventive measures was developed, but was not broadly embraced. The concept of a natural resources account (NRA), which France, Norway, Sweden and other countries have explored for the last several years, is now being studied by the UN and the OECD. In this system, the existing volume and quality of the natural resources are assessed to quantitatively express how the balance changes from the beginning to the end of each year. At the present time, the NRA is not considered a substitute for the GNP indicator, but can be used to complement the GNP in the economic policy decisions. 1.5.3
Environmental Ethics
The 1989 Tokyo Conference on Conservation of the Environment called for a review of a wide variety of economic and social activities which evolved in developed countries, including the life-styles, and appealed for the establishment of a code of environmental ethics. It was requested that:
"All countries, especially industrialized countries, should recognize the need to make their socioeconomic activities and life-styles environmentally sound." "[All countries] endeavour to carry out socioeconomic activities in a manner which has less burden on the global environment, such as the promotion o f resource conservation and saving of energy." "[All countries] will pursue the awareness and education programs o f the global environmental protection, since it is indispensable to obtain the understanding and cooperation of the public in all walks of life."
"Efforts be made so that economic society may be managed in a manner with few loads on the global environment, such as the promotion o f resource- and energysaving." "Dissemination and enhancement be promoted for conservation o f the global environment, as the understanding and cooperation of each segment o f the people are indispensable." To create a resource- and energy-saving society, it is necessary that each State promotes and strengthens the above measures. Furthermore, it is important to give policy guidance to entrepreneurs so that they may conduct their businesses with a full understanding of the relationship between their own activities and the environment.
SUMMARY AND CONCLUDING REMARKS 1.6
15
SUMMARY AND CONCLUDING REMARKS
Global environmental problems, when taken singly or in combination, pose challenges to the geoenvironmental engineering profession. In particular, the safe disposal of hazardous solid and liquid wastes is paramount. Concerns should be focussed on the requirements and criteria needed to establish what is safe, with respect to public health and the environment. Fundamental to the achievement of sustainable development is the development and use of environmentally responsible land disposal and waste management systems. Thus, if we consider the environmental policies of the Arche G-7 summit, the 1989 Tokyo Conference on Conservation of the Global Environment, and the 1992 Rio Declaration on the Environment and Development, it becomes clear that the proper practice of geoenvironmental engineering will be critical. The geoenvironmental engineering profession will need to provide modelling capability of the various interactive processes (e.g., soil-pollutants), the criteria for effective waste disposal and management, and technologies for various remedial measures. The problem of waste management requires solutions that satisfy criteria designed to protect both human health and the environment. This requires attention not only to waste water quality and treatments needed to increase the quality of the discharge streams, but also to groundwater protection. A proper and effective solution will inevitably require a thorough understanding of the: (1) composition and characteristics of substrate materials, (2) composition of wastes and issues arising from the handling of the waste materials, (3) interaction mechanisms between the waste and substrate materials, (4) basic principles of the treatment techniques, and (5) immediate and long term adverse health effects of the chosen remedial measures.
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CHAPTER TWO
SUSTAINABLE DEVELOPMENT
2.1
INTRODUCTION
The need for sustainable development has been recognized for many years. However, it is only within the past few years, as discussed in Chapter 1, that it gained prominence as an important concept and philosophy upon which economic development and environmental management should be based. Sustainable development has been defined by various organizations as: (1) The 1987 World Commission on Environment and Development (WCED) defined sustainable development as: "Development that meets the needs of the present without compromising the ability offuture generations to meet their own needs." (2)
The 1989 Tokyo Conference on Conservation of the Global Environment stated that: "Sustainable development calls for a review of not only the conventional framework of the worm economy, such as trade, direct investment, international financing and official development aid, but each country's domestic economic, financial and monetary policies."
(3)
The 1992 Rio Conference on Environment and Development stated, in Principle # 3, that: "The right to development must be fulfilled so as to equitably meet developmental and environmental needs of present and future generations." and Principle # 4 further stated that: "In order to achieve sustainable development, environmental protection shall constitute an integral part of the development process and cannot be considered in isolation from it."
Sustainable development recognizes that environment and development are highly interdependent and should not be dealt with as separate issues. The basic elements of sustainable development are (Muschett, 1997): (1) population stabilization, (2) new technologies and technology transfer, (3) efficient use of natural resources, (4) waste reduction and pollution prevention, (5) integrated environmental management systems, (6) determining environmental limits, (7) refining market economy, (8) education, (9) perception and attitude changes, and (10) social and cultural 17
18
SUSTAINABLE DEVELOPMENT
changes. Fundamentally, sustainable development aims for the satisfaction of human needs, the maintenance of ecological integrity, the achievement of equity and social justice and the provision of social self-determination. The real challenge lies in finding ways of putting sustainable development into practice.
2.2
APPROACHES TO SUSTAINABLE DEVELOPMENT
It is desirable that approaches used by engineers and other disciplines in an effort to achieve sustainable development fit into a multi-disciplinary planning and design environment and have the ability to cross over disciplinary lines and form the common foundation of a unified process. To date, sustainable development has been examined from many different angles by various disciplines, as discussed below. 2.2.1
Economics and Sustainability
In view of Principle # 4, environmental protection is considered as an integral part of the development process. This is different from the traditional pattern of making economic decisions and then correcting the environmental impacts which may result (Muschett, 1997). This can be illustrated with the use of Figure 2.1 in which the natural system includes the ambient physical environment, ecosystem and natural resources. The economic system refers to the factors of production of goods and services. Utilization of the natural system by the economic system results in a decrease in the natural resources, and produces additional environmental problems, such as solid wastes, air and water pollution, and greenhouse gases. The importance of these impacts upon the natural system varies geographically, depending on the existing states of both the natural environment and the economy. There is increasing recognition that the economy and the environment cannot be viewed entirely separately. Several authors in the field of ecological economics (Hanley, 1995; Norgaard, 1984; Perrings, 1987; Common and Perrings, 1992) have argued that dynamic interdependencies between economic and ecological systems imply that strict economic optimization may lead to a decrease in the stability of ecosystems, producing undesirable feedbacks on the economic system. The world view sees the parameters of both systems (economic and ecological) co-evolving, so that decision-making based on one system only is unlikely to be a sufficient stand-alone decision mechanism. Economic sustainability involves the consumption of interest rather than capital and is defined by Goodland and Daly (1995) as: "the amount one consumes during a period and still be as well off at the end o f the period." Though this definition has focussed solely on man-made capital, the principle can be broadened to include natural capital. Rather than view the economy as an isolated system, we must begin to see it as linked to the environmental system. Therefore, the economic sustainability can be defined as the maintenance of capital in general, both man-made and natural. The concept of growth, defined as the increase in size by the assimilation or accretion of materials, comes from a time when the economy was seen as an isolated system. Measures of growth such as the gross domestic product (GDP) were questioned and criticized, since they measure only
APPROACHES TO SUSTAINABLE DEVELOPMENT
19
economic flows without considering how much of these flows result from the liquidation of manmade capital or natural resources. It is clear that many activities that were considered positive in terms of GDP actually represented something negative in terms of human well-being. One recent example that is often cited is the Exxon Valdez oil spill, off the coast of Alaska, which had a very positive effect on the GDP while devastating the coastal environment.
Minerals
SYSTEM ~.
.
SYSTEM1
Solid Waste
Figure 2.1. Conceptual interactions between economic and natural systems.
The debate over the future of the fundamental tools of economics continues today, with the field of environmental economics receiving increasing attention. The basic approach of environmental economics seeks to integrate the environmental system into a broader economic system in which current economic principles will still apply. Much research is being done on internalizing environmental values which are external to the classic economic system, as well as finding ways of placing monetary values on intangible non-market (and non-marketable) components of the environment. Techniques such as contingent valuation (CV) (Johansson et al., 1995), which attempt to place a monetary value on environmental goods based on peoples' willingness to pay, as determined by means of specially designed surveys, are gaining much credibility. The CV technique was used to establish the cost of environmental damages caused by the Exxon Valdez oil spill for litigation purposes. Though such techniques are attractive, since they theoretically allow one to include environmental costs in cost-benefit analysis, the practice has many critics who do not believe that all social and environmental effects can or should be monetized (Hanley, 1995; de Laet, 1990). Nonetheless, the practice of contingent valuation is growing rapidly, and should not be dismissed.
20
2.2.2
SUSTAINABLE DEVELOPMENT
Environmental Sustainability
Resource scarcity in many areas highlights the need to live within the limitations of the natural environment. We depend on natural capital for resources (inputs) and as a sink for wastes. Environmental sustainability, in the interest of human welfare, seeks to protect the raw materials, and ensure that their capacity as a waste sink is not exceeded (Goodland and Daly, 1995). Though non-renewable resource use is by definition unsustainable, one approach is to ensure the depletion rate does not exceed the rate at which renewable substitutes can be created (El Sarafy, 1991). With the maintenance of natural capital approach in mind, Goodland and Daly (1995) draw some practical rules of thunab, which they call the input/output rules, to guide economic development. These rules are:
(1)
(2)
2.2.3
Input rules: (a) Renewable: harvest rates of renewable resources inputs would be within the regenerative capacity of the natural system that generates them; and (b) Non-renewable: depletion rates of non-renewable resource inputs should be equal to the rate at which renewable substitutes are developed by human invention and investment. Part of the proceeds from liquidating non-renewable resources should be allocated to research in pursuit of sustainable substitutes (El Sarafy, 1991). Output rule: Waste emissions from a project should be within the assimilative capacity of the local environment to adsorb, without unacceptable degradation of its future waste absorptive capacity or other important services.
Social Sustainability
It has long been recognized that investments in human capital, such as education, health, and nutrition are an important part of economic development. Though, social capital has not gained the same recognition, it is nonetheless required for social sustainability. The key components of social sustainability are human rights, education, employment, women's empowerment, transparency of decision-making, fiscal accountability, participation as well as what some call moral capital. Moral capital, which is generally more difficult to measure, includes such factors as social cohesion, cultural identity, diversity, sense of community, love, tolerance, compassion, humility, patience, forbearance, fraternity, fellowship, institutions, pluralism, commonly accepted standards of honesty, laws, discipline, etc. In order that this capital not depreciate, it must be maintained and replenished by shared values and equal rights, and by community, religious and cultural interactions (Goodland and Daly, 1995). Ways in which social sustainability can be promoted include: (1) encouragement of systematic community participation, (2) emphasizing full cost accounting and cradle-to-grave pricing, including social costs, (3) promotion of qualitative improvement of social organization patterns and community well-being over quantitative growth of physical assets, and (4) use of resources in ways that increase equity and social justice while reducing social disruptions.
APPROACHES TO SUSTAINABLE DEVELOPMENT
2.2.4
21
Land Sustainability
Land sustainability can be achieved by integrating the ecological, economical, and social objectives (Serageldin, 1993; World Bank, 1992), as shown in Figure 2.2. Ecologists stress the preservation of the integrity of the ecological systems that are critical to the overall stability of our global ecosystem, and deal in measurement units of physical, chemical and biological entities. Economists seek to maximize human welfare within the existing capital stock and technologies, and use economic units (i.e., money or perceived value) as a measurement standard. Sociologists emphasize that the key actors in sustainable development are people with a range of needs and desires, and use units which are often intangible, such as well-being and social empowerment.
ECOLOGICAL OBJECTIVES El Ecosystem integrity ICI Carrying capacity El Biodiversity D Global issues
ECONOMIC OBJECTIVES I:1 Growth I:1 Equity I::1 Efficiency SUSTAINABLE DEVELOPMENT
SOCIAL OBJECTIVES I:1 Empowerment I:1 Participation a Social mobility m Cultural identity !~1 Institutional dev.
Figure 2.2. Integrated set of objectives for sustainable development (Serageldin, 1993).
Sustainable solutions for land development fall at the intersection of the spheres, as shown in Figure 2.3, that represent the three key ingredients for sustainable development (Campbell and Heck, 1997). Sustainable development occurs only when management goals and actions are simultaneously ecologically viable, economically feasible, and socially desirable. These imply environmental soundness and political acceptability. Imbalance among the three components, due to failure in one or more of the spheres, will likely result in failure to achieve sustainable development (Zonneveld, 1990).
22
SUSTAINABLE DEVELOPMENT
COLOGICAL VIABLE
Figure 2.3. Sustainable solution for the development of sustainable land.
2.3
SUSTAINABLE DEVELOPMENT AND THE AMBIENT ENVIRONMENT
2.3.1
Assimilative Capacity in Environmental Management
The concept of sustainable yield was originally developed for harvesting ecological renewable resources, such as forests and fisheries, at the rate at which nature (assisted by human management) was able to replenish. In practice, this is a kind of dynamic equilibrium since natural factors such as climate and ecological productivity vary from year to year. It should also be noted that when the harvest rate exceeds the replenishment rate, a new state of equilibrium with a lower resource base is reached (Muschett, 1997). There are limits to what nature will permit without damaging the ecological system and resource base. Similarly, environmental scientists have come to recognize that the physical, chemical and biological characteristics of the ambient environment determine the ability to accept, dilute, diffuse and transform pollutants. This assimilative capacity limits the amount of pollution tolerable without causing damages. This principle holds whether we are considering a very localized leachate plume from a landfill site or in a river, a regional air pollution problem or a global climate change (Muschett, 1997). In general, as the geographical scale increases, the complexities and interactions of natural processes also increase. Therefore, to achieve sustainable development, we must consider the assimilative capacity of the environmental system, which in turn determines the carrying capacity of the supporting population and economic activity and the resulting pollutant emissions. A general framework for environmental quality management is shown in Figure 2.4 (Muschett, 1997). There is often an iterative process which examines different strategies, the resulting spatial patterns of discharge of pollutants, and the modelled ambient concentrations in terms of the assimilative capacity. Ultimately, one or more environmental strategies are selected and implemented.
SUSTAINABLE DEVELOPMENT AND THE AMBIENT ENVIRONMENT
)
23
Measure ambient environmental quality
NO
Is allowable limit
Inventory sources of pollutants Determine ambient assimilative capacity Determine environmental management strategies Model ambient concentration with strate_qies YES
s allowable limit ._exceeded? _ ,
~
~
N
O
Implement environmental management strategies
Figure 2.4. General framework for pollutant management.
2.3.2
Water Quality Management and Sustainable Development
Depending upon the amount of pollutant discharged into a water body and the assimilative capacity, the resulting concentrations of the pollutants in the water and in the tissues of aquatic organisms will determine whether the body of water is fit or unfit for human consumption, aquatic life, commercial fishing, recreational purpose or industrial use. There are two constraints that the environmental system poses in relation to water quality. First, river flows tend to be extremely variable from season to season, and coupled with withdrawals for human use, there is a severe upper limit on assimilative capacity. Second, the increasing bioaccumulation of pollutants in successively higher levels of the food chain also severely limits the allowable concentration of pollutants in the water. The combination of these limiting ambient conditions, together with a high density population and economic activity, can make sustainable development very difficult to achieve.
24 2.4
SUSTAINABLE DEVELOPMENT ENVIRONMENTAL IMPACT ASSESSMENT
The term environmental impact, which has often been cited as a requirement in the assessment of the performance of many civil facilities, is now becoming a requirement for many kinds of activities associated with engineering projects. According to Therivel (1992), environmental impact assessment (EIA) is "the process of predicting and evaluating an action's impact on the environment, the conclusions to be used as a tool in decision-malang." The practice of EIA began in the United States, following the National Protection Act (NPA) of 1969, with its prime focus on dealing with environmental issues in the public forum (Beanlands, 1985).
2.4.1 Guiding Principles Many would argue that in the context of impact assessment, ecosystems have no intrinsic importance other than the extent to which they are valued by man (Beanlands, 1985). Thus, in an impact assessment, the ecological repercussions of a project are usually translated into effects on the physical and biological resources that are valued by humans for commercial, recreational or aesthetic reasons (Beanlands, 1985). These are often referred to as valued ecosystem components (VECs). Ecologists, however, are aware of the more profound changes that could affect the function of natural systems. The involvement of biologists and ecologists in environmental impact assessment has led to what is referred to as the ecosystem approach. In environmental planning, management and assessment, the ecosystem approach implies the following: (1) examining the combined effects of multiple stresses in a defined place or time, (2) understanding the carrying capacity, tolerance level, or assimilative capacity of the ecosystem, (3) knowledge of the extent to which the ecosystem is already under stress, and (4) considering broader temporal and spatial boundaries (e.g., regional level rather than just the project level). In an effort to assess the various approaches to EIA, Gardner (1990) reviewed much of the literature and defined eight main principles of sustainable development. These principles can equally be applied to the planning and design process, given that environmental and social considerations need to be considered from the outset. The eight principles can be subdivided into two categories: (1) substantive, and (2) process-oriented. The four substantive principles of sustainable development are: (1) satisfaction of human needs, (2) maintenance of ecological integrity, (3) achievement of equity and social justice, and (4) provision for social self-determination. The four process-oriented principles of sustainable development are: (1) goal-seeking, (2) rational, (3) adaptive, and (4) interactive. It has become quite evident that the engineering profession must adapt and evolve with current conditions. Perhaps the single most important indication of this is the overwhelming consensus in the EIA literature that environmental factors need to be considered along with technical and economic factors from the beginning of the planning process (Molget, 1996). Such integration can have many benefits. For the proponent, these can include reduced cost and time of project implementation, cost saving modifications in project design, increased project acceptance and avoided impacts and violations of laws and regulations. Other results, such as reduced environmental damages, avoided health costs, maintenance of biodiversity, avoided cleanup/treatment costs, minimized land use/resource-use conflicts and the achievement of socioecosystem objectives can all represent benefits for the population. Achieving sustainable development through engineering practice will require that the
ENVIRONMENTAL IMPACT ASSESSMENT
25
substantive principles be reflected in the goals, values or objectives used to evaluate design choices or compare design alternatives. The process-oriented principles will be used as a framework for evaluating the suitability of an engineering methodology for achieving sustainable development. Though a suitable methodology does not automatically lead to sustainable engineering practice, it would facilitate rather than hinder the achievement of this goal. 2.4.2
Attributes of a Successful E I A
Despite the problems that have been encountered, or perhaps as a result of these problems, a lot can be learned from the EIA processes developed over the past twenty five years. Although much work remains to be done in this field, many techniques have been developed to predict and evaluate environmental impacts, as well as to bring together the knowledge of various disciplines. This experience can be most helpful since many of the attributes of a good approach to EIA are equally valid for an integrated planning and engineering process. Both need to consider environmental and social factors as well as technical and economic ones, and both involve communications among many specialists and stakeholders. In fact, many practitioners of EIA claim that it should be a continuous process which begins at the planning stage and extends through project design, environmental assessment, construction (or execution) to post-project monitoring. It is likely, however, that more emphasis will be placed on feasibility at the planning and design stages. Many textbooks now exist on the subject of impact assessment, and these should be consulted for specific examples and case studies. For the purpose of this book, however, we have compiled a list of desirable attributes of EIA processes which adhere to the principles of sustainable development (Erickson, 1979 and 1994; Molget, 1996; Sadar, 1994; Jain et al., 1993; Wright and Greene, 1987; Whitney and Maclaren, 1985; Rosenberg and Resh, 1981): (1) Comprehensive: should cut across disciplinary lines and account for the complex interrelationships that exist in the natural world between the physical, biological and human environments; (2) Inclusive:should seek the involvement of all the relevant specialists at an early stage of planning and, where required, representatives of public and private interests; (3) Objective:should provide unbiased measurements and predictions, free from political influence or other interests; (4) Comparative:should identify project-induced changes and effects as separate from those which would occur under existing natural conditions. Should also permit comparisons between project variants, altemative projects, including the option not to proceed; (5) Selective:critical impacts should be identifiable as early in the process as possible. Unimportant impacts should not be allowed to dissipate efforts or hinder decision-making. (6) Repeatable:should be repeatable and permit an iterative process of design modification and impact prediction; (7) Integrated:should begin in the earliest stages of planning and design, be useful during the EIA process, and continue through construction and post construction; (8) Dynamic:should be flexible and adaptable, allowing for changes and modifications along the way, including the possibility of a shift in the main objectives; (9) Understandable:should not be overly complex; figures, relationships and values should be readily understandable; (1O) Open:should not be a secretive or closed process. Secret and closed processes only serve to
26
(11)
2.4.3
SUSTAINABLE DEVELOPMENT feed public suspicion. By being forthright and honest, promoters have a much greater chance of having their project accepted by the public. Private promoters may have industrial secrets to protect which may hinder this openness. Also, their investors are their primary concern. Public utilities, on the other hand, should not have this impediment. In fact, since they exist to serve the good of the population as a whole, public utilities have a greater responsibility to being open with the public; Honest: should avoid tokenism. Taking into account environmental concerns in engineering must be a completely honest process in order to be constructive. Promoters, forced by legislation to perform environmental assessments, often only pay lip service to the environment, with the sole purpose of obtaining a permit to construct. Rosenberg and Resh (1981) describe three ways in which tokenism is manifest: (a) Assessment which plays no role in the fundamental yes/no decision for a project. In this case, rather than addressing the rationale which justify a project, the assessment assumes the project will go ahead and concentrates instead on mitigation. To this category can be added the assessment which assumes a project will be built in some form or other, and therefore concentrates on comparing the impacts and the benefits of project variants; (b) Assessment that is simply justification for existing engineering designs or management decisions; and (c) Assessment which is treated as necessary legal hurdles to overcome before starting a project. Evaluative Tools
Evaluative tools include techniques such as benefits and damage estimation (BDE), costbenefit analysis (CBA) and cost-effectiveness analysis (CEA), all of which rely a great deal on the accounting system used for natural capital (Stokoe, 1991; Hamel et al., 1986). In the context of sustainable development, the usefulness of analytical approaches such as CBA is becoming increasingly questionable. Many economists now believe that CBA is an inappropriate technique for projects with ecological impacts. Peter Soderbaum, a Swedish institutional economist, said: "For those who consider evolutionary processes and the paths that ecosystems take over time, the idea that non-monetary impacts at different periods o f time can somehow be pressed together into one value ...... is absurd." (Hanley, 1995)
2.5
ENGINEERING FOR SUSTAINABLE DEVELOPMENT
If it is accepted that the engineering profession has a responsibility to respect the principles of sustainable development, then every engineer should acquire and maintain an understanding of the goals and issues related to sustainability and conduct his/her work in a manner which supports sustainability. Engineering for sustainable development cannot be achieved by engineers alone, but must evolve within the framework of a larger planning process in which many disciplines and other parties cooperate. The following steps are recommended for the planning and design process leading to projects or activities which may impact on the environment (Molget, 1996); they are of particular
ENGINEERING FOR SUSTAINABLE DEVELOPMENT
27
relevance to projects or activities for which an environmental impact assessment may be required by law.
Step 1: Define the Objectives Defining the overall objectives should be the first task brought for discussion. This task involves many stakeholders, such as local communities potentially affected by the project or activity, the proponent, and other businesses. It is also essential to have the involvement of the decisionmakers at this point -- those who will decide in the end which option best meets the stated objectives. In this process, sustainability must also be declared as a principal objective. Objectives which relate to economic, social and environmental sustainability are (Goodland and Daly, 1995): (1) Social objectives: empowerment, participation, equity, poverty alleviation, social cohesion, population stability, and institutional development; (2) Economic objectives: development for all countries, growth for less developed countries, efficiency, poverty alleviation, and equity; (3) Environmental objectives: ecosystem integrity, conservation of carrying capacity, climatic stability, and conservation of biodiversity.
Step 2: Adopt a Cooperative Approach In order to progress effectively through the planning and design process, respective disciplines and stakeholders must be able to cooperate. Adopting a structured cooperative approach can help prevent the process from being paralysed by disciplinary bias or by conflicting views. There are three common approaches used in EIA processes for obtaining input from all the relevant disciplines (Holling, 1978): (1) interdisciplinary team, (2) modelling workshop, and (3) study tasks (individual discipline). Employing such structured techniques in planning and design will foster efficient and productive results from the process. This methodology recommends using one of the first two approaches to benefit fully from the involvement of all participants. When required, the study tasks approach may be used in addition to the primary approach. The following briefly outlines some of the advantages and disadvantages of these approaches (Holling, 1978).
Interdisciplinary Team It attempts to promote communication among disciplines by having specialists work together as a team. Individuals normally work in the same location. There is a better chance of having crossdisciplinary effects considered using this approach. The disadvantages are that it is costly to maintain a full team of qualified specialists and the process can get bogged down in details, in view of scientists' tendency to break things down into components and sub-components.
Modelling Workshops Using this approach, small teams of qualified specialists, methodologists and decisionmakers are brought for short and intense working sessions. Between sessions, analysts create and update models in order to structure and relate the information ensuing from the sessions. There are many advantages to this approach: (a) it is cost effective, (b) provides a better chance of having cross-disciplinary effects considered, (c) better chance of sequestering highly qualified individuals
28
SUSTAINABLE DEVELOPMENT
for a short period of time, thus avoiding unnecessary detail since emphasis is on the essentials, and (d) it is an effective way of beginning problem analysis, maintaining focus and forcing judgement. To be most effective, professionals, specialists, methodologists familiar with analysis and modelling, and decision-makers all need to be present at the modelling sessions.
Study Task In this approach, studies, reports, and/or statements are requested from different specialists regarding the probable impact of a given development decision on their specific area of expertise. These requests usually take the form of consulting contracts. Some coordination is required in administrative matters, data gathering and preparation of a final report. Drawbacks to this approach are: (a) cross-disciplinary interactions are often omitted, (b) lack of transparency, and (c) potential for selective presentation of information by the promoter, since they decide which contract/research will be done by whom and what results will be made public. While no single approach is ideal for all situations, each has advantages and weaknesses that should be noted when selecting a particular approach for a given situation. For the purpose of crossdisciplinary communication and cooperation, the first two approaches are preferable.
Step 3: Develop Options to meet Objectives The search for alternative means of fulfilling the objectives should be a multi-disciplinary multi-stakeholder exercise. Individuals that would normally be involved in the analysis of options should also be involved at this earlier stage of defining potential options. Options can arise from efforts of individual disciplines (such as engineers), but can also be created or enriched through the creative exchange between all parties. In this way, engineers involved in preliminary design work will have the benefit of input and feedback from the various key participants.
Step 4: Identify the Effects of each Option This step involves analysing the consequences of each altemative design, plan or action. Activities include identifying the benefits as well as the potential environmental and social impacts of each option and developing criteria and indicators for evaluation. This stage may also involve data collection as well as modifications to options and/or the addition of new options. Once different solutions are put forward which meet the objective(s) to varying degrees, discussion should focus on identifying the various effects, both positive and negative. What is often lacking at this point is a common language and a framework which will keep the discussion focussed, and which will aid in progressing towards better design and choices. Knowledge from the different disciplines needs to be integrated and studied on a common ground. A goal should be to define both the sources of environmental impacts as well as the mechanisms involved in their occurrence. This knowledge is essential in order to predict impacts. Using this approach, all environmental impacts are then attributed to some physical change resulting from the development. Physical change is defined as a temporary or permanent alteration in the physical (or chemical) environment. The physical environment is comprised of land, water and air, but excludes those elements better described as part of the biological or human environments. Repercussions of physical changes on the biological or human environments are referred to as impacts. By linking impacts and physical changes, the causes of these impacts are better understood,
ENGINEERING FOR SUSTAINABLE DEVELOPMENT
29
and engineers can readily identify problem areas in the design. At this stage, fundamental design choices can still be revisited, and change made to eliminate, reduce or compensate for various impacts.
Step 5: Evaluate the Options This step involves the most difficult task of evaluating the various options in preparation for the selection of the most appropriate course of action. It is important that all relevant disciplines and stakeholders have an input in this important step. What is required at this stage is a consistent framework that will help structure the process toward making complex decisions. Quantitative decision analysis techniques provide this structure and can aid in making trade-offs between competing attributes. The challenge remains in developing a model that will do so without constraining the decision making process by removing or restricting judgements that can be made. Multi-attribute decision analysis and multi-criteria analysis are techniques that can be helpful at this stage.
Multi-Attribute Tradeoff Analysis Techniques such as cost-benefit analysis (CBA) require attributes to be quantified on a common scale -- usually in monetary units. In the case of environmental attributes, it is difficult and often impossible to design monetary values. Multi-attribute tradeoff analysis (MATA) avoids this problem. The first step in MATA (Litchfield et al., 1994; Andrews, 1992) is to identify the most important attributes (or criteria) and determine a scale upon which to measure the performance of each attribute. Analysts and decision-makers should be involved in this determination. An advantage with this technique is that environmental attributes can be physical measures, such as km 2 or tons, or subjective measures such as a range from very high impact to no impact. Analysts can then measure the performance of each option based on each attribute. Normally, the next step would be to determine the relative importance of each attribute in order to be able to rank the options.
Multi-Criteria Analysis Simply put, multi-criteria analysis is an action that provides the tools to assist the decisionmaker in resolving a problem in which several, often contradictory, points of view need to be considered (Vinke, 1989). It is a way of thinking (in most situations there are usually many points of view, and aspects of criteria that are often conflicting); it is a communication tool, involving structured interaction between analysts and decision-makers); and it is a method of analysis. The procedure can be described in four steps (Brunelle and D'Avignon, 1996): (1) Definition of the problem and of potential alternative actions (choice, ranking or ordering); (2) Analysis of the consequences of each action, determination of the criteria, and the evaluation of each alternative action with respect to each criterion (construction of a performance table); (3) Global performance modelling and aggregation of the performances (criteria to retain, aggregation of the performance of the actions based on these criteria, relative importance of the criteria, etc.); and (4) Multi-criteria synthesis (analysis of results, robustness, discussion) and identification of the best action. The steps described are not necessarily successive and it is possible to go back to revise the
30
SUSTAINABLE DEVELOPMENT
analysis based on new information, new criteria or new options.
Step 6: Select the Most Appropriate Option The responsibility for selecting the most appropriate course of action belongs to the decisionmakers. Sustainable development favours democratic, political decision-making that is locally initiated and participatory. The role of engineers and other disciplines should be to provide a range of options which are designed and optimized to meet the objectives as efficiently as possible, to assess the options with respect to the criteria and objectives, and to present the resulting analyses to the decision-makers in such a form as to give a clear picture of the various aspects involved in the ultimate decision. To make value judgements at any point in this process which go beyond professional judgement is to transgress the professional's boundary of responsibility. Practitioners need to avoid making choices or value-based decisions which might in some way limit or skew the analysis on which the decision-maker relies when making the ultimate decision. The process does not end with the decision. As with EIA, there should be follow-up and monitoring during the execution of the project and throughout the life of the project. If the outcome or the impacts do not evolve as predicted, future action may be necessary, plans may need to be adapted in order to remain on course towards sustainability. The preceding methodology reflects most of the process-oriented principles of sustainable development, i.e., goal-seeking, relational, adaptive and interactive, defined by Gardner (1990). It places emphasis on overcoming some of the more common obstacles to sustainable development by integrating social and environmental criteria right from the beginning of the multi-disciplinary multistakeholder process, and by fostering communication. This methodology is discussed in Chapter 22, which deals with the development of a remedial action plan for a hypothetical case study.
2.6
SUMMARY AND CONCLUDING REMARKS
It is apparent that if sustainable development is to be achieved, and the requirements of the environmental impact assessment process satisfied, environmental and social considerations as well as technical and economic criteria must be considered from the very beginning of the planning process. This will require leadership on the part of the engineering profession, since it has traditionally dominated the planning process, as well as an evolution towards a more multidisciplinary approach to design that involves the input of biologists, sociologists, anthropologists, etc., at the earliest stages. Engineers should move swiftly to define a new ethic, a new role, new training, and new approaches to the profession. A successful approach to sustainable development must: (1) foster full integration of social and environmental factors along with technical and economic ones from the beginning of the planning process, (2) cross over disciplinary lines and facilitate multi-disciplinary cooperation, (3) produce the analysis and information required to make informed decisions, and (4) leave the subjective societal choices to the decision-makers. Sustainable development is a relatively young area of research. The idea of engineering for sustainable development is even more recent. Further research should include an examination of current engineering school curricula to ensure that students are exposed to the central issues,
SUMMARY AND CONCLUDING REMARKS
31
methodologies and thought processes of the humanities and social sciences, as well as course work dealing specifically with issues relating to engineering and sustainable development. A similar professional development course should be designed and made available to practicing engineers.
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CHAPTER THREE
SOURCES, CHARACTERISTICS, AND MANAGEMENT OF WASTES
3.1
INTRODUCTION
The perception of waste as a throw-away material and as a problem differs from one individual to another. Consumers, as collective and individual representatives of the public, have widely differing opinions and appreciation of waste. On the other hand, professionals such as engineers, lawyers, health scientists, and politicians tend to have highly focused views of waste material. Regardless of one's perception, the problem of waste management needs to be properly addressed, to avoid overloading the carrying capacity of the soil, and to avoid polluting the groundwater. The term waste management is used to describe several distinct processes: the elimination or reduction of waste, the recycling or reuse of waste material, the treatment or destruction of waste, i.e., physically destroying, chemically detoxifying or otherwise rendering the waste permanently harmless, and land disposal of wastes. Environmental regulations have not necessarily channeled industry efforts toward waste minimization -- the optimum approach to waste management. Rather, recycling, reuse, and treatment seem to be industry's preferred waste management options, even though such methods pose more environmental risks than waste minimization. Numerous laws designed to protect human health and the environment from hazardous wastes have been enacted over a period of many years. In order to adequately assess environmental risks, industry, government, and the general public must have a thorough understanding of these laws and regulations.
3.2
SOURCES OF WASTES
A clear definition of what constitutes a waste material is necessary in order to promulgate meaningful regulations. In many instances, through recycle and reuse, waste generated by one industry (or source) could constitute a resource material for another industry (user). Defining what constitutes a waste material becomes difficult when one attempts to distinguish among the different types of wastes. The different ways of categorizing waste materials are: (1) by the medium to which they are released, i.e., air, water or soil, (2) according to their physical characteristics, i.e., gaseous, liquid or solid, (3) type of hazard that they pose, i.e., ignitable, corrosive or reactive, (4) type of risk that they pose to human health, i.e., chronic effects such as carcinogenic, mutagenic and teratogenic, or acute effects on the nervous-, gastro-, or neuro-system, and (5) according to their origin, e.g., mine tailings, municipal waste or industrial waste. Although the first and second categories of classification can provide estimates of wastes 33
34
WASTE
generated, they overlook the primary processes that produce the waste. The primary processes are key factors for implementation of technology designed to obtain reduction of waste material at source. The third and fourth categories reflect the regulatory philosophy of governmental waste management programs. With such a mode of control, only the regulated (hazardous or radioactive) wastes are tracked or managed. Information on movement or management of unregulated wastes (primarily contributed by municipal solid waste) is difficult to obtain. Monitoring and proper management of unregulated wastes are critical since they could contain hazardous or toxic substances. Municipal solid waste is known to contain a whole series of hazardous household products such as paints, solvents, household cleaners, insecticides, pesticides, batteries and aerosols which, when land disposed, may produce a leachate containing various toxic elements and compounds. A reasonable way for classifying wastes is according to the sources of origin of the waste material or stream. The advantage of this method of classification is that it takes into account the processes that generate the wastes, and thus permits identification of areas where waste generation can be minimized or suitably recycled. More importantly, this method of classification permits one to track the source, such that proper and safe management and disposal of the wastes can be implemented. 3.2.1
Municipal Solid Waste
Municipal solid waste encompasses the heterogeneous mass of throw-away from residential and commercial sources, open areas and treatment plants. The materials comprising the municipal solid waste may be classified under broad categories of food wastes, rubbish, ashes and residues, demolition and construction wastes, treatment plant wastes, and miscellaneous items, as shown in Table 3.1 (Tchobanoglous et al., 1993). Based on the source, municipal waste is composed of (Tchobanoglous et al., 1993): (1) household and cleaning products (40%), (2) personal care products (16.4%), (3) automotive products (30.1%), (4) paint and related products (7.5%), (5) pesticides, insecticides, and herbicides (2.5%), and (6) others (3.5%). The composition of municipal solid waste can vary significantly due to differences in climate, social attitudes, and economic characteristics. The leachate produced by municipal solid waste is usually a highly complex liquid mixture of soluble, organic, inorganic, bacterial constituents and suspended colloidal solids in a principally aqueous medium. It contains products of degradation of the organic material and soluble ions which may pollute surface and groundwater. A typical composition of municipal solid waste leachates is shown in Table 3.2. (Naylor et al., 1978). The quantity of leachate produced depends on the composition of waste, dissolution process, and combined physical, chemical and biological activities. These unregulated hazardous wastes are exempt from any regulation (RCRA, 1976) simply because of the small quantities generated. However, the total amount of these wastes when disposed of at a municipal landfill may be a significant source of groundwater pollution.
SOURCES OF W A S T E S
35
Table 3.1: Sources of municipal solid wastes Types of solid waste
Source Residential
Food wastes, paper, cardboard, plastics, textiles, leather, glass, tin cans, aluminum, other metals, ashes, street leaves, special wastes (including bulky items, consumer electronics, batteries, oils, tires), household hazardous wastes.
Commercial
Paper, cardboard, plastics, wood, food waste, glass, metals, special wastes, hazardous wastes, etc.
Construction and demolition
Wood, steel, concrete, dirt, etc.
Municipal services
Special wastes, rubbish, street sweepings, landscape and tree trimmings, catch basin debris, general wastes from parks, beaches, and recreational areas.
Treatment plant sites
Treatment plant wastes, principally composed of residual sludge.
Table 3.2: Range of leachate composition in solid waste landfills Solid waste landfills Domestic landfills USA and U K ~ Michigan, USA 2 Croydon, UK 3 Variable (mg/1) (mg/1) (mg/1) Alkalinity (as CaCO3) BOD (5 days) COD TOC Ca Mg Na K Fe C1 SO4 Ammoniacial nitrogen Organic nitrogen TDS pH US EPA (1973);
2
Nil - 20,850 9 - 54,610 Nil - 89,520
7,000 - 7,800 13,800 - 16,800 46,600 - 52,400 9,250 - 10,300 2,230 a 727 a 1,440 a 680 a 1,510 a 1,325- 1,375 265 - 411 336 - 340 22 - 33 5,661 - 6,086 5.8 - 6.2
5 - 4,080 16.5 - 15,600 N i l - 7,700 2.8 - 3,770 0.2 - 5,500 3 4 - 2,800 1 - 1,826 Nil - 1,106 ..... Nil - 42,276 3.7 - 8.5
Kunkle and Shade (1976);
3
1,400 - 6,400 21 - 340
40 - 690 70 - 308 3 9 0 - 1,600 2 4 0 - 800 9 - 72 3 6 0 - 1,800 10 - 350 140 - 990 8 - 40 ..... 6.3 - 8.2
Davison (1969); ~ Average value.
36 3.2.2
WASTE Pesticide Wastes
Pesticides are plant or crop protection agents. They are designed to combat and control pests on cultivated plants and bacterial diseases. They typically include substances to control vegetational competitors (weeds) as well as fungal plant diseases. Classification of the main chemical classes of pesticides is shown in Table 3.3. The diverse nature of the pesticide industry and the wide distribution of the products make it difficult to analyse and assess the pollution impact of the specific active ingredients and their finished formulations. Pesticide wastes result from the periodic cleaning of equipments, spills, area washroom, drum washing, air pollution control devices, and area runoff. Wash waters and stream condensates from cleaning operations are the source of liquid wastes from the formulation lines and filling equipment. Steam cleaning condensates and rinse water from other processing units such as mix tanks, drum washers and air pollution control equipment are also sources of pesticide wastes. The scrubber waters themselves are a waste stream with area wash-down, leaks and spills making up the remaining principal sources. Many pesticide wastes are aqueous solutions or suspensions of organic compounds. Some pesticide wastes are generated in the production of dieldrin, methyl-parathion, dioxin, aldrin, chlorodane DDD, DDT, 2,4-D, endrin, guthion, heptachlor, and lindane. Inorganic-based wastes result from the production of arsenic, arsenate, and mercurial compounds.
Table 3.3: Classification of the main chemical classes of pesticides Pesticide Type
Class of Chemicals
Insecticides
Chlorinated hydrocarbons, phosphoric acid esters, carbamates, natural and synthetic pyrethroids, dinitrophenols, microbials (disease microorganisms and their metabolic products), botanicals.
Herbicides
Phenoxy aliphatic acids, substituted amides, nitroanilines, substituted ureas, carbamates, thiocarbamates, heterocyclic nitrogen, dinitrophenols.
Fungicides
Inorganic fungicides (based sulfur, copper or mercury), organic fungicides (e.g., dithiocarbamates), systemic fungicides (e.g., benzimidazoles), antibiotics ~roduced by microorganisms).
3.2.3
Mining Wastes
Most mining wastes are associated with procurement of host rock or soil material, and processes leading to the production of metals such as aluminum, iron, copper, gold, lead, molybdenum, silver, tungsten, uranium, zinc, or non-metals such as coal, asbestos, gypsum, barite, potash mineral, salt mineral, bitumen (from tar sands or shales), quartz, lime, sand, gravel and
SOURCES OF WASTES
37
stones. For example, without taking into account the tar sands industry, the Canadian mining industry is currently producing more than 500 million tonnes of solid waste a year in the metal, uranium, and coal industries. Waste water associated with mining and milling operations also constitutes a significant amount of waste in the form of liquid streams. Solid wastes generated from mining activities include the overburden material of waste rock, which typically consists of soil, debris and unconsolidated material, and waste rock which contains no ore or ore which cannot be economically removed. The milling process, i.e., concentration and preparation of the ore for subsequent stages of processing, generally removes unwanted ore constituents, alters the physical properties of the ore (e.g., particle size and moisture content) and increases the desired mineral concentration (i.e., ore upgrade). Solid wastes generated in the milling process consist of the unwanted minerals and other soil constituents, and are generally disposed off as a slurry (tailings) into mined-out pits or specifically designed tailings ponds. The tailings can be considered as a hazardous waste material because of their acidic and heavy metals content. Liquid wastes generated in a mining operation are generally the hydrologic drainage from the mining site due to percolation from waste rock and mine tailings, and surface runoff. The drainage water generally constitutes a pollution problem because of its acidic nature and high concentrations of heavy metals. This is known as acid mine drainage and poses a significant environmental impact which may persist even long after mine closure.
Table 3.4: Waste stream characteristics of electroplating industry Segment of Industry (max. cone. mg/1) Common Precious Pollutant metals metals Electroless Anodizing parameter plating plating plating Copper Nickel Chromium Zinc Cyanide Fluoride Cadmium Lead Iron Tin Phosphorus TSD Silver Gold Palladium Platinum Rhodium
272 2,954 526 252 150 142 22 25 1,482 103 144 10,000 --------. .
.
----. . . . . . . 10 --. . . . . . . . . . . . 144 10,000 176 25 2.2 6.5 . . . .
48 47 . . . .
. . . .
.
. 12 18 . . . . . . . . 109 40 . . . . . . . . . .
----79 .
.
. . . .
. . . . 33 924
. . . .
. . . . .
79
. 78 ---
. . . .
Coatings
. . . . .
. . . .
126
168 7 53 5,300
38
3.2.4
WASTE
Electroplating and Metal Finishing Industry
The electroplating industry can be classified into three principal segments: plating, metal finishing, and the manufacture o'f printed circuit boards. The plating segment can be further subdivided into common metal electroplating, precious metal electroplating and electroless plating. Sub-segments of the metal finishing category include: anodizing, chemical conversion coating, chemical milling, etching, and immersion plating. Because of the heavy metal content of most wastes from electroplating and metal finishing operations, many wastes from this industry may be hazardous; appropriate tests should be run to determine whether the waste liquids are hazardous. Waste stream characteristics of various pollutants found in each sub-segment of the electroplating industry are shown in Table 3.4 (US EPA, 1979). The concentrations shown are the maximum reported values.
3.2.5
Metal Smelting and Refining Industries
Production of iron and steel generates significant amounts of wastes, as shown in Table 3.5, that can be broadly categorized under: (a) iron oxide-containing wastes, (b) acid wastes, and (c) oilcontaining wastes. In iron ore production, pig iron is first extracted from the ore and then sent to a smelter to produce steel. In the pig iron production process, the dust generated is captured by wet dust cleaners. Smoke-vent dust is also generated and captured in the same manner as in the production of steel in open hearth furnaces. Steel finishing involves a number of necessary processes, designed to impart desirable surface and mechanical characteristics to the steel. These processes generate considerable amounts of liquid wastes. Water leaving the wet dust cleaners usually contains anywhere from 100 mg/1 to 10,000 mg/1 of suspended solids, depending on the furnace burden, furnace size, operating methods employed and type of gas-washing equipment. For older installations, gravity sedimentation in either circular or rectangular basins is the preferred method of gas-washer water treatment. In modern installations, clarifiers are used in place of gravity sedimentation ponds. The sludge from the settling basins or clarifiers is sent to a sintering plant while the effluent is sent for secondary treatment prior to discharge. The blast furnace fine dust and sludge contain trace elements such as Zn, Pb, S, Na, and K that are both undesirable and difficult to remove. The finer materials are often difficult to handle and assimilate in normal operations. Airborne particulate matters from the handling of these products become an important environmental problem. Also, the oil present in some of these wastes can vaporize and lead to operating difficulties with pollution control equipments. Pickling solutions represent one of the most environmentally hazardous by-products in steel production. Sulphuric acid represents about 90% of the acid used in pickling steel. Pickling solutions contain free acids, ferrous sulfate, undissolved scale and dust, and the various inhibitors and wetting agents, as well as dissolved trace elements. Spent sulphate pickling solutions are discharged at temperatures of 170 to 190 ~ and can amount to 100 - 1000 gal/day in a large mill. The concentration of acids and their types largely depend on the type of iron being produced. Finishing operations following pickling processes generate effluents which contain rolling oils, lubricants and hydraulic oils. The oils are both in free and emulsified states. Soluble oils which come from cold reduction mills, electrolytic tin lines, and a variety of machine shop operations, can
SOURCES OF WASTES
39
also be present in the effluents. The effluents from a typical mill contain about 200 mg/1 of oil, of which 25% is stable emulsion. The oil bearing waters represent an environmental problem when discharged in receiving waters. For example, emulsified oil can contribute to BOD. Free oil is objectionable in streams due to its wide-spread surface films.
Table 3.5: Potentially hazardous wastes generated by the metal smelting and refining industry. Product or Activity
Waste Stream
Hazardous Constituents
Coking
Ammonia still lime sludge
Oil and grease, cyanide, naphthalene, phenolic compounds, arsenic, heavy metals. Oil and grease, phenol, naphthalene, pyrites, polyaromatics, nitrogen, heterocycles, heavy metals. Cr, Pb, Cd. Cr, Pb, high pH. Cr. Cr. Cr, Pb. Cr, Pb, Mn. Cr, Pb. As, Pb, Cd, Cu, Se, Zn. As, Pb, Cd, Hg. As, Cd, Se, Zn. Pb, Cd, Zn. Pb, Cd, Zn. Cu, Pb, CN, F.
Decanter tank tar sludge
Steel electric furnace Steel finishing Ferro-chromium-silicon Ferro-chrome Ferro-manganese Gray and ductile iron Copper smelting Lead smelter Zinc smelter
Aluminum smelting
Emission control dusts Spent pickle liquor Sludge from lime treatment Emission control dust or sludge Emission control dust or sludge Emission control dust or sludge Emission control dust or sludge Acid blow-down slurry Surface impoundment solids Waste water treatment sludge Electrolytic anode slimes Cadmium plant leach Spent cathodes
Non-ferrous metals (copper, zinc, lead and nickel) production records show that, by and large, production of nickel and lead far exceed the production of the other non-ferrous metals. Wastes associated with smelting and refining of non-ferrous metals, however, are largely produced at copper and zinc facilities. Copper smelting and refining process generates smelter wastes. High copper concentrations are generated from the acid plant blow-down thickener (up to 380,000 mg/1) and from the electrostatic precipitator dust (240,000 to 280,000 mg/1). Available data also show that the lead and zinc concentration in the acid plant sludge and the electrostatic precipitator dust can exceed 10,000 mg/1. The slag produced in primary and secondary copper recovery processes will generate leachates which can contain low concentrations of metals. Higher concentrations of metals, however, can be leached out of the fine dust.
40
WASTE
Lead is mined principally as a co-product of zinc or as a by-product of poly-metallic ores. The processes involved in lead metal production include sintering, charring in a blast furnace, and further refining to remove impurities. The primary wastes generated by lead smelting include blast furnace slag, solids in the slag-granulation slurry, and wastes associated with scrubbing of SO2 gases. Typical kinds of constituents found in lead smelting wastes include metals, with apparent high concentrations of Cd, Cu, Pb, and Zn in almost all the slags and sludge. 3.2.6
Pulp and Paper Wastes
Wood is by far the most important raw material for the production of pulp and paper. Its main components are cellulose (--20%), lignin (-25%) and extra-actives (-5%). In the various steps in the production of paper, each of the processes generates effluent, sludge and atmospheric emissions. The various steps are generally classed as wood preparation, pulping, bleaching and paper making. Bark material constitutes a source of wastes. It is an essentially non-fibrous material which decreases the quality of the wood and is therefore an unwanted layer. It is usually removed by the method of wet or dry barking. The effluent properties of the two methods show that at least twice the amount of suspended solids is generated by the wet process, for the same amount of output flow. However, since the dry barking process obviously generates significantly less effluents for the same amount of product output, the suspended solids in the effluent discharge from dry barking is about 4 times less than that produced in the wet barking process, for the equivalent product output. The results show that the BOD measured in the wet process tends to be about one order of magnitude larger. The general pH values for the effluents vary around 5.5, with the wet process effluents showing slightly higher values. Historically, bark was buried or dumped into the nearest river. This obviously created considerable environmental concerns and problems. Presently, bark is incinerated in an energy recovery system for steam production. However, combustion of bark polluted with chlorophenols or other pesticides represents an important source of polychlorinated dibenzopdioxins (PCDD) and polychlorinated dibenzofurans (PCDF) rejects in the atmosphere. The two main types of pulping are chemical and mechanical in nature. It is the chemical pulping operations that generate the hazardous waste streams through the use of chemicals to separate the fibers from the lignin in the wood. The kraft of sulphate pulping process generates sludge high in chromium, lead, and sodium. Fortunately, a large proportion of the plants using this process recycle many of their wastes. Pulp processing and cleaning to remove dirt and foreign matter such as silvers, knots, grit, bark, sand, uncooked chips etc., from the pulp, is usually accomplished by (a) coarse screening, (b) fine screening, and (c) centrifugal cleaning operations. Fiber rejects from kraft mills, including screening and cleaning rejects, are discharged into the sewer or de-watered for landfill or incineration. Lignin derivatives are the major color constituents in paper and have to be removed to produce the desired whiteness through chlorine or oxygen bleaching of the pulp. The bleach plant effluent BOD constitutes the principal source of toxicity in most bleached kraft mills. More importantly, the principal source of toxicity in chlorine-based bleaching process is chlorine itself. This problem can be solved by replacing chlorine with chlorine dioxide.
SOURCES OF WASTES Table 3.6: Chemical analysis of primary and secondary treatment sludge from the pulp and paper industry. Primary and secondary De-inking sludge Pretreatment Board Constituent sludge from semisludge #f sludge #2 sludge from mill ( mg/l) chemical pulping Paper coating sludge Water (%) Solids ('YO) Ash (%) COD Phenol PCB Oil and grease Total nitrogen Aluminum Cadmium Calcium Chloride Chromium Copper Iron Lead Magnesium Manganese Nickel Phosphorous Potassium Sodium Sulfate Zinc
90 - 96 4 - 10 1 - 2.5 60,000 - 120,000 5 < 13 1 1,400
---
1.5 4,000 - 15,000
---------
120
---
250 25
---
1,600 1,400 120 260
77 22.4
40 60 40
86 14
-----
Priman & secondary sludge
---
400,O 00
---
----_-_ 21,300 32 4,390 332 20
---
538 32 1,170 16 2.3 310 114 146 0.03 151
--180 330 1,500 1,300
79 62 2,400 380
--47
42
WASTE
It is useful to note that in many mills, concern for the environment has generated the use of oxygen in lieu of chlorine, resulting in the reduction of about one-half the BOD and COD contents of the effluents. Moreover, installation of oxygen-bleaching sequence decreases the chlorine content of the waste water. The paper making process contributes primarily to the effluent suspended solids and fow, with very little contribution to BOD or toxicity. In a news print mill effluent, the suspended solids typically vary from 1 to 50%, while the BOD is generally less than 5%. Typical waste characteristics from the pulp and paper industry are shown in Table 3.6 (US EPA, 1979).
3.2.7
Petroleum Refining Wastes
Transforming crude oil into petroleum products involves processes that can be divided into four major categories: separation, conversion, treatment and blending. The separation process involves the physical separation of different classes of molecules. Separation, for example, removes inorganic salts from crude oil, separates the products of each process and removes more impurities from the products. Separation is achieved by atmospheric distillation and vacuum distillation. The conversion process includes processes such as cruding, reforming and alkylation in which the natural fractions of the crude are used as raw materials to produce new marketable types. Conversion changes some less desirable hydrocarbons into more marketable types and is achieved by cracking processes. There are four cracking methods, namely, catalytic cracking, hydro-cracking, vis-breaking and coking. After primary separation, most crude fractions contain some impurities. Treatment processes are used to increase the number of products that can be made from heavy distribution fractions or reduced crude, as well as to improve the quality of many products. The method of treatment can be divided into hydro-treating, chemical treating, gas treating and treatment by physical means. Blending is the final step in the production of finished petroleum products to meet quality specifications and market demand. The blending operation involves accurate proportioning of the base stocks along with proper mixing to produce a homogenous product. Improvement of the properties of the products is achieved by the use of additives. For example, tetraethyl lead is added to gasoline to increase the octane number. Anti-oxidants, anti-icing agents and metal de-activators are used to inhibit gum formation. The major use of water in petroleum refining is for steam generation and heat transfer. Raw refinery waste contains large quantities of free as well as emulsified oil. Water soluble hydrocarbons (e.g., phenolic compounds) present in petroleum also occur in the waste water. Trace heavy metals can be found in waste water depending on the type and quantity of corrosion inhibitor used in noncontact cooling water circuit. On the other hand, the sources of solid wastes are highly variable. The possible sources of solid wastes show that the types of wastes vary from process solid wastes emanating from storage procedures to general waste obtained in regular plant maintenance. Table 3.7 (Stewart, 1978) shows some typical concentrations and solid waste sources. It is significant to note that almost all are classified as hazardous solid waste.
SOURCES OF WASTES Table 3.7: Ranges of concentrations and total quantities for some refinery solid waste sources Cooling Silt from Leaded Non-leaded Crude tower Coke storm water tank product tank tank Parameter sludge fines runoff bottoms bottoms bottoms (mg/l) Phenols Cyanides Selenium Arsenic Mercury Beryllium Vanadium Chromium Cobalt Nickel Copper Zinc Silver Cadmium Lead Molybdenum Ammonium salts Ben-a-pyrene Oil (wt. %)
0.6 - 7.0 0 - 14 0 - 2.4 0.7 - 21 0 - 0.1 trace 0.12 - 42 181 - 1,750 0.38 - 7 0.25 - 50 49 - 363 118 - 1,100 0.01 - 1.6 0.06 - 0.6 1.2 - 89 0.25 - 2.5 0.07 - 14 0 - 0.8 0.07 - 4
Total weight (tondyear)
0.1 - 0.13
Spent line from boiler feed water treatment
trace 0 - 1.3
6.3 - 13.3 0.48 - 0.95 1.1 - 2.2 1.0 - 10 0.23 - 0.36 trace 2 5 - 112 32.5 - 644 11.0- 11.3 30 - 129 14.8 - 41.8 60 - 396 0.4 - 0.6 0.1 - 0.4 20.5 - 86 6.3 - 7.5 1.o 0.03 - 2.5 2.2 - 5.5
2.1 - 250 1.7 - 1.8 trace 0 - 14.7 1.5 - 22.4 0.1 - 3.1 63 - 455 trace 0.1 1 - 0.94 0.41 - 0.04 trace 0.03 - 0.49 1.0 - 9.8 9.1 - 34.6 9.0 - 13.7 12.7 - 13.1 26.5 - 71 5.9 - 8.2 235 - 392 12.4 - 41 110 - 172 6.2 - 164 1190 - 17,000 29.7 - 541 0.05 - 1.7 0.5 - 0.7 4.5 - 8.1 0.25 - 0.4 12.1 - 37.3 158 - 1100 0.5 - 118 0.25 - 18.2 ..... 0.2 0.02 - 0.4 0.3 - 0.9 18.9 - 21 45.1 - 83.2
6.1 - 37.8 0.01 - 0.04 5.8 - 53 5.8 - 53 0.07 - 1.53 trace 0.5 - 62 1.9 - 75 3.8 - 37 12.8 - 125 18.5 - 194 22.8 - 425 0.03 - 1.3 0.025 - 0.42 10.9 - 258 0.03 - 95 2.0 0 - 0.6 21 - 83.6
0.05 - 3.6 0 - 1.28 0.01 - 9.2 0.01 - 2.3 0 - 0.5 trace 0 - 31.6 0.03 - 27.9 0 - 1.3 0.13 - 26.2 0.22 - 63.2 2.0 - 70 0.05 - 0.7 0 - 1.3 0.01 - 7.3 0 -0.05 trace trace 0.04 - 0.5
0.06 - 4.2
2.7
0.2 - 1.3
0.1 - 0.26
28.5 - 214.7
0.4 - 2.7 trace 0.01 - 1.6 0.2 - 10.8 0 - 0.2 0 - 0.2 400 - 3,500 0.02 - 7.5 0.2 - 9.2 350 - 2,200 3.5 - 5 0.2 - 20 0.01 - 3 0.02 - 2 0.5 - 29 0.1 - 2.5
.....
34.7 - 77
44 3.2.8
WASTE Paint and Allied Industries
The paint and allied industries utilize many organic and inorganic raw materials, some of which are present in the wastes. There is no waste stream in the sense of wastes and by-products of production. The wastes come mainly from the packaging of raw materials, air and water pollution control equipment, off-grade products and spills, most of which is reclaimed and reused except for paints absorbed on the final cleanup material. Coatings containing significant amounts of toxic metals are reworked, and the wastes contain little or no metallic residues. In the formulation of paints and coatings, a number of metal compounds are used as pigments; oil and polymer resins are used as bases and solvents are used as thinners. These ingredients become part of the wastes as spoiled batch or spills. Such waste constitutes about 6% of production. Toxic chemicals usage is restricted, and thus a proportionally small amount of toxic substances (mainly mercury and lead) reach the waste stream from this source. Potentially hazardous materials in paints include: inorganic metals such as arsenic, beryllium, cadmium, chromium, copper, cobalt, lead, mercury, selenium, asbestos and cyanides, and organic compounds such as halogenated hydrocarbons, and pesticides.
3.3
HAZARDOUS WASTE
Defining what constitutes a hazardous waste is a complex process since many factors are potential contributors to the hazardous nature of a waste, and what can be viewed as hazardous to one individual may not be to another. For instance, the unsafe handling and transportation of waste products or substances can result in their becoming hazardous. A waste may have a low degree of toxicity and yet highly flammable, explosive, or reactive. Wastes which have a tendency to bioaccumulate are more likely to produce undesirable chronic effects than highly flammable or explosive materials; therefore an assessment of the acute versus chronic toxicity of waste is also needed in any attempt to define hazardous waste. 3.3.1
United States of America
Over the years, numerous attempts have been made by government, industry, and the general public to define hazardous waste. In the Resource Conservation and Recovery Act (RCRA) of 1976, US Congress defined the term hazardous waste as: "a waste, or combination o f wastes, which because o f its quantity, concentration, or physical, chemical, or infectious characteristics may (1) cause, or significantly contribute to, an increase in mortality or an increase in serious irreversible or incapacitating reversible illness or (2) pose a substantial present or potential hazard to human health or the environment when improperly treated, stored, transported, or disposed o f " Although the US Congress defined the term hazardous waste in RCRA, the Environmental Protection Agency (EPA) was left to develop the regulatory framework that would identify those wastes that must be managed as hazardous wastes under Subtitle C. In Code 40 of Federal
HAZARDOUS WASTE
45
Regulations Part 261, EPA then specified that a solid waste is hazardous if it meets any of the four conditions: (1) exhibits any of the characteristics of a hazardous waste based upon analysis, (2) has been listed as a hazardous waste, (3) is a mixture containing a characteristic or listed hazardous waste and a nonhazardous waste, unless the mixture is specifically excluded or no longer exhibits any of the hazardous waste characteristics, and (4) is not specifically excluded from regulation as a hazardous waste. Furthermore, the by-products of the treatment of any hazardous waste are also to be considered hazardous unless they have been specifically excluded. In establishing these criteria for the identification of a hazardous waste, EPA selected four characteristics as inherently hazardous in any substance: (1) ignitability, (2) corrosivity, (3) reactivity, and (4) toxicity. In selecting these characteristics, EPA used two criteria. The first criterion was that the characteristic be capable of being defined in terms of physical, chemical, or other properties (Verschueren, 1983; Weast, 1975). The second criterion was that the properties defining the characteristic be measurable by standardized and available testing protocols. The second criterion was adopted because the primary responsibility for determining whether a solid waste exhibits any of the characteristics rests with the regulators. EPA believed that unless generators were provided with widely available and uncomplicated methods for determining whether their wastes exhibit the characteristics, the identification system would prove unworkable. For this reason EPA refrained from adding carcinogenicity, bioaccumulation potential, and phototoxicity to the characteristics.
IGNITABILITY A waste exhibits the characteristics of ignitability if a representative sample of the waste has any of the following properties: (1) it is a liquid, other than an aqueous solution containing less than 24% alcohol by volume, and has flash point less than 60~ as determined by ASTM Standards D93-79 or D-93-80 or D-3278-78, (2) it is not a liquid and is capable, under standard temperature and pressure, of causing fire through friction, absorption of moisture, or spontaneous chemical changes and, when ignited, bums so vigorously and persistently that it creates a hazard, (3) it is an ignitable compressed gas as defined in Code 49 of Federal Regulation, and (4) it is an oxidizer as defined in Code 49 of Federal Regulation. Ignitability is the characteristic used to define as hazardous those wastes that could cause a fire during transport, storage, or disposal. Examples of ignitable wastes include waste oils and used solvents.
C ORR OSIVITY A waste exhibits the characteristic of corrosivity if a representative sample of the waste has either of the following properties: (1) It is aqueous and has a pH less than or equal to 2 or greater than or equal to 12.5, as determined by a pH meter using EPA test method, (2) it is a liquid and corrodes steel at a rate greater than 6.35 mm per year at a test temperature of 55 ~ C, as determined by the test method specified in NACE (National Association of Corrosion Engineers) Standard TM01-69. Corrosivity, as indicated by pH, was chosen as an identifying characteristic of a hazardous waste because wastes with high or low pH can react dangerously with other wastes or cause toxic pollutants to migrate from certain wastes. Examples of corrosive wastes include acidic wastes and used pickle liquor from steel manufacture. Steel corrosion is a prime indicator of a hazardous waste since wastes capable of corroding steel can escape from drums and liberate other wastes.
46
WASTE
Table 3.8: Maximum concentration of pollutants for RCRA toxicity characteristics Maximum Maximum Pollutant concentration Pollutant concentration (m~;/1) (mg/1) Arsenic Barium Benzene Cadmium Carbon tetrachloride Chlordane Chlorobenzene Chloro form Chromium o-Cresol m-Cresol p-Cresol 2,4-D 1,4-Dichlorobenzene
1,2-Dichloroethane 1,1 -Dichloroethylene 2,4- Dinitrotoluene Endrin Heptachlor Hexachlorobenzene
5.0 100.0 0.5 1.0 0.5 0.03 100.0 6.0 5.0 200.0 200.0 200.0 10.0 7.5 0.5 0.7 0.13 0.02 0.008 0.13
Hexachloro- 1,3-butadiene Hexachloroethane Lead Lidane Mercury Methoxychlor Methyl ethyl ketone Nitrobenzene Pentachlorophenol Pyridine Selenium Silver Tetrachloroethylene Toxaphene Trichloroethylene 2,4,5-trichlorophenol 2,4,5-TP (Silvex) Vinyl chloride
0.5 3.0 5.0 0.4 0.2 10.0 200.0 2.0 100.0 5.0 1.0 5.0 0.7 0.5 0.5 2.0 1.0 0.2
REA CTIVITY A waste exhibits the characteristic of reactivity if a representative sample of the waste has any of the following properties: (1) it is normally unstable and readily undergoes violent change without detonating, (2) it reacts violently with water, (3) it forms potentially explosive mixtures with water, (4) when mixed with water, it generates toxic gases, vapours, or fumes in a quantity sufficient to present a danger to human health or the environment, (5) it is a cyanide- or sulfide-bearing waste which, when exposed to pH conditions between 2 and 12.5, can generate toxic gases, vapours, or fumes in a quantity sufficient to present a danger to human health or the environment, (6) it is capable of detonation or explosive reaction if subjected to a strong initiating source or if heated under confinement, (7) it is readily capable of detonation or explosive decomposition or reaction at standard temperature and pressure, and (8) it is a forbidden explosive as defined in Code 49 of Federal Regulation 173.51. Reactivity was chosen as an identifying characteristic of a hazardous waste because unstable wastes can pose an explosive problem at any stage of the waste management cycle. Examples of reactive wastes include water from TNT operations and used cyanide solvents.
HAZARDOUS WASTE
47
TOXICITY The Toxicity Characteristic Leaching Procedure (TCLP) is designed to identify wastes likely to leach hazardous concentrations of particular toxic constituents into the groundwater as a result of improper management. During the TCLP test, constituents are extracted from the waste to simulate the leaching actions that occur in landfills. If the concentration of the toxic constituent exceeds the regulatory limit, the waste is classified as hazardous. If the extract from a representative waste sample contains any of the pollutants listed in Table 3.8 at a concentration equal to or greater than the respective value given, the waste exhibits the toxicity characteristic. Where the waste contains less than 0.5% filterable solids, the waste itself is considered to be the extract. 3.3.2
Canada
Definitions of hazardous material and wastes are found in two Federal regulations, the Canadian Environmental Protection Act (CEPA) and the Transportation of Dangerous Goods Act and Regulations (TDGA). The TDGA defines dangerous goods as: "any product, substance or organism included by its nature or by regulations in any of the classes listed in the Schedule." The Schedule contains a list of 9 primary classes of dangerous goods which are: explosives, gases, flammable liquids, flammable solids, oxidizing substances -- organic peroxides, poisonous (toxic) and infectious substances, radioactive materials, corrosives, and miscellaneous products. Details of these classes can be found in CEPA and TDGA, and will not be repeated here. Based on the Canadian Toxic Chemical Management Program (TCMP), toxic chemicals are defined as: "those substances which, when released into the environment, or thereafter if chemically transformed through combination or otherwise, could pose a significant threat to natural ecosystem or to human health or well-being." Chemical substances under this definition have the following characteristics: (1) the ability to become widely dispersed in air, land and water, (2) the capability of causing biological changes at trace concentrations, (3) the ability to become more toxic when combined in the environment with other chemicals, and (4) the ability to become irreversible once released into the environment, with effects that are largely irreversible.
3.3.3 European Community According to the Council Directive of 1991, the European Community (EC) defines hazardous wastes as: "those wastes which possess one or more of the following properties: explosive; oxidizing," flammable; irritant," harmful; toxic or posing a serious health threat if inhaled, ingested or coming into contact with the skin; carcinogenic; corrosive; infectious; teratogenic; mutagenic; substances which release toxic gases when
48
WASTE coming into contact with water, air or an acid," substances which may yield, after disposal another substance which is hazardous; ecotoxic or presenting a danger to the environment, either immediately or with time."
3.3.4
Germany In the German Chemicals Act (ChemG, 1990), hazardous waste is defined as: "substances and compounds or their derivatives which are inclined to alter the quality o f the natural balance o f the water, soil or air, climate, animals, plants, or microorganisms in such a way that they may cause risks for the environment, either immediately or at some later time."
In this definition, the term natural balance, which can be described as "the nexus of relationships amongst the ecosystems and with their environment", is especially problematic. In the German Plant Protection Law, it is defined as "soil, water, air, animal and plant species as well as the interaction between them." The term cannot be clearly defined in scientific terms. Environment, on the other hand, is understood to mean the totality of all factors (both abiotic and biotic) which have a direct, active effect on the extemal world.
3.3.5
The Netherlands
The definition of hazardous wastes in the Netherlands is based on the EC definition and states that: "hazardous wastes are defined as toxic and dangerous wastes as described in EEC directives 78/319/EEC pB L 84/83 and PCB's as described in EEC directives 76/403/EEC pB L 108/76."
In addition, the Netherlands has recognized the fact that some new substances may prove to be hazardous and will have to be included in the list of hazardous waste. Consequently, the Dutch authorities have added to the original definition the following statement "Hazardous wastes are chemical wastes and other substances designated by the Minister." Hazardous waste is therefore regarded as chemical waste. However, the list of toxic and hazardous wastes is somewhat broader than the list of chemical wastes.
3.4
HAZARDOUS WASTE MANAGEMENT
3.4.1
US Regulation Disposal Philosophy
The RCRA legislation is an incremental step in the control of hazardous waste since it defines and regulates the wastes. The legislation was criticized for focusing on the new and existing generators, and transportation of hazardous waste but failing to address the need for control of abandoned sites. As a result of that criticism, the 1980 Comprehensive Environmental Response,
HAZARDOUS WASTE MANAGEMENT
49
Compensation and Liabilities Act (CERCLA) was enacted for abandoned sites. The goal of the RCRA was to substantially eliminate the damage incidents occurring under the past state-of-the-art technology and management system. The goals were to investigate newer and standardized technology in an upgraded management system. These goals would be reached utilizing the pathway approach or cradle-to-grave management of hazardous waste cycle from generation to treatment, storage and disposal. The underlying philosophy of the approach was to: (a) create a system to develop adequate treatment and disposal capacity for hazardous waste, (b) minimize the cost to society, (c) protect humans and the environment, (d) recover materials and energy, and (e) assign costs to the generators. US EPA prioritized the waste management (disposal) options under RCRA as follows: (1) waste reduction, (2) waste separation and concentration, (3) waste exchange, (4) energy and material recovery, (5) waste incineration and/or treatment, (6) secure storage and/or disposal. The Act relies on market mechanism to realize the disposal goals since waste management strategies all imply cost savings. EPA and the complimentary State Authorities are responsible for identifying, defining, and enforcing the management system designed to protect the integrity of air, land and water pathways. The generators, transporters and handlers of hazardous waste are allowed as much freedom as possible in choosing options that yield compliance.
STANDARDS FOR GENERATORS Generators are required to evaluate their wastes to determine if they are hazardous under RCRA. If the wastes are found to be hazardous, they are subjected to a variety of reporting, recordkeeping and labeling requirements. The generator must also comply with the hazardous waste manifest system that tracks the substances shipped off-site for treatment, storage or disposal. It is the generators who are responsible for determining that the waste subsequently arrives at the hazardous waste facility designated and that it is authorized to handle the waste. STANDARDS FOR TRANSPORTERS A transporter must have an EPA identification number and comply with the hazardous waste manifest system for the tracking of hazardous waste shipments. The transporter must comply with the Department of Transportation's hazardous materials transportation regulation on labeling, marketing, packaging and placarding (all of which have been incorporated into the RCRA regulation). REQUIREMENTS FOR TREATMENTS, STORAGE AND DISPOSAL ~SD) FACILITIES Facilities must have either an interim status or an RCRA permit. Interim status is granted to qualifying facilities pending the issuance of a RCRA permit. During that interim status the facility must comply with Phase I standards in Part 265 of the regulations. These regulations are not as rigid or elaborate as the full permit regulations. The interim status regulation emphasizes reporting, record-keeping and maintenance as well as financial responsibilities rather than technical or facility design standards. Interim status TSD facilities may submit a Part B RCRA permit application for a full permit or wait until requested to do so. Development and completion of a Part B RCRA is expensive and time consuming. As a result of the effort put into the applications, significant time is required by EPA to process and review them. The result is that for a new facility, the full permit applications must be made before the facility is built.
50
WASTE
The Part B (Part 264) standards are particularly more stringent for land disposal than the interim equivalent. Specific design and/or performance standards are also set for most types of facilities. For example, groundwater monitoring requirements are also set for land disposal facilities. Once issued a permit, the TSD facility must comply with the Phase II standards in Part 264. As a final note, TSD facilities must comply with various reporting, record-keeping, inspection, closure, post-closure care and financial requirements outlined in the standards.
WASTE MINIMIZA TION Even without incentives or Federal programs, the economics of waste minimization is becoming attractive to hazardous waste producers. With the enactment of the ban on land disposal of untreated waste, the cost of land disposal of waste has dramatically increased. Also with the ban, the amount of handlers per shipment has increased and with the increase comes additional risk of accident. Accidents prior to the receipt of the waste by the ultimate disposal facility are the responsibility of the producer to cleanup. The increase in paper work alone is a lever to reduce the amount of waste generated. That paper work takes the form of transfer and tracking of documents as well as generation and inspection of reports, and contingency plans. The contingency plans, which are a requirement under Title III of the Superfund Amendment, are time-consuming and laborious. Public disclosure has had an effect on waste generation since the required reporting of spills and the sharing of plant operation involving hazardous materials have encouraged plants to stop using the chemicals. The Pollution Prevention Act (PPA) of 1991 builds on the RCRA requirements for hazardous waste generators and owners of TSD facilities to verify on a biannual manifest the steps they have taken to reduce the volume of toxic waste generated. Most of the other EPA regulations on waste minimization has taken the form of guidance to help the producers reduce the production of hazardous waste or inquiries to the producers on what they need to help them reduce waste generation. An Audit performed to gather information is required to make an informed judgment on waste minimization and is the first step to reaching the reduction goal. RECYCLING AND RECLAMATION The Waste Minimization Opportunity Assessment Manual distributed by EPA emphasizes, as does the RCRA waste management priorities, that reduction is preferable to waste recycle or reclamation but at the same time supports recycling and reclamation options. The guidelines suggest that early in the assessment program, waste reduction can be accomplished through simple means. Later methods identified by the guideline require large expenditures and should be evaluated with a feasibility study. A waste reduction feasibility study includes a technical evaluation of the procedure contemplated (i.e., pilot tests) as well as an economic evaluation. The economic evaluation is suggested to be carried out using standard methods of profitability including payback for equipment, retum on investment and net present worth. EPA, under RCRA provisions, regulates recycling activities, especially those that appear to be closer to disposal than production (e.g., placing the waste on land as a fertilizer). A list of recycled materials that remain hazardous is included in the RCRA regulations (e.g., spent solvents, sludge from waste treatment residues). In order to qualify as a non-hazardous recycled material, the waste must undergo a chemical reaction in the production process that makes it inseparable from another
HAZARDOUS WASTE MANAGEMENT
51
substance. HAZARDO US WASTE PRETREA TMENT Land disposal is the least favoured method of ultimate disposal of hazardous waste (1984 amendment). EPA has reinforced this position by imposing a land disposal ban for all untreated hazardous waste (1990 Amendment). As with most legislations there are exceptions. The main exceptions discussed here are with materials that became subject to the general pretreatment landfill ban regulations after June 1990. In order to be exempt, the waste must qualify according to one of the following: (1) the disposer must be able to prove that waste will not migrate from disposal site for as long as it is hazardous, (2) a special exemption existed until August 1990 for underground injection waste. Some wastes that qualified for the injection exemption are petroleum wastes, and steel pickle liquid, (3) an exemption existed for wastes contaminated with soil and debris which have a treatment standard based on incineration until June 1989, (4) some prohibited landfill wastes can be treated by surface impoundment. The regulation demands that surface impoundment facility should comply with minimum design specifications. Above all, in these regulations, it is reinforced that surface impoundment is for the purposes of treatment rather than storage. Exemptions were given to specific generators of hazardous waste such as: (1) farmers disposing of pesticides on their land, (2) small quantity generators. To qualify for the exemption, the generator had to produce less than 100 kg of non-acute hazardous wastes or less than 1 kg of acute hazardous wastes, and (3) grand-fathered generators. Those generating wastes identified as hazardous after November 8, 1984 for which EPA has not issued land disposal prohibitions or treatment standards. The regulation control for land disposal of waste is based on minimum treatment standards. That is, the waste is treated to a minimum level before land disposal and the successfulness of that treatment is regulated. In this manner, the quality control of the ultimate disposal system (i.e., the landfill) is placed on the incoming waste rather than the processes within the landfill. Minimum treatment standards for the wastes are either concentration or technology-based. In order to further define types of waste in these two classes of wastes, EPA refers to dilute solutions (probably high in volume) as waste waters and more concentrated wastes as non-wastewaters. Wastewaters are formally defined as having less than 1% total suspended solids (TSS) and less than 1% total organic carbon (TOC). Non-wastewaters are defined as wastes that do not meet the wastewater requirements. The concentration-based wastes treatment standards are predicted utilizing the best available technology. The reference to best available technology allows substitution for an EPA recommended treatment technology with any other treatment except dilution. The published maximum concentration tables fall into two categories: Constituent Concentration on Waste Extract (CCWE) tables and Treatment Standard Expressed as Waste Concentrations (TSWC) tables. The CCWE are to be used to determine if the waste can be disposed of without pretreatment. CCWE standards require that an extract, treatment residue or a concentration of the waste using Toxicity Characteristic Leaching Procedure (TCLP) be below the listed values. The waste may be landfilled after meeting the requirements or after the listed treatment or an EPA-sanctioned alternate is performed. TSWC standards identify the restricted wastes and their hazardous constituents that may not be exceeded by the waste or a treatment residual (not an extract) if that waste is to be disposed of on land (Jessup, 1992). There are some wastes that are prohibited from land disposal no matter what concentration or treatment. For the situation of a combined waste, the waste component
52
WASTE
with the most stringent treatment standard is the pretreatment standard requirement for the entire combined wastes. HAZARDOUS WASTE LANDFILLS
As discussed previously, a landfill can no longer be the final resting place of an untreated hazardous waste. The most important environmental consideration of a hazardous waste landfill is for the neighbouring waters, both underlying groundwater and adjacent surface waters. In order to minimize the effect of the landfill of treated hazardous waste on the environment, strict requirements on the testing and interim storage of waste are required. Interim storage is required since the treated hazardous waste must be stored until the results of the TCLP testing on the treated waste are available. The global requirement is that the waste must be sheltered from the weather and sampled both before and after treatment. Groundwater permit standard for hazardous waste facilities (40 CFR 264) apply to all land disposal-based treatment systems (i.e., surface impoundments, waste piles, land treatment units, and landfills) that accept hazardous waste. The groundwater monitoring program, as a result of the 40 CFR 264 standard, requires the drilling and monitoring of a number of wells with the intent to: (a) collect water samples at varying depths to accurately represent the uppermost aquifer, (b) sample the unpolluted background water, (c) sample the quality of the water passing the compliance requirements, and (d) detect the migration of pollutants in the uppermost aquifer as well as net migration from the facility when in compliance. The allowable hazardous waste constituents for a landfill are written into the facility's permit. The normally quoted levels do not exceed the background levels set for 14 drinking water constituents set under the Safe Drinking Water Act. For new facilities, 40 CFR 264 outlines the following main criteria:
Figure 3.1. Schematic diagram of composite liner system for surface impoundment for disposal of hazardous waste (US EPA, 1989).
HAZARDOUS WASTE MANAGEMENT
(1) (2)
(a)
53
Location Standards Seismic Considerations: hazardous waste treatment, storage, and disposal facilities must not be located within 60 m of a Holocene fault. Flood Plains: A hazardous waste management facility located within a 100-year flood plain must be designed, constructed, operated, and maintained to prevent washout of any hazardous waste by a 100-year flood. Salt Dome Formation, Salt Bed Formation, Underground Mines, and Caves: The placement of any non-containerized or bulk liquid hazardous waste in any of these formations is prohibited. These requirements may have some exceptions, as outlined in 40 CFR 264.
Groundwater Protection Standard Facilities must be designed to ensure that hazardous constituents specified under 40 CFR 264 entering the groundwater from a regulated unit do not exceed the concentration limits under 40 CFR 264 in the uppermost aquifer underlying the waste management area beyond the point of compliance. Design Criteria A surface impoundment unit used for disposal of hazardous waste must have two or more liners and a leachate collection and removal system between such liners. The liner system must include a top liner made of geomembrane. The bottom liner must be a composite system, as shown in Figure 3.1. The requirements for a hazardous waste landfill are similar to those for a surface impoundment. The system must have a leachate collection system, a top flexible membrane liner (FML), a leachate collection, detection and removal system, and a composite liner system, as shown in Figure 3.2.
Figure 3.2. Schematic of a double liner and leachate collection system for a hazardous waste landfill (US EPA, 1989).
54
WASTE Closure Criteria
A landfill undergoing closure must be covered with a final cover that minimizes long term migration of liquids through the closed landfill. In addition, it must function with minimum maintenance, promote drainage, and minimize erosion of the cover, accommodate settling, and have hydraulic conductivity less than or equal to that of any bottom liner system or natural subsoil present. A multilayer cover system for a hazardous waste landfill closure is shown in Figure 3.3.
Vegetation/soil top layer
m
Drainage layer FML liner
......
Compacted soil layer
Solid waste
. . . . . . . i~~;ii i!j ii L: :.i i':ii(iiiiili{i,::;!:)i!iii! ~#
z~
'~
9 d
@'
're
'. '
-
;.~
"
:"
..:~:-',
~
. . . . .
0.66m
~ / : ' "i
ii'
~L ~<>> ....
~
... ~,.+
.. >:"_" _.
_"
...
~"
...
"
....'.:>~
... ~"
...'>~
Figure 3.3. A multilayer cover system for hazardous waste landfill closure (US EPA, 1989)
HAZARDOUS WASTE INCINERATION
The incineration of RCRA hazardous waste can only be accomplished under rigid control criteria, which include: (1) Waste feed chemical analysis to demonstrate the destructive capabilities of the incinerator; (2) The incinerator can only bum wastes permitted (i.e., waste the facility has proven capability of destroying) or wastes exempt from the hazardous waste incineration standard (i.e., waste with no hazardous constituents that are classified hazardous by virtue of their ignitable or corrosive nature); (3) The incinerator must meet performance criteria. These criteria include: (a) Principle organic hazardous constituent (POHC) destruction and removal efficiency (DRE) of 99.99%; (b) If the incinerator burns dioxins and furans, the combustion efficiency must be 99.9999% for each designated POHC; (c) The emissions of hydrogen chloride (HC1) must not exceed the larger of either 1.8 kg/hour or 1% of the HC1 in the stack gas before entering the pollution control equipment; (d) Particulate emissions are less than 180 mg/m3;
HAZARDOUS WASTE MANAGEMENT
(4) (5) (6) (7) (8)
55
Facility specific operation conditions with respect to controlling fugitive air emissions (i.e. non-venting or re-venting systems); The plant must continually monitor combustion temperature, waste feed rate, gas velocity and carbon monooxide levels; Specified minimum inspection obligations; New incinerators are required to go through a period of shake down and a trial burn; Closure stipulations that require the site to be cleared of all hazardous waste and residuals (i.e. ash, process water, etc.).
A RCRA final permit is required before an incinerator can start up. The permit outlines four conditions of operation or phases for the incinerators life: (1) shake down, (2) trial burn, (3) follow up, and (4) permanent operation. In the final permanent operation phase, the operating conditions must guarantee continued compliance with the standards.
3.4.2
Canada
In Canada, the Federal legislation that provides the complimentary cradle-to-grave tracking of hazardous waste is the Canadian Environmental Protection Act (CEPA). This legislation is further reinforced with the Transportation of Dangerous Goods Act (TDGA). Like the US initiative, the purpose of the CEPA is to control and protect the environment through procedures and strive to manage hazardous wastes from their manufacture to ultimate disposal as well as publish the state of the environment in Canada report (annually). Also, like the American legislation, the Canadian Act stresses minimization, recycling and conservation of resources. The main direction, at the Federal level, for publishing national guidelines for hazardous waste landfills and the incineration of hazardous waste has come from the Canadian Council of Ministers of the Environment (CCME). The CCME guidelines for hazardous waste land disposal (April 1991) are essentially the same as the American Hazardous and Solid Waste Amendments (HSWA) -- a 1984 amendment to the 40CFR 264 specification. With respect to pretreatment, the CCME guidelines specifically reference that fluids and materials containing free liquids be prohibited from hazardous waste landfills. A more pro-active stance on hazardous waste landfilling has been taken by some provinces (e.g., Ontario and Quebec). In Ontario, the Ontario Waste Management Corporation Act mandates the crown corporation to minimize, recycle and dispose of hazardous waste. In Quebec, Bill 60 modifies the provincial legislation Act to promote recycling. The CCME has been more successful at a federal level in adopting criteria for the performance of incinerators and issuing toxic metal, acid gas trace organic emission levels.
3.4.3
European Community
In March of 1990, the Commission of the European Communities adopted a waste policy resolution that applies to all wastes. This policy provides the EC with a general strategy for the management of its wastes. Briefly, it recommends the reduction of waste at its source and in this respect plans a proposal to deal with packaging. It encourages the use of materials which are biodegradable, reusable and recyclable, and manufacturing methods that are the least harmful. It promotes clean products and technologies and controlled recycling and reuse of waste where its production is unavoidable. It encourages the ongoing rehabilitation of waste disposal facilities,
56
WASTE
discourages land disposal of wastes without pretreatment, in an effort to reduce the volume and toxicity of wastes, and encourages incineration for volume reduction and energy recovery. It advises the member states to ensure a suitable infrastructure for the handling of wastes as close as possible to their points of generation, and stresses the importance of the adequate monitoring of wastes. It recognizes the need for the community to gather information on the volume of waste produced, the type of waste produced, the disposal and treatment facilities available, and the final point of disposal. It recommends that the polluter be responsible for the cost of waste disposal and finally it recommends that the transporting of waste be minimized and controlled as per the Basel Convention, which only permits trans-boundary movement if the country producing the waste lacks the necessary technical capabilities to dispose of it. In December 1991, a Council directive specific to hazardous waste was put forward by the Commission of the EC providing guidelines to the member states and directing them to draw up their own plans for the management, disposal and recovery of hazardous wastes, within the EEC's guidelines. These guidelines cover the identification and record-keeping of all hazardous wastes discharged at tipping sites, the prohibition of mixing different hazardous wastes without a permit, the encouragement to separate mixed waste if economically and technically feasible and the necessity of ensuring that all hazardous waste is correctly packaged and labeled.
Table 3.9: Classification of wastes accordin~ to leachate concentration Parameter Concentration (mg/1) Inert waste Hazardous waste pH TOC Arsenic Lead Cadmium Chromium Copper Nickel Mercury Zinc Phenols Fluoride Ammonium Chloride Cyanide Sulphate Nitrate AOX Chloinated solvents Chlorinated pesticides
4.00 - 13.00 < 40.00 < 0.1 < 0.4 < 0.1 < 0.1 < 2.0 < 0.2 < 0.02 < 2.0 < 10.0 < 5.0 < 50.0 < 500.0 < 0.1 < 1000.0 < 3.0 < 0.3 < 0.001 < 0.0005
4.00 - 13.00 40.00- 200.00 0.20 - 1.0 0.40- 2.0 0.10 - 0.5 0.10 - 0.5 2.00 - 10.0 0.20- 2.0 0.02 - 0.01 2.00 - 10.0 20.00 - 100.0 10.00 - 50.0 200.00 - 1000.0 1200.00- 6000.0 0.20 - 1.0 200.00 - 1000.0 6.00- 30.0 0.60- 3.0 0.02 - 0.1 0.001 - 0.005
HAZARDOUS WASTE MANAGEMENT
57
LAND DISPOSAL OF HAZARDOUS WASTE The commission of the EC considers land disposal as the last resort because a number of member states in the past have encountered serious soil and groundwater pollution problems. One of their main objectives, therefore, is to minimize the waste sent to the landfill by encouraging prevention and recycling. Another of their goals is to ensure uniform standards so the environment will be equally protected everywhere. Separate landfills for hazardous waste, inert waste, and municipal and other compatible wastes which are not hazardous are specified. However, not all hazardous waste is acceptable for land disposal. Hazardous wastes which cannot be land-disposed are: (a) all liquid wastes unless they are compatible or beneficial, (b) explosive, oxidizing and flammable wastes, as they could cause accidents, and (c) infectious wastes, which could pose a health threat. Other hazardous wastes may be disposed of in a hazardous waste landfill if the leachate concentration of the waste is within the range of values given in Table 3.9. Hazardous wastes that do not fall within the specified range have to be treated before they can be disposed of in a hazardous waste landfill. If treatment is not possible, codisposal, sometimes, can be performed. Codisposal is the mixing of two different wastes that will interact beneficially. It is not the mixing of, for example, a hazardous waste with another waste to dilute the waste that is hazardous, as this is not acceptable. Hazardous wastes which are not suitable for codisposal are the following: (1) Hazardous wastes not acceptable for land disposal; (2) Acid tars; (3) Immiscible organic solvents or liquid wastes containing greater than 1% immiscible organic chemicals; (4) Solvents containing greater than 10% water miscible organic chemicals; (5) Wastes that cause violent reactions with soil and organic material; (6) Asbestos; (7) Wastes containing: (a) PCBs and PCTs with concentrations greater than 50 ppb; TCDDs for isomer 2,3,7,8 with concentrations greater than 10 ppb; (b) PCNs with concentrations greater than 50 ppm; (c) PAHs with concentrations greater than 20 ppm; (d) Free cyanides with concentrations greater than 10 ppm; (e) Pesticides with concentrations greater than 2 ppm; (f) (g) Chlorinated hydrocarbons including chlorophenols with concentrations greater than 1 ppm; Organometallic compounds. (h) If neither treatment nor codisposal of a hazardous is possible, then the waste may be disposed of in a mono-landfill. A mono-landfill is a landfill site or part of a landfill site where only one type of waste is disposed, wastes of the same origin and composition, and those wastes producing similar leachates. Wastes of unknown characteristics and leachate composition will require sampling. To protect the soil and groundwater, lining of the base and side of hazardous and municipal waste landfills should be at least 3 m thick and the maximum hydraulic conductivity of the lining material should be 1.0x 10.3 m/sec. As inert waste is not expected to not undergo any major physical, chemical or biological changes, the lining material of inert waste landfills has no limit value for the
58
WASTE
hydraulic conductivity. It is the responsibility of the landfill site operator to accept or reject wastes at a site, and he or she will also be responsible for instituting corrective measures, should any adverse affects develop. The operator is responsible for the site for up to 10 years after closure. The cost of disposal, which is to be paid for by the polluter, must cover the landfill costs during construction, operation and closure.
3.5
SUMMARY AND CONCLUDING REMARKS
The management of hazardous waste, i.e., generation, storage, transport and disposal, offers several challenging issues if the requirements for protection of human health and the environment are to be met. Three basic sets of issues can be identified: (a) what are the rules, standards and criteria governing safe protection requirement? (b) what is the best disposal technology, and how can this be obtained and evaluated? (c) perception of governmental policy, regulations, and conformance requirements. It is difficult to address each of these issues without considering other pertinent aspects of the overall problem of waste disposal management. Land disposal of hazardous waste is still the lowest cost of the available technologies. However, there is growing economic incentive to use recycle and recovery techniques. Processes that produce large quantities of waste have been modified or abandoned in favour of those that produce less waste and, hence are more cost effective in the long run. While it is possible to reduce the volume of hazardous waste in manufacturing, it is simply not possible to eliminate it entirely because of continued and varied demand for goods and services. Waste exchange-reuse procedures offer potential disposal solutions for many industries. However, a major stumbling block to large scale adoption of waste exchange has been the lack of means to match supply and demand. Secrecy and competitiveness in industry have also hampered the free flow of information and have caused regulatory backlash. Any strategy for disposal of hazardous waste must take into account all the costs of the numerous components in the cradle-to-grave route to ultimate disposal: (1) generator handling and storage costs prior to treatment, (2) transportation to the treatment process, (3) operation of the treatment process, (4) handling and storage of the waste treatment residue, (5) transportation of residues for final disposal, and (6) final disposal. There is a need for improved long term planning in the management of hazardous waste. While secure land disposal is the least desirable alternative for managing hazardous waste, there will always be a need for some disposal of the residual hazardous waste in the environment. Land disposal techniques, no matter how well engineered, are ultimately vulnerable to various types of failure that can release pollutants into the environment. The planning process must ensure that the design of short term treatment and disposal techniques is compatible with the technologies that will be used to minimize the failure of the ultimate disposal sites. Minimization, incineration, recovery, and treatment offer the greatest growth potential for hazardous waste management in the future.
CHAPTER
FOUR
SOIL S Y S T E M
4.1
INTRODUCTION
As discussed in Chapter 1, the soil system, referred to as the geosphere, is an integral part of the land environment. Composed of soil constituents, water and air, it is the living environment for and from which human, animals and plants extract most of their food and energy. In this respect, the soil system is probably the ecosystem that receives the highest degree of anthropogenic stresses. From the geoenvironmental engineering viewpoint, the primary concerns relate to the problem of land disposal of waste and the decontamination of polluted soils. Evaluation of the effectiveness of a good barrier system, for land disposal, can usefully benefit from a closer consideration of how the pollutants are partitioned within the soil system. The pollutant retention mechanisms vary with soil constituents (mineral, amorphous material, soil organic matter, carbonates, etc.). A precise knowledge of a soil's retention mechanism can be used to estimate: (1) The potential for geoaccumulation, i.e., the tendency of a pollutant to persist in the soil for a long time; (2) The potential bioavailability of pollutants; bioavailability refers to the fraction of the total pollutant that is available for uptake, from polluted soils, by biota; and (3) The appropriate decontamination process, i.e., physical, chemical, biological, or electrical. The following development of the soil system will be concerned with attributes and characteristics which are pertinent to its utility as a waste retention and/or decontaminating agent.
4.2
SOIL PHASES
4.2.1
Gas Phase
The gas phase in soils is called soil air. It is located in the pore spaces, defined as the space in soil that is occupied by the gas and liquid phases. Soil air is composed of the same type of gases commonly found in the atmosphere. Biological activity in soil, however, may cause the percentage composition of soil air to differ considerably from that of atmospheric air [781 ml nitrogen (N2), 209 ml oxygen (O2) , 9.3 ml argon (Ar), and 0.31 ml carbon dioxide (CO2) in one litre of dry air]. Wellaerated soil contains 180-205 ml O2 per liter of soil air, but may drop to 100 ml/1 at one metre below the soil surface after a rainfall (i.e., flooding of the soil). The fractional volume of CO2 in soil air is typically 3-30 ml/1, but can approach 100 ml/1 at one metre depth after the flooding of the soil. The high CO2 content of soil air, relative to that of the atmosphere, can have a significant impact on a soil's acid buffering capacity, i.e., the ability of the soil to resist changes in its pH. 59
60
SOIL SYSTEM
The dissolution of soil air gases into the soil solution is an important factor in the cycling of chemical elements in the soil environment and the remediation of polluted soil by vapour extraction. Gaseous species are partitioned between soil air and soil water. At equilibrium, the relation between the concentration of gas in soil solution and the partial vapour pressure in soil air is given by Henry's law:
[4.1]
KH = [Aaq]/eA
where K/4 is Henry's constant (mol/m3), [Aaq] is the concentration of gas A in the soil solution (mol/m3), and PA is the partial vapor pressure of gas A in the soil air (atm). Eq. [4.1 ] is only valid when the gas concentration in the soil solution is low. The importance of Eq. [4.1] lies in its use as a method for: (1) classifying the potential volatilization of organic chemicals, (2) modelling the partitioning of organic chemicals in unsaturated soils, and (3) defining the limits for using soil vapour extraction process to decontaminate soils polluted with organic chemicals. Table 4.1 lists values of KHat 25~ for several uncontaminated soil air gases. For example, from Table 4.1, K, = 34.06 mol/m 3 at 25~ for CO). Assuming the partial pressure of CO2 in soil air is 0.03 atm; then, according to Eq. [4.1 ], the concentration of [CO2 (aq)] in the soil solution is 1.02 mol/m 3.
Table 4.1: Values of Henry's constant of various ~ases in soil at 25~
x.
K. Gas CO 2 CH 4 NH 3
N20
mol
m 3 atm 1
Gas
34.06 1.50 5.76• 104 25.55
NO 02 SOz HzS
mol
m -3
atm ~
1.88 1.26 1.24• 103 1.02•
Aeration has a significant effect on many biochemical reactions in soils. Many of the decomposition reactions of soil organic matter are influenced by the soil's aeration status. When adequate amount of air (02) is present, aerobic reactions (oxidation) prevail; when aeration is poor, anaerobic reactions (reduction) predominate. It is noteworthy that soil properties in oxidized and reduced states are markedly different. Thus, the solubility of pollutants in soils is dependent on the oxidation-reduction state, and has an important impact on the mobility of pollutants through soils. In well-aerated soils, organic matter will decompose into CO2 and H20, and release nutrient elements. The carbon dioxide produced will react with soil water to form carbonic acid, and alter the soil acidity. The increase in soil acidity will enhance the dissolution of soil minerals. Other acids that are produced by aerobic decomposition of organic matter include humic, nitric, and sulfuric acids. These acids also contribute to the solubilization of soil minerals. Such dissolution is disadvantageous
SOIL PHASES
61
for the design of land disposal systems since it increases soil hydraulic conductivity and reduces the ability of soils to retain pollutants. Reduction processes, which prevail in anaerobic conditions, contribute to the reduction of many soil pollutants, e.g., reduction of iron to Fe 2+, manganese to Mn 2+, sulphur to SO32-, and nitrate to nitrite (NO2). Through a series of successive reduction, a nitrogen compound may eventually be reduced to N2 gas. This process, known as denitrification, is often applied to remove nitrates from polluted soil, as discussed in Chapter 21. Soil aeration is affected by compaction, soil moisture and temperature. When the soil pores are destroyed by compaction, not only will air be deficient, but also, the soil becomes dense. As the soil moisture content decreases, the soil air content increases, reaching its maximum (15% 02) at the field capacity. Field capacity refers to soil moisture content at a moisture tension of 0.3 bars (Brady, 1984). With increased soil drying, the air content remains constant. As the temperature increases, the 02 concentration is increased. 4.2.2
Fluid Phase
The fluid phase in soil constitutes between one- and two-thirds of the total soil volume. Soil water exists principally in the condensed phase. Soil water is a repository for dissolved solids and gases and, for this reason, is commonly referred to as soil solution. The term soil moisture is frequently used to refer to soil water or soil solution. Soil moisture can be found in the macro- and micro-pores. When water saturated soil drains under the influence of gravity, some water is retained in the macro-pores in the form of a thin film around the soil particles. The micro-pores, on the other hand, remain saturated. Evaporation and consumption by soil organisms lower the moisture content in soils; at low moisture content, moisture exists as thin films and as wedges at the contact points of soil particles. Many of the chemical properties of water are attributed to its dipolar molecular structure. Water exhibits a high dielectric constant, which promotes the dissociation of many compounds in water. Thus, water is an excellent solvent for a number of chemical compounds. It is the most important transporting agent for nutrient elements and pollutants in soils. Many of the dissolved materials are in ionic forms. If a compound with a general formula A,B x is in contact with soil water, it dissociates into its ionic components An+ and B~-, where A is the metal cation with charge n and B is the anion with charge x. As in the case with protons (H+), which exist a s H3O+, a metal ion cannot exist by itself. In soil water, the metal reacts with water molecules, forming a hydration shell. The number of water molecules attracted to the metal ion depends on the coordination number of the cation. The hydrated cation carries the original number of positive charges, and is denoted A(H20)• n+. Water is held in the pore spaces by forces of attraction exerted by the soil matrix. These forces expressed in terms of matric, osmotic, and pressure potentials, are collectively known as soil water potential. The matric potential and the osmotic potential are the forces that bind the soil water to the soil solids and the soil solutes, respectively. The pressure potential, which results from pressure differences in soils, is responsible for the retention of water in soil pores (capillary forces) and on the surface of soil particles (adsorption). Several units have been used to express differences in energy levels of soil water. One is the height in centimetres of a unit water column whose weight just equals the potential under consideration. The pF unit is commonly used to characterize soil water potential. It is defined in
62
SOIL SYSTEM
terms of the logarithm of the height, in cm, of water, i.e., pF = lOgl0 cm H20. Practical values of soil water potential range from 1 cm of water (pF = 0) for saturated soil to 108 cm of water (pF = 7) for oven dry soil at 105-110~ A second is the standard atmosphere pressure at sea level, which is 760 mm Hg or 1020 cm of water. The unit termed bar approximates that of a standard atmosphere. At field capacity, the soil water potential is 0.3 bars, which corresponds to a pF value of 2.54. Energy may be expressed per unit of mass (joules/kg) or per unit of volume (newton/m2). Soil water potential has a negative value because of the matric and osmotic forces, which reduce the free energy level of the soil water. This means that soil water cannot move freely. The larger the negative value of soil water, the smaller the amount of water present in the soil. Generalized soil water potential curves for soils with different grain sizes are shown in Figure 4.1. Kaolinite clay holds much more water at a given potential level than does sandy clay or fine sand. Likewise, at a given moisture content, the water is held much more tenaciously in the kaolinite than in the other two soils. Soil texture clearly exerts a major influence on soil moisture retention. Soil structure also influences soil moisture-energy relationships. A well-granulated soil has more total pore space than one with poor granulation or one that has been compacted. The reduced pore space may result in a lower water-holding capacity. The compacted soil also may have a higher proportion of small- and medium-sized pores, which tend to hold water with greater tenacity than do larger pores.
Figure 4.1. Soil moisture variation with soil water potential for different soils.
The soil moisture-soil water potential relationship during drying differs from the relationship during wetting. This phenomena, known as hysteresis, is illustrated in Figure 4.1. Hysteresis is caused by a number of factors, including the nonuniformity of soil pores. As soils are wetted some of the smaller pores are bypassed, leaving entrapped air that prevents water penetration. Likewise, as a saturated soil dries, some of the macro-pores that may be surrounded only by micro-pores may
SOIL PHASES
63
not lose their water until the soil water potential is low enough to remove the water from the smaller pores. Soil water is generally classified into free water, capillary water, and hygroscopic water as shown in Figure 4.2. These classes are defined as: (1)free water: water held between the maximum retention capacity (pF 0) and the field capacity (pF 2.54). Saturated soil contains free water, (2) capillary water: water held between the field capacity (pF 2.54) and the hygroscopic water (pF 4.5). Hygroscopic water refers to the maximum amount of water adsorbed by soils from the atmosphere. Unsaturated soils contain capillary water, and (3) hygroscopic water: water held at pF 4.5.
Figure 4.2. Classification of soil moisture.
4.2.3
Solid Phase
The solid phase in soils consists of both inorganic and organic fractions. The inorganic fractions, derived from the weathering products of rocks, range in size from tiny colloids (< 2 ~tm) to large gravel and rocks (> 2 mm), and include many soil minerals, both primary and secondary. The inorganic fractions exert a tremendous effect on the physico-chemical properties of soils and the ability to retain chemicals. Separated according to size, the inorganic soil fraction can be divided into three major soil groups: sand, silt and clay. Sand grains are irregular in size and shape, and are not sticky and/or plastic when wet. Their presence in soil promotes a loose and friable condition which allows rapid water and air movement. They are chemically inert and do not carry electrical charges. Thus, they have low water-holding and cation exchange capacities. Silt particles are intermediate in size and possess characteristics between those of sand and clay. Some silt particles may be capped or coated with clay films as a result of weathering. Such particles may, therefore, exhibit some plasticity, stickiness, and adsorptive capacity for water and cations. Clay is the smallest particle in soil and has colloidal properties. It carries a negative charge and is chemically the most active inorganic soil
64
SOIL SYSTEM
constituent. The presence of clay contributes to: (1) high water-holding capacity, (2) large specific surface area, and (3) high cation exchange capacity. The inorganic soil fraction is composed of soil minerals, and is, therefore, also referred to as the mineral fraction of soils. Minerals are inorganic in nature, possessing definite physical characteristics and chemical compositions. Soil minerals can be grouped into primary and secondary minerals. Primary minerals are minerals that have been released by weathering from rocks in a condition that is chemically unchanged. These minerals constitute the sand fraction of soils. Secondary minerals are derived from the weathering of primary minerals. They are present in the clay fraction of soils. Organic components include plant and animal residues at various stages of decomposition, cells and tissues of soil organisms. Organic components, although normally present in much smaller quantities than inorganic components, may significantly alter a soil's properties.
4.3
MINERAL COMPOSITION
4.3.1
Primary and Secondary Minerals
The composition of soil minerals is variable and depends on the composition of the rocks from which they were derived. Rocks are mostly composed of the elements 02, Si, A1, Fe, Ca, Mg, Na, and K. These are, therefore, the elements usually found in soil minerals. Silicates and oxides are probably the most common soil minerals. The silicon (Si) in soil silicates is present in the form of silica tetrahedral, which constitute the basic units of the clay mineral. On the basis of the arrangement of the silica tetrahedral (SiO4) in the crystal structure, 15 common silicates, listed in Table 4.2, are formed. The first six silicates in Table 4.2 are primary minerals since they are typically inherited from parent material, as opposed to precipitated through the weathering process. The key structural entity in these minerals is the Si-O bond, which is a more covalent and, therefore, stronger bond than typical metal-oxygen bonds. The relative resistance of any one of the minerals to decomposition by weathering can be correlated positively with the Si/O molar ratio of its fundamental silicate structural unit. This is because a larger ratio means a lesser need to incorporate metal cations into the mineral structure for the purpose of neutralizing the oxygen anion charge. To the extent that metal cations are so excluded, the degree of co-valency in the overall bonding arrangement will be greater and the mineral will be more resistant to decomposition in the soil environment. For the first six silicates shown in Table 4.2, the Si/O molar ratios of their fundamental structural units are as follows: 0.5 (quartz and feldspar, SiO2) , 0.40 (mica, SizOs), 0.36 (amphibole, Si401~), 0.33 (pyroxene, SiO3), and 0.25 (olivine, SiO4). The decreasing order of the Si/O molar ratio is the same as the observed decreasing order of resistance to chemical weathering in the sand or silt fractions of soils. The minerals epidote, tourmaline, zircon, and rutile, shown in Table 4.2, are found to be highly resistant to weathering in soil environment. With reference to Table 4.2, listed minerals from kaolinite to gypsum are secondary minerals since they result from the weathering transformations of primary silicates. Often these secondary minerals are of clay size and exhibit poorly ordered atomic structure. Variability in their composition through the substitution of ions into their structure (isomorphous substitution) is also frequently noted in soils. The secondary silicates, smectite and vermiculite, bear a net charge on their surfaces,
MINERAL COMPOSITION
65
principally because of their variability in composition. Kaolinite and the secondary metal oxides below it in the list also bear a net surface charge due, however, to proton adsorption and desorption, not compositional variability. Metal oxides such as gibbsite and geothite tend to persist in the soil environment longer than the secondary silicates. This is because Si is more readily leached than A1, Fe, and Mn, unless significant amounts of soluble organic matter are present to render the metals more mobile.
Table 4.2: Common soil minerals (Bohn et al., 1979) Name
Chemical Formula
Importance Abundant in sand and silt Abundant in soil that is not leached extensively Source of K in most temperate-zone soils
Pyroxene Olivine
SiO 2 (Na,K)A102[SiO2]3 CaA1204[SiO212 KzA12OS[Si2OS]3A14(OH) 4 K2AlzOs[Si2Os]3(Mg, Fe)6(OH)4 (Ca, Na, K)z.3(Mg, Fe, A1)5(OH)2 [(Si, A1)40,112 (Ca, Na, Fe, Ti, A1)(Si, A1)O3 (Mg, Fe)2SiO4
Easily weathered to clay minerals and oxides Easily weathered Easily weathered
Epidote Tourmaline Zircon Rutile
Ca2(al, Fe)3(OH)Si3Ol2 NaMg3A16B3Si6027(OH, F)4 ZrSiO 4 TiO2
Highly resistant to chemical Highly resistant to chemical Highly resistant to chemical Highly resistant to chemical
kaolinite Smectite Vermiculite Chlorite
Si4A14Olo(OH)8 Mx(Si, A1)8(A1, Fe, Mg)4020(OH)4 Mx(Si, A1)8(A1, Fe, Mg)402o(OH)4 Mx(Si, A1)8(A1, Fe, Mg)402o(OH)4
Abundant Abundant Abundant Abundant
Allophane
Si3A14012.nH20
Imogolite
Si2A14Olo.5H20
Gibbsite Goethite Hematite Ferrihydrite Calcite Gypsum
AI(OH)3 FeO(OH) Fe203 FeloO15.9HzO CaCO 3 CaSO4.2H20
Abundant in soils derived from volcanic ash deposits Abundant in soils derived from volcanic ash deposits Abundant in leached soils Most abundant Fe oxide Abundant in warm regions Abundant in organic soils Most abundant carbonate Abundant in arid regions
Quartz Feldspar Mica Amphibole
Note: M represents interlayer cations.
weathering weathering weathering weathering
in clays due to weathering in clays due to weathering in clays due to weathering in clays due to weathering
66
SOIL SYSTEM
Organic matter is an important constituent of the solid fraction of soils. The structural complexity of soil organic compounds precludes the making of a simple list of component solids like that shown in Table 4.2. 4.3.2
Trace Elements in Soil Minerals
One of the most important aspects of the variability in composition of soil minerals is their content of trace elements. A trace element is any chemical element whose mass concentration in a solid phase is less than or equal to 100 mg/kg (Sposito, 1989). Soil minerals bearing trace elements serve as reservoirs for the elements, releasing them slowly into the soil solution as weathering of the minerals occurs. The bioavailability of an element depends on the rate at which it is transformed from a solid phase to a soluble chemical form. Soil physico-chemical properties such as pH, redox potential, and moisture content will affect the rate of this transformation and, thus, control element solubility. In this manner, the weathering rate of soil solids containing toxic elements (e.g., Cd) will determine, in part, the potential hazard to human, plants and animals. Trace elements in secondary soil minerals and soil organic matter are listed in Table 4.3. The chemical phenomenon underlying trace element occurrence is called coprecipitation, which is defined as the simultaneous precipitation of a chemical element with other elements by any mechanism and at any rate. The three broad types of coprecipitation are inclusion, adsorption, and solid-solution formation.
Table 4.3: Trace elements coprecipitated with secondary soil minerals and soil organic matter Solid Fe and A1 oxides Mn oxides Ca carbonates Illites Smectites Vermiculites Organic matter
Coprecipitated trace elements B, P, V, Mn, Ni, Cu, Zn, Mo, As, Se P, Fe, Co, Ni, Zn, Mo, As, Se, Pb P, V, Mn, Fe, Co, Cd B, V, Ni, Co, Cu, Zn, Mo, As, Se, Pb B, Ti, V, Cr, Mn, Fe, Co, Ni, Cu, Pb Ti, Mn, Fe A1, V, Cr, Mn, Fe, Ni, Cu, Zn, Cd, Pb
If a trace element forms a pure solid phase with atomic structure different from the host mineral, then two morphologically distinct solids will occur together. This kind of association is termed inclusion with respect to the trace element. For example, CuS often occurs as an inclusion, a small separate phase, in the primary silicates. If there is only limited structural compatibility between a trace element and the corresponding major element in a host mineral, coprecipitation can produce a homogeneous mixture of the two elements at the host mineral-soil solution interface. This mechanism is termed adsorption because the mixed solid phase is restricted to the interfacial region, which can change as the host mineral
SOIL M1NERAL TRANSFORMATIONS
67
continues to precipitate from the soil solution. Well known examples of adsorption are the incorporation of oxy-anions like borate, phosphate, or molybdate into secondary metal oxides, and of transition metals like Fe or Ni into soil organic matter. Finally, if structural compatibility is high and free diffusion of a trace element within the host mineral is possible, a major element in the host mineral can be replaced uniformly throughout by the trace element. This kind of homogeneous coprecipitation is called solid-solution formation. It is enhanced if the size and valence of the substituting element are comparable to those of the element replaced. Solid-solution formation occurs, for example, when secondary alumino-silicates precipitate and incorporate metals like Ni, Cu, and Zn to replace A1 in their structure or when calcium carbonate precipitates with Cd replacing Ca in the structure.
4.4
SOIL MINERAL TRANSFORMATIONS
The continual input and output of percolating water, biomass, and solar energy alters the composition of soils with the passage of time. The changes in clay fraction mineralogy observed during the course of soil development are shown in Table 4.4. These changes, known as JacksonSherman weathering stages, can be classified as early, intermediate, and advanced. Soils in stage 1 weathering may contain some gypsum and halite. Soils containing significant amounts of olivine are representative of stage 3, and biotite mica is representative of stage 4. This sequence is in agreement with their oxygen-silicon ratios and weathering resistance. Soils with minerals representative of stages 1 to 5 are considered to be in the early weathering stages. Such soils are often the least weathered, and are primarily found in regions where limited water restricts chemical weathering. Soils of the intermediate weathering stages (stages 6 to 9) include most soils of humid temperate regions. Quartz is often abundant in these soils and is representative of weathering stage 6. Hydrous mica, vermiculate, and montmorillonite are typically transformed in soils, and accumulate as fine-sized particles in the clay fraction. Soils of the advanced weathering stages include the intensely weathered soils of humid tropics. Soils dominated by minerals of weathering stages 10 to 13 may have lost all or most of the original minerals of the parent material. They may consist mainly of stable minerals that have been synthesized during weathering. These minerals are kaolinite, gibbsite, hematite, and anatase (TiO2).
4.5
CRYSTAL CHEMISTRY OF SILICATES
When atoms combine, a bond involving a redistribution of valence electrons is formed. The type of bond formed is a function of the electronic structure of the combining atoms. Ionic or electrostatic bonding occurs between ions of opposite charge, such as Na + and C1-. Such ions are formed by the complete loss or gain of electrons to form positive or negative ions having an electron structure like an inert gas. Ionic bonding forces are strong, and solid ionic compounds have high melting points. Ionic bonding forces are also undirected, that is, they are exerted uniformly in all directions. The valence of a given ion is shared by all neighboring ions of opposite charge. The number of such neighbors is determined by their size relative to the size of the central ion. Ionic bonds predominate in many inorganic crystals, including the silicate minerals.
68
SOIL SYSTEM
Table 4.4: Representative minerals associated with weathering stages (Jackson and Sherman, 1953) Weathering stage
Representative minerals
Typical soil group
EARLY WEATHERING STAGES Gypsum Soils dominated by these minerals, in Calcite, dolomite the fine silt and clay fractions, are Olivine-hornblende mainly weathered soils. They are of the Biotite desert regions, where limited water keeps Albite, microcline, orthoclase chemical weathering to a minimum.
INTERMEDIATE WEATHERING STAGES Quartz Soils dominated by these minerals are in Hydrous mica, muscovite the fine silt and clay fractions. They are Vermiculite mainly of temperate regions. Montmorillonite
10 11 12 13
ADVANCED WEATHERING STAGES Kaolinite Soils dominated by these minerals are in Gibbsite the clay fraction. They are intensely Hematite, geothite weathered soils of the warm and humid Analase, zircon equatorial regions. The soils are frequently characterized by acidity.
Covalent bond or shared electron pair bonding is common between identical atoms or atoms having similar electrical properties, such as H2O, F 2, and C H 4. Covalent bonding is the sharing of electron pairs between the combining atoms so that each atom attains the inert gas electronic structure. Covalent bonding is strong, but bonding is directional. Covalently bonded molecules have little tendency to ionize. Bonding within ionic radicals, such as SO42, is frequently covalent. Hydrogen bonding occurs between hydrogen and two atoms of high electro-negativity, such as F, O, and N. The H-bond is essentially a weak electrostatic bond but is nevertheless important in crystal structures of oxy-compounds, such as the layer silicates. Summed over many atoms, the individually weak H-bonds can strongly bond adjacent structures. van der Waals bonding is the weak electrostatic force between residual charges on molecules. Residual charges may result from natural dipoles of non-symmetrical molecules, polarization dipoles, or vibrational dipoles, van der Waals forces are generally obscured by stronger ionic and covalent bonding forces but may dominate the properties of some molecules. Although differences in the types of bonds described above are rather clear-cut, bonding in most crystals is not. For example, the Si-O bond in silicates is intermediate between the extremes of purely ionic and purely covalent bonding. The degree of ionic nature of the Si-O bond is
CRYSTAL CHEMISTRY OF SILICATES
69
sufficient, however, to apply the rules of ionic bonding to silicate structures. The internal bonding in silicates is predominantly ionic. As a result, forces are undirected and ionic size plays an important part in determining crystal structure. The crystal radii of common ions in the silicates are given in Table 4.5. The ionic radius of oxygen is much larger than that of most cations found in silicates. The oxygen ion constitutes nearly 50% of the mass, and over 90% of the volume, of most common silicate minerals. Hence silicate structures are largely determined by the manner in which the oxygen ions pack together. An ion in a crystal surrounds itself with ions of opposite charge. The number of ions that can be packed around a central ion depends on the ratio of radii of the two ions and is called the coordination number of the central ion. Ions are held together more or less rigidly in a crystal structure, determined by geometry and by electrical stability. More than one structure may meet the necessary requirements, but the most stable form will be the one in which the ions have the lowest potential energy. The requirement of electrical stability means that the sum of positive and negative charges must be equal. However, ions of opposite charge are not paired off to achieve neutrality. Instead, the positive charge of a cation may be considered to be divided equally among surrounding anions. The number of anions around each cation is determined by the coordination number, or radius ratio, of the cation and anions, rather than by the charge of the cation.
Table 4.5: Crystal ionic radii of selected cations, and their coordination number with oxygen (Bohn et al., 1979) Ion radius Ion Si 4+ A13+ F e 3+
Mg 2+ Fe 2§ Na § Ca > K+ NH4+ 02.
(A)
0.42 0.51 0.64 0.66 0.74 0.97 0.99 1.33 1.43 1.32
Coordination number with oxygen Observed Predicted 4 4,6 6 6 6 6,8 8 8,12 8,12 . . . . .
4 4 4 6 6 6 6 8 8
Predicted and observed coordination numbers of oxygen anion with common cations are given in Table 4.5 (Bohn et al., 1979). The Si 4§ cation occurs in fourfold or tetrahedral coordination. Aluminum is generally found in sixfold or octahedral coordination but may also occur in tetrahedral coordination. The tetrahedral and octahedral coordination units are basic to the atomic lattices of most layer silicate minerals. The tetrahedral structure, shown in Figure 4.3, consists of four 02. ligands coordinated around Si 4§ giving the ionic unit (SiO4) 4". The electrostatic bond strength (ion charge divided by number of bonds to the ion) for the tetrahedral unit is one. In fourfold
70
SOIL SYSTEM
coordination, the hole between the four O 2" ranges from 0.29 to 0.52 ,~. The radius of Si 4+, in fourfold coordination, is about 0.42 A, indicating some distortion from the ideal tetrahedral unit. The basic octahedral structure, shown in Figure 4.3, consists of six OH- groups coordinated around a central cation. The electrostatic bond strength is 1/2 or 1/3, depending on the charge of the central ion. Sixfold coordination yields an eight-faced structure, hence the name octahedral. The hole between O 2 ligands in this configuration has a theoretical radius of about 0.61 ,~. Ions commonly found in octahedral coordination in layer silicates include A13+ (radius 0.51 A in this coordination), Mg 2§ (0.66 ]k) and Fe z+ (0.74 A).
Figure 4.3. Silica tetrahedron and octahedrons of aluminum and magnesium.
4.6
STRUCTURAL COMPONENTS OF SOIL CLAYS
Many soil clays are structurally related. Thus, learning the basic components of soil clays facilitates an understanding of their nature and differences. Knowledge of clay structure is essential for understanding how clays affect the physical and chemical properties of soils, such as changes in hydraulic conductivity due to intrusion of pollutants, mobility of pollutants, soil adsorption capacity, buffering capacity, etc.
STRUCTURAL COMPONENTS OF SOIL CLAYS
4.6.1
71
Silica, Gibbsite, and Brucite Sheets
Many soil clays are alumino-silicates. As already noted, silicon-oxygen tetrahedra share oxygen atoms to form sheets. The apical oxygen has one excess negative charge, and the silica sheet has the formula Si2052-. The gibbsite sheet, shown in Figure 4.4, is composed of aluminum in six coordination with hydroxyl. An individual octahedron is AI(OH)63-. The sharing of hydroxyls by adjacent octahedra forms an octahedral or gibbsite sheet with the composition A12(OH)6. Two of the three potential spaces for aluminum in the sheet are filled, the mineral is referred to as dioctahedral. Although gibbsite sheet is an integral constituent of many alumino-silicate clays, it also exists by itself in soils.
Figure 4.4. Silica, gibbsite and brucite sheets.
In 9 a manner similar to aluminum and hydroxyl, magnesium and hydroxyl form an octahedral sheet called the brucite sheet, as shown in Figure 4.4. The magnesium is in six coordination with hydroxyl, and the sheet has the composition Mg3(OH)6. All three of the potential spaces for cations are filled with magnesium to produce a tri-octahedral mineral. Both gibbsite and brucite are octahedral and have electrically neutral structures. They differ in that gibbsite is dioctahedral and brucite is trioctahedral.
72
SOIL SYSTEM 1:1 LAYER SILICA TES
The structural unit of the kaolin group is formed by the superposition of a tetrahedral sheet upon an octahedral sheet. Such minerals are referred to as 1:1 layer silicates. The apical oxygens of the tetrahedral sheet are shared by the octahedral sheet, forming a common plane of oxygen ions within the structure, as shown in Figure 4.5. In the shared plane, two-thirds of the oxygen ions are shared between Si and A1. The remaining one-third of the oxygen ions have their charge satisfied by H § to form OH groups. The upper surface of kaolin is a layer of closely packed OH groups, but the bottom surface is composed ofhexagonally open-packed oxygens and OH groups recessed within hexagonal openings. Kaolinite [AI2Si2Os(OH)4] is the layer silicate mineral that represents the kaolin group. Silicon is apparently the only cation in the tetrahedral sheet of kaolinite, but A13§ or Mg 2+ may occupy the octahedral positions. If A13+is in octahedral coordination, the mineral is kaolinite or one of its poorly crystallized polymorphous forms (dickite or nacrite). With Mg 2+ in octahedral coordination, the mineral is antigorite [Mg3Si2Os(OH)4]. Halloydite is a form of kaolinite in which water is held between structural units in the basal plane, yielding a c-spacing of 10 ~, when fully hydrated. Most kaolin structural units, however, are held together in the basal plane by hydrogen bonding between oxygen ions of the tetrahedral sheet and hydroxyl ions of the octahedral sheet.
Figure 4.5. Schematic structure of kaolinite.
2:1 LA YER SILICA TES In 2:1 layer silicates, the unit layer is one octahedral sheet sandwiched between, and sharing oxygen atoms with, two tetrahedral sheets. The unit layers then stack parallel to each other in the cdimension. The atomic arrangement for illite mineral is shown in Figure 4.6.
STRUCTURAL COMPONENTS OF SOIL CLAYS
73
The various 2:1 minerals are differentiated by the kinds and amounts of isomorphous substitution in both the tetrahedral and octahedral sheets, which can lead to localized charge within the crystal. This excess charge must be balanced by other cations, either inside the crystal or outside the structural unit. The magnitude of charge per formula unit, when balanced by cations external to the unit layer, is called the layer charge. Typical layer charges of 2:1 minerals are shown in Table 4.6.
Figure 4.6. Schematic structure of illite.
The magnitude of a layer charge plays a dominant role in determining the strength and type of bonding in the basal plane. If the layer charge is zero, as in pyrophyllite, the basal planes of adjacent unit cells are bonded together by van der Waals forces. If the layer charge is negative, adjacent basal planes are bonded electrostatically by cations located between the unit layers. The greater the layer charge, the stronger the interlayer bond. Smectites of low layer charge have a weak bond, enabling polar molecules, such as water, to get between the basal planes and cause the minerals to expand. In minerals of high layer charge, such as the micas, the ionic bond is so strong that polar molecules cannot get between the basal planes, and the minerals are non-expanding. Vermiculites are intermediate in layer charge and also intermediate between mica and smectite in their expansion properties. Within a given mineral group, specific minerals are defined by the predominant ion in octahedral coordination, as shown in Table 4.6.
74
SOIL SYSTEM
Table 4.6: Classification of layer silicate minerals Layer charge per unit Mineral group Tetrahedral Octahedral Sheet sheet Pyrophyllite Talc Smectites Vermiculites Micas l
0 0.25 - 0.6 0 0.6 - 0.9 1
0 0 0.25 - 0.6 0 0
Octahedral cation A13+
Mg 2+
Pyrophyllite Beidellite Montmorillonite Vermiculite Muscovite
Talc Saponite Hectorite Vermiculite Biotite 1
Mg and Fe are in octahedral coordination; K is in the interlayer position.
In sepiolite and palygorskite, the 2:1 layers do not form continuous sheets, but form fibres six (sepiolite) or four (palygorskite) silicon tetrahedra wide. Simplified formulas are [Mg 4Si6 O~5(OH)2.6H20] and [(Mg, A1, Fe)4Si8 020. nH20], respectively. The symbolic structure of these minerals is shown in Figure 4.7. Sepiolite may occur as the pure Mg end member, but most sepiolites and all natural palygorskites contain some aluminum and usually some exchangeable cations.
Water k'.x'.'CZZZ].'.x'.'4 Water I f . f l Z Z Z b , ' . f . ~
Wa ter ~,xJ,,~,E~lYJ,~ Water ~ 7 , ] ~ ~ F 7 , 1
Water
Wa ter
Water rT~,eJ~,~,,~77~ Water
Figure 4.7. Symbolic structure ofpalygorskite and sepiolite. The 2:1 layers form chains rather than a continuous sheet.
2:1:1 LAYER SILICA TES The chlorite mineral group is closely related to the micas and has about the same layer charge. In chlorite, the interlayer potassium of mica is replaced by positively charged octahedral brucite [Mga(OH)6 ] sheet. The brucite sheet develops a positive charge when the Mg 2+ is partially replaced by A13+, yielding the basic unit [Mg2AI(OH)6] +1 that fits into the interlayer position of the
PROPERTIES OF LAYER SILICATES
75
2:1 layer silicates. This is referred to as a 2:1:1 type classification. Chlorites are non-expanding minerals with low cation exchange capacity.
4.7
PROPERTIES OF LAYER SILICATES
4.7.1
Kaolins
The kaolinite crystal consists of repeating layers, each layer consisting of a silica sheet and an alumina sheet sharing a layer of oxygen atoms between them, as shown in Figure 4.5. Each layer is three oxygen atoms thick. The layers are held together by hydrogen bonding between hydroxyls from the alumina sheet on one face and oxygens from the silica sheet on the opposite face. These forces are relatively strong, preventing hydration between layers and allowing many layers to build up. A typical kaolinite crystal may be between 70 to 100 layers thick. Kaolinite occurs commonly in soils, often as hexagonal crystals with an effective diameter of 0.2 to 2 ~tm. Hydrogen bonding between adjacent unit layers prevents expansion of the mineral beyond its basal spacing of 7.2 A. Surface area is limited to external surfaces and, hence, is relatively small, ranging from 10 to 20 m2/g. Kaolinite is a coarse clay with low colloidal activity, that is, low plasticity and cohesion, and low swelling and shrinkage.
Table 4.7: Summary of selected properties of solid phase components Cation Mineral Layer Exchange Surface Component type charge Capacity area (meq/100~) (m2/~) Kaolinite Montmorillonite Vermiculite Mica Chlorite Organic matter
1:1 2:1 2:1 2:1 2:1:1 ..........
0 0.25-0.6 0.6-0.9 1.0 -~1.0
1-10 80-120 120-150 20-40 20-40 100-300
10-20 600-800 600-800 70-150 70-150 800-900
c-spacing
A 7.2 variable 10-15 10 14 ....
The ideal unit formula for kaolinite [A12Si2Os(OH)4] has an Si/A1 ratio of one which suggests little or no isomorphous substitution. Most of the cation exchange capacity (1 to 10 meq/100g) of kaolinite can be attributed to the dissociation of OH groups on clay edges. The cation exchange capacity of kaolinite is highly pH-dependent, suggesting that isomorphic substitution is not the predominant source of charge. A summary of some selected properties of kaolinite is shown in Table 4.7.
76
4.7.2
SOIL SYSTEM
Hydrous Mica (Illite)
Mica minerals have repeating layers of an alumina sheet between two silica sheets, with shared oxygen to give a unit of four oxygen atoms thick. The layers are bonded together by potassium ions which are just the right size to fit into the hexagonal holes of the silica sheet, as shown in Figure 4.6. The bonding, via potassium ions, between adjacent unit layers prevents expansion of the mineral beyond its basal spacing of 10 A. The potassium ions exist in twelve coordination, bonding six oxygens from one silica sheet to the adjacent six oxygens of the silica sheet of the next layer. Negative charge, to balance the potassium cations, arises from the substitution of aluminum for silicon in the silica sheet (isomorphous substitution). A typical unit formula for mica is K[AI2(Si3A1)OI0(OH)2]. Despite the relatively large layer charge (-- -1) of the mica, its cation exchange capacity is only 20 to 40 meq/100g. Its total surface area is about 70 to 120 m2/g and is restricted to external surfaces, as indicated in Table 4.7. Soils containing illite have properties intermediate between kaolinite (low activity) and montmorillonite (high activity). Illite occurs widely in temperate and in arid regions.
Figure 4.8. Schematic representation of montmorillonite structure.
4.7.3
Montmorillonite
A typical unit formula is Nax[(A12_xMgx)Si40~0(OH)2] in which Na § is the chargecompensating exchangeable cation. Montmorillonite minerals have the same layers as micas, discussed in the previous section. However, soil montmorillonites exhibit imperfect isomorphic
PROPERTIES OF LAYER SILICATES
77
substitution, with some A13+ substituting for Si4+ in the tetrahedral sheet and with Fe 2+(as well as Mg 2+) substituting for A13+ in the octahedral sheet. There are no potassium ions to bond the layers together, and water enters easily between layers, as illustrated in Figure 4.8. The distance of separation of the layers on hydration can be controlled if certain organic liquids rather than water are used. Montmorillonite saturated with glycerol will show layer spacings of 17.7 ~,, of which 10 A is the thickness of the layer and 7.7 A of the glycerol. Typical cation exchange capacities for montmorillonite range from 80 to 120 meq/100g, as indicated in Table 4.7. The cation exchange capacity is only slightly pH-dependent. The lower layer charge allows the mineral to expand freely, exposing both internal and external surfaces. Such expansion yields a total surface area of 600 to 800 m2/g, with as much as 80% of the total surface area due to internal surfaces. Montmorillonite has high colloidal activity, that is, high plasticity and cohesion, and high swelling and shrinkage. Montmorillonite normally occurs as a fine clay with irregular crystals having an effective diameter of 0.01 to 1 gm. The combination of high specific surface area, cation exchange capacity and swelling potential of montmorillonite makes it attractive for use as a waste barrier material. The interlayer spacing, which can include several water layers, will, however, not respond in a similar fashion in the presence of certain organic pollutants. In essence, when water saturated montmorillonite is exposed to certain organic pollutants, penetration of the organic pollutants, known as intercalation phenomenon, into the saturated montmorillonite occurs easily. 4.7.4
Vermiculites
Vermiculites occur extensively in soils formed as a product of weathering or hydrothermal alteration of micas. The layer structure of vermiculite resembles that of the mica from which the mineral is derived. Due to weathering, the interlayer K + in the micas is replaced by Mg 2+, and the c-spacing expands, in most cases, to 14-15 ,&. An idealized unit formula is [Mg(H20)6]n[(Mg , Fe)3(Si4_n, A1,)O10(OH)2], with the hydrated magnesium cation Mg(H20)62+ serving as the exchangeable cation. The layer charge in vermiculite gives rise to a cation exchange capacity of 120 to 150 meq/100g, which is considerably higher than the exchange capacity of montmorillonite. As with montmorillonite, the cation exchange capacity is only slightly pH-dependent. Vermiculite swells less than montmorillonite because of its higher layer charge. The total surface area of vermiculite ranges from 600 to 800 m2/g. The mineral, with a basal spacing of 10 A, is non-swelling when saturated with K + or NH4+ ions. A summary of some selected properties of vermiculite is shown in Table 4.7.
4.7.5
Chlorites
Chlorites occur extensively in soils and are examples of 2" 1"1 layer silicates. The positively charged mica-like sheet restricts swelling, decreases the effective surface area, and reduces the effective cation exchange capacity of the mineral. An idealized unit formula is [A1 Mg2(OH)6]x [Mg3(Si4.xAlx) O10(OH)2]. Substitution in such classical chlorites is in the tetrahedral layer, with the brucite sheet serving as the interlayer cation, as illustrated in Figure 4.9. The repeating layer has a thickness of 14 A. The layer charge of the 2"1 portion of the mineral is variable but is similar to that of mica. Cation exchange capacity ranges from 10 to 40 meq/100g, and total surface area from 20
78
SOIL SYSTEM
to 150 m2/g. A summary of some selected properties of chlorites is shown in Table 4.7.
si
/
Si
"ZJ _Y
si
/
Si
Figure 4.9. Schematic representation of typical chlorite structure.
4.7.6
Sepiolite and Palygorskite
In sepiolite and palygorskite, the 2:1 layers do not form continuous sheets, but form fibers six (sepiolite) or four (palygorskite) silicon tetrahedra wide, as shown in Figure 4.7. Simplified formulae are [Mg 4Si60!5(OH)2.6H20] and [(Mg, A1, Fe)4Si8020. nH20], respectively. Sepiolite may occur as the pure Mg end member, but most sepiolites and all natural palygorskites contain some aluminum and usually some exchangeable cations. Cation exchange capacity ranges from 20 to 30 meq/100g, and total surface area from 170 to 370 m2/g.
4.7.7
Mixed-layer Clays
The structures of2:1 clays and chlorites are closely related. It is not surprising, therefore, that same minerals contain more than one type of interlayer behaviour. Some layers in a crystal may be of the smectite type and some of the illite type, giving a mixed-layer illite-smectite. Also, regions of gibbsite or brucite may occur between the layers of a smectite or vermiculite, giving what is known as a mixed-layer chlorite-smectite or a hydroxy-interlayer smectite. The different layers may be distributed randomly or may exhibit several types of ordering, making precise identification of mixed-layer structure difficult.
4.7.8
Soil Clays
As mentioned previously, soil clays often differ appreciably in properties from those of the pure minerals described above. Soil clays are usually less well ordered and smaller in size than the pure minerals and often overlap neighboring particles or sheets. Inter-stratifications of various layer silicates are common, and the mineralogy of soil clays is rarely simple or uniform. Coating of iron and aluminum oxides and organic matter on most layer silicates further complicate the mineralogy
SOIL ORGANIC MATTER
79
of soil clays. Such coatings can drastically alter mineral properties by decreasing cation exchange capacity and surface area values and by restricting the swelling and collapsing of expansible minerals. Oxide coatings, however, magnify anion exchange and other properties associated with positively charges surfaces.
4.8
SOIL ORGANIC MATTER
Soil organic matter is an accumulation of partially decayed and partially re-synthesized plant and animal residues. Such material is in an active state of decay, being subjected to continued attack by soil microorganisms. Consequently, much of it is rather transitory and must be constantly renewed by addition of plant residues. The organic matter content of surface mineral soils is usually only about 0.5 to 5% by weight. Soil organic matter can exert a profound effect on the physical and chemical properties of the soil. Physically, it improves aggregation of soil particles, resulting in the development of a stable soil structure. Chemically, it increases the cation exchange capacity, and the water holding capacity of soils. Biologically, soil organic matter is the main source of food and energy for soil organisms. The accumulation of organic matter in soil is strongly influenced by temperature and the availability of oxygen. Since the rate of biodegradation decreases with decreasing temperature, organic matter does not degrade rapidly in colder climates and tends to build up in soil. In water and in waterlogged soils, decaying vegetation does not have easy access to oxygen, and organic matter accumulates. Non-humus organic matter includes those materials that are undecomposed (original tissue) or only partially decomposed. Non-humus substances include carbohydrates and related compounds, proteins and their derivatives, fats, lignins, tannis, and various decomposition products. Non-humus organic matter may also include roots and tops of plants. The degradation products of non-humus materials undergo enzymatic and chemical reactions to form new colloidal polymers called humus. Humus is a generic term for the water-insoluble material that makes up the bulk of soil organic matter (Stevenson, 1994). Humus is composed of a base-soluble fraction (humic and fulvic acids) and an insoluble fraction (humin), and is the residue from the biodegradation (by bacteria and fungi) of plant material. The bulk of plant biomass consists of relatively degradable cellulose and degradation-resistant lignin, a complex polymeric substance that is second only to carbohydrates in natural abundance (Sarkanen and Ludwig, 1971). Humic materials in soil strongly adsorb many solutes in soil water and have a particular affinity for polyvalent catioo~ and interact with the clay minerals. Both the humus and non-humus fractions of soil organic matter are important to the soil environment. Non-humus material provides short-range effects, such as sources of food and energy for microorganisms. Humus provides long-term effects, such as maintaining good soil structure and increasing soil cation exchange, pH-buffering, and water holding capacity. Thus, humus reduces bioavailability.
80 4.9
SOIL SYSTEM CHARGE DEVELOPMENT IN SOILS
The two properties that most account for the reactivity of soils are surface area and surface charge. Surface area is a direct result of particle size and shape. Most of the total surface area of a mineral soil is due to clay size particles and soil organic matter. Charge development in soils is associated with these two fractions, although the sand and silt size fractions may contribute some cation exchange capacity if coarse grained vermiculite is present. A charge develops in soils through isomorphic substitution and ionization of functional groups on the surface of solids that make up the soil matrix. These two mechanisms give rise to the constant surface charge minerals and the constant surface potential (pH dependent charge) clay minerals. The separation is not a rigid one because a single soil mineral can exhibit both types. 4.9.1
Constant Surface Charge Minerals
A perfectly formed crystal lattice would possess no excess charge at the surface because all atoms in the crystal would be electrically balanced. Imperfections in the lattice structure, however, cause an excess of positive or negative charge, which is then compensated for by the accumulation of oppositely charged ions (counter ions) at the crystal surfaces. Such an imperfection, for instance, might be the substitution of the trivalent aluminum atom in a silicate sheet, which would lead to an excess of negative charge at the particle surface. The substitution of trivalent aluminum for divalent magnesium would lead to an excess of positive charge at the surface. This type of substitution is called isomorphous substitution. As this defect occurs in the interior of the crystal lattice, the resulting charge imbalance is permanent and cannot be influenced by external factors such as the pH of the ambient solution. Hence, we have a constant surface charge mineral. 4.9.2
Constant Surface Potential Minerals
In this general type, surface charge is created by the adsorption of ions onto the surface, the net charge being determined by that ion which is adsorbed in excess. The charging process requires the presence of these ions, called potential determining ions, in the ambient solution in quantities sufficient for adsorption. The primary source ofpH-dependent charge is considered to be the gain or loss ofH + from functional groups on the surfaces of soil solids. The functional groups include hydroxyl [-OH], carboxyl [-COOH], phenolic [-C6H4OH], and amine [-NH2]. The charge that develops from functional groups depends largely on the pH of the ambient solution, which regulates the degree of protonation or deprotonation of the functional group. The soil solids that contain functional groups capable of developing pH-dependent charge include layer silicates, oxides and hydrous oxides, and soil organic matter.
Protonation of Exposed OH Groups Exposed OH groups are present on the surface of A1 octahedral sheets. They are prevalent in 1:1 types of clays, oxides, and amorphous soils. These OH groups are in contact with the soil solution and tend to protonate due to the addition of H+ ions. This process contributes to the oversaturation of the OH groups with protons, thus rendering the clay surface positively charged, as
CHARGE DEVELOPMENT IN SOILS
81
illustrated by the following reaction: - A I - O H + H* -~ - A I - O H H + neutral positively charged octahedral octahedron
[4.2]
Protonation of exposed OH groups occurs only at low pH, because acid conditions are required for the supply of the extra proton. The positively charged octahedron will contribute to the increased mobility of cations through the soils, hence increasing the potential of polluting the groundwater.
OH
/
/
Si
AI
(+1/2)
~
H
(+1)
~ +
/
H
Acid
(-1)
Si
(+1/2) OH
+2OH" ~
AI
/ (+ 1/2) OH
0
/
Si OH
/
OH
~
(-1/2) 0 § 2H20
/ AI
OH (-1/2) Neutral pH
OH
(-1/2)
Basic
Figure 4.10. Representation of pH-dependent charge at kaolinite edges.
Deprotonation o f Exposed O H Groups
Since OH groups are in contact with soil water, they tend to dissociate (deprotonate), and release their protons, as illustrated in the following reaction: - A I - O H -~ - A I - O - + H + neutral negatively charged octahedral octahedron
[4.3]
The dissociation of H + leaves one non-neutralized negative charge in the octahedron. Such a dissociation reaction occurs at high pH, and decreases at low pH. The magnitude of the negative charge also increases and decreases accordingly with the change of pH. Therefore, this type of
82
SOIL SYSTEM
negative charge is called pH-dependent or variable charge. Figure 4.10 illustrates the pH-dependent charge at kaolinite edges. The acidity of OH groups can be characterized by using the dissociation constant, pKa, with a value of 5.0 assigned to the AI(OH2)+i group, 7.0 to the (A1-OH-Si) +~ group, and 9.5 to the SiOH group (Sposito, 1989). The high pKa value for SiOH groups indicate that their deprotonation occurs only at high pH. Thus, variations in pH-dependent charge of layer silicates are more likely associated with reversible protonation and deprotonation of exposed A1OH groups. Hydroxyl ions exposed on planar surfaces of minerals are also characterized by high pKa values and contribute pH-dependent charge only at high pH. pH-dependent charges are more important for kaolinite than for smectites, illite and vermiculites. As a rule of thumb, only 5 to 10% of the negative charge on 2:1 layer silicates is pHdependent, whereas 50% or more of the charge developed on 1:1 minerals can be pH-dependent.
Zero Point of Charge The zero point of charge (ZPC) is the pH at which a mineral has no charge, or has equal amounts of negative and positive charges. It is similar in meaning to the isoelectric point. As previously discussed, at high pH values the mineral carries a negative charge, which decreases with a decrease in pH. When the pH is continuously decreased, a point will be reached at which the negative charge equals zero. The pH at which this occurs is the ZPC. Typical values of ZPC, of selected minerals, are shown in Figure 4.11. The ZPC is a specific characteristic of the clay mineral, and its value differs from one mineral to another. When the net charge is zero, at the ZPC, clay particles in soil water will not repel each other but will tend to aggregate and form larger particles. This in turn will contribute to an increase in soil hydraulic conductivity and transport of pollutants through soils. In contrast, negatively charged clay particles repel each other, resulting in dispersion and a decrease in soil hydraulic conductivity.
4.10
SURFACE FUNCTIONAL GROUPS The common surface functional groups on inorganic solids are discussed below (Sposito,
1989).
Lewis Acid Site The combination of metal cation and water molecule at an interface is a Lewis acid site, with the metal cation identified as the Lewis acid. For example, at the periphery of gibbsite mineral, water molecules are bound to A13+ions, which result in a positive charge. Lewis acid sites can exist also on the surface of geothite if peripheral Fe3+ions are botmd to water molecules there. Thus, any metal hydrous oxide, as well as the edge surfaces of clay minerals like kaolinite, can expose Lewis acid sites to the soil solution. These surface functional groups are very reactive, since the positively charged water molecule is unstable and is exchanged readily for an organic or inorganic anion in the soil solution, which then can form a more stable bond with the metal cation.
SURFACE FUNCTIONAL GROUPS
83
Hydroxyl Group The inorganic surface functional group of greatest abundance and reactivity in soil clays is a hydroxyl group that is exposed on the outer periphery of a mineral. This kind of OH group is found on metal oxides, oxy-hydroxides, and hydroxides on clay mineral and on amorphous silicate minerals like allophane. In the case of soil organic matter, the surface functional groups are organic molecular units. But in general they can bound to either organic or inorganic solids, and they can have any molecular structural arrangement. The main functional groups are hydroxyl [-OH], carboxyl [-COOH], phenolic [-C6H4OH], and amine [-NH2].
Hematite
Kaolinite
Amorphous Iron
Gibbsite to to
2.1
1
2
. I
1
3
4
8.5~ J
5
6
I
7
8
pH Figure 4.11. Zero point of charge (ZPC) values of selected minerals.
4.11
S U M M A R Y AND CONCLUDING R E M A R K S
Soil is a multi-component system consisting of solid, liquid, and gaseous phases, and living organisms. The solid phase of soils consists of both inorganic and organic components. Inorganic components exert a tremendous effect on the physical and chemical properties, such as cation exchange capacity and surface area, and on the overall suitability of soil as a barrier for waste containment. The organic components, although normally present in much smaller quantities than inorganic components, may significantly alter soil properties. The variability of these separate soil components and pore fluid chemistry will impact on the nature of solid-pore fluid interaction mechanisms, adsorption capacity, and fluid transport properties such as hydraulic conductivity, diffusion and dispersion. These mechanisms and properties are important in evaluating the fate of chemical substances in the terrestrial ecosystem and determining the proper clay mixture for designing waste containment barrier systems. From a soil cleanup viewpoint, evaluation of the effectiveness of a decontamination procedure can be achieved from a closer consideration of how the pollutants are retained in the organic and inorganic solid phases.
This Page Intentionally Left Blank
CHAPTER
FIVE
SOIL-WATER-POLLUTANT INTERACTION
5.1
INTRODUCTION
The importance of soil-water-pollutant interaction in relation to geoenvironmental engineering design can be fotmd in the control of pollutant transport (mobility) through soils and in the changes of soil properties. Of significance to the problem of waste containment is the need to provide a low hydraulic conductivity barrier to the transport of waste leachates. From the development of soil-water-pollutant interaction, various physico-chemical parameters such as soil type, density, void ratio, water content, specific surface area, cation exchange capacity, pollutant concentrations and solution pH will impact on hydraulic conductivity. All soil components contribute in some measure to cation exchange between the solid surface and bulk solution. The source, extent, and quality of the contribution of each soil component to the cation exchange capacity is varied and complex. Reactions between components alter the extent and quality of cation exchange. Also, reactions are strongly influenced by the pH of the soil-waterpollutant system, and the electrolyte content of the soil solution. Studies of adsorption processes take into account the mechanisms of binding of the pollutant to the soil surfaces. If the mechanism of adsorption is to be understood, it is necessary to have information about: (1) kinetic aspects of the interaction, specially the rates of transfer of pollutants to the clay surface, (2) interactions which take place at the surface, (3) initial and final equilibrium conditions, and (4) composition and structure of the final adsorption complex. Therefore, the following development of soil-water-pollutant interaction will be concerned with attributes and characteristics which are pertinent to waste containment barrier systems and decontamination of polluted soils.
5.2
ADSORPTION MECHANISMS
Adsorption is the net accumulation of matter at the interface between a solid phase and an aqueous solution phase. The matter that accumulates at an interface is the adsorbate. The solid surface on which it accumulates is the adsorbent. A molecule or an ion in the soil solution that can potentially be adsorbed is termed adsorptive. Adsorption on soil particle surfaces can take place via the following mechanisms (Sposito, 1984):
The Inner-Sphere Surface Complex An inner-sphere surface complex is defined as the complex that is obtained when no water molecule is interposed between the surface functional group and the ion or molecule it binds. An example of inner-sphere surface complex is shown in Figure 5.1 (Sposito, 1984) where K + ion is 85
86
SOIL- WATER- POLLUTANT INTERACTION
coordinated with 12 oxygen atoms bordering two opposing siloxane cavities. The layer charge in soil vermiculite is large enough that each siloxane cavity in a basal plane of the mineral can complex a K § cation. Moreover the ionic radius of K § is almost equal to that of a cavity which gives Kvermiculite surface complexes great stability in soils, and is the molecular basis for potassium fixation.
Figure 5.1. Inner-sphere surface complex of K § on vermiculite.
Inner-sphere surface complexes involve either ionic or covalent bonding, or some combination of the two. Since covalent bonding depends significantly on the particular electron configurations of both the surface group and the complex ion, inner-sphere surface complexation is termed specific adsorption. Adsorption in this case occurs in the siloxane surface, which is defined as the plane of oxygen atoms on the surface of a 2:1 layer silicate, as illustrated in Figure 5.1. This plane is characterized by a distorted hexagonal symmetry with its constituent oxygen atoms. The functional group associated with the siloxane surface is the hexagonal cavity formed by six comer-sharing silica tetrahedra, as shown in Figure 5.2. This cavity has a diameter of about 0.26 nm and is bordered by six sets of electron orbitals emanating from the surrounding ring of oxygen atoms. The reactivity of the siloxane cavity depends on the nature of the electronic charge distribution in the layer silicate structure. If there are no neighboring isomorphic cation substitutions to create local deficits of positive charge in the underlying layer, the siloxane cavity will function as a very mild electron donor that can complex only neutral, dipolar molecules such as water molecules. The complexes formed are not very stable, an example being the easily reversed entrapment of a water molecule having one of its hydroxyl groups directed into a cavity perpendicularly to the siloxane surface. If isomorphic substitution of A13§by Fe 2§ or Mg 2+ occurs in
ADSORPTION MECHANISMS
87
the octahedral sheet, the resulting excess negative charge on the nearby siloxane cavity makes it possible to form reasonably strong complexes with cations. If isomorphic substitution of Si 4+ by A13+ occurs in the tetrahedral sheet, the excess negative charge is located much nearer to the surface oxygen atoms, and much stronger complexes with cations and molecules become possible because of this localization of charge.
Figure 5.3. Outer-sphere surface complex of Ca(H20)62§ on montmorillonite.
88
SOIL- WATER- POLLUTANT INTERACTION
The Outer-Sphere Surface Complex An outer-sphere surface complex is defined as the complex that is obtained when at least one water molecule is interposed between the surface functional group and the ion or molecule it binds. Outer-sphere surface complexes involve electrostatic bonding mechanisms and, therefore are less stable than inner-sphere complexes. This type of adsorption is called non-specific adsorption. An outer-sphere complex with a Ca + cation is shown in Figure 5.3 (Sposito, 1984) for the two-layer hydrate of Ca-montmorillonite. Two opposing siloxane cavities complex a Ca 2§ cation solvated by six-water molecules in octahedral coordination.
Diffuse-Ion Swarm (Layer) If a solvated ion does not form a complex with a charged surface functional group, but instead neutralizes surface charge, it is said to be adsorbed in the diffuse-ion swarm. This adsorption mechanism involves ions that are fully dissociated from surface functional groups and are free to move in the soil solution. The diffuse-ion swarm involves electrostatic bonding. Accordingly, there is a weak dependence on the electron configuration of the surface group and the adsorbed ion. This type of adsorption is also termed non-specific adsorption. A typical diffuse-ion swarm is shown in Figure 5.4.
Figure 5.4. Diffuse-ion swarm.
ADSORPTION MEASUREMENTS 5.3
89
ADSORPTION MEASUREMENTS
Adsorption is studied experimentally in soils by means of two basic laboratory operations: (1) reaction of the soil with a solution of known composition at fixed temperature and applied pressure for a prescribed period of time, and (2) chemical analysis of the reacted soil, the soil solution, or both, to determine their composition. The reaction in step 1 can take place either with the solution mixed uniformly with the soil particles (batch process) or with the solution in uniform motion relative to a column or pad of soil particles (leaching column test). The reaction time should be long enough to permit a detectable accumulation of the adsorbate, but short enough to avoid unwanted side reactions, such as redox, precipitation, or dissolution reactions. In batch processes, the chemical analysis in step 2 is usually carried out after the isolation of the soil from the reacted solution by centrifugal or gravitational force. Some of the reactant solution will always be entrained in the soil in this kind of separation. In leaching column test, the composition of the effluent solution is analysed to determine the changes caused by adsorption. The moles, Cf, of chemical species i adsorbed per kg of dry soil is calculated with the equation: Cs i = (C o` - Ce i) w c
[5.1]
where Co' is the total moles of species i per kg of dry soil in the reactant aqueous solution ( initial input concentration), Ce ~ is the chemical species i in the supernatant solution, i.e., equilibrium solution (batch test), or in the effluent solution (leaching column test) in moles per kg of water, and wc is the gravimetric water content, i.e., total mass of water per mass of soil (kg water per kg of dry soil).
5.4
METAL CATION ADSORPTION
5.4.1
Metal Cation Adsorption by Soil
Metal cations are adsorbed onto soil particle surfaces through the various interaction mechanisms described earlier. The relative affinity that a given metal cation has for a soil adsorbent depends in a complicated way on the soil solution composition. But, as a first approximation, the selectivity of a soil for an adsorptive metal cation can be rationalized in terms of inner-sphere and outer-sphere surface complexation and diffuse-ion swarm concepts. The relative order of decreasing interaction strength among the three adsorption mechanisms is: inner-sphere complex > outer-sphere complex > diffuse-ion swarm. The electronic structure of the metal cation and surface functional group are important for the formation of the inner-sphere surface complex. For the diffuse-ion swarm only the metal cation valence and surface charge are critical to determining adsorption affinity. The outer-sphere surface complex is intermediate, in that valence is probably the most important factor. As a rule of thumb, the relative affinity of a soil adsorbent for a free metal cation will increase with the tendency of the cation to form inner-sphere surface complexes. For a series of metal cations of a given valence, this tendency is correlated positively with the ionic radius for the following reasons:
90 (1)
(2)
SOIL- WATER- POLLUTANT INTERACTION For a given valence, z, the ionic potential (z/r), where r is the ionic radius, decreases with increasing ionic radius. This trend implies that metal cations with larger ionic radii will create a smaller electric field and will be less likely to remain solvated in the face of competition for complexation by a surface functional group, and A larger ionic radius implies a larger spread of the electron configuration in space and a greater tendency for a metal cation to polarize (distort) in response to the electric field of a charged surface functional group. This polarization is the necessary prerequisite for the distortion of the electron configuration leading to covalent bonding.
Given these considerations, the relative adsorption affinity series the basis of ionic radius is: Cs + > Rb + > K + > Na + > Li + > Ba R+> Sr 2+ >
C a 2+ >Mg 2+ >
(selectivity sequence),
Hg 2+ > Cd 2+ >
on
Z n 2+
With respect to transition metal cations, however, ionic radius is not adequate as a single predictor of adsorption affinity since electron configuration plays a very important role in the complexes of these cations (e.g., Mn 2§ Fe 2§ Ni 2§ ). Their relative affinities tend to follow the IrvingWilliams order: Cu 2+ > Ni 2+ > Co 2+ > Fe 2+ > Mn 2+
Figure 5.5(a). Amounts of lead, copper, zinc, and cadmium retained as a function of pH for Domtar Sealbond, for a mixed metallic ion solution at total concentration of 4 cmol/kg soil.
METAL CATION ADSORPTION
91
The effect of pH on metal cation adsorption is principally the result of changes in the net proton charge on soil particles. As pH increases, surface charge decreases toward negative values, and the electrostatic attraction of a soil adsorbent for a metal cation is enhanced. If a soil is reacted with a series of aqueous solutions containing a metal cation at the same initial concentration but having an increasing valence, the amount of metal cation adsorbed will increase with pH. The relationship between metal cation adsorbed and the solution pH is characterized by a sigmoid shape known as an adsorption edge. Adsorption edges of Pb, Cu, Zn, and Cd are shown in Figure 5.5(a) for Domtar Sealbond (an illitic soil ) and Figure 5.5(b) for Avonlea bentonite (Yong, Mohamed and Warkentin, 1992). Often these curves are characterized numerically by the value of pHs0, the pH at which one-half the maximum adsorption is achieved. It is observed typically that pHs0 correlates negatively with the relative affinity of the soil for the metal cation. For example, from the results shown in Figure 5.5(a), pHs0 is higher for Cd than Zn, Cu, and Pb, respectively.
Figure 5.5(b). Amounts of lead, copper, zinc, and cadmium retained as a function of pH for Avonlea Bentonite, for a mixed metallic ion solution at total concentration of 4 cmol/kg soil.
5.4.2
Metal Cation Adsorption by Soil Constituents
Determination of metal cation adsorption by soil constituents is most often accomplished through selective sequential extraction (SSE) of the contaminants from soil samples (Yong, Mohamed and Warkentin, 1992). The basic use of SSE is its use of appropriate chemical reagents in a manner that releases the different heavy metal fractions from the soil solids by destroying the bonding between the metals and the soil solids. It should be noted that this method of analysis is not precise. However, it provides a qualitative appreciation of the capability of the various soil constituents to retain heavy metals.
92
SOIL- WATER- POLLUTANT INTERACTION
Figure 5.6(b). Amounts of Pb retained by the various soil constituents and mechanisms as a function of pH for Avonlea bentonite.
The presence of carbonates in a soil contributes significantly to the retention capacity of the soil. This is demonstrated in Figure 5.6(a) for Domtar Sealbond (an illitic soil) which contains about 15% carbonates. At low pH values (pH < 4), dissolution of carbonates occurs and, hence the
ADSORPTION EQUILIBRIUM
93
retention of lead by carbonates will be minimal. For a soil system which is a mono-mineral clay, e.g., montmorillonite, we would expect cation exchange to be the dominant mechanism for Pb retention, as shown in Figure 5.6(b). When the pH of the soil solution is greater than 4, lead is retained in the two representative soils by hydroxide, carbonate and as exchangeable cations. As the pH of the system is reduced (below the precipitation pH of lead), we would expect that the dominant mechanism for lead retention would be via sorption mechanisms. When the pH of a soil polluted with lead reaches values close to 4, we can anticipate that precipitated and hydroxy species will form. In the case of Pb, PbOH § and Pb(OH)2 ~ are formed. It has been observed that the pH at which the lead begins to be retained by the hydroxides in the soil is lower than the pH at which the lead begins to precipitate, and also lower than the pH at which formation of metal hydroxy species begins to occur. For example, in Figure 5.6(a), retention of lead in the soil, in the hydroxide form, begins when pH is greater than 3. Calculations from the MINTEQ Geochemical Model (Phadungchewit, 1990) show that lead begins to form hydroxy species at about pH 6.
5.5
ADSORPTION EQUILIBRIUM
The relationship between the adsorbed concentrations, Cs~, and the equilibrium concentrations, C j, at fixed temperature and applied pressure is portrayed in an adsorption isotherm. Adsorption isotherms are convenient for representing the effect of adsorptive concentration on soil surfaces, especially if other variables such as pH and ionic strength are controlled along with temperature and pressure. The four categories of adsorption isotherms commonly observed in studies of soils are the S-, L-, H-, and C-curve isotherms.
Figure 5.7. Adsorption isotherm of copper for a clay loam soil.
94
SOIL- WATER- POLLUTANT INTERACTION S-Curve Isotherm
The S-curve isotherm is generally characterized by an initially small slope that increases with adsorptive concentration. This behaviour suggests that the affinity of the soil particles for the adsorbate is less than that of the aqueous solution for the adsorptive. For example, in copper adsorption on a clay loam at pH 5.1 and 25 ~ C, shown in Figure 5.7 (Holford et al., 1972), the Scurve is thought to result from competition for Cu 2§ ions between soluble organic matter and the soil particles. Once the concentration of Cu2§ exceeds the complexing capacity of the organic ligand, the soil particle surface gains in the competition and begins to adsorb copper ions significantly. Thereafter, the isotherm takes on its characteristic S-shape. In some instances, especially when organic compounds are adsorbed, the S-curve isotherm is the result of cooperative interactions among the adsorbed molecules. These interactions cause the adsorbate to become stabilized on a solid surface and, thus produce an enhanced affinity of the surface for the adsorbate as its concentration increases. L-Curve Isotherm
The L-curve isotherm is generally characterized by an initial slope that does not increase with the concentration of adsorptive in the soil solution. This type of isotherm is the resultant effect of a high relative affinity of the soil particles for the adsorbate at low concentration. In addition, as the concentration of the adsorbate increases, the amount of the remaining adsorbing surface is decreased. For example, in lead adsorption on kaolinite and bentonite soils shown in Figure 5.8 (Yong, Mohamed and Warkentin, 1992), the isotherm is concave to the concentration axis because of the high relative adsorption affinity.
Figure 5.8. Adsorption isotherm of lead for kaolinite and Avonlea bentonite.
ADSORPTION EQUILIBRIUM
95
H-Curve Isotherm
The H-curve isotherm is characterized by a large initial slope suggesting a very high relative affinity of the soil for an adsorbate. This condition is usually produced either by inner-sphere surface complexation or by significant van der Waals interactions (Sposito, 1984). Figure 5.9 shows the results of low cadmium concentration adsorption by kaolinitic soil at pH 7.0 and 25 ~ C (GarciaMirayagaya and Page, 1978). The observed H-curve isotherm is believed to be the result of specific adsorption (Sposito, 1984). Large organic molecules and inorganic polymers (e.g., aluminum hydroxy polymers) provide examples of H-curve isotherms resulting from van der Waals interactions.
Figure 5.9. Cadmium adsorption isotherm for Boomer loam (kaolinitic) soil.
C-Curve Isotherm
The C-curve isotherm is characterized by an initial slope that remains independent of adsorptive concentration until the maximum possible adsorption is achieved. This kind of adsorption can be produced either by a constant partitioning of an adsorptive between the interfacial region and the soil solution, or by a proportionate increase in the amount of adsorbing surface as the concentration of an adsorbate increases. An example of C-curve isotherm is shown in Figure 5.10 (Chiou et al., 1979). The figure shows the adsorption isotherms of 1,1,1 trichloroethane, 1,1,2,2 tetrachloroethane, and 1,2 dichloroethane in porous media. The L-curve isotherm, shown in Figure 5.8, is by far the most commonly encountered in soils. The mathematical description of this isotherm almost invariably has involved either the Langmuir equation or Freundlich equation. The Langmuir equation has the form:
96
SOIL- WATER- POLLUTANT INTERACTION
abC ~
C~s --
[5.2] 1 + bCe ~
where Cs' is the moles of chemical species i adsorbed per kg of dry soil, Ce~is the chemical species i in the supernatant solution, i.e., equilibrium solution, and a and b are material parameters. The parameter a represents the value of C,' that is approached asympototically as Cj becomes large. The parameter b determines the magnitude of the initial slope of the isotherm. The Freundlich isotherm equation has the form: Cs
= g d
Ce
[5.3]
where Kd and n are material parameters that depend on the constituents, nature of soil and interaction mechanisms established within the system. These two parameters can be estimated by plotting log C] versus log C] for the range of adsorptive concentrations over which Freundlich equation applies. Then, log K d a n d n are calculated as the y intercept and slope respectively, of the resulting straight line. It should be noted that adsorption isotherm equations cannot be interpreted to indicate any particular adsorption mechanism and should be regarded as c u r v e - f i t t i n g m o d e l s without particular molecular significance. However, they can be used as a predictive tool under limited conditions.
Figure 5.10. Adsorption isotherms of 1,1,1 trichloroethane, 1,1,2,2 tetrachloroethane, and 1,2 dichloroethane.
MOLECULAR ADSORPTION MODELS 5.6
97
MOLECULAR ADSORPTION MODELS
Adsorption isotherms are often interpreted with the help of molecular adsorption models. These models are mathematical representations of adsorption based on hypotheses about the interaction between an adsorptive and an adsorbent that result in a particular arrangement of an adsorbate on a surface. Ideally, the hypotheses underlying a molecular adsorption model are developed from spectroscopic information about the adsorptive-adsorbent interaction and the structure of the adsorbate. Usually, however, the underlying hypotheses come from less complete information that is derived from experimental data on the effects of pH, ionic strength, competitive adsorption, and aqueous solution composition on the adsorption isotherm. The basic features of molecular adsorption models are perhaps best appreciated by the following detailed considerations. 5.6.1
Electric Double Layer Structure
The Double layer on a Flat Layer Surfaces In Chapter 4, it was shown that the clay crystal carries a net negative charge as a result of isomorphous substitution. The net negative charge is compensated by cations which are located on the layer surfaces. In the presence of water, these compensating cations have a tendency to diffuse away from the layer surface since their concentration is smaller in the bulk solution. On the other hand, they are attracted electrostatically to the charged layers. The result of these opposing trends is the creation of a distribution of the compensating cations in a diffuse electrical double layer on the exterior layer surfaces of a clay particle, as shown in Figure 5.4. The compensating cations between the layers are confined to the narrow space between opposite layer surfaces. The compensating cations act as the counter-ions of the double layers, and, like all counter-ions, they are exchangeable for other cations. At the same time, negatively charged anions will be repelled by electrostatic forces, with diffusion forces acting in the opposite direction so that there is a deficit of anions close to the surface. The electric double layer on the clay layer surfaces has a constant charge which is determined by the type and degree of isomorphous substitution, as discussed in Chapter 4. Therefore, the layer surface charge density is independent of the presence of electrolytes in the suspension. The Double Layer on the Edge Surfaces of Clay Particles The plate-like clay particles have flat layer surfaces as well as an edge surface area. The atomic structure of the edge surfaces is entirely different from that of the flat layer surfaces. At the edges of the plates, the tetrahedral silica sheets and the octahedral alumina sheets are disrupted, and primary bonds are broken. On such surfaces, an electric double layer is created by the adsorption of potential determining ions. This double layer may become more positive with decreasing pH, and its sign may be reversed with increasing pH. Despite the fact that the net charge on the clay particles is always negative, the existence of a positive edge double layer may not be excluded. It may be concluded that the double layer structure of the clay particle is complicated by the fact that two crystallographically different surfaces are exposed by the plate-like particles, each carrying a different type of electric double layer.
98
5.6.2
SOIL- WATER- POLLUTANT INTERACTION
Gouy- Chapman Model
On the basis of the above discussion, a simple model was developed by Gouy (1910), and independently by Chapman (1913), relating the density of charge on the surface to the electrical potential across the double layer. The model was based on the following assumptions: (1) The adsorbent surface is a uniform plane of charge density; (2) The adsorptive ions are point species that interact mutually and with the adsorbent through the coulomb force. Their only mechanism of adsorption is the diffuse-ion swarm; and (3) The aqueous solution phase is a uniform continuum of dielectric constant in which the pointion adsorptive is immersed. When equilibrium is established in the double layer, the average local concentration of ions at a distance x from the surface can be expressed as a function of the average e l e c t r i c p o t e n t i a l at that distance according to Boltzmann's theorem: - = no
EXP
[5.4]
z
_ z+e
n * :
no* E X P
17~
[5.5]
kT
where n' and n are the local concentrations of the positive and negative ions, and n o a n d G are the concentrations of positive and negative ions far away from the surface, i.e., in the pore fluid. When the surface potential, ~ , is negative n' > n and when the surface potential is positive n- > n +. The concentrations are expressed as number of ions per cm 3. The valences of the ions are z + and z-, e is the elementary charge, k is Boltzmann constant, and T is the absolute temperature. The local density charge, p, is given by: P
: z*en+
- z-en-
[5.6]
The local charge density and the local electric potential are also related by Poisson's equation as: d2~ dx 2
[5.7]
where c is the dielectric constant. Combining Eqs. [5.4] to [5.7], we obtain the fundamental differential equation for the double layer:
MOLECULAR ADSORPTION MODELS
d2~
_
sinh
8nnze
NX 2
99 [5.8]
kT )
E
It is convenient to rewrite this equation in terms of the following dimensionless quantities: y -
ze~
;
ze~
Z-
kT
o.,
~ - r.x
[5.9]
kT
where K2 ._ 8 n n e 2Z 2
cm
-2
[5.10]
ekT
Then, Eq. [5.8] becomes a2y
d~
sinh y
-
[5.11]
Integrating once with the boundary conditions =
~;
dy
d~
_
O;
y = 0
[5
12]
we obtain dy
a~
_ -(2 cosh y - 2) 7 - -2 sinh
[5.13]
The second integration with the boundary conditions
-0,
~' -~'o,
y = Z
[5.14]
1 ) e -~ 1 ) e -~
[5.15]
yields e y/a = e z/2 + 1 + (e z / 2 -
e z/2 + 1 - ( e z / 2 -
This equation describes the decay of the potential as a function of the distance from the
100
SOIL- WATER- POLLUTANT INTERACTION
surface at a given surface potential and electrolyte concentration. It approximately represents an exponential decay. It is of interest to consider an approximation which is valid for small surface potentials less than or equal to 25 mV. The fundamental differential equation becomes d2~
_ K2~;
9 = 1ti exp (- r,x)
dx 2
[516]
o
9
The decay of the potential with distance from the surface is now purely exponential. The centre of gravity of the space charge coincides with the plane r,x = 1 or x = 1/K. Hence, 1/K may be called the thickness of the double layer. It is equal to the characteristic length in the Debye-Huckel theory of strong electrolytes. Finally, the total double layer charge can be computed from the potential function as follows:
o
=
-
f
=
0
4~
f 0
dx 2
=
-
4~: - ~
=o
[5.17]
Apparently, the surface charge is determined by the initial slope of the potential function. The result is:
oo =
(;)'
2n kT -i
sinh
ze ~o 2kT
[5.18]
where 0 o is surface charge density (esu cm2), n is counter ion concentration in the equilibrium solution (ions cm3), E is dielectric constant, k is Boltzmann constant (ergs ~ ), T is absolute temperature (~ z is counter ion valence, e is electronic charge (esu), 1tlo is surface potential (mV), tlJ is surface potential variation with distance (mV), and x is distance (cm). Eq. [5.18] shows that the charge on the particle surface is dependent upon the potential difference across the electrical double layer, electrolyte concentration, valence of the counter-ion, dielectric constant (static permittivity) of the medium, and temperature. Sample Problem 1" A flat surface has a constant charge density of 11.7 gC/cm 2. Compute the surface potential at the following electrolyte concentrations, assuming that Gouy theory is valid: 10 -5, 10-3, and 10! MofNaC1, and 10-5, 10-3, and 10l MofMgSO4. Calculate 1/K for each electrolyte concentration. Solution" n - (molarity x 10.3 • Avogardo's number) = M x 10 .3 x 6.02 • 1023 ions/cm 3 kT = 0.4 • 10 -~3 ergs at room temperature e =4.80 • 10 ~~ esu
MOLECULAR ADSORPTION MODELS
101
= 80 = 11.7 gtC/cm2 = (11.7x10 6 / 3.33 x 10 "1~ esu/cm 2 = 3.5 • 104 esu/cm 2 For
1 0 .3
M of NaC1 we find: 1/K =
8~ne2z 2
80 04 10
8~•215
= 0.0093•
-l~
(4.80•176
-
2 •
I
10 -6 cm - 100 ~,
From Eq. [5.18],
[ze~~ : 3.5x104[ sinh [ 2kT
2•215
1017•215
• 10 -13
I
=31.7
Hence, the parameters sinh [ze ~o/ 2 kT], z, e, and kT are known. Then, the surface potential, ~o, can be determined. The calculated surface potential is 208 mV. For the other concentrations and the divalent electrolyte, the calculated results are shown in Table 5.1.
Table 5.1" The dependence of 1/K on the concentration of the outer solution and the valence of the counter-ion Monovalent ions Divalent ions n 1/K Surface potential 1/K Surface potential (mole/l) (m) (mV) (m) (mV) 10.5 10 .3
10 .7 10 .8
322 208
0.5• 10 .7 0.5 • 10.8
10 -1
10 .9
94
0.5• 10 .9
161 104 47
S a m p l e P r o b l e m 2: A flat surface of an illitic clay has a cation exchange capacity of 25 meq/100g of dry soil and specific surface area of 15 m2/g. Compute the thickness of the diffuse ion layer and the surface potential at the following electrolyte concentrations, assuming that Gouy theory is valid: 0, 20, 60, 80, and 100% of ethanol ( C 2 H 6 0 ) in distilled water by volume.
(1)
Solution: Calculation of ethanol concentrations from ethanol fractions: For distilled water, cation (H +) valence is 1 and molecular weight of solvent is 18 g/mole. For
102
SOIL- WATER- POLLUTANT INTERACTION
ethanol, cation valence is 1, molecular weight of ethanol is 46 g/mole, and density is 0.789 g/cm 3. n equals the number of i o n s / c m 3 = Molarity x 10 .3 • Avogardo's number. Assume that distilled water has a pH = 7; hence [H +] = 1 x 10 .7 M. For ethanol, molarity depends on its solubility in water. The following equation applies:
[C] = [Et] + [E,-]
[5.19]
where [C] is total concentration of ethanol added (moles/litre), [E,] is non-soluble ethanol concentration, C2HsOH, (moles/l), and [E,-] is soluble (ionized) portion of ethanol, C2H50 - (moles/l). Electro-neutrality in solution requires that: [H30 ] = [Et- ] + [OH-]
[5.20]
The equilibrium constant of ethanol is calculated as:
Ket = [/-/301 [Et] = 1.26x10-16 [E,]
[5.21]
The equilibrium constant of water is calculated as:
Kw
= [/-/3 O + ]
[OH-]
= 10 -14
[5.22]
[C] is calculated as follows: [C] = g solute / g wt. Solute x liters of solution = (% ethanol x 1000 ml x 0.789 g/mole) / (46 g/mole x 1 liter ethanol solution) From Eqs. [5.19] to [5.22]' Kw
[Et] = [H3 O+] -
[H30 +]
[5.23]
Substituting Eq. [5.23] into Eq. [5.19], we obtain [E,] = [C] - [H30* ] +
Kw [H30 "]
Substituting Eqs. [5.23] and [5.24] into [5.21], we obtain
[5.24]
MOLECULAR ADSORPTION MODELS
103
Xe, = [/-/30+] [/43 0+] - [/43o+] [5.25] If] -
[93 O+] + In3 O+]
Solving for [H3 O+ ] and substituting into Eq. [5.24], we obtain the molarity of ethanol in solution, [E,-]. The calculated concentrations are shown in Table 5.2.
Table 5.2: Calculated concentrations, diffuse ion layer thicknesses and surface potentials Ethanol fraction (%)
Dielectric constant
[C] (mol/1)
[H3 O+ ] (mol/1)
1/K (m)
0 20 40 60 80 100
84 74 62 50 40 32
0 3.43 6.86 10.29 13.72 17.15
1.00x 10 .7 1.02x10 -7 1.04x 10 -7 1.06x10 7 1.08x 10 -7 1 . 1 0 x l 0 "7
9.82x 10 .7 9.12x10 -7 8.26x 10 -7 7.35x10 "7 6.51 x 10 -7 5.77x10 "7
(2)
(3)
Surface potential mV 567 570 574 579 584 589
In the following, the calculation procedure, for distilled water, is illustrated. Determination of diffuse layer thickness: z =1 e = 4.8 x 10 1~ esu kT = 0.4 x 10 13 ergs at r o o m temperature n = 1 x l 0 -T x 6.023 x1023 x 1 0 .3 = 6.023 x 1013 H + / c m 3 1/K = { (84 x 0.4 x 10-13)/(8 x 3.14x 6.023 x 1013 x (4.8 x 10-10) 2 x (1)2)} 1/2 = 98 x 10 -6 cm Determination of surface potential: Surface charge density = Cation exchange capacity / specific surface area = 0.25 (meq/g) / 15 (m2/g) = 1.67 x 10 .2 m e q / m 2 = 1.67 x 10 .9 e q / c m 2 Charge on particle surface = surface charge density x Avogardo's n u m b e r x elementary charge = 1.67 x 10 .9 ( e q / c m 2) x 6.023 x 1023 x 4.8 x 10 l~ (esu) = 4.818 x 105 esu/cm 2
104
SOIL- WATER- POLLUTANT INTERACTION
[.ze~~ : 4.818x105[ n 2kT 2x6.023x1013x84•
sinh [
-13
I
: 42.45x103
Therefore, the surface potential, go = 1.89 • 103 (esu) x 300 • 103 (mV/esu) = 567 mV Similar calculations were performed for the various ethanol concentrations. The results are shown in Table 5.2. It can be seen from the table that as the concentration increases, the diffuse ion layer thickness decreases and the surface potential increases.
Applicability of Guy-Chapman Model The consequences of applying Eq. [5.18] to both constant surface charge and variable charge minerals are discussed below.
Constant Surface Charge As discussed in Chapter 4, surface charge is controlled by lattice defects in the interior of the crystal, so that double layer potential, concentration of electrolyte, and counter-ion valence are not able to influence the sign or magnitude of the surface charge. From Eq. [5.18], it is seen that if the electrolyte concentration or valence is increased, or if the dielectric constant of the medium is increased, the electrical double layer potential must be reduced. This is accomplished by a reduction in the distance that the double layer extends into the equilibrium solution, i.e., the thickness of the double layer is reduced or, as commonly expressed, the double layer is compressed.
Constant Surface Potential In this system, the surface potential is controlled by the adsorption of potential determining ions, which in turn depends on the activity of those ions in the equilibrium solution. Thus, assuming a certain activity in solution, the surface potential shown in Eq. [5.18] is constant. It is assumed that the concentration of the potential determining ions has a negligible effect on the value of concentration. Thus, if the electrolyte concentration or valence, or the dielectric constant, is increased, the effect is to increase the charge on the particle surface. For kaolinite and oxide minerals, with H + and O H as the potential determining ions, the potential is governed by the H § and O H activity in solution, i.e., by the pH. The Nernst equation
go = ~kT In ( -all- / ] : 2.303kT (PHo_PH) e ~ an,,.) e
[5.26]
relates surface potential to pH, where pHo is the pH at which the surface potential is zero. Thus, by combining Eqs. [5.18] and [5.26], we have 1
oo =
2nrkTe ~ sinh 1.15 z (pHo- pH)
[5.27]
MOLECULAR ADSORPTION MODELS
105
Eq. [5.27] is an important equation, for it shows that at some point on the pH scale, the surface charge can be zero. There is, therefore, a pH at which the kaolinite surface has zero point of charge (ZPC), as discussed in Chapter 4. sinh x -~ x when sinh [ 1.15 z (pHo- pH)] is less than one. By substituting for the Debye K (Eq. [5.10]), we obtain: ~
= cK ~o = ~K/2"303kT 4~: e (pH-pH)
[5.28]
The surface charge given by Eqs. [5.18] and [5.27] is equal if ze V = ze [2.303kT ( p H - p H ) = 1.15 z ( p H - p H ) _<1 2kT o 2kT[ e
[5.29]
Eq. [5.27] indicates that over a small range o f p H n e a r p H o ( 0.5 pH units when z =2 and 1.0 pH unit when z = 1), the surface charge is linearly related to (n)v2 and to z. Beyond the specified p H limits, a sinh function would be followed. At this point, it is appropriate to discuss the various parameters in Eq. [5.27] that affect the sign and magnitude of charge on a variable-charge soil colloid. pHo This is the pH value at which equal amounts of H + and O H have been adsorbed onto the hydroxylated surface so that the net surface charge from this source is zero. It is an important parameter in a variable-charge system because it determines the sign of the net surface charge. Thus, if the actual pH of the system is less than pH o (i.e., more acidic than pHo), then the surface is net positively charged and, conversely, the surface is net negatively charged when pH is greater than pHo. Iron and aluminum oxides have relatively high pH o values, usually between pH 7 and pH 9 depending on their composition and degree of crystallinity, so that soils whose colloidal fraction is dominated by these oxides usually exhibit an anion exchange capacity. Silica and organic matter, on the other hand, have low pH o values and therefore increase the cation exchange capacity of variablecharge soils. For such constituents to influence the overall soil charge characteristics, they do not have to be present in large amounts. For instance, coarse textured soils containing less than 5% iron and aluminum oxide have been shown to have pH o values as high as 6, presumably because the oxides coat the siliceous particles and, therefore constitute a large portion of the surface (Gillman and Bell, 1976). pH o corresponds to a point of maximum chemical stability, and at this pH the surface potential is zero. Thus, highly leached siliceous or organic soils with low pH o tend to be more acidic than highly leached soils with high pH o. This tendency of the soil pH to drift toward pH o is termed isoelectric weathering (Mattson, 1932). It should be possible to increase the cation retention capacity of a soil by lowering pH o. This could be accomplished by introducing onto the colloid surface an anion that would impart more negative charge to the surface. Wann and Uehara (1978) demonstrated that increasing the amounts of phosphate to a soil caused the pH o to decrease in a linear manner with an increase in surface net negative charge.
106
SOIL- WATER- POLLUTANT INTERACTION
Soil pH Soil pH influences the magnitude of the net surface charge. For a given counter-ion valence and electrolyte concentration, the value of ( pH o - pH) determines the sign and magnitude of the net surface charge. Thus, raising the pH of a soil by liming increases the soil's cation exchange capacity. It is usually difficult, however, to raise the pH of a variable-charge soil above pH 6.5 if it has high buffering capacity associated with a high specific surface area. A liming material such as calcium carbonate introduces hydroxyl ions into the system by hydrolysis of the carbonate ion, and in many constant-charge soils this causes an increase in pH up to and even surpassing the neutral point. But, a hydroxylated variable-charge surface will release protons to neutralize OH, thus buffering the soil against pH change. Consequently, to make (pH o - pH) more negative, it is sometimes easier to reduce the value of pHo than to raise pH. Electrolyte Concentration Eq. [5.27] predicts that the net surface charge is directly proportional to the square root of the electrolyte concentration. This means that pollutant transport in a variable-charge soil causes the cation exchange capacity to increase for at least as long as the electrolyte concentration remains high. However, the increase of electrolyte concentration decreases the surface potential and increases the surface charge. The lowering of the surface potential is indicated by a change in solution pH. When the surface charge is negative, pH decreases with electrolyte concentration and increases when the surface charge is positive. If there is no change in pH when cations are added to a suspension that is relatively free of salt, one may assume that the net surface charge is zero.
Figure 5.11. Potential distribution as a function of distance from charged surfaces.
MOLECULAR ADSORPTION MODELS
107
Counter-ion Valence Theory indicates that the surface charge and surface potential vary with the valence of the counter-ion. Thus, when pH is one unit higher than pH o the negative charge is increased by more than three-fold, as indicated by the ratio (sinh 1.15• 1.15x l ) = 3.5, when a divalent cation replaces a monovalent cation. This explains why many soils adsorb more cation equivalents from a divalent electrolyte than from an equal concentration of a monovalent electrolyte, all other parameters being equal. Dielectric Constant of the Medium From theory, the surface charge is directly proportional to the square root of the dielectric constant of the solvent solution. Absolute Temperature From theory, the surface charge is directly proportional to the square root of the absolute temperature. However, an increase in temperature results in a decrease in the dielectric constant. Therefore, when temperature increases, the surface charge decreases due to dielectric constant decrease. 5.6.3
Stern Model
The Gouy-Chapman model has only limited use because even with moderate surface potentials (e.g., 250 mV), high values are predicted for the amount of counter-ions adsorbed into the diffuse layer. This is because the Gouy-Chapman model does not account for the actual size of the counter-ions, which are treated as point charges. The Stem model shown in Figure 5.11 attempts to correct this deficiency by allowing the ions to approach the surface to within a certain minimum distance (a few angstroms). Thus, there are counter-ions in a space of width "6", which is defined as Stern layer, between the surface and the first layer of the adsorbed counter-ions. The electrical potential drops linearly from the surface potential to the potential at distance "5". The remainder of the double layer consists of a diffuse layer of counter-ions, as described in the Gouy-Chapman model, in which the potential decays exponentially from the potential at distance "8" to zero. The charge in the Stem layer was derived by Stem (1924) as:
Nze 01
=
[ ze~ 8 + 1 + ~NA E X P n Mw kT
s
[5.30]
where N is number of adsorption sites available per c m 2 of surface, NA is Avogardo's number, Mw is molecular weight of the solvent, ~6 is potential on the border between Stem and Gouy layers (Stem potential), and ~s is specific adsorption potential of the counter-ions at the surface. For the molecular condenser, we find o -
E1
4~z8
(tI'o - ~8)
[5.31]
108
SOIL- WATER- POLLUTANT INTERACTION
where el is dielectric constant of the medium in the field of the molecular condenser, and 8 is the thickness of the Stem layer. At this point, it should be noted that Stem introduced the concept that some ions might be adsorbed into the compact layer by forces other than those that are purely electrostatic and that the energy required for this is accounted for by the specific adsorption potential. The charge in the diffuse layer is given by 1
ze
02 =
sinh
[5.321
2kT 8
and electrical neutrality requires that [5.33]
(Jo = (J1 + 02
Sample Problem 3: Compute the Stem layer and Gouy layer charge and the Stem potential for a fiat double layer with a constant surface charge of 11.7 gC/cm 2 at a concentration of 10-1, 10-3, 10.5 MNaCI, and 10.3 M MgSO4, when the specific adsorption potential is 0 and 0.1 eV. The following values may be assumed: (1) The thickness of the Stem layer is equivalent to two mono-molecular layers of adsorbed water, or 5 A; (2) The dielectric constant within the Stem layer varies from 3 to 6; and (3) The number of adsorption spots on the surface, N, equals 1015 per cm 2. Solution o = 11.7 gC/cm2 = (11.7x106/3.33 x 10-1~ esu/cm 2 = 3.5 x 104 esu/cm 2 By using Eq. [5.31 ] /
o
\
- ~ 8 = 3"5x104 / 4 ~ x 5 x 1 0 - 8 ] = 3.67x10-3 ~, 6 ]
let
Y8
-
ze~ kT '9
esu = 1100 mV
2n;kT 3;1 X =
B
= NlzeM n NA
Since 1 is small with respect to the exponential term in the denominator, Eq.[ 5.30] becomes:
oI=B ExP[YSs
[5.34]
MOLECULAR ADSORPTION MODELS
109
Assuming that the specific adsorption is zero, then for 10 .3 NaC1 B
= {1015 x
1 x 4.80 x 10 -lo x 18
x 10 -6 x
6.02 x 1023 }/{6.02 x 1023}
=8.5 and Eq. [5.32] becomes O2
[5.35]
= X sinh [ - ~ /
where X
= [{2 x 80 x 0.4 =1110 and Eq. [5.33 ] becomes 3
x 10 -13 x 10 -6 x
6.02
x 10 23} / { 3 . 1 4 } ] 1/2
x104 = 1 , , 0
from which Y8 = 6.9, and the surface potential at the Stem layer = 86 mV. The ratio of charges in Stern and Gouy layers is about 50:50.
Table 5.3" Stem and Gouy potential variations with concentrations
Concentration
Stem potential (mV) Monovalent Divalent
Ratio of Stem: Gouy potentials Monovalent Divalent
104 M 10 -3 M 10-1M
300 185 70
37 " 63 38 " 62 40"60
144 86 30
50 " 50 50" 50 53 " 47
In the same manner, the values shown in Table 5.3 are obtained for the three indicated concentrations of monovalent and divalent electrolytes. For the case of specific adsorption, the specific adsorption potential = 0.1 eV = 0.1 x 1.6x 10 -12 erg. The calculated Stem potential for 10.3 MNaC1 is 105 mV, which results in a ratio of 88:12, that is, an increase of adsorption in the Stern layer. 5.6.4
Other Models
There have been a number of further modifications to double layer theory in attempts to more fully explain experimental observations. Perhaps the most notable is that of Grahame (1947) who,
110
SOIL- WATER- POLLUTANT INTERACTION
working with mercury electrode, postulated that anions are specifically adsorbed into the Stem layer when they lose some of their water of hydration whereas the hydrated cations are only electrostatically attracted to the surface. Further modifications to Grahame model were introduced by Yong and Mohamed (1992), and Yong, Mohamed and Warkentin (1992). These modifications will not be discussed.
NH 3
\
NH 3
I /
NH 3 ~{
M n+
/
NH 3
I \ NH 3 ~{
NH 3
(A) Complexation R " ~ ~
~i{
o-COO-
Organic ions Metal ion with charge n Organic ions
mechanism
M n+
Humic acid molecule
,.--
M~
_
Chelate ring
(B) Chelation mechanism
Figure 5.12. Interaction of naturally occurring organic matter with metal ion: (a) complexation mechanism; (b) chelation mechanism.
5.7
ORGANIC POLLUTANT-SOIL ORGANIC MATTER INTERACTION
Before we proceed with the nature of the interaction process, comlexation and chelation mechanisms need to be explained. Complex formation is the reaction of a metal ion and a ligand, through electro-pair sharing (Murmann, 1964; Mellor, 1964). The resulting product is called a metal coordination compound. The metal ion is the electron-pair acceptor, and the ligand is the electronpair donor. The metal ion serves as the central ion, and the organic ions are coordinated around it, as shown in Figure 5.12(a). The number of ligands bonded to the central atom in a definite geometry is called the coordination number. Some of the organic ligands can bind the metal ion with more than one donor functional group, as shown in Figure 5.12(b). This type of bonding forms a heterocyclic ring called a chelate ring. The process of formation of a chelate ring is called chelation.
ORGANIC POLLUTANT-SOIL ORGANIC MATTER INTERACTION
111
Several naturally occurring soil organic acids are capable of complexing metal ions. The reaction occurs mostly with the transition metals, A1, Fe, Cu, Zn, and Mn, and is often considered a special case of an adsorption process. This kind of adsorption is quite different from the regular coulombic adsorption of cations in an electric double layer. The metal ion adsorbed in a complex reaction cannot be exchanged rapidly in the traditional manner of cation exchange reactions. The bonds in a complex compound are covalent bonds, hence they are stronger bonds than the electrostatic bonds in cation exchange reactions. However, an exchange of the complexed metal is still possible, but such an exchange depends on many factors (e.g., soil pH, affinity of the metals for the ligand, and stability of the complexes). The exchange will take place more rapidly between transition metals. For example, an exchange of complexed A1 by free Fe occurs more easily than by Na ions. The exchange by Na ions is very difficult since the Na ion cannot occupy the center position of the A1 ion in the complex.
NH3 §
NH3 §
I CH3 " " C - - C O O H
NH 2
I ~
I
CH3 - " C - - COO--"'q["~m~3H3 --' C '-- C O O -
l
I
I
H
H
H
C ation
Acidic condition pH < Z P C
D ipola r
Neutral condition PH = Z P C
A nion
Alkaline condition PH > Z P C
Figure 5.13. Amino acid ion species formation with changes in soil pH.
The presence of functional groups, such as carboxylic and amino groups, in humic compounds gives them the ability to exist either as a cation, anion, or dipolar (Bjerrum, 1923). The dominant ion species present in the soil solution depends on the soil reaction, as shown in Figure 5.13 for amino acid. The ion species in which the amino acid exists govern the interaction with soil. In an acidic soil, or at pH values below ZPC, amino acids are usually positively charged. As such, the cationic form of amino acid can be attracted to the clay surface by cation exchange. At neutral soil pH, or soil pH close to ZPC, amino acid are dipolar. Although most soils under field conditions are in this pH range, the pH at the clay interface is generally lower than the pH of the bulk solution. When present as a dipolar, amino acid will interact through ion dipole reactions. The positive pole (NH3+) can be attracted directly to the negatively charged clay surfaces. On the other hand, the negative pole of amino acid (CO0-) can also undergo interactions with metal cations adsorbed on the clay surfaces
112
SOIL- WATER- POLLUTANT INTERACTION
or in solution. In alkaline conditions, amino acids are negatively charged and possess anion characteristics. Hence, they have the capability to react with positively charged clay surfaces and metal ions in the soil solution. Organic compounds that contain fimctional groups that become positively charged when in protonated form can also react with soil humus by cation exchange. Basic amino acids, which contain two protonatable NH 2 groups, are good examples of these positively charged compounds. Even without the development of a net positive charge, protonated functional groups like COOH and NH can form hydrogen bonds with electronegative atoms such as O, N, and F. For example, the C=O group in phenyl carbamate can form hydrogen bond (...) with the NH in soil organic matter as C=O...HN. Humus contains carboxyl, hydroxyl, carbonyl, and amino groups in a broad variety of molecular environments that lead to a spectrum of possibilities for hydrogen bonding with organic compounds. Much of the molecular framework of soil organic matter is not electrically charged. This nonionic structure can react strongly with the uncharged part of an organic compound through van der Waals attractions. The van der Waals interaction involves weak bonding between polar units, either permanent (like OH and C=O) or induced by the presence of neighboring molecules. The van der Waals interaction between nonionic compounds, or nonionic portions of compounds, and soil organic matter is often stronger than the interactions between these compounds and soil water (Sposito, 1989). Water molecules in the vicinity of large non-polar molecules are not attracted and so cannot orient their very polar OH bonds in ways that are compatible with normal structure of liquid water (hydrophobic effect). The resultant disorder of this situation produces a low water solubility of the non-polar molecule and a chance for it to react with soil organic matter through van der Waals interactions. This reaction is usually described by a distribution coefficient, Kd:
q Kd -
Ce
[5.36]
where Cs is sorbed concentration (moles per kg of soil), and Ce is equilibrium aqueous solution concentration (moles per kg of solution). The extent of adsorption can be reasonably estimated if the organic carbon content of a soil is known (Karickhoff et al., 1979; Karickhoff, 1984):
Kd = Koc foe
[5.37]
where Koc is a proportionality constant, and foe is the fraction organic carbon content of the soil. This approach works reasonably well for a wide range of soils, providing the soil organic content is sufficiently high (e.g.,fo,. is greater than 0.001). For lower carbon content soils, adsorption of non-polar organic onto soils can cause errors in estimating Koc (Chiou et al., 1985). Ko~ values for many organic compounds are unknown, hence numerous researchers have developed correlation equations to relate Ko~to more commonly available chemical properties such as solubility or octanolwater partition coefficient (Chiou et al., 1982 and 1983; Schwarzenbach and Westal, 1981; Karickhoff, 1981; Kenaga and Goring, 1980), as shown in the following equations:
SOIL ORGANIC MATTER-SOIL MINERALS INTERACTION
113
log K c : 3.64 - 0.55 log S
[5.38]
log K C : 1.377 + 0.544 log K w
[5.391
log Kr : 1.963 + 0.681 log BCF
[5.40]
where S is solubility in moles/kg of solution, Kow is octanol-water partition coefficient, and BCF is
bioconcentration factor. Unfortunately, the linear equilibrium approach to adsorption is not adequate for some realworld situations. For example, trichloroethylene (TCE) adsorption onto glacial till shows a change in K d of more than 50-fold (Johnson et al., 1989; Myrand et al., 1989; Mckay and Trudell, 1978). In addition, for many very hydrophobic organic, adsorption and desorption occur over time scales of many months (Karickhoff and Morris, 1985; Witkowski et al., 1988; Coates and Elzerman, 1986; Wu and Gschwend, 1986).
5.8
SOIL ORGANIC MATTER-SOIL MINERALS INTERACTION
The coating of soil minerals by humus plays a major role in the cycling of chemical elements and in the formation of soil aggregates. Humus bound to clay minerals appears to be relatively stable against biodegradation and so presents a reactive surface to dissolved solutes in the soil solution (Sposito, 1989). The partial dissolution of humus coatings and the release of soluble organic compounds, by soil fauna and flora, increase anion desorption from clay minerals. Thus, soil minerals serve as both a substrate for humus and a source of metal ions for forming soluble humus complexes. The interaction mechanisms between soil organic matter and soil minerals are (Sposito, 1989): (1) CationExchange: This involves, for example, the exchange of the protonated amino group, NH3+, with monovalent exchangeable metal cation initially bound to a clay mineral surface. (2) Protonation: This refers to the association between an organic functional group and a surface-bound proton. The mineral surfaces in soils can develop acidity in a variety of ways (proton exchange, dissociation of hydroxyls, hydrolysis of solvated metal cations, etc.). The surface acidity offers the possibility that proton selective organic functional group like NH 2 and C=O can be bound through a protonation reaction. This mechanism is expected to be most important under conditions of low pH or low water content in soils. (3) Anion Exchange: This involves, for example, the exchange of a carboxylate group (COO-) with a univalent, exchangeable anion (e.g., CI or NO3 ) bound to a protonated surface hydroxyl (OH2+). This mechanism is not observed often because of the weakness of the electrostatic bonds involved. However, it should be prominent in acidic soils containing metal oxides. (4) Water Bridging: This involves complexation with a water molecule. For example, the carboxylate groups in humic substances associate with montmorillonite through cation bridging when monovalent exchangeable cations are present and through water bridging
114
SOIL- WATER- POLLUTANT INTERACTION when bivalent exchangeable cations are present.
(5) (6)
(7)
5.9
Ligand Exchange: This is specific to bond formation between a carboxylate group and either A1 or Fe(III) in a soil mineral bearing hydroxyl groups. This mechanism involves stronger chemical bonds than those involved in anion exchange or in the water bridging mechanism. Hydrogen Bonding: This occurs between organic functional groups and either clay minerals or metal hydrous oxides. It does not appear to be significant in humus reactions with soil minerals, probably because the electro-negativity of mineral surface oxygen usually is not large enough to form strong H bonds. van der Waals Interactions: For polymeric humus material, the van der Waals interactions with the atoms in the mineral surface can be quite strong and relatively long-ranged. The effects of van der Waals interactions are apparent when conditions in the soil solution suppress the ionization of acidic functional groups on large organic molecules, i.e., at ZPC.
INFLUENCE OF POLLUTANTS ON SOIL HYDRAULIC CONDUCTIVITY
The influence of various organic and inorganic chemicals (pollutants) on soil hydraulic conductivity is addressed in this section.
5.9.1
Influence of Inorganic Chemicals
Inorganic acids may dissolve some of the constituents of a clay soil structure. It has been reported that acid dissolves aluminum, iron, alkali metals and alkaline earth (Grim, 1953). Since clay minerals contain alumina in large quantities, they are susceptible to partial dissolution by acids. The solubility of clays in acids varies with the nature of the acid, the acid concentration, the acid to clay ratio, the temperature and duration of treatment (Grim, 1968). The action of an acid on clay is enhanced when the acid has an anion about the same size and geometry as a clay component. This in turn would permit even weak acids to dissolve clays under some conditions. The study conducted by Pask et al., (1954) in which several clay minerals were boiled in acid showed that the percent of dissolution of alumina was 3% for kaolinite, 11% for illite, and greater than 33% for montmorillonite. The role and effect of ionic species and concentration in the soil pore water on the development of soil hydraulic conductivity were demonstrated in the experiments of Pavilonsky (1985) using compacted montmorillonite specimens and nitric, hydrochloric and sulfuric acids, over periods of 200 to 700 days. The thickness of the diffuse ion layer, discussed previously, which depends on the type of chemical species, concentration, dielectric constant of the solvent, temperature and pH, influences the hydraulic conductivity of soils. The results are shown in Table 5.4. With reference to the ratio kI/kw in the third column, kI refers to the final value of the hydraulic conductivity of a specimen permeated with acid while kw refers to the hydraulic conductivity of a compacted specimen permeated with water. The information in the table shows that permeation with acids leads to an approximate 5 to 12-fold increase in the hydraulic conductivity of the montmorillonite specimen in comparison to that determined using water as the permeant solution. Also, a decrease in acid concentration has little effect on the kI / kw ratio. The differences or modifications in soil hydraulic conductivity in relation to the chemistry of the permeating fluid can be attributed to: (1) extraction of lattice aluminum ions from the octahedral sheets of the clay
INFLUENCE OF POLLUTANTS ON SOIL HYDRAULIC CONDUCTIVITY
115
mineral, (2) ion exchange on the surface of the clay minerals due to replacement of naturally adsorbed cations of lower valence (Na +, K § Ca 2§ Mg 2+) by the extracted aluminum ions whose valence is 3. This results in a reduction in the thickness of the diffuse ion layer, and (3) increase in effective pore space and a decrease in the tortuosity factor, resulting therefore in an increase in the soil hydraulic conductivity.
Table 5.4: Variations in clay hydraulic conductivity with inorganic pollutants Permeant
Clay type
ACIDS 7.0% HNO3 0.7% HNO3
Bentonite
5.0%
k// kw
H2SO 4
3.65% HC1 0.36% HC1
NaOH NaOH
CATIONS Valence Effect
Cation Size Effect
5.7 to 7.7 11.6 6.9 10.2 10.9
kf/ kw
BASES
4.0% 0.4%
Hydraulic conductivity variations
Bentonite
0.3 0.2
Montmorillonite Bentonite Bentonite
kca2+/kna + = 28
kca2+/k,f =
8 kFe3+/k,f = 33
at void ratio of 2.5 at void ratio of 2.0 at void ratio of 2.0
Bentonite
kx+/ k,,a+ = 5
at void ratio of 2.0
Source: Yong, Mohamed and Warkentin (1992).
In contaminant leachate solutions containing inorganic bases, mineral dissolution is a significant problem. Inorganic bases increase the net negative charge on clay surfaces, and can also dissolve silica (Grim, 1953). Since clay minerals contain large quantities of silica in their tetrahedral sheets, they are susceptible to particle dissolution by inorganic bases. Replacement of exchangeable sodium by calcium ions could alter the hydraulic conductivity of bentonite considerably. Mesri and Olson (1971) showed a 28-fold increase in the hydraulic conductivity coefficient of montmorillonite clay, for the same void ratio, while results of Sridharan et al., (1986) indicated a 8-fold increase for a similar exchange reaction. This increase in hydraulic conductivity has considerable consequence in the specification of a particular type of bentonite to be used in waste containment systems. The increase in hydraulic conductivity due to exchange of calcium ions for sodium ions can be attributed to a reduction in the diffuse double layer thickness due to replacement of the monovalent sodium ions by divalent calcium ions, and formation of quasi-
116
SOIL- WATER- POLLUTANT INTERACTION
crystals between exchangeable calcium ions and a pair of opposing siloxane cavities (Sposito, 1984). The replacement of monovalent sodium ions by trivalent ions has a more pronounced effect on the hydraulic conductivity of bentonite, the increase being 33-fold. This can be attributed to a reduction in the diffuse double layer thickness due to: (1) replacement of the monovalent sodium ions by trivalent ferric ions, and (2) hydrolysis of the trivalent ferric ions, hence forming ion hydroxy species. The hydroxy species form a coating around the quasi-crystals and influence soil aggregation via electrostatic bonding. Replacement of monovalent sodium ions by monovalent potassium ions leads to an approximate 5-fold increase in the hydraulic conductivity of bentonite at a given void ratio. The increase in hydraulic conductivity is attributed to a reduction in the thickness of the diffuse ion layer arising from higher adsorption of potassium ions in the Stem layer and partial fixation of the cation in the hexagonal holes in the surface silicate layer. The influence of the permeant salt concentration on the changes in the hydraulic conductivity of a soil can be quite significant. In the example shown in Figure 5.14, the hydraulic conductivities of compacted natural clay specimens are shown for different accumulated pore volumes during leaching tests, and for leachate with different pore salt concentrations. The compacted clay permeated with distilled water shows initial decrease during the first few pore volumes, with subsequent relatively constant hydraulic conductivity, as a result of the removal of the natural salts in the soil through leaching of the specimen by distilled water, i.e., decreasing the ion concentration. At dilute ion concentrations, the replusive forces become dominant. Fluid transport through the soil pores must overcome the increased energy field established by the replusive forces. An increase in the replusive energy field could induce self-detachment of soil particles, resulting in a decrease in the hydraulic conductivity. This problem is characteristic of dispersive clays.
Figure 5.14. Influence of salt concentration on hydraulic conductivity of a natural soil; reference hydraulic conductivity before permeation = 3.5x 10-I~ m/s.
INFLUENCE OF POLLUTANTS ON SOIL HYDRAULIC CONDUCTIVITY
117
Increasing the cation concentration in the diffuse ion-layer will result in a decrease of the replusive forces. This can be illustrated from the experimental results shown in Figure 5.14 for clay specimens leached with NaC1 solutions. Hydraulic conductivity increases with increasing NaC1 concentrations in the permeating fluid. The increase in hydraulic conductivity can be largely attributed to the significant decrease in replusive inter-particle forces due to the increase in Na ions in the pore fluid. Another possible consequence of the introduction of high concentrations of NaC1 could be the disruptive effects of Na ions on water structure (Lutz and Kemper, 1959). At high solution concentration, Na ions would compete with the original water structure in order to satisfy their separate hydration shell requirements. This in turn could cause a special kind of water structure breakdown, i.e., destruction of water structure or reduction of its thickness around the clay particles, hence hydraulic conductivity of the soil is increased. The situation for bentonite-sand mixtures such as those used in waste containment systems is, on one hand, easily explained in terms of changes in the replusive forces when cation concentrations are changed (Figure 5.15). On the other hand, it is somewhat more difficult to explain if homogeneous mixing of the bentonite with the sand is not achieved. The results of changes in cation concentration may not easily be perceived when non-homogeneous mixing occurs since preferential flow in soil voids with fewer clay particles will occur.
Figure 5.15. Influence of lead concentration on hydraulic conductivity of 90% sand and 10% bentonite mixture.
5.9.2
Influence of Organic Chemicals
When a compacted clay soil is permeated by an organic chemical, changes in interlayer spacing occur. For organic chemicals with dielectric constant lower than that of water, the individual clay particles will contract as a result of thinner interlayer spacing, thus providing an opportunity for
118
SOIL- WATER- POLLUTANT INTERACTION
the clay particles to orient themselves. Changes in the hydraulic conductivity of the clay could result. The effects may be significant for swelling soils such as bentonite, but are of less significance for non-swelling soils that are compacted to a high density. Historically, hydraulic conductivity variations have been generally explained in terms of differences in the physical variables which control soil hydraulic conductivity, e.g., particle size, shape, and geometric arrangement of clay particles, which in turn describe the geometrical configuration of the pore system. If hydraulic conductivity of soils permeated with organic chemicals is predicted according to the intrinsic permeability concept, the hydraulic conductivity of the organic permeant solution should be dependent on the ratio between solution density and viscosity, as shown in Table 5.5. As the ratio increases, hydraulic conductivity increases, hence organic chemicals should have permeated more rapidly through the clay specimens than the distilled water. Fernandez and Quigley (1988) have demonstrated that for pure organic solvents, viscosity is not the controlling parameter of hydraulic conductivity.
Table 5.5: Variation of hydraulic conductivity with or~;anic solvents Solvent
Hydraulic conductivity (m/sec) Kaolinite Illite Natural soil
Water Acetic acid Aniline Acetone Heptane Xylene
3.0 6.0 2.0 2.3 1.4 7.0
x 10-6 • 10.6 x 10.6 x 1 0 -6
x 10.6 x 107
9.4 2.5 7.0 1.3 5.1 2.0
x 10.7 x 10-6 x 10.7 x 1 0 "7
• 10.7 x 10"7
7.0 x 10.7 2.1 x 10.6 1.1 x 1 0 "7 2.6 x 10.7 3.3 • 10-7
Dielectric constant 78.5 6.2 6.9 20.7 1.9 2.6
Solvent density/ viscosity 1 0.875 0.226 2.390 1.660 1.074
Source: Yong, Mohamed and Warkentin, 1992.
Organic molecules permeating the clay-water system can be considered to move by diffusion and advection through the macro-pores where, at each stage along the way, partitioning between the aqueous phase and soil aggregates occurs. Molecules weakly absorbed by soil aggregates tend to move quickly through the aqueous channels. Hydrophobic substances such as heptane, xylene and aniline, are highly partitioned at any instant onto the soil aggregates, and would consequently be expected to develop resultant soil-heptane, soil-xylene, etc. hydraulic conductivities lower than those of soil-water and soil-acetone. Hydraulic conductivity is not a unique soil property because of the complex physical, mineralogical and chemical interactions. Moreover, we need to pay specific attention to hydraulic conductivity with respect to the type of permeant, specially in the case of organic chemicals, because of partitioning, as demonstrated in Table 5.5 for a very limited example. In the results shown in Figures 5.16 and 5.17, hydraulic conductivity tests were conducted under a standard hydraulic gradient of 20 for each test sample. The soil samples were moulded with distilled water and compacted in standard proctor test at maximum dry density and optimum moisture content. The organic chemicals were introduced only after the passage of one pore volume
INFLUENCE OF POLLUTANTS ON SOIL HYDRAULIC CONDUCTIVITY
119
of distilled water. The partitioning effect suggests that there may be a relationship between the hydraulic conductivity and the octanol-water partition coefficient (kow) which accounts for the tendency of permeant molecules to escape from the aqueous phase and adsorb into octanol, kow simulates the hydrophobic adsorption mechanism between organic chemicals and soil organic matter. Figure 5.16 shows the relationship between the relative hydraulic conductivity (ks/kw) and log kow, where kI is the final hydraulic conductivity with the organic solution, and kw is the hydraulic conductivity with distilled water. In general, the hydraulic conductivity to an organic chemical decreases as the log kow increases. Because kow is a measure of escaping tendency of the organic chemical from water, those substances least compatible with water should move most slowly through the soil. The more positive the kow, the smaller the hydraulic conductivity. In other words, the more hydrophilic the organic chemical, the more rapidly it moves through the soil.
Figure 5.16. Relative hydraulic conductivity variations with log octanol-water partition coefficient of organic pollutants for kaolinite, illite and a natural clay soil from Quebec.
The relationship between the relative change in hydraulic conductivity and the octanol-water partition coefficient can be expressed by:
kf _ _ 4.856
6.678
+
[kow- 10.292] k,,,
1 + exp-
r 2 = 0.9122 [5.41]
-5.885
Chemical analysis of pore fluids of soils permeated with organic chemicals have shown considerable decrease in pore fluid cation concentration. The decrease in cation concentration could
120
SOIL- WATER- POLLUTANT INTERACTION
result in an increase in the replusive forces between particles, thereby promoting soil particle dispersion. This in turn contributes to a decrease in the hydraulic conductivity. The calculated replusive energies indicated that as log kow increases, the replusive energy increases and, consequently, the hydraulic conductivity decreases.
Figure 5.17. Relative hydraulic conductivity variations with molecular weight of organic pollutants for kaolinite, illite and a natural clay soil from Quebec.
As a general rule, as the molecular weight of the organic chemical increases, the clay soil hydraulic conductivity, to the organic chemical, decreases, as shown in Figure 5.17. This is due to the fact that more water molecules will be displaced since the large molecule has more points of contact with the active clay surface. In the adsorption of long chain molecules, van der Waals interactions are important since these forces are additive and tend to reorient the organic molecules for maximum contact points with clay surface. This explanation accords well with the observation that molecular weight can be used as a measure of organic substance hydrophobicity. The greater the molecular weight, the higher the tendency of organic substance to be hydrophobic. Movement of the molecule can therefore be expected to be slower through the aqueous channels in the clay soil. The relationship between the relative change in hydraulic conductivity and molecular weight of organic chemicals, Mw, can be expressed by: k~ J - - 25.366 kw
+
26.335 1 + exp[- [Mw-1 167"569]]6.946 ]
r 2 - 0.8773 [5.42]
The dielectric constant of a permeant has been widely recognized as a critical parameter
SUMMARY AND CONCLUDING REMARKS
121
affecting the hydraulic conductivity of clays. Liquids having low dielectric constant may decrease the thickness of the diffuse double layer around the clay particles, leaving a large space for permeant flow (at constant void ratio), resulting in large values of hydraulic conductivity. Aqueous solutions containing increasing amounts of water soluble ethanol and dioxane exhibit a steadily decreasing dielectric constant with increase of hydrocarbon content (Fernandez and Quigley, 1988). However, this trend is not directly reflected by the measured hydraulic conductivity values of water-compacted clays permeated with dilute solutions of organic chemicals. The reported results by Fernandez and Quigley (1988) indicate no significant increase in kI for aqueous solutions containing up to 70% organic chemicals. This concentration level was also clearly defined in a very broad review of the literature by Mitchell and Madsen (1987). It is suggested that the strong affinity of double layer cations for water results in the exclusion of less polar liquids, preventing the significant double layer decrease that apparently takes place with concentrated water soluble organic chemicals. In addition, the large values of kinematic viscosity in the dilute to moderate concentration range could be responsible for a decrease in hydraulic conductivity when one might expect an increase. The application of static effective stresses to simulate landfill loading (vertical pressure 160 kPa) appears to compensate for the dominant effects of a lower dielectric constant. In fact, the sample permeated with concentrated ethanol, which has a kinematic viscosity nearly twice that of water, actually exhibited a lower kI than did pure water (kw).
5.10
SUMMARY AND CONCLUDING REMARKS
The various soil-water-pollutant interaction mechanisms depend on the type of soil constituents and the manner in which they are packed together to form soil mass, and the nature of the pollutants constituting the leachate. Laboratory studies of adsorption of soil mixtures provide only microscopic descriptions of the processes involved without reference to the detailed microscopic structure and behavior of matter. Analysis of the many studies on the adsorption of various kinds of pollutants by different types of soils shows that the transmission properties of the soils will change with accumulation of pollutants. Accordingly, if predictions or analysis of transport of pollutants are to be made, it is necessary to include considerations of the sorptive characteristics of the different types of soil constituents and soil structure. Many regulatory bodies in the various countries, states, provinces, etc., use a hydraulic conductivity criterion in the design of a clay barrier for waste containment. From the previous analysis, we begin to realize that when the surface active soil constituents interact with pollutants, the various interaction processes may alter the properties of the soil material and influence the hydraulic conductivity. To be of practical value, it is important to be able to relate behavior in the laboratory to performance in the field. The properties of soils in the field are subject to continual changes because of pressure, concentrations, temperature, pH, and biological activities. Extrapolations of interpretations of adsorption data from the laboratory to field conditions should take into account these changes, as well the inevitable differences in the application and mixing procedures in the two systems, differences in the amounts of solvent present, and the biological and chemical degradation of the adsorptive and adsorbate species in soil.
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CHAPTER SIX
F A T E A N D E F F E C T S OF P O L L U T A N T S
6.1
INTRODUCTION
The human race has produced literally hundreds of thousands of unnatural chemicals, as discussed previously in Chapter 3. This ever-increasing array of chemical agents is regarded as a potential threat to humanity and its living environment. Many of these man-made chemicals, which have found their way into the biosphere, have been identified as toxic or potentially harmful chemicals, as discussed in Chapter 3. It is important to be able to evaluate their fate in the natural environment via determining the quantity, or level of exposure at which individuals may be adversely affected. This chapter deals with the fate of chemicals in the environment, the relationship of their physical-chemical properties to fate, their persistence in the biosphere, their partitioning in the biota, and toxicological and epidemiological forecasting based on physico-chemical properties. The reader should realize that a comprehensive and systematic study of the fate of pollutants in the environment is an integration of diverse disciplines such as soil science, chemistry, physics, biology, geology, and hydrogeology.
6.2
POLLUTANT PATHWAYS
The movements of pollutants between the four ecosystems (atmosphere, air; hydrosphere, aquatic medium and water; lithosphere, land and terrestrial medium; and biosphere, plant and animals excluding man) have a profound effect on its bioavailability. The rates of transfer are important and their magnitude can affect the environment livability. The natural processes that promote these exchanges are ever present and are responsible for the magnitude and direction of the exchanges, be they desirable or not. Once a pollutant enters one of the mobile phases (i.e., air or water), it disperses rapidly because of fluid movements. Movement within a phase is termed inter-phase mass transfer, diffusion, or dispersion. Inter-phase mass transfer is important to the movement of pollutants between various phases of the ecosystem. People and other organisms that constitute the biosphere reside, to varying degrees, within the other three ecosystems. The direct and indirect routes of pollutants through the environment and eventually to the human organism are shown in Figure 1.3 of Chapter 1. Humans encounter potentially harmful substances by direct contact with pollutants contained in food, food additives, medicines, cosmetics, the workplace, home, and so on. There are, however, several indirect contact modes of a more subtle nature. The continual intak r 9f air and water is an indirect source of many pollutants. The pathways for entry of pollutants into various ecosystems are 123
124
FATE AND EFFECTS OF POLLUTANTS
shown in Figure 1.4 of Chapter 1. The biosphere, as a possible source of food, is another indirect source made even more critical due to bioaccumulation effect, which is defined as the increase in concentration of a chemical with time due to ingestion into the organism without any discharge of the chemical. Medicines, cosmetics, and other similar personal goods made of living matter constitute another indirect contact source. Chemicals and energy are inevitably placed into the environment in a non-natural manner. The eventual assimilation of this disturbance is largely dependent on the transport processes present in the natural setting.
Table 6.1: Classification of the environment by medium and compartments Medium Air medium Aquatic medium Terrestrial medium
6.3
Compartments Atmosphere; airborne particles; clouds; biota. Surface waters; sediments; organic suspended solids; groundwater; biota. Soil; soil pore fluid; soil atmosphere; vegetation layer and other biota.
ENVIRONMENTAL FATE
Exposure is understood to mean the concentration of a chemical substance to which humans, animals, plants or ecosystems are exposed at a certain time (Korte, 1987). The exposure of a chemical substance primarily depends on its fate in the environment. A distinction can be made between the following three processes (1) transport/translocation (mobility), (2) metabolism (transformation), and (3) mineralization (degradation). In the case of transport or translocation, chemical substances spread from the site to other places within a compartment or move to a new compartment may continue to have an effect. Classification of the environment by medium and compartment is shown in Table 6.1. New substances (metabolites) are created by transformation processes, either through photolysis, in the body of an animal or plant, or through microorganisms. Their fate and effect are often very different from those of the initial substance (parent compound). In other words, metabolites may be more toxic than the original chemical; but their damaging effect may also be neutralized (e.g., methylation of heavy metals). Heavy metals function as atoms. Consequently, once they are in the environment, they are not degradable like organic chemicals, whose molecules are broken down into water, CO2 or other small molecules. Heavy metals and organic chemicals, however, can be bound to soils, as discussed in Chapter 5, thus affecting their bioavailability.
6.3.1
Surface Water Compartment
In water bodies such as rivers and lakes, environmental conditions are relatively uniform over great distances. This is true in a horizontal direction, but less so for vertical stratifications, e.g., discontinuity layer in lakes. Ecotoxicological tests, for the aquatic medium, are generally performed
ENVIRONMENTAL FATE
125
on a body of water without sediment or suspended particles. The organisms' chances of surviving in water depend on chemical factors such as the concentration of dissolved substances (oxygen, nutrient salts or organic molecules) and physical factors such as temperature, water movement and light penetration. The number of species living in the aquatic medium is higher than in the terrestrial medium. The exposure situation of a chemical in the aquatic medium is determined primarily by the following factors: (1) the presence of reactive groups of chemicals, e.g., complexing agent, (2) the kind and quantity of suspended organic matter, (3) grain size distribution of the solid substances, (4) microbial activity, (5) temperature and pH, and (6) hydrological factors such as water quantity and flow rate.
6.3.2
Soil Compartment
Soils are a part of terrestrial ecosystem. The bedrock, the regional climatic conditions and the influence of biological factors determine whether and what type of soil evolves, as discussed in Chapter 4. Consequently, there is an extremely wide variety of soil types with significant differences in composition and characteristics, which are constantly changing, even without human influence. In soils, biomass is produced by plants as well as decomposition and mineralization of soil organic matter (Beck, 1993). Hundreds to thousands of animals and microbe species are involved in this process. Microbes react more flexibly to stress situations than animals. This capacity for adaptation is also manifested by their ability to degrade and mineralize a large number of chemicals, including man-made substances such as pesticides. Because ecotoxicological studies have been conducted in aquatic systems for a few decades now, there is a relatively sizeable database available on them. The situation in the terrestrial medium is significantly worse. Moreover, there is a growing number of indications of a dramatic increase in the level of stress on soil due to application of chemicals (Winteringham, 1988). Soils differ from surface waters in the following ways: (1) soils are more complex and nonhomogenous than bodies of water, (2) since soil is composed of different phases, i.e., solid, air, and water, many organisms interact with these phases with varying degrees of intensity depending on the species, age, season, etc., and (3) the chemical reactions in soils are different, depending on the type and composition of silicate mineral, amorphous, carbonate, sulfide, and organic matter. This element of complexity makes prognoses on the fate of chemicals in soils much more difficult than in surface waters, especially over the course of time. In general, however, a gradient of stress is observable, with the concentration of chemicals decreasing from top to bottom. The situation is further complicated by adsorption, bonding or incorporation of the substance to soil components as well as by climatic influences, as discussed in Chapter 5.
6.3.3
Sediment Compartment
The formation of sediments is about as complex as that of soils. They act as sinks, i.e., adsorption of chemicals (Luoma and Ho, 1993). Despite the possibility of remobilization, sediments can document states of stress due to accumulation of chemicals in certain strata. Most of the information already provided for soils is also valid for sediments, which are often referred to as underwater soils. Thus, here as well, the content and composition of the organic matter is probably the single most important factor in the fate of an organic chemical, and for bonding in particular. As
126
FATE AND EFFECTS OF POLLUTANTS
far as the state of exposure is concerned, however, several peculiarities should be born in mind: (1) There is a constant exchange of matter between the sediment and open water, although substances move from the water to the sediment in much higher quantities than they do in reverse. As a result, the stress gradient in sediment usually runs from top to bottom, (2) Sediments may have extremely different thicknesses. In many cases, there may be a shortage of oxygen, even at only a few centimeters depth, i.e., there are anaerobic conditions. This has a considerable influence on the metabolism and degradation of chemicals, since such processes take a completely different course, qualitatively and quantitatively, in the absence of oxygen, and (3) Abiotic (non living) processes, such as fluctuations in pH value, may trigger remobilization.
6.4
BIOAVAILABILITY
(1)
Various definitions of bioavailability are given in the literature as: Ramamoorthy and Baddaloo (1991):
"the fraction of the total chemical that is available for uptake by biota from the encompassing environment such as water, sediment, soil, suspended soil;" (2)
Kettrup et al., (1991 ):
"the percentage of external exposure (adsorbed or re-sorbed by the organism) which enters the metabolism and is thus available at a possible site;" (3)
Landrum and Robbins (1990):
"the fraction of total contaminant in the interstitial water and on the sediment particles that is available for bioaccumulation" whereas bioaccumulation is defined as "the accumulation of contaminant via all routes available to the organism;" and (4)
Sposito (1989):
"an essential or a toxic element is bioavailable if it is present as, or can be transformed readily to, the free ion species, if it can move to plant roots on a time scale that is relevant to plant growth and development, and if, once adsorbed by the root, it affects the life cycle of the plant." The phrase "affects the life cycle of the plant" means "produces growth and development" in the case of an essential element whereas for a toxic element it means "produces phototoxicity" due to light-induced chemical reactions. These definitions indicate that partitioning of pollutants between liquid and solid phases is a key element for understanding bioavailability. The available amounts of pollutants is highly dependent on: (1) the physico-chemical characteristics of the substance, (2) the composition or
BIOAVAILABILITY
127
characteristics of the surrounding medium, (3) the absolute quantity of chemicals in the medium, (4) the possibility of uptake or contact between substance and organism, and (5) the physiological status of the organism. This clearly indicates that bioavailability is not a substance-specific parameter which could be defined with a single test. Instead, the chemical concentration at a target site (e.g., in a test substrate) as a whole can be measured by residue analysis. However, no distinction is made between bound and available quantities of a substance. No procedure exists to date which could be used to assess the available portion based on absolute measured concentration. For this reason, the concentration measured in the open body of water, i.e., without suspended solids or sediment, is considered as the available quantity in the aquatic medium. This approach cannot be applied to the soil, however, because of heterogeneity and high percentage of bound components. Consequently, selective sequential extraction procedure is important for determining the speciation of pollutants in various forms, as discussed in Chapter 5. Hence, correlation factors between the extraction efficiency of certain solvents and the bioaccumulation in plants or earthworms can be determined (Sheppard and Evenden, 1992).
Figure 6.1. Lead availability and fixation as related to pH.
6.4.1
Availability of Heavy Metals
The presence of carbonates in a soil contributes significantly to the formation of unsoluble complexes which become unavailable to plants. This is demonstrated in Figure 6.1 for the Domtar Sealbond (illitic soil) which is known to possess some carbonates. At low pH values (pH -~ 4), dissolution of carbonates occurs, and availability increases. When the pH of the soil solution is larger than 4, heavy metals are retained in soils by hydroxide, carbonate and as exchangeable cations. The
128
FATE AND EFFECTS OF POLLUTANTS
residual fraction, shown in Figure 6.1, is considered to be retained within the lattice of silicate minerals, hence it is unavailable to plants.
6.4.2
Availability of Inorganic Phosphates
Phosphate anions can be attracted to soil constituents with bonds that are insoluble and becoming unavailable to plants. This process is called phosphatefixation (Tan, 1993). Acidic soils usually contain significant amounts of soluble and exchangeable A13§ Fe 3§ and Mn 2+ions. At pH 3-4, solubility of the A1 and Fe hydroxy-phosphates is considered very low. However with increasing pH levels, solubility of these phosphate compounds increases and reaches a maximum at approximately pH 6.5. Above this pH level, the A1- and Fe-hydroxy compounds decrease again in solubility, as shown in Figure 6.2 (Brady, 1974). On the extreme alkaline range, pH 8, Ca-phosphate is in an insoluble form. By decreasing the pH, this compound also becomes slightly soluble, and maximum solubility is reached at pH 6.5. It appears that at pH 6.5, the soil may contain the maximum amounts of phosphate that can be solubilized from all the insoluble inorganic phosphate forms present in soils. The forms of phosphate ions present also depend on soil pH. In acidic conditions, the H2PO4 ions prevail, in alkaline conditions the HPO42 ions are dominant, and at pH 6.5, H2PO4 , HPO4 2", and PO43- can exist in combination in the soil solution (Tan, 1993).
Figure 6.2. phosphate availability and fixation as related to pH (Brady, 1974).
BIOAVAILABILITY 6.4.3
129
Availability of Organic Chemicals
Organic chemicals can be bound to humic substances as well as to clay minerals, as discussed in Chapter 5. The bonding processes are determined by a large number of soil and substance characteristics, the most important of which are (1) type and relative quantity of clay content, (2) particle size, (3) ion exchange capacity, (4) pH and conductivity of soil solution, (5) redox potential, and (6) composition and quantity of soil organic matter. A precise knowledge of these factors can be used to estimate the potential for geoaccumulation, i.e., the tendency of a substance to persist in the soil over the long term, with a simultaneously low bioavailability. This potential can be determined by evaluating the adsorption behaviour, as discussed in Chapter 5. It should be noted that tests on adsorption do not offer any indication of a possible remobilization of bound residues. The use of radio-labeled (non-reactive) substances in fate studies clearly showed that a considerable portion of substances in all classes of chemicals are so thoroughly bound to soil particles (usually humic substance or clay particles) that they cannot be extracted with organic solvents. The hazard potential of these non-extractable residues can be very high, for example, the release of aluminum ions when soil pH is low. However, in the case of pesticides, the nonextractable portion may be as high as 90% (the average range is 20% to 70%) of the applied quantity, depending on the substance and the soil type (Calderbank, 1989). With time, as shown in Figure 6.3, these residues also undergo microbial or abiotic degradation, albeit in a very slow process (Kloskowski and Fiihr, 1988). The minimal quantities of metabolites produced as a result should have little if any ecotoxicological effects. This is also true in cases where the active ingredient may be slowly released as the soil's organic content ages. Sudden releases of large quantities are a different matter if, as in the case of aluminum, they lead to a dramatic change in the environmental conditions.
Figure 6.3. Bioavailability of a pesticide in soil (Kloskowski and Fiihr, 1988).
130 6.5
FATE AND EFFECTS OF POLLUTANTS EFFECTS OF POLLUTANTS
The effect of pollutants can be observed from different levels of organization of biological systems. The transition from individual to the population opens up new dimensions of interpretation (Moriarty, 1988): (1) Sub-lethal Effects: It lowers the birth rate over the long term. For example, it can have a much stronger influence on the development of a population than acute (short term) mortality. In the case of species with a mortality which is naturally high, the pollutant is only an added mortality factor. For species which take longer to grow, however, lower birth rates pose a much greater threat. A high mortality, on the other hand, can increase the chances for reproduction of the remaining individuals (Harwell and Harwell, 1989); (2) Individual Species: Pollutants have an effect on individual organisms, regardless of their number. In other words, like most physico-chemical factors, they are an environmental factor which does not depend on the density of individuals affected by them. However, an indirect impact, e.g., an alteration in the competitive or nutrient relations, may indeed be densitydependent; (3) Environmental Conditions: They influence the effects of pollutants. For example, climatic factors alter the availability of pollutants for a population or parts of population; and (4) Genetic Variability: Just as the individuals of a species react in different ways to pollutants, populations of the same species also react in different ways to pollutants. This makes experimental results more difficult to replicate. 6.5.1
Uptake of Pollutants
A substance is taken up either directly (i.e., orally, through inhalation or dermal contact) or indirectly through the food and drinking water. Pollutants must be adsorbed in order to take full effect. In animals, this generally happens through the skin, the stomach or the respiratory system. For plants, some chemicals are made to stick to the leaf surface while the others are taken up via the roots and transported to the leaves. Insecticides, for example, occasionally have adverse effects on plants, and fungicides may be extremely toxic to soil organisms. Organisms can accumulate pollutants in the body up to a level at which there is an observable effect. For example, there have been several cases, especially in the 1950s and 1960s, in which persistent pesticides (especially DDT) accumulated in food. Accumulation factors, i.e., the ratio between the concentration in the surrounding medium and the body tissue, of up to tens of thousands are possible (e.g., in mussels or fish) (Phillips, 1993). 6.5.2
Types of Pollutant Effects
The most direct and obvious effect on organisms is the mortality of affected individuals after short term exposure (acute effect), and diminished reproduction and growth after long term exposure (chronic effect). The prognosis or interpretation of direct biological effects is difficult due to the following aspects, among others (Harwell and Harwell, 1989): (1) differences in the sensitivity of target organisms to chemicals, (2) the relative significance of the directly affected species with regard to their ecological role in the ecosystem or their aesthetic, economic or other value for human beings, (3) the influence of physico-chemical environmental conditions on the direct-dose response
PARAMETER IDENTIFICATION
131
relationships, and (4) the interactions with other chemicals stressors in the target organism. The direct effects at the individual level can have a vast range of indirect effects on the level of populations, and ecosystems (Harwell and Harwell, 1989), such as: (1) trophic effects, i.e., changes or interventions in the food webs and resulting shifts in the species composition, (2) destructive impact on the habitat, often resulting in a reduction of the spatial heterogeneity and the number of ecological niches, (3) changes in competitive relationships, such as a decrease in competitive pressure due to a decline in one species which shared a resource with another; at the level of the ecosystem, this mechanism is a mean of compensating for effects at population level, (4) effects on populations which play a role in the relationships between two species, e.g., parasites, pollinators, etc., and (5) resistance formation as a frequently observable consequence of selection by the pesticide as an environmental factor. Normal genetic variability in a population opens up the prospects of an evolutionary adaptation to continuing chronic exposure (seven years to life time) to chemicals, especially in species with short succession of generation. In much the same way as direct biological effects, indirect effects may be influenced to a critical degree by the presence of several chemicals and interactions with the physico-chemical environmental conditions. Moreover, indirect biological interactions and ecosystem processes may be subject to considerable time lags between exposure, the direct or primary biological reaction, and finally the indirect effect. The direct and indirect biological reactions to a chemical are integrated at ecosystem level. This combination of mechanisms leads to qualitatively new reactions to chemical stress: (1) direct and indirect effects on key species which play a decisive role in the structure or function of the ecosystem, (2) changes in the community structure which are produced by the change in the relative abundance of species involved and changes in the interspecific links, and (3) impact on ecosystem processes such as production, decomposition and the nutrient cycle. These process-related effects are frequently characterized by changes in the photosynthesis rate of the dominant plants or by changes in microbial activity which control the decomposition. Effects on ecosystem level are strongly influenced by the kind of nutrient cycle and the physicochemical environmental conditions.
6.6
PARAMETER IDENTIFICATION
In order to assess the ecotoxicological hazard potential of pollutants, the characteristic features of pollutants must be determined. Various parameters are identified: (1) physico-chemical parameters for describing the features of the chemical substance itself, which, depending on the prevailing environmental conditions, largely determine the fate and effect, (2) fate parameters for describing the occurrence of a substance in the various environmental compartments. The transition from the physico-chemical parameters is flexible, and (3) effect parameters are biological measurement variables which describe the effects on organisms. These effects may be directly or indirectly manifested at the various levels of organization of a biological system. In the context of our discussion, a distinction is made between test parameters and test procedures, primarily for practical reasons. Most of the tests commonly employed today are based on only a very small number of measurement variables which are selected because they are easy to measure. The values identified for a given parameter form the initial basis for assessing the effect of a pollutant at the level of organization from which the parameter was derived. The results obtained
132
FATE AND EFFECTS OF POLLUTANTS
at this level must be extrapolated to predict the effects at other levels, which invariably gives rise to uncertainties. The fate parameters are subject to the same principle as effect parameters in this respect, making it necessary to organize the ecotoxicological test procedures in the form of a tiered test program, as discussed in section 6.7. The following sections provide a definition of the most important parameters which are necessary to assess the ecotoxicological hazard potential of pollutants. Generally speaking, parameters which are measured in the laboratory, e.g., to determine the acute effect, are easier to quantify than field parameters. Standardized physico-chemical parameters can be defined much more precisely than biological parameters. The biological parameters must then be differentiated according to whether they are toxicological parameters at the level of individual or system parameters at higher levels of organization (population, ecosystem). Toxicological parameters are often used to assess the effects on human beings, although many of the actual tests are conducted on laboratory animals. The degree to which a parameter lends itself to immediate interpretation declines as the complexity of the level of organization increases, making it necessary to introduce additional, e.g., socioeconomic, criteria to assess the environmental risk.
6.6.1
Physico-chemical Parameters
Currently there is no general consensus as to which physico-chemical parameters are necessary for the ecotoxicological assessment of a chemical substance. One model favoured by the US EPA, for example, recommends the following list of 13 measurements, which can be arranged in various groups depending on the environmental chemical and medium involved (Donaldson, 1992): (1) ionization (pK~), (2) Henry's law constant (He), (3) vapour pressure (Pv), (4) water solubility (Sw), (5) soil-water partition coefficient (Kd), (6) soil-water partition coefficient, organic carbon normalized (Koc), (7) octanol-water partition coefficient (Kow), (8) oxidation rate (Kox), (9) direct aqueous photolysis rate (Kph), (10) aqueous photolysis reaction quantum yield (Ox), (11) molar absorptivity (Et), (12) hydrolysis r a t e (Khyd), and (13) biotransformation rate (Kb,o). Water solubility, hydrolysis, vapor pressure, Henry's law constant, octanol-water and adsorption coefficients, discussed in Chapter 5, are most often employed to predict the environmental fate of pollutants. These parameters are discussed as follows. With regard to water solubility, the biggest problem is that under the artificial conditions of a physico-chemical test, the value measured often proves to be much different from that which occurs in laboratory tests or even in the field. Hydrolysis is understood to mean the splitting of a chemical molecule through the uptake of water. Depending on the polarity of a chemical bond, hydrolysis requires a high temperature and interaction with hydrogen ions. Vapour pressure is a measure of the tendency of a substance to evaporate. Despite the use of standardized methods, the data reported in the literature may diverge by as much as a factor of two for the same pesticide. Henry's Law Constant is of critical importance in assessing the rate at which a chemical passes from water to air. This coefficient is defined as the ratio between the concentrations of a substance in the gas and liquid phases. It is often given as the ratio between the saturation vapour pressure and the maximum water solubility [atm m 3 moll]. Values of Henry's law constants for most hazardous waste compounds range from 10.7 to > 10.3 atm m 3 mol l, as shown in Table 6.2 (Lyman et al., 1982). When the value of Henry's law constant is high, the resistance of the liquid phase dominates over the gas phase. Such compounds are highly volatile. For compounds with
PARAMETER IDENTIFICATION
133
Henry's law constant between 10-5 and 10-3 atm m 3 m o l -l, both the liquid and gas phase resistance are important. Volatilization of compounds with constants in this range is less rapid than for compounds in a higher range, but is still significant (Lyman et al., 1982).
Table 6.2: Relationship between Henry's law constants, half-life, and volatilization potential. Henry's law constant (Hc) Half-life [atm m 3 mol -~] tv2 (hr) Transfer mechanism < 10-7 10 -7
< H c (
>
1 0 -5
10-5 < Hc < 10 -3
> 10-3
3000
120 - 3000
6- 120
1-6
The substance is essentially nonvolatile The gas phase resistance dominates the liquid phase resistance by a factor of 10 at least; therefore, substances volatilize slowly. Liquid phase and gas phase resistance are both important. Volatilization for compounds in this range is less rapid than for compounds in a higher range of Hc, but is still a significant transfer mechanism. The resistance of the liquid phase dominates. Thus, transfer is liquid-phase controlled and these substances are highly volatile.
Table 6.3: Relationships between water solubilities and Ko~ for different chemicals Equation log S
= -1.370 log Kow + 7.26
log S log S log (l/S) log (1/S) log (1/S)
= = = = =
-0.922 log Kow + 4.184 -1.490 log Kow + 7.46 1.294 log Kow - 0.248 0.996 log Kow - 0.339 1.237 log Kow +0.248
Units of S
Chemical classes represented
ktmol/1
Mixed classes; aromatics and chlorinated hydrocarbons. Mixed classes; pesticides. Mixed classes; pesticides. Alkenes. Benzene and benzene derivatives. Alkenes
mg/1 ~tmol/1 mol/1 mol/1 mol/1
The octanol-water partition coefficient is the ratio of the concentrations of a substance in octanol relative to that in water. It is a measure of the polarity of organic pollutants. The higher the numerical value, which is usually given as logarithm, the stronger the tendency of the chemical to accumulate in the fatty tissue of organisms. In various guidelines, a log Kow of approximately 2.7 to 3 is regarded as the threshold value for this tendency. Chemicals with low log Kow values, i.e., less than about 2, have high water solubilities, small adsorption coefficients and small bioconcentration factors (BCF). Chemicals with high log Kow values, i.e., greater than about 3, are very hydrophobic
134
FATE AND EFFECTS OF POLLUTANTS
with low water solubilities and high sorption coefficients. The water solubility (S) of a chemical can be estimated from Kow values. A number of equations have been developed to correlate these two parameters. The equations generally take the form of: l o g S : a + logKow + b
[6.1]
where a and b are empirical regression equation constants. Examples of such relationships are shown in Table 6.3. The adsorption coefficient describes the partitioning of the substance between the soil and the aqueous phase, normalized to the organic carbon content of the soil, as discussed in Chapter 5. The Koc can be predicted from Kow or water solubility. Typical relationships between Koc, Kow, and water solubility (S) are shown in Table 6.4. Substances with low adsorptivity have very low Koc values whereas extremely high-bond chemicals such as dioxins achieve values of several hundred thousand. Koc values of more than 10,000 suggest a low anticipated mobility in soil, with little or no probability of a hazard to groundwater (Esser et al., 1988). It must be borne in mind that the physico-chemical data are invariably based on measurements taken under artificial conditions. A substance may behave very differently depending on the kind of tests conducted in the laboratory (fate or effect), thus yielding values which deviate significantly from the physico-chemical values.
Table 6.4: Relationships between Koc, Kow, and aqueous water solubilit~r (S). Equation log Koc = log Koc = log Koc = log Koc =
- 0.550 log S + 3.64 - 0.557 log S + 4.277 0.544 log Kow + 1.377 0.937 log Kow - 0.006
log Koc = 1.00 log Kow - 0.21
6.6.2
Units of S
Chemical classes represented
mg/1
Wide variety, mostly pesticides. Chlorinated hydrocarbons. Mostly pesticides. Aromatics, polynuclear aromatics, triazines and dinitroaniline herbicides. Mostly aromatic or poly-nuclear aromatics.
~tmol/l
Fate Parameters
In general, the physico-chemical characteristics of a chemical substance are used to determine its fate in the environment. However, in soils, persistence and mobility are the two parameters that used to quantify the fate of a chemical substance. These parameters are discussed in the following sections.
Persistence Persistence refers to the degree to which a chemical substance remains in its original form.
PARAMETER IDENTIFICATION
135
It depends on the mobility as well as the degree of possible biologically mediated transformation processes, i.e., biodegradation, oxidation-reduction, and polymerization. The less a chemical leaches, for example, the more time there is for microbial degradation. With respect to degradation, a distinction is made between primary degradation, in which the test substance is chemically altered, and mineralization, in which the substance is completely broken into small molecules such as water or carbon dioxide. A distinction must be made between the persistence of biological activity and the persistence of residues. For example, paraquat herbicide may be biologically active in the soil for no more than one day or so whereas its residues have a virtually unlimited persistence, due to irreversible sorption (Heitefuss, 1987). Hazardous wastes are generally classified as persistent or non-persistent, as shown in Table 6.5 (Portier, 1985).
Table 6.5: Hazardous wastes cl~sification based on persistence and its potential hazard Typical compounds
Potential hazard NON-PERSISTENT ORGANIC WASTES
Oil, low molecular weight solvents, some biodegradable pesticides (organo-phosphates, carbamates, triazines, anilines, ureas), waste oils, most detergents.
Toxicity problems, primarily to environment and biota, at the source or point of release. Toxic effects occur rapidly after exposure (acute and subacute).
PERSISTENT ORGANIC WASTES High molecular weight chlorinated and aromatic hydrocarbons, some pesticides (chloronated insecticides like hexachlorobenzene, DDT, DDE, lindane), PCBs, and phthalates.
Immediate toxic effects (acute and subacute) may occur at the source or point of release. Long term chronic toxicity may result. Transport of organic waste from the source can result in widespread contamination and bioconcentration in the food chain. Environmental transport may expose biota to lower levels of the pollutant, resulting in chronic toxicity.
Persistence can be quantified in a number of different ways. The degradation time is the most widespread method for quantifying persistence of a pollutant in the terrestrial medium; it is defined as the amount of time it takes for 50% (tl/2) of the applied concentration (day 0 = 100%) of a substance to be broken down. The relationships between Henry's law constant, volatilization rate and degradation time (half-life) are shown in Table 6.2. The t~n values are calculated from the residue-analysis measurements using empirically determined first or second order ftmctions. The first order function is described as follows:
136
FATE AND EFFECTS OF POLLUTANTS dc _
dt
_
kTC
[6.2]
where c is concentration at time t, t is time, and kr is first order reaction rate constant. The integrated form of Eq. [6.2] is ln/~)
= krt
[6.3]
where Co is concentration at time zero. When 50% of the original concentration has degraded, Co/~S equal to 2; the corresponding time is given by the following expression: tin -
In 2 0.69 kr - kr
[6.4]
Mobility
Mobility is understood to mean the migration behaviour of a substance within or between the soil, water, air, and respective compartments, primarily as leaching of a chemical from the surface down to the groundwater. Mobility of pollutants in soils can be estimated via the following methods. Leaching Test As part of the American Nuclear Society leachability test (ANS 16.1), the leachability index is recommended as a standard method for evaluating the relative mobility of pollutants. The leachability index is expressed as:
LX-log
~-~)
[6.5]
where LX is the leachability index, [3 is a constant with an assigned value of 1 cm2/sec, and D e is effective diffusion parameter (cm2/sec). The leachability index compares the relative mobility of pollutants on a uniform scale which varies from very mobile for values of 5 (De - 10.5 cm2/sec) to immobile for values 15 (D e = 1015 cm2/sec) or greater. Physico-chemical Parameters The mobility of pollutants can also be estimated based on octanol-water partition coefficient, solubility and vapour pressure as:
PARAMETER IDENTIFICATION
MI =
S,P, v logl/ K ow,i
137
[6 61
where MI is the mobility index, S, is solubility, p v vapour pressure, and Kow,, is octanol-water partition coefficient. Pollutants with mobility index greater than 5 are considered very mobile while pollutants with mobility index less than - 10 are considered immobile.
6.6.3
Effect Parameters
Pollutants have a direct effect on individual organisms or affect them indirectly through changes in the surrounding environment. In standard ecotoxicological studies, especially in the laboratory, animals are used far more frequently than plants because animals are easier to breed and manage, and they seem to be more sensitive to pollutants. As far as duration is concerned, test procedures, as well as the measurement parameters employed in them, are generally subdivided into acute and chronic, as shown in Table 6.6. Acute tests are usually employed for pollutants with high concentrations, whereas chronic tests are used when pollutant concentration did not show any acute effects, or only a low mortality (Hamburger, 1983).
Table 6.6: Definition of effect tests used in the laboratory Test type
Parameter
Duration
Acute
Lethal (mortality)
Short term (a few days to 2 weeks); usually without feeding.
Chronic
Sub-lethal (reproduction & growth)
Long term In small organisms most of the individual life cycle; in larger animals, up to one-third of the mature life cycle; in mammals, 1.5 to 2 years.
6.6.4
Biological Parameters
In the case of toxicological parameters, a distinction is made between those with a direct (acute or chronic) damaging effect and those used for accumulation.
Direct Damage Effect The mortality or acute toxicity in individuals is a standard parameter for the assessment of a pollutant. In mammals (including human beings), a distinction is made between oral or dermal application (concentration indicated in mg/kg body weight) and application through inhalation (concentration indicated in mg/1 inhaled air). In tests with other animals and plants, on the other
138
FATE AND EFFECTS OF POLLUTANTS
hand, exposure primarily takes place via the surrounding substrate (water or soil). In such tests five concentrations are usually selected to determine a dose-response ratio of mortality over several levels of concentration. The exposed population will typically experience no deaths at low doses, a few deaths as the dose increases, and more deaths with higher doses until all the organisms die. Most plots of log dose versus cumulative mortality will show the classic non-linear S-shaped curve as shown in Figure 6.4. The dose at which 50% of the organisms remained alive is referred to as median lethal dose (LDs0). The acute oral LDs0 in rats for selected chemical compounds are shown in Table 6.7 (LaGrega et al., 1994). The equivalent for inhaled substances is the median lethal concentration (LCs0), where the dose is expressed as the concentration of the substance present in a volume of inhaled air.
100, 9O 80
~
7o
~
60
E
50
Q
~
40
~ E
3O
o~
~
LDso
20
10 0
Dose (mg/kg) Figure 6.4. Dose-response relationship.
Table 6.7: Acute oral LD~o in rats for selected chemical compounds Substance
LD~o (mg/kg)
Substance
LDso (mg/kg)
Ethanol Table salt Aspirin Lindane Sodium fluoride
13000 3800 1500 88-270 180
DDT Nicotine Sodium arsenate Parathion Dioxin (TCDD)
113-118 50-60 40 2 0.02-0.05
PARAMETER IDENTIFICATION
139
Tests which are conducted to assess the ecotoxicological hazard potential generate data of varying quality. The transferability of the data depends on the extent to which the tested species, with its mode of life, physiological and morphological characteristics, is representative of the organisms exposed to the substance in the field. The quality of the results of the individual tests is also determined by the characteristics of the test procedure itself, such as the number of replicates and test organisms and specially the statistical methods used to evaluate the raw data.
Bioaccumulation Bioaccumulation refers to the deposits of pollutants in the organism, in the fatty tissues, in a concentration exceeding that which is found in the organism's environment. A distinction has to be made between the bioconcentration (uptake from the ambient medium) and biomagnification (input via the food supply). The extent of this input strongly depends on the biochemicalphysiological processes which differ from one species to another as well as on the biotransformation (metabolism) and elimination processes. The bioaccumulation parameters used to quantify the environmental hazard are shown in Table 6.8. The bioconcentration factor (BCF) is the standard for measuring accumulation. In experiments, it is usually determined for aquatic organisms (e.g., fish or mussels). Standardized procedures do not exist for terrestrial organisms. BCF values higher than 100 are considered ecotoxicologically problematic, especially if other risk factors are also involved. The degree of accumulation strongly depends on the condition of exposure, in addition to characteristics of the organism (i.e., volume of fatty tissues). If the log Kowvalue is higher than 2.7 to 3.0, there is a strong indication of a higher bioaccumulation potential in the case of pesticides. It is important to bear in mind that accumulation will again begin to decline if the log Kowvalues are higher than 6 (Rombke and Moltman, 1995).
Table 6.8: Bioaccumulation parameters Parameter Molecular weight log Kow
BCF
Description Characteristic of substance; substances < 700 (g/mole) have a propensity for accumulation. Relationship of concentration in octanol and water; the higher the value (especially 2.7 to 3), the greater the tendency for bioaccumulation. The value can be assessed on the basis of water solubility. Value for the ratio of the concentration of a substance in an organism to the concentration in the ambient medium at steady state. It can be estimated from log Koc: log Koc = 0.681 log BCF + 1.963
140 6.7
FATE AND EFFECTS OF POLLUTANTS TIERED TEST PROGRAM
The tiered test program is designed to minimize the amount of time and expense that has to be invested in tests. Semi-field and field tests are much more elaborate and costly to conduct than laboratory tests. If a simple fate and effect assessment test is sufficient to tell that a substance poses only a relatively low environmental risk, then there is no need to conduct any further tests. The OECD (Organization for Economic Co-operation and Development) test guidelines (1984) are the most widely recognized procedure in the world. The tiered procedure for a fate test (leaching) and an effect test (toxicity in earthworms) is shown in Table 6.9 (OECD, 1989).
Table 6.9: Tiered procedure for a fate test (leachin$) and an effect test (toxicity in earthworms) Test sta~e
Test description
Monitored values
LEACHING BEHAVIOUR Laboratory Semi-field Field
Small glass column filled with standard soil. Lysimeter studies with large intact soil cores. Monitoring in the field ( well and ground water).
Amount of applied quantity found in the leachate. Amount of applied quantity found in the leachate. Amount of substance detected.
TOXICITY IN EARTHWORMS Laboratory
Acute mortality for compositeworm in artificial soil. Chronic reproduction test with compost-worm in artificial soil. One-year sampling of natural population after application.
LCs0 > 1000 mg/kg
NOEC reproduction per soil concentration. Discrepancies in abundance Field and dominance spectrum compared with control area. LCs0 = 50% lethal concentration; NOEC = No Observed Effect Concentration. Semi-field
6.8
SUMMARY AND CONCLUDING REMARKS
Evaluation of the fate of pollutants involves the assessment of the distribution of contaminants in soil phases through characterization of the pollutants contained in the soil and the manner in which these are held within the solid phase constituents, i.e., soil organic matter, silicate minerals, amorphous, and carbonates. Geochemical factors such as pH and oxidation-reduction conditions will control the chemical forms and mobility of pollutants. Due to changes of pH, the mobility and bioavailability of the fixed pollutants will be affected. In the case of organic pollutants,
SUMMARY AND CONCLUDING REMARKS
141
biological transformations are considered to be very important in the distribution of organic pollutants in solid fractions. Both biological and geochemical factors exert considerable influence on the persistence and mobility of pollutants in soils. In evaluating the toxic effects of pollutants, ecotoxicology can provide a relatively precise indication at the level of individuals, e.g., single species tests in the laboratory. The practical relevance of the laboratory results is rather limited. However, valid assessment can be obtained from field tests. As we can see from a critical analysis of ecotoxicological assessment data, the number and range of uncertainties increase in population to the level of organization. Inadequate methodology is one of the reasons for this, but the increasing complexity of ecological systems is also a major factor. The fact is that ecological uncertainties do exist and cannot be completely eliminated. Generally speaking, they are attributable to the following causes: (1) insufficient data and an inadequate understanding of ecological processes and systems due to deficient methods, (2) potential indirect effects in ecosystem which lead to unpredictable events, (3) wide range of variability in the system's environmental conditions, and (4) the need to extrapolate from special conditions of the respective ecosystem or test substance. Consequently, when assessing the ecotoxicological hazard potential, an attempt must be made to weigh a large number of different data to yield an overall result that will permit the respective chemical to be classified in an understandable way. The core of the problem here is that there is no way to prove that a substance is not hazardous. Nor will it be possible to eliminate this contradiction through the increasing use of models. Detailed analysis for quantifying the hazard and risk of pollutants is given in Chapter 10.
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CHAPTER
SEVEN
SITE I N V E S T I G A T I O N
7.1
INTRODUCTION
The establishment of an effective remedial action plan for uncontrolled hazardous waste sites must integrate all the pathways involved in the transport of pollutants through the environment and to receptors through a proper site characterization study. The basic aspects of a remedial action selection process are: (1) Assessment of the nature and extent o f pollution: Prior to the selection of any remedial program for an existing site, it is necessary to establish the extent of pollution via sampling of soil, groundwater, surface water, and biota. Monitoring of these various media will provide a detailed analysis of the nature and extent of pollution; (2) Collect site specific data: Site specific characteristics are of high importance to the selection of an appropriate remedial action. It is therefore advisable to collect as much specific information as early as possible in the selection process. The data should also include the quantity and quality of waste material, characteristics of site cover, the climate of the areas, subsurface geology, proximity to various receptors, existing land use etc; (3) Determine remedial options: Various remedial techniques are available for each environmental medium. Remedial techniques range from passive and extraction systems to direct or indirect on-site and off-site treatment processes; and (4) Selection of a remedial action: This can be established through detailed environmental impact assessment. This chapter outlines an approach to site investigation at polluted sites.
7.2
SITE INVESTIGATION APPROACH
Most, if not all, successfully engineered works begin with planning and site investigation. From the initial empirical development of soil mechanics, to Terzaghi's founding theories in soil mechanics and investigative procedures, we have reached to the point where systematic evaluation and measurement of the physical properties of soils is the basis for today's site investigation. Site characterizations have gradually evolved over the past forty years, from borehole studies for roads, dams and buildings in the 1950s to literally hundreds of borings and monitoring wells for hazardous and solid waste investigations in the 1980s. There seems to be the belief that the larger the amount of data procured during a field investigation, the greater the probability that a project will be completed successfully without complication or delay. The procedures used for the gathering of information for subsurface characterization through drilling, sampling and logging have changed 143
144
SITE INVESTIGATION
very little within the last 50 years. The basic technology for field collection of subsurface samples for field investigation was summarized by Hvorslev in the late 1940s and can still be considered valid today. To establish a cost-effective approach to field investigation, we need to take a look at the philosophy of the classic site investigation approach developed by Peck (1969). The approach, known as the observational method, is divided into six phases: (1) conduct an investigation of sufficient scope to establish the general characteristics of the site, (2) assess the most probable site conditions and the deviations from these probable conditions, (3) develop a design based upon the most probable site conditions, (4) determine what course of action should be taken if the conditions deviate from predictions, (5) measure and evaluate actual conditions during construction, and (6) modify the design as needed to suite actual situations. Even though Peck (1969) developed this method for a specific geotechnical application, this observational philosophy can be used to define the proper location for monitoring systems or to establish the rate and extent of environmental pollution (Sara, 1993). The implementation of this approach to field investigation relies upon the geotechnical engineer's knowledge, experience and judgement. The phased approach to site investigation involves the determination of site characteristics based upon: (1) a thorough review of existing literature and available technical information for the site under investigation, (2) preliminary site aerial and ground reconnaissance, (3) development of initial regional and site-specific geologic conceptual models, and (4) design of field investigation based upon the initial conceptual models. Investigations must be precise enough to establish target monitoring zones and identify limiting design criteria.
7.3
PHASE I INVESTIGATIONS
Phase I investigations are preliminary in nature and are designed to furnish a comprehensive overview of available site information. The objectives of phase I investigations are to: (1) identify potential subsurface pollutants and environmental concerns at the site, (2) develop a conceptual model, and (3) establish a frame work for phase II investigations. 7.3.1
Collecting Information
The following sources of information are generally required for identifying the potential subsurface pollution and various environmental concerns.
Sources of Information on Site History Information about current and historical land use and activities can indicate the existence or the likelihood of subsurface pollution at the site. The types and possible sources of information are summarized in Table 7.1 (CCME, 1994). The key issues to be considered during phase I investigations are: (1) The nature of known or suspected pollution." It is important to identify the potential pollutants and their physico-chemical, fate and exposure parameters, as discussed in Chapter 6. In addition, it is required to determine the relevant environmental regulations for soil, air, and water mediums;
PHASE I INVESTIGATION
145
Table 7.1: Sources of information for site history Sources of information
Type of information
Availability
Owner and regulatory agency files
Site operational and environmental history
Site owner and government agencies (environment and regulatory)
Land use and ownership history
Site activities and operations
Municipal tax records and directories, and title searches
Aerial photographs
Land use history, physical and drainage features
Government agencies (natural resources, national and local archives)
Archival records
Historical photographs and operational history
National and local archives, corporate files
Site plans and engineering drawings
Site layout and features
Corporate and municipal files
Historical maps and fire insurance plans
Land use and industrial process areas
National and local archives
Reports
Site history and practices
Present and former employees, local residents, and local historians
Industrial activities and processes
Manufacture, use, storage, and disposal of chemicals
Corporate archives, historical and contemporary trade joumals and texts
(2)
(3) (4)
The sources or possible sources of pollution." Some of the obvious potential sources of pollution are waste management practices and disposal facilities. Also, the transfer of bulk chemicals and petroleum products to train and trucks often generate uncontrolled spillage on the ground surface with a significant cumulative subsurface pollution. Although the pollution potential of some practices and facilities can be readily recognized, the potential of others may be masked and only available in reports such as inappropriate disposal practices of used solvents and solvent sludges, past chemical usages, and current processes at the facility; The extent o f pollution: The extent of known or suspected pollution will depend on the physico-chemical and fate properties, volume of chemicals introduced to the subsurface, and the nature of the subsurface environment; and Health and safety: A detailed health and safety plan should be specific to the site and consider risks likely to be present at the site. Some common risk sources are electrical utilities, unstable slopes, dangerous debris, aesthetic impact, temperature, radiation, disease, infection, explosion, fires, etc.
146
SITE INVESTIGATION
Geologic and Hydrogeologic Information Reports on previous investigations at or in the vicinity of the site can provide information regarding bedrock elevation, overburden descriptions and thicknesses, regional settings, and structural features (i.e., fractures in bedrock). Soil type can be useful for inferring subsurface material properties such as hydraulic conductivity. Aerial photographs can provide an excellent assessment of geology and surficial drainage features. Geophysical logs of water supply wells can assist interpretation of subsurface conditions for the water well industry. The hydrogeologic setting of any site plays a critical role in the movement of water and pollutants in the subsurface. In phase I investigation, guidelines for collecting and synthesizing information on the subsurface setting are developed. Potential recharge and discharge zones, depth to the water table, general groundwater flow directions, surface drainage patterns are some of the features that can be identified in phase I investigation.
Hydrologic Information The focus of the investigation will be on surface water, its location, movement, quality, and connection to groundwater. Much information about sources and direction of flow of water can be estimated from topographic maps. More details can be found in more specialized water resources reports. Some of the existing information may also be useful at later stages of investigation. For example, the flow rates of nearby streams and rain fall data might be useful for calculating the amount of water moving through the subsurface. 7.3.2
Field Reconnaissance
Once the literature search has been completed, a site visit must be conducted in order to substantiate conclusions drawn from the literature findings. During this phase, the following important observations are made: (1) Site terrain and its surroundings to assess accessibility for geological surveys and drilling equipments; Site logistical considerations to determine presence of underground utilities, need for (2) excavation/drilling clearance, access routes across private property and availability of clean water; Site geological conditions to determine the consistency of the available background data with (3) the regional pattern of geology and locations of bedrock; Site topography, drainage and vegetation to determine locations of liquid waste discharge (4) and stressed vegetation areas; Status of waste, particularly its mobility and degree of exposure; (5) Status of monitoring devices in the study area, notably their condition, depth of penetration, (6) and the groundwater level; and Site climate to determine precipitation and temperature at study site. Based on precipitation (7) conditions, surface water and groundwater conditions, and wind and erosion potentials are determined. Physical processes such as rates of reaction, volatilization, microbial degradation and transformation processes are highly dependent on temperature conditions at the site.
PHASE I INVESTIGATION 7.3.3
147
Development of a Conceptual Model
After collecting the necessary information, a conceptual model can be developed and updated as new data becomes available. The model should include the following basic components: (1) The geologic setting at and in the vicinity of the site: The conceptual model will distinguish between various geologic layers in terms of their hydraulic characteristics, and will attempt to indicate the significance of the various layers in influencing the groundwater flow system, and their potential control of the migration of pollutants in the subsurface; (2) The regional and local surface and groundwater flow systems: The conceptual model should identify the interaction between the groundwater and surface water systems in the vicinity of the site, and also indicate the inter-relationships between the regional and local groundwater flow systems. It should also incorporate topographic and stratigraphic information into the schematic diagrams of the groundwater flow system; (3) Identify impact of human activities on water flow and pollutant migration at the site: For example, buried pipelines, utilities, and sewers and their associated coarse-grained backfill often provide conduits for the preferred flow of non-aqueous phase liquids and groundwater through the subsurface. Groundwater pumping wells in the vicinity of a site may also alter hydraulic gradients and modify groundwater flow system; (4) Identify the natural and preferential pathways for pollutant migration: These pathways might include high hydraulic conductivity layers and lenses in the geological materials or fractures in clays and rocks; (5) ldentify pollutants characteristics: It is important to include the pollutant characteristics in the development of the conceptual model to ensure that potential areas of occurrence and migration can become the focus of site monitoring and investigation program; and (6) Identify potential receptors for evaluating the degree of environmental impact: Receptors may include people, plant, animals, and aquatic organisms. 7.3.4
Establishing The Work Plan
The collation and evaluation of existing background information and field visit during phase I investigation are essential for the development of a comprehensive work plan or program for subsequent stages of the site assessment process. The work plan should account for special physical features at the site. For example, low hydraulic conductivity layers on a site may protect deeper zones from near-surface pollution. Inappropriate drilling techniques may jeopardize the integrity of such features and create increased pollution. Some investigative techniques are more successfully applied in certain geologic environment than in others, thus geologic conditions have a significant influence on the selection of appropriate methods of investigation. Special characteristics of the pollutants at a site should also be considered in the development of a work plan. Examples of these considerations are: (1) suitability of the overall approach to site investigation, i.e., avoid making subsurface pollution worse, (2) suitability of using geophysical techniques during site investigation, (3) compatibility between pollutants and monitoring well materials, and (4) suitability of drilling, monitoring well installation or sampling techniques. The selection of an appropriate technique is discussed in the following sections.
148 7.4
SITE INVESTIGATION PHASE II INVESTIGATIONS
Phase II investigations consist of a site characterization phase (phase IIa) and groundwater monitoring well installation (phase lib). A variety of investigative techniques are available for the collection of data for the characterization of sites. The actual site investigation will include direct and indirect methods. Direct methods include, but are not limited to, boreholes, piezometers and geotechnical analysis of soil samples. Indirect methods include aerial photography, ground penetrating radar and earth conductivity and resistivity geophysical studies. The investigator should combine direct and indirect methods to produce an efficient and complete characterization of the site under investigation. Regardless of the method employed, the intended purpose of soil sampling at polluted sites is to determine whether hazardous pollutants exist in concentrations which exceed levels established as being adequately protective of both humans and the environment. Specifically, soil sampling efforts may be designed and conducted to: (1) determine the extent to which soils act as either sources or sinks for air or water pollutants, (2) determine the risk to human health and/or the environment from pollution by selected pollutants, (3) determine the presence and concentration of specific pollutants in comparison to background levels, (4) determine the concentration of pollutants and their spatial and temporal distribution, (5) measure the effectiveness of control or removal actions, (6) determine the potential risk to plants and animals from specific soil pollutants, (7) identify pollutant sources, transport mechanisms or routes, as well as potential receptors of pollutants, (8) obtain measurements for validation or use of pollutant transport models, and (9) meet the legislative provisions and intent of environmental laws. The extreme complexity and variability inherent to most soils necessitates that a multitude of sampling and monitoring approaches be incorporated into subsurface investigations. Both field and laboratory tests are necessary to comprehend the presence and behaviour of polluted soils. Field testing primarily provides information regarding soil characteristics, groundwater flow conditions, and pollutant migration. Laboratory testing supplies analytical data on the type and quantity of a pollutant present in the subsurface. Groundwater level monitoring is also essential to any subsurface investigation program. In order to properly determine temporal and area variations in groundwater flow and distribution in both the saturated and unsaturated zones within the subsurface, proper groundwater monitoring instrumentation and correct placement methods must be utilized. The correlation of the aforementioned requirements into a phased investigation for the determination of soil and water regime characteristics will facilitate an already difficult task, by eliminating many of the variables consistent with the environmental conditions. A brief and basic outline of the various techniques will be presented, with the recognition that literally volumes of material could be and have been written on each precise method and technique.
7.5
GEOPHYSICAL TECHNIQUES
Geophysical techniques are used to better understand subsurface conditions and delineate the extent of pollution. The basis for using geophysical approach to investigate groundwater pollution is in the fact that dissolved electrolytes form ions, which enable water to conduct electrical currents and, up to a limit, the conductivity of the solution is proportional to the amount of dissolved
GEOPHYSICAL TECHNIQUES
149
electrolyte. Geophysical techniques include ground penetrating radar (GPR), electromagnetic conductivity (EM), electrical resistivity surveys, and seismic surveys. It is important to understand that any geophysical method which is successful at one site may or may not have the same success at another site. A professional with sufficient experience to know the limitations of each of these techniques should be consulted prior to developing a large geophysical reconnaissance program. The following discussion of the various geophysical techniques places emphasis only on the basic principles and use limitations. 7.5.1
Ground Penetration Radar
Ground penetration radar (GPR) uses high frequency radio waves in the range of 10 MHz to 1000 MHz. GPR operates at frequencies where the displacement currents, which depend on the dielectric permittivity, dominate, as opposed to conventional electromagnetic (EM) surveying where the conduction currents dominate. An antenna, typically 0.25 m to 1 m in length and in direct contact with the ground, emits a short burst of electromagnetic radiation that propagates downward into the subsurface. Although radar systems are characterized by their frequency of operation (e.g., 50, 100, and 500 MHz), the emitted signal is not a continuous sinusoid. The pulse width (z) in seconds is related to the system frequency (/), in Hz, asf=l/z. For example, a 100 MHz system has a pulse width of 10.8 sec. The downward propagating pulse depends on the conductivity (o) and the polarization (e) of the subsurface materials. The resulting subsurface profile shows features such as bedding, voids and fractures. There are depth limits to this technique, with signal attenuation becoming more important at depths where subsurface materials have low electrical conductivities, i.e., where pore fluids are present in quantity. What this means is that optimal conditions for this technique are sandy or rocky soils in unsaturated (vadose) zones or bedrock with low hydraulic conductivity where water is not permanently present. Clay-rich sediment, which by their nature retain more water than most sediments, and other dump or saturated sediments will generally yield poor results. However, reliable data are obtainable in settings where pore fluids have a low specific conductance, e.g., where significant amounts of petroleum products are present. 7.5.2
Electromagnetic
The application of electromagnetic techniques (EM) to the measurement of terrain conductivity is not new. Excellent descriptions of this approach can be found in the literature (Keller and Frischknecht, 1966; Wait, 1962). A typical induced current flow in homogeneous half space is shown in Figure 7.1. The system consists of a transmitter coil (TC), energized with an alternating current at an audio frequency, and a receiver coil (RC), located at a short distance away from the transmitter. The transmitter coil (TC) creates a time-varying primary magnetic field (Hp) that induces currents in the subsurface. These currents generate a secondary magnetic field Hs which is sensed, together with the primary field, Hp, by the receiver coil (RC). In general, this secondary magnetic field is a complicated fimction of the inter-coil spacing, s, the operating frequency, f and the ground conductivity, o. Under certain constraints, technically defined as operation at low values of induction number, the secondary magnetic field is a simple function of these variables:
150
SITE INVESTIGATION Hs
it~ktoOS 2 ezt
4
[7.1]
where Hs is secondary magnetic field at the receiver coil, lip is primary magnetic field at the receiver coil, co = 2gf, f is frequency (Hz), ~to is permeability of free space, cr is ground conductivity (mho/cm), s is inter-coil spacing (m), and i = q-1.
Figure 7.1. Induced current flow in homogeneous half space.
The ratio of the secondary to the primary magnetic field is linearly proportional to the terrain conductivity, a fact which makes it possible to construct a direct-reading, linear terrain conductivity metre by simply measuring this ratio. Given H/Hp the apparent conductivity, a a, indicated by the instrument is given, from Eq. [7.1 ], by: aa
[7.2]
-
CO~oS
The MKS units of conductivity are the mho (Siemen) per m or, more conveniently, the mmho p e r m.
Interpretation of Electromagnetic Data In order to determine the instrmnent response as a function of depth (EM-31 or EM-34 developed by Geonics Ltd., McNeill, 1980), consider a homogeneous half space on which a
GEOPHYSICAL TECHNIQUES
151
transmitter is located on the surface, as shown in Figure 7.2, for vertical and horizontal dipoles. Fixing our attention on a thin layer of thickness dz at a distance z (where z is the depth divided by the inter-coil spacing, s), it is possible to calculate the secondary magnetic field in the receiver coil arising from all the current flow within this or any other horizontal thin layer.
l *Jz) Tc
Figure 7.2(a). Relative response versus depth for vertical dipole. ~v (z) is the relative contribution to H~ from material in a thin layer dz located at depth z.
~
s
,.2
r
k...._ Figure 7.2(b). Relative response versus depth for horizontal dipole.
152
SITE INVESTIGATION
One can thus construct the functions ~v(Z) and ~/~(z) which are the relative contributions to Hs from material in a thin layer dz located at a distance z, as shown in Figure 7.3. It is interesting to see that for ~v(Z), the material located at a distance of approximately 0.4 s gives maximum contribution to the secondary magnetic field but the material at a distance of 1.5 s still contributes significantly. At zero distance, the terrain makes a very small contribution to the secondary magnetic field and therefore this coil configuration is insensitive to changes in near surface conductivity. For the case where both transmitter and receiver dipoles are horizontal co-planar rather than vertical coplanar, the relative contribution from the material near the surface is large and the response falls off monotonically with depth. Since the apparent conductivity (Eq. [7.2]) is given in terms of the secondary magnetic field at the receiver, the functions ~v(z) and ~ z ) also give the relative contribution from material at different distances to the apparent conductivity indicated by the instrument meter. The integral of either function from zero to infinity gives the total secondary magnetic field at the receiver coil from a homogeneous half space which is directly related to the electrical conductivity of the half space.
~"- "
9~*
Vertical RR (Y1) 9 9 9
(b
1.5
-
9 9
?,,_
~,",,,,
(3 O.
a:
:,,,
9
9 9
%
9
.....
Horizontal RR (Y1)
. . . . .
Vertical CR (Y2)
.................
Horizontal CR (Y2)
0.8
0.6
~
?
0.4
~, 0
0.5 0.2
0
0
I
I
I
0.5
1
?.5
Normalized
2
0
d i s t a n c e (z/s)
RR = Relative Response; CR = Cumulative Response Figure 7.3. Response versus depth for vertical and horizontal dipoles.
For multi-layered earth response, cumulative response functions Rv(z) and RH(z) are defined as" oo
R (z) = f a, Az)az; z
oo
R.(z) = f %(z)az z
[7.3]
GEOPHYSICAL TECHNIQUES
153
with graphical representations shown in Figure 7.3 (McNeill, 1980). To calculate the instrument reading on an arbitrarily layered earth, we simply add the contribution from each layer independently, weighted according to its conductivity and depth. For example, assume that we have a two-layer case, as shown in Figure 7.4.
Layer 1
G~
Layer2
a2
I
Z
Figure 7.4. Two layer earth model.
The contribution from the upper layer is given by: %
= o111
- Rv(Z)]
[7.4]
since all the material below the surface yields a relative contribution of unity or 100% to the meter reading. Conversely, all of the material in layer 2 adds a contribution given by: o
= ozRv(z)
[7.51
and the actual instrument reading will therefore be the sum of these two quantities o a -- Ol[1 - Rv(z)] + ozRv(z)
[7.6]
If the earth is three-layered, as shown in Figure 7.5, the same procedure is employed to determine the instrument response. In this example, the calculations are performed for different middle layer thickness. Since the electromagnetic technique produces results in conductivity units, it can be of immediate use at the site to: (1) Determine the water quality: The recorded conductivity values provide an indication of the relative difference in dissolved ion contents, hence provide a measure of water quality, as shown in Table 7.2.
154
SITE INVESTIGATION [_..q
S
C
......
Layer 1
ol
Layer 2
o= = 2 m S / m
Layer 3
03 = 20 mS/m
For z2 = 1; For z 2 1.5; =
= 20 mS/m
C
A
.J. z 1
u
A h Z2
oa = o, [ 1 - R (z,) ] + o2 [ R (z,)- R (z2) ] + o 3 [ R (z2) ] oa = 20 [1- 0.70] + 2 [0.70 - 0.44] + 20 x 0.44 = 15.3 mmho/m oo = 20 [1- 0.70] + 2 [0.70 - 0.32] + 20 x 0.32 - 13.2 mmho/m
Figure 7.5. Response calculation for three-layered system.
Table 7.2: Typical values of specific conductivities for various water types. Electrical conductivit), (mmho/cm)
Water type Natural groundwater Groundwater polluted by: Municipal landfill leachate Leachate from a fly ash disposal dump Sea water
0.14 5.99 4.61 36.30
(2)
Quantify the soil salinity: Salinity classes, based on electrical conductivity of a saturated extract, as defined by US EPA, are shown in Table 7.3.
(3)
Determine the ionic strength of soil pore fluid: The relationship between electrical conductivity and ionic strength of pore fluid solution is given by (Kayyal and Mohamed, 1997): o = 0.072971 + 0.172302xi ~
[7.7]
where o is electrical conductivity in mmho/cm, I is ionic strength of the cations in the equilibrium solution in mmole, which is given by (Benefield et al., 1982):
GEOPHYSICAL TECHNIQUES
I
C l
:
155
L
--, --)IC,Z 2 2 i=l
[7.8]
is concentration, and Z, is oxidation number.
Table 7.3: Salinity classes of soils Electrical conductivity (mmho/cm)
Class Non-saline Slightly saline Moderately saline Very saline Extremely saline
(4)
0- 2 2-4 4-8 8 - 16 > 16
Determine the concentration of exchangeable cations and heavy metals: This can be obtained with the use of the following relationships (Kayyal and Mohamed, 1997): (a) For exchangeable cations: o : 0.121249
(b)
+ 0.003155xC
0"940930
[7.9]
where o is electrical conductivity in mmho/cm, and C is the equilibrium concentration of the total exchangeable cations (ppm). For heavy metals: o = 0.004773
+ 0.001894xC
0"941766
[7.10]
where o is electrical conductivity in mmho/cm, and C is the equilibrium concentration of the total heavy metals in solution (ppm).
7.5.3
Surface Resistivity
Surface Resistivity (SR) methods differ fundamentally from electromagnetic methods because they require injection of current through electrodes into the ground. The EM methods induce currents in the ground. Resistivity methods can provide information on the surface in areas of moderately high resistivity where secondary EM fields would be small. This, coupled with their low cost and fairly simple equipment requirements, constitutes the primary advantage of the resistivity methods. Resistivity surveys use four or five electrodes, usually metal stakes spaced and driven into
156
SITE INVESTIGATION
the ground at 200 mm intervals along a straight line. A current is supplied through two electrodes by means of a battery or small generator, and a voltage drop or potential difference is measured between the other two. The most common electrode arrangement for constant-depth, transverse surveys is called the Wenner spacing (Figure 7.6), a traverse being made by stepping the last electrode ahead by a constant distance a . The Schlumberger spacing, shown in Figure 7.7, is advantageous for deep surveys. Essentially, the investigation depth is progressively increased by stepping the two outermost electrodes outward by increasing distances, a. The flow of electrical current through soil is analogous to the flow of water through soil, voltage being the driving force, analogous to water head. The current paths and voltage contours therefore may be represented by a f l o w net. As the electrode spacing a is increased, the ratio of measured voltage (V) to current (/) decreases in proportion if the soil is homogeneous, and not in proportion if the soil is layered, due to changing distortion of the flow net. Resistivity p is a property of a conductor, defined as electrical resistance per unit length of a unit cross-sectional area. For example, resistivity could be measured directly on a cubic block of soil 1 m on a side, by applying a voltage to two opposite sides. From Ohm's law, the resistance R = VII. Resistivity would therefore be p = R x (cross-sectional area/length) = R x 1 m~/1 m. If V is in volts and I is in amps, p will be in ohm. m, abbreviated f~.m.
0
II II
oro~176176
"
/ {,;......... E1
,
i m
,
,
"-
~
l
~-
~
,.
,
~--
...........
l
:
:
~
ql
I-
~'
:
:
"
,,-
.:
:
-
......
%'... _~"q~ -
~
~, _,,,|,,,'%
~
~'~,.
m
tia, ,i es ...'}~'"
......... ,/, .................
,r
,,"
Figure 7.6. Wenner electrode spacing used in resistivity surveys for homogenous material.
The soil used in resistivity measurements is, generally, far from being a cube. It can be shown by use of calculus that the effective resistivity for any spacing between two adjacent electrodes are: (1) For Wenner arrangement:
GEOPHYSICAL TECHNIQUES
157
V P = 2=a-I (2)
[7.11]
For Schlumberger arrangements: [7.12]
where V is in volts, I is in amps, a and b are spacings between electrodes in Wenner and Schlumberger arrays, respectively, and 9 is in ohm. m.
L II. Ground surface
n
I
I
Flow lines-
-
;-
,
_i
i
9
~
:
~.
~h.,,
~*.
d
..........
*~.,,,.
%,, %
,,,,,,"
:
=
",
"-..
"% ',,,.
~
~~
:
%
,. ~ - ~ '
I
-a
_ ~ _
~t~ . . . . i
~"
. . . . . . . .
],4
"-.~..,
%- . . . . , - ~ " "
-'.... -..... v
.~-...%
. . . . . ... . .
=J"
,'~_ ~.
n
.
aam-----mm . . .
a
;,,'......... s,....
:
,t,"
:
=
,,,:'
:
..-.~_
.
$
...~
9 | " " " ~-~..;""" I* ;,*. . . . . . . . . . . .,. . . . . . . . . ....',' . ..-~
.... . ,
.~ . . . . ~
oo 9
"~#.. .......
_,,
" .
:
.
,
~9
....,,". . . . .
I ~ q u t,;,~_'tlo~noat tmes
'
II
Figure 7.7. Schlumberger electrode spacing for a two-layer system.
Interpretation o f Resistivity Data A method of interpretation proposed by Moore (1961) is to plot ~P versus spacing a. If 9 is constant with depth, the result is a straight line with slope 91/a, where 9~ is the resistivity of the upper layer. As the spacing is increased and a different material is encountered, a second line, with slope 9Ja, is formed. The intersection of the two lines indicates depth of the boundary between the two layers. A method devised by Barnes (1952) treats the soil as layers which act as resistors in parallel. In this case it can be shown that the conductance in Siemens, S, or mho, or reciprocal of the resistance in ohm, of any layer may be found from:
158
SITE INVESTIGATION
Zn=EZn-~Zn_
'
[7.13]
where L, is defined as the conductance or reciprocal resistance of layer n, ~L, is the conductance of all layers down to and including layer n, and ~L,_I is the conductance of all layers down to but not including layer n. For calculations, we must therefore convert total resistance readings to total conductances, ~L. Successive ~L, and ~L,_1 values are then subtracted to obtain individual L~ values. The layer conductance L, is then converted to resistivity by use of a modified Wenner equation: 2 ~a n
Pn-
Ln
[7.14]
where a, is the incremental increase (m). The electrical potential measured between electrodes depends on the electrical properties of the geologic materials which in tum depend on the resistivity of the pore fluid and the amount of pore water. Most soil and rock materials are highly resistive while water is highly conductive. Because of concentrations of some solute, contaminant plumes frequently appear as a highly conductive layer. Resistivity variations as a function of ion concentration and lithology are shown in Figure 7.8 and Tables 7.4 and 7.5.
Figure 7.8. Resistivity variations with concentration for solution and saturated sand/gravel material.
GEOPHYSICAL TECHNIQUES
159
Table 7.4: Resistivity variations with lithology (Culley, 1975) Lithology Permeable Clayey soils Sandy soils Loose soils River sand and gravel Impermeable Clay Glacial till
Resistivity range ( ohm. m)
100 - 800 8000 - 10000 1000 - 100000 100 - 8000 1 - 100 10 - 8000
Table 7.5" Summary of typical resistivity measurements of fluids and bulk soil fluid mixtures (Campanella and Weemes, 1990) Water type
Resistivit' (ohm. m) Pore fluid
Drinking water Sea water McDonald farm clay Colebrook site clay Railway site clay B.C. Highway strong pit clay Mcdonald farm sand Colebrook site sand Strong pit sand Typical landfill leachate 100% Ethylene Dichloride (ED) 50% ED/50% 150 ohm. m fluid in Wedron 7020 sand 30% ED/70% 150 ohm. m fluid in Wedron 7020 sand 17% ED/83% 150 ohm. m fluid in Wedron 7020 sand
>15 0.2 0.3 18.2
1.5-6
Bulk solid
1.5 25 8 35 5-20 70 115
0.5-10 20400 696 335 273
If we compare the polluted saturated sand/gravel resistivity values (Figure 7.8) with values for lithology (Table 7.4) we find that materials having the ability to allow pollutant migration are at least one or more orders of magnitude more resistive than even a relatively dilute 5 ppm solution of cupric sulphate. Thus, the cupric sulphate pollutant can be clearly discernable from various lithology
160
SITE INVESTIGATION
of permeable materials. Furthermore, based on measured resistivity values, leachate corrosion potential, i.e., degree of hazard, can be estimated. The standard shown in Table 7.6 has been adopted in the United Kingdom (QJEG Working party, 1988).
Table 7.6: Classification of corrosion potential based on resistivity values Resistivity (ohm. m) < 10 -
>
10 100 100
Corrosion potential Severe Moderate Slight
Since the surface resistivity technique is limited to measurements at fixed electrode points, which must be removed prior to each new survey, it is slower than the EM technique, which is not limited by fixed points. Meaningful resistivity results cannot be obtained in areas which contain buried debris (landfill), dry sand, frozen ground, or near fences, or railroad tracks and underground pipes. 7.5.4
Seismic Surveys
In the seismic exploration method, a sound pulse is introduced into soil, and the time of first arrival of the vibrations at several different horizontal distances is recorded. The sound is introduced by a small explosive charge or by artistically banging a steel plate or ball into the ground with a sledge hammer; the latter technique is limited to survey depths of about 15 m or less. The sound is detected by a special vibration detector called a geophone. In some instruments, the hammer below starts an electric timer and the geophone stops it, and the time may be read directly from the timer. Most engineering seismographs use only one geophone, and either the geophone or the hammer is moved successively to different locations along a straight line. At each position the test is repeated to give the time-distance relationships, as shown in Figure 7.9. The reciprocal slope of the line is velocity in m/sec. As the survey distance is increased, usually by increments of 3 m, the velocity, indicated by a change in slope of the curve, increases. Figure 7.9 shows two velocities labelled vi and v2. The first (lower) velocity represents soil whereas the second velocity represents bedrock. The buried higher velocity material shortens the arrival times by a process of refraction. Refraction also may occur through a buried pipe or boulder, or laterally through a building foundation, sometimes leading to erroneous interpretation. Furthermore, refraction will not shorten arrival times if the surface layer has a higher velocity, therefore seismic exploration cannot be conducted through a pavement or frozen soil layer. The approximate depth D~ to a higher velocity layer can be calculated by assuming a rectangular travel path equal to L+2D l, where L is the distance between the source and the geophone.
GEOPHYSICAL TECHNIQUES
161
Dividing L and D~ by the respective velocities gives the total travel time: L
t =
+
2Dl
~_
Timer ITI
II
L
\
-}~
illl Hammer ....
o,
[7.15]
VI
V2
~sseajtl|.l~
"sss,.
I~
t
Geophone 7-
..~
~
?
1. ~
~.-
"" .........................'1~ ,,~
Ground surface
9 ,
,
D 1
,
C
.
V2 L - 2 D1,,tani
I
',
V2 (D
._E I--
4-
Horizontal 'clistance (L)
C1
Figure 7.9. Principle of seismic refraction.
The last term is a delay time due to twice traversing the lower velocity v~ layer. The delay time, tl, which may be read directly on a t versus L plot by extrapolating to L = 0, is given by: 2D l t~ -
from which
[7.16]
162
SITE INVESTIGATION tlV 1
D 1 "-
[7.17]
2
Eq. [7.15] is fairly accurate when v~ and v2 differ widely. However, when the difference is smaller the true travel path is trapezoidal, decreasing the refraction travel and delay times. By using the laws of refraction, one can show that the precise expression for delay time is: 2Dlcosi t~ :
[7.18] V1
from which fly 1
D1 -
[7.19]
2cosi
where i is the angle of incidence, and sin i = v, /Y2" A more convenient expression uses the intersection distance C~ whereby the times for the direct and refracted waves are equal. That is: C I V1
-
C l
+
V2
2DlCOSi
[7.20]
VI
Solving for D 1 and substituting for cos i gives: C1 /
1;2 - Yl
D 1 = - - ~ ~ V2 + V 1
7.5.5
[7.21]
Borehole Logging
This technique includes a variety of methods involving the lowering of a tool into the borehole (Keys and MacCary, 1971; Keys 1988). A summary ofborehole log applications is shown in Table 7.7. The tool measures the physical properties of the geological materials and the response of the system to induced disturbance. Common logging tools include calliper, resistivity, neutron, gamma, and sonic tools. Logging can proceed in both cased or non-cased boreholes, though most measurements can be made only when the hole has not been cased. Most of the logging methods are effective in distinguishing between sand and clay and are useful in locating zones of high hydraulic conductivity (Kwader, 1986). Resistivity logging is effective in identifying soil and rock types, geologic correlations, soil and rock porosity, and pore fluid resistivity. Natural gamma logging can assist in positioning wells
GEOPHYSICAL TECHNIQUES
163
and casings, by providing information on clay and shale content, grain size, pore fluid resistivity, and soil and rock identification. Gamma-gamma logging provides a means to place cementing materials for the well casings and to determine total porosity or bulk density. Neutron logs can provide estimates of moisture content above the water table, total porosity below the water table, specific yield of confined aquifers, the location of the water table outside the casing, chemical and physical properties of the water, and the rate of moisture infiltration. Temperature logs help to" (1) provide the chemical and physical characteristics of the water, (2) source and movement of the water in the well, and (3) dilution, dispersion and movement of the waste.
Table 7.7: Summary ofborehole log application Required information 9 Lithology and stratigraphic correlation 9 Porosity or bulk density 9 Porosity or true resistivity 9 Clay or shale content 9 Permeability 9 Specific yield of unconfined aquifers 9 Grain size 9 Location of water level or saturated zones 9 Moisture content 9 Infiltration 9 Direction, velocity, and path of ground water flow 9 Dispersion, dilution, and waste movement 9 Chemical and physical characteristics of water 9 Construction of existing wells 9 Screen setting 9 Cementing 9 Corrosion 9 Casing leaks and/or plugged screen
Logging technique 9 Electric, sonic, calliper logs, nuclear logs. 9 Sonic and Neutron logs. 9Normal resistivity logs. 9 Gamma logs. 9 No direct measurement of logging. 9 Neutron logs. 9 Electric logs in association with formation factor. 9 Electric, temperature, conductivity, neutron, gamma-gamma logs. 9Neutron logs. 9 Radioactive tracers. 9 Single and multi-well racer techniques, point dilution and single-well pulse. 9 Fluid conductivity, temperature logs, gamma logs. 9 Fluid conductivity, neutron chloride logging, multi-electrode resistivity 9 Gamma-gamma, calliper, perforation, borehole television. 9 Lithology logs. 9 Calliper, temperature, gamma-gamma, acoustic. 9 Calliper, collar locator. 9 Tracer and flow-meter.
164 7.5.6
SITE INVESTIGATION Video Cameras
They can be used for visual inspection and to provide a visual record of the wall of the borehole. They are particularly useful for inspecting the casing for corrosion, damage, or leaks, and also are used in uncased rock holes for locating fractures and fracture zones.
7.6
HYDROGEOLOGICAL INVESTIGATIONS
Field hydrogeological investigations are conducted to: (1) detect and identify subsurface pollutants, (2) measure the concentrations of pollutants, (3) characterize the geological environment through which pollutants migrate, and (4) determine the direction of groundwater flow. The choice of the most suitable well installation procedure depends on: (1) the expected geology of the site, (2) type of liquid waste and its anticipated effect on the drilling mud and well materials, and (3) the effect of the installation procedures on the reliability of the water quality data. 7.6.1
Drilling methods
Several drilling methods for sampling and installing groundwater monitoring wells are available (Hvorselv, 1965; GeoTrans, 1989). In choosing the best method, site geology, along with size and type of well materials, are critical. A required well material may influence available drilling methods. Well casing and screens constructed from PVC or thin-wall stainless steel are best installed using rotary or auger drilling methods. Conversely, cable tool drilling requires the use of black steel or galvanized steel casing. Often, a combination of two or more drilling techniques is used to complete monitoring wells. Commonly used drilling methods are briefly discussed below.
Hollow-stem Auger Hollow-stem auger is one of the most desirable drilling methods for constructing monitoring wells. No drilling fluids are used and disturbance of the geologic materials penetrated is minimal. Depths are usually limited to no more than 50 m. Typically, auger rigs are not used when consolidated rock must be penetrated. In formation where the borehole will not stand open, the monitoring well can be constructed inside the hollow-stem auger prior to its removal from the hole. The diameter of the well, that can be built, is limited to 100 mm. The hollow stem auger has an added advantage in offering the ability to collect continuous in-situ geologic samples without removal of the auger sections.
Solid-stem Auger Solid-stem auger is most useful in fine grained, unconsolidated materials that will not collapse when unsupported. The method is similar to the hollow-stem except that the auger flight must be removed from the hole to allow the insertion of the well casing and screen. Cores cannot be collected when using a solid stem auger. Hence, geological sampling must rely on cuttings that come to the surface. This, in turn, creates a possibility of cross contamination and makes it difficult to have a precise logging.
HYDROGEOLOGICAL INVESTIGATIONS
165
Cable-tool Drilling Cable-tool drilling is one of the oldest methods used in the water well industry. The method offers many advantages for monitoring well construction. With the cable-tool, excellent formation samples can be collected and the presence of thin permeable zones can be detected. As drilling progresses, a casing is normally driven and this provides an ideal temporary casing within which to construct the monitoring well.
Air-rotary Drilling In air-rotary drilling, air is forced down the drill stem and back up the borehole to remove the cuttings. This technique has been found to be particularly well suited to drilling in fractured rock. If the monitoring is intended for organic compounds, the air must be filtered to ensure that oil from the air compressor is not introduced into the formation to be monitored. Air-rotary should not be used in highly polluted environments because the water and cuttings blown out of the hole are difficult to control and can pose a hazard to the drill crew and observer. Where volatile compounds are of interest, air-rotary can volatilize them and cause waste samples to be unrepresentative of insitu conditions. The use of foam additives to aid removal of the cuttings presents the opportunity of organic contamination of the monitoring well.
Air-percussion Rotary or Down Hole Hammer Air-rotary with percussion hammer increases the effectiveness of air-rotary for materials likely to cave and highly creviced formations. Addition of the percussion hammer gives air-rotary the ability to drive casings through the porous material. The capability to construct monitoring well inside the driven casing, prior to its being pulled, adds to the appeal of air-percussion. However, the problems with contamination and crew safety must be considered.
Reverse Circulation Drilling This technique has limited application for monitoring well construction. Reverse circulation rotary requires that large quantities of water be circulated in the borehole and up the drill stem to remove cuttings. If permeable formations are encountered, significant quantities of water can move into the formation to be monitored, thus altering the quantity of the water to be sampled.
Hydraulic Rotary Hydraulic or mud rotary is probably the most popular method used in the water well industry. In hydraulic rotary technique, a drilling mud (usually bentonite) is circulated down the drill stem and up the borehole to remove cuttings. The mud creates a wall on the side of the borehole that must be removed from the screened area. With small diameter wells, the drilling mud is not always completely removed. The ion-exchange potential of most drilling muds is high and may effectively reduce the concentration of trace metals in water entering the well. In addition, the use of biodegradable, organic drilling muds, rather than bentonite, can introduce organic compounds to water samples from the well.
7.6.2
Sampling Methods
There are many altemative techniques for the collection of samples. The most appropriate technique depends on the type of material being sampled and the drilling technique being used.
166
SITE INVESTIGATION
Drill Cutting Samples During any drilling operation, the cutting action of the drill bit produces fragments of the geologic material being penetrated. In a preliminary drilling program, these cuttings can be used to provide basic information on mineralogy, grain size, and stratigraphy. Core Samples For laboratory studies of hydraulic characteristics and diffusion through undistributed samples, continuous core sampling is preferable. The most common coring technique is the split spoon sampler. This device consists of a hardened steel drive shoe screwed onto a hollow steel tube that is split down the middle. The top of the tube is connected to a head assembly that attaches to the drill rod. The core tube is 0.45 "or" 0.6 m in length and is commonly between 0.05 and 0.10 m in diameter. With the advancement of a borehole installation, a sample can be obtained at the specified depth. The sampler is driven into the geological material by a hammering device on the surface. The number of blows required to advance the sampler its full length is recorded. This technique is generally referred to as standard penetration test (ASTM D1586-84). The hollow stem auger rig provides the most efficient and reliable method of collecting split spoon samples. 7.6.3
Well Installation Techniques
After the test bore hole is drilled and the subsurface material is sampled, a monitoring well device can be installed in the bore hole for groundwater sampling and measuring piezometric levels. The most common techniques for monitoring well installations are discussed in the following sections.
Drive Point Wells The simplest and least expensive technique to install a monitoring well is to drive the well screen and casing down to the desired depth with a hammering device. Drive point monitoring wells commonly range in diameter from 15 to 30 mm. They have been successfully installed in soft soils up to 30 m. The length of the screen commonly ranges between 0.2 and 1.0 m. Several other drive point monitoring techniques are described by Desaulnier (1983). Individual Wells The most common type of monitoring well is the individual well installed in a drilled borehole. These types of monitoring wells commonly range from 0.02 to 0.1 m in diameter. They are typically constructed of steel, stainless steel, PVC, or Teflon, depending upon the requirements for chemical sampling. The screen and filter pack should ensure that formation water can pass easily into the monitoring well. The placement technique is as follow: (1) place the selected well screen and casing down into the borehole to the required depth, (2) install permeable filter pack material around and slightly above the well screen to allow groundwater from the adjacent formation to flow freely to the well screen, (3) place a sealing material above the filter pack to isolate the well screen from the rest of the borehole, (4) backfill the annulus above the seal with a grouting material, and (5) install a protective cover over the well casing at ground level for security and to prevent precipitation from entering the well.
HYDROGEOLOGICAL INVESTIGATIONS
7.6.4
167
Monitoring Well Design Components
A monitoring well is built to give access to the groundwater so that a representative sample of water can be withdrawn and analysed. The design components of a monitoring well must not materially alter the quality of the water being sampled. An understanding of the chemistry of suspected pollutants and the geological setting in which the monitoring well is to be constructed play a major role in the drilling technique and well construction materials used. The major components that need to be considered in monitoring well design are shown in Figure 7.10 and discussed in the following sections.
Figure 7.10. Monitoring well components.
Diameter
The diameter of a monitoring well was generally based on the size of the device (bailer, pump, etc.) used to withdraw water samples. This practice worked well in very permeable formations, where an aquifer capable of furnishing large volumes of water was present. However, monitoring wells are quite often completed in very marginal water-producing zones. Hence, a pumping one or more well volumes of water from a well built in low-yielding materials may present a serious problem if the well has a large diameter. In addition, when hazardous liquid wastes are present in the groundwater, the purged water must be properly disposed. Therefore, the quantity of water pumped from the well should be minimized for reasons of safety, as well as disposal cost. For these reasons, a 50 mm inner diameter wells have become the standard in monitoring well technology. For cases whereby monitoring is followed by treatment of groundwater and polluted soils, large diameter wells used for monitoring can be used as a supply well to remove polluted water for treatment. Also, since large diameter wells have higher strength, they are often used for deep monitoring.
168
SITE INVESTIGATION
Casing and Screen Material The type of material used for a monitoring well can have a distinct effect on the quality of water sample collected. The material of choice should neither adsorb nor leach chemical constituents that would change the representativeness of the samples collected for plume detection. Figure 7.11 is a schematic representation of material selection procedures for a plume detection program.
I Material selection I
I
Subsurface pollution condition
Unknown
IHigh
Polytetrafluoroethylene Stainless steel Polyvinyl chloride
High organic or
inorganic Suspected
inorganic s u s p e c t e d
I
I
_~
]
I
NO
~ YES
Polytetrafluoroethylene I Stainless steel Figure 7.11. Schematic representation of material selection procedures for plume detection.
The types of materials that are generally used are: (1)
(2) (3)
Polyvinyl chloride (PVC): Because of PVC's low cost and easy installation, it has been used extensively for casing and well screens. PVC is inert to chemical reactions in nearly natural environments. However, PVC deteriorates when it comes in direct contact with low molecular weight ketones, aldehydes, and chlorinated solvents. Generally, as the organic content of a solution increases, direct attack on the polymer matrix or solvent adsorption or leaching may occur. Teflon: Teflon is considered to be the most inert well material. However, because it is expensive, it is used where no chemical interferences can be tolerated. Galvanized steel casing: Galvanized casing can be superior to PVC because it is inert to organic chemicals and more durable if the well must be driven into the formation. The galvanic coating inhibits rust formation which otherwise decreases the life of the well. Also,
HYDROGEOLOGICAL INVESTIGATIONS
(4)
169
galvanized casing can increase iron, manganese, zinc, and cadmium concentrations in water. Steel casing may contribute to sample contamination due to increase of iron and manganese concentrations. Therefore, PVC casing is preferred for monitoring groundwater polluted by heavy metals. Stainless steel casing: Stainless steel is inert to virtually all pollutants. However, at low pH, stainless steel may release chromium into groundwater. Also, it may catalyse some organic reactions. The primary disadvantage is its high cost.
Sealing Materials When drilling a hole using rotary, auger or jetting methods, the final borehole diameters are larger than the well casing. In order to prevent the flow of polluted groundwater into the well, the annuals between the well casing and the borehole wall is grouted with bentonite, cement or a bentonite/cement mixture. Some concerns related to each grouted material are discussed below. (1) Bentonite grout: Bentonite slurry mixture is used generally in drilling mud and also to act as a borehole seal after the well is completed. The bentonite structure is alumino-silicate sheets bonded through cation bridging, as discussed in Chapter 4. Bentonite clay has appreciable ion exchange capacity, which may interfere with the chemistry of collected samples when the seal is adjacent to the screen or well intake. (2) Cement grout: It is used to seal the annulus, particularly after setting casing in a hole drilled with rotary methods. Cement is more permeable to groundwater than bentonite, thus cement is sometimes considered unsuitable as a grouting material. However, cement is rigid and allows for better integrity around the well casing. When improperly placed, cement grout has been known to seriously affect the pH of sampled water. Upon curing and weathering, cement grout may undergo shrinkage and cracking. (3) Bentonite cement mixture: Bentonite and cement mixtures are often used to create a slurry for grouting. After setting, the grout is slightly weaker than pure cement and little more permeable than bentonite. Variations in the mix can enhance the structural strength or the impermeability of the grout. Screen Length and Depth of Placement The length of well screen and the depth at which it is placed in the ground depend on: (1) the behaviour of the pollutants as it moves through unsaturated and saturated zones, and (2) the goal of the monitoring program. When monitoring an aquifer, used as a water supply, the entire thickness of the water-bearing formation could be screened. However, when specific depth intervals must be sampled at one location, vertical nesting of wells is common. This technique is often necessary when the saturated zone is too thick to adequately monitor with a long screened section. Screen lengths of 0.3-0.6 m are common in detailed plume geometry investigations. Monitoring of non-aqueous phase liquids (NAPLs) demands special attention. In particular, light NAPLs, i.e., with densities less than water, will float on the groundwater surface. Monitoring wells constructed to detect floating pollutants should contain screens that extend above the zone of saturation so that these lighter substances can enter the well. The screen length and position must accommodate the magnitude and depth of variations in water table elevation. Location and Number Monitoring well locations and the number of wells in the monitoring program are closely
170
SITE INVESTIGATION
linked and determined by the purpose of the monitoring program. Most dissolved constituents migrate vertically through the unsaturated zone beneath the area of activity and then, upon reaching the saturated zone, move horizontally in the direction of groundwater flow.
Figure 7.12. Typical monitoring well network.
A typical monitoring well placement scheme is shown in Figures 7.12(a) and 7.12(b). Well "A" is the background monitoring well and is located far enough up-gradient from the site to ensure that the landfill has no effect on the hydraulic conductivity of the soil at the well. Well "B" is located on site, and is placed in a location where migrating pollutants can be detected. The "B" well will also serve as a first indicator of the effectiveness of the remedial action program. If water quality in the monitoring well does not indicate a steady improvement over time, it will indicate the need for further remedial action considerations. This well must be very carefully constructed and sealed, in order to prevent vertical migration of pollutants down the well casing. Well "C" is located downgradient from the site, at a position close enough to detect changes in groundwater quality as soon as possible. These wells should also, over a period of time, show a similar trend of improvement of groundwater quality. These wells should be screened over the entire distance of the aquifer to ensure that a leachate plume is not passing under the well system. Site geology, site hydrology, pollutant characteristics, and the size of the area under investigation all help determine where and how many wells should be constructed. Certainly, the more complicated the geology and hydrology, the more complex is the transport of pollutants. The
HYDROGEOCHEMICAL INVESTIGATIONS
171
larger the area being investigated, the greater is the required number of monitoring wells. 7.6.5
Well Decontamination Procedures
Prevention of contamination of monitoring wells must be considered at all phases of water well construction, from the initial soil boring stages to the final water sampling and water measuring stages. The most used techniques for well decontamination are: (1) Steam cleaning: The pressurized steam frees residual soil materials, washes them from the augers, and strips organic chemicals from the metal surface. Surfactants are added to the makeup water to more completely remove oil and grease from the drilling equipment. (2) Heat treatment: Heat is occasionally used to remove residual organic pollutants. Equipments such as augers, bits and wrenches are stripped of organic chemicals using an open flame which is less preferred for safety aspects. (3) Sand blasting: Sand blasting is sometimes used to strip soil materials from augers, bite and tools. After sand blasting, equipment is normally decontaminated using a steam cleaner and solvent rinses.
7.7
H Y D R O G E O C H E M I C A L INVESTIGATION
7.7.1
Subsurface Environment
Sampling of subsurface tends to change the chemical equilibrium, hence new reactions could be developed. The potential geochemical effects of drilling methods, materials used for well construction and sampling devices, and sampling methods must all be integrated when developing a sampling program. The sensitivity of a chemical system to disturbance depends on a number of physical and chemical environmental parameters. The major geochemical parameters that characterize the subsurface may include: (1) pH and alkalinity, (2) redox potential, (3) salinity and dissolved constituents, (4) soil matrix, (5) temperature and pressure, and (6) microbial population.
Table 7.8: Soil pH classes (US Soil Survey Staff, 1991). Class Ultra acid Extremely acid Very strongly acid Strongly acid Medium acid Slightly acid
pH <3-5 3.5 - 4.5 4.5 - 5.0 5.1 - 5.5 5.6 - 6.0 6.1 - 6.5
Class Neutral Mildly alkaline Moderately alkaline Strongly alkaline Very strongly alkaline
pH 6.6- 7.3 7.4- 7.8 7.9-8.4 8.5 - 9.0 >9.0
172
SITE INVESTIGATION
pH and Alkalinity pH and alkalinity are the main variables that affect solution composition and precipitation reactions. According to Barcelona et al., (1988), sampling methods and material may affect subsurface pH. For example, when using cement as a grouting material, pH may increase by 4 to 5 units. Also, during well purging, pH may increase or decrease by 0.1 to 5 units. These changes in pH will in turn affect geochemical processes in the subsurface environment. A US Soil Survey Staff (1991) identified 13 pH classes for soils, as shown in Table 7.8. Subsurface geochemical processes that may be affected by pH are : (1)
Acid-base reactions: strong acids and bases tend to flocculate and disperse clay structures,
(2)
Adsorption-desorption: pH strongly influences adsorption because hydrogen ions play an
respectively;
(3) (4)
(5) (6) (7)
(8) (9)
active role in both chemical and physical bonding processes. Mobility of heavy metals is strongly influenced by pH. Adsorption rates of organic and natural clays are also pHdependent; Precipitation-dissolution: acidic solutions tend to dissolve carbonates and clays while highly alkaline solutions tend to dissolve silica, as discussed in Chapter 5. Generally, higher pHs increase cation exchange capacity of clays; Complexation: strongly influences position of equilibria involving complex ions and metal chelate formation; Oxidation-reduction: redox systems generally become more reducing with increasing pH; Biodegradation: high to medium pH and low Eh environments will generally restrict bacterial populations to sulfate reducers and heterotrophic anaerobes (Baas-Becking et al. 1960); Salinity: pH-induced dissolution increases salinity while pH-induced precipitation decreases salinity; Temperature: pH-driven exothermic (heat-releasing) reactions increase fluid temperature while pH-driven endothermic (heat-consuming) reactions decrease fluid temperature; and Pressure: is affected only when pH-induced reactions result in a significant change in the volume of reacting products.
Alkalinity indicates the buffer capacity or resistance to change in pH. A solution with a high buffer capacity has a large resistance to change in pH. Since carbonate buffering is common to most natural waters, the solution pH may be quite sensitive to volatilization of CO2 during sampling operations.
Redox Potential The oxidation-reduction potential, or Eh, is an expression of the intensity of redox conditions in a system. It is measured in volts or millivolts (mV) as the potential difference between a working electrode and the standard hydrogen electrode. Positive readings in natural water generally indicate oxidizing (aerobic) conditions, and negative readings indicate reducing (anaerobic) conditions. The following classes are used by US EPA (1991): (1) highly oxidated for Eh greater than +400 mV, (2) intermediate for Eh ranging from +400 to -100 mV, and (3) highly reduced for Eh less than-100 mV. Near-surface soils and sediments have Eh values of + 299 mV or lower while surface water bodies are generally around +499 to 600 mV because they are often in equilibrium with oxygen in
HYDROGEOCHEMICAL INVESTIGATIONS
173
the atmosphere. Most redox reactions in the subsurface are microbially mediated, as discussed in Chapter 21. The measurement of the major by-products of these reactions may be a better indicator of the strength of the reducing environment than Eh measurements or calculated equilibrium potentials.
Salinity and Dissolved Constituents Total dissolved solids (TDS) content can be qualitatively estimated in the field by measuring specific conductance. During well development, if purging or sampling groundwater is mixed with water of different salinity or chemical composition, the results may be precipitation-dissolution and redox reactions that significantly change the inorganic chemistry of a sample. Geochemical sampling of wells is problematic because of these effects. The more saline the water, or the more different in chemical composition the two waters, the greater the bias that can be introduced to geochemical samples. Salinity classes, based on electrical conductivity of a saturation extract, are defined in US EPA (1991), as shown in Table 7.9.
Table 7.9: Salinity classes (US EPA 1991) Class Non-saline Slightly saline Moderately saline Very saline Extremely saline
Electrical conductivity (dS/m or mmho/cm) 0- 2 2-4 4-8 8 - 16 > 16
Soil Matrix The mineralogy and particle size distribution of the saturated and unsaturated zones strongly influence geochemistry of subsurface waters. As particle size decreases, the surface area increases, providing more opportunities of chemical reactions between solids and water. A particularly important chemical parameter of solids is the cation exchange capacity which is a function of mineralogy, particle size, and previous geochemical history. It may be a good measure of the potential attenuation of pollutants by ion exchange or sorption reactions. Characterization of clay mineralogy can provide considerable insight into subsurface geochemistry, as discussed in Chapters 4 and 5. Temperature and Pressure Temperature and pressure directly influence the rate of chemical reactions. As pressure increases, the amount of dissolved gases in solution tend to increase. Consequently, sampling methods that allow gases and volatile organic carbons (VOC) to de-gas to the atmosphere may tend to underestimate concentrations. The deeper the sampling, the greater is the potential for errors resulting from pressure changes.
174
SITE INVESTIGATION
Microbial Activity Groundwater contains diverse populations of microorganisms. The main limitation to microbial growth in the subsurface is low levels of nutrient and dissolved organic carbon, as discussed in Chapter 21. Most organic pollutants are readily degraded under aerobic conditions and any pollutant loading that adds more than traces of contaminants will rapidly deplete the available natural oxygen supply. Whether a specific pollutant will be degraded depends on geochemical conditions and on the presence of microorganisms that are capable of adaptation. Redox potential and water chemistry can provide considerable insight into subsurface microbial activity. Nitrogen, ammonia, hydrogen sulfide, and methane in groundwater are all initiators of microbial activity. Presence of carbon dioxide also may indicate microbial activity. However, the presence is more difficult to interpret because carbon dioxide may come from inorganic sources such as calcium carbonate and dolomite. 7.7.2
Sampling Considerations
The goal of a sampling program is often to avoid underestimating a particular impact either in terms of concentration or partial distribution. Characterization of geochemical variability is necessary to identify potential chemical problems that may affect the choice of a specific remedial technique. Sample location and frequency are among the most critical aspects of a sampling program. Good vertical and horizontal resolution of hydrogeologic conditions are essential before choosing sample locations.
Sampling Location The spatial distribution of pollutants is a major concern in a sampling program. The intensity and number of samples depend on concentration variability. To obtain representative samples, the screen length should be at least 1.5 m. There are two broad designs for sampling: (1) grids in which samples are taken from a matrix of squares or quadrants at a site, and (2) transects in which samples are taken at specified intervals along a line. Grids presume an aerial or dispersed source of some kind, and transects presume a preferential source. Grids can be used to estimate short range correlation. Transects along the path of groundwater or pollutant movement provide the best way to look at long range correlation. The combination of the two strategies coupled with the initial analysis of selected solid samples at alternate grid or transect locations can be quite effective. For soils, at least 5 percent of sampling points should be duplicated to help determine the sampling variability.
Sampling Frequency The estimated ranges of sampling frequency, in months, are shown in Table 7.10. Also, sampling frequency can be estimated by using (Barcelona et al., 1985):
f
nxd_ kwX i
[7.22]
wheref is the frequency of sampling, n is effective porosity, d is distance along flow path, kw is soil hydraulic conductivity with reference to water permeation, and i is hydraulic gradient.
HYDROGEOCHEMICAL INVESTIGATIONS
175
Table 7.10: Estimated ranges of samplin~ frequency in months (US EPA, 1991) Parameter
Water quality 9Trace Constituents (< 1.0 mg/1) 9Major constitutents (1.0 to 1000 mg/1) Geochemical 9Trace constitutents 9Major Constitutents Contaminant indicator 9Total organic carbon (TOC) 9Total organic halogen (TOX) 9Conductivity .pH
Pristine background conditions
Contaminated Up-gradient
Down-
gradient
2-7
1-2
2-10
2-7
2-38
2-10
1-2 1-2
<2 7-14
1-5 1-5
2 6-7 6-7 2
3 24 24 2
Sample Type and Size Sampling program must account for the nature of the subsurface condition. For example, if the subsurface has obvious fractures and channels, samples should be obtained from both affected and apparently non-fractured areas for comparison. Soil sample quantities of less than 100 g tend to be unrepresentative (Williams et al., 1989).
Vadose Zone Sampling Simple techniques for surface sampling include the hand auger, brace and bit, and post hole diggers. The most commonly used core sampling devices are split spoons or Shelby tubes that provide a continuous or driven core during drilling operations. Sampling continuously or ahead of hollow-stem drilling augers are good ways to obtain unpolluted and minimally disturbed soil samples. Suction lysimeters are used to sample pore water in the vadose zone. The zone of sampling influence with a suction lysimeter is about 0.1 to 0.2 m for a 24 hour period (Morrison and Lowery, 1990). In some instances, longer suction sampling may extend the influence to 0.5 m. Soil gas sampling generally involves driving a probe into the subsurface. Typically, the probes are driven by hand or some kind of pneumatic or electric hammer. Soil gas is obtained by applying a vacuum that brings the soil gas into the vicinity of the tip of the probe. Samples are collected in fluorocarbon bags or syringes and analysed on site or in a laboratory. At least 5 percent of air-filled porosity is required to pull a vacuum to obtain samples.
Groundwater Sampling Efforts to develop reliable sampling protocol and optimize sampling procedures require
176
SITE INVESTIGATION
particular attention to: (1) method of sampling effects the integrity of groundwater samples (Barcelona et al. 1985; Stolzenberg and Nichols, 1985), (2) potential error involved in well purging, (3) delivery tubing exposures, (4) sample handling, and (5) the impact of sampling frequency on both the sensitivity and reliability of chemical constituent monitoring results. The following steps are necessary for achieving a successful sampling protocol. STEP 1:
Water level measurements: These measurements are needed to estimate the amount of water to be purged prior to sample collection. STEP 2
Wellpurging: The removal of stagnant water from the monitoring well is referred to as well purging. Generally, it is recommended to remove 3, 5, or 10 well volumes. However, these recommendations can cause time delays and unnecessary pumping of excess contaminated water because they largely ignore the hydraulic characteristics of individual well and geologic settings. It is also recommended that pumping continues until well purging parameters (e.g., pH, temperature, specific conductance, redox potential) stabilize to plus or minus 10 percent over at least two successive well volumes pumped. The advantage of using the same pump to both purge stagnant water and collect samples is the ability to measure pH and specific conductance in an in-line flow cell. Since pH is a standard variable for aqueous solutions that is affected by de-gassing and depressurization (i.e., loss of CO2), in-line measurements provide more accurate and precise determinations than discrete samples collected by grab sampling mechanisms. STEP 3 Sample collection and handling: The order in which samples are taken for specific types of chemical analyses should be decided by the sensitivity of the samples to handling and the need for specific information. For example, samples for organic chemical constituents determinations are taken in decreasing order of sensitivity to handling errors (i.e., volatile organic, dissolved gases, large volume samples for organic compound determinations) while the inorganic chemical constituents, which may require filtration, are taken afterwards (i.e., alkalinity, acidity, trace metals, sensitive inorganic species, NO2,NH4 +, S, major cation and anions). Table 7.11 shows the US EPA recommended sample handling and preservation procedures for a detective monitoring program. STEP 4 Quality assurance~quality control (QA/QC): The use of field blanks, standards, and spiked samples for field QA/QC performance is analogous to the use of laboratory blanks, standards, and procedural or validation standards. The fundamental goal of field QC is to ensure that the sample protocol is being executed faithfully and that situations that might lead to an error are recognized before they seriously impact the data. Field blanks and standards enable quantitative correction for bias, which arise due to handling, storage, transportation, and laboratory procedures. Spiked samples and blind controls provide the means to correct combined sampling and measurement accuracy for the actual conditions to which the samples have been exposed. All QC measures should be performed for at least the most sensitive chemical constituents for each sampling.
HYDROGEOCHEMICAL INVESTIGATIONS
177
Table 7.11: Recommended sample handling and preservation for a detective monitoring program (modified from Scalf et al., 1981). Parameter
Well purging .pH 9Specific conductance 9Temperature 9Redox potential Pollutant indicators 9pH 9Specific conductance 9Total organic carbon 9Total organic halogen
Volume required ml/sample)
F/L
Container (material)
Preservation method
Maximum holding period
T,S,P,G T,S,P,G T,S,P,G T,S,P,G
None None None None
T,S,P,G T,S,P,G G, T G, T
None None Dark, 4~ Dark, 4~
10 100
G,S T, G,P
Dark, 4 ~ C 4 ~ C/None
< 2 4 hrs < 2 4 hrs
1000 50 100 50 400
T, P T,P,G T,P, G T,P,G T, P, G
pH<2; HNO3 4~ 4~ C 4~ C 4 ~ C/H2SO4; pH2 4 ~ C/H3PO4; pH4
6 months 1-7 days 24 hrs 7 days 1-7 days
4 ~ C/HNO3; pH2
6 months 7 days
4~
24 hrs
50 100 1000 1000
50 100 40 500
F F L L
Water Quality 9Dissolved gases (02, CH4, C O 2 )
9Alkalinity/acidity 9Fe, Mn, Na, K, Ca, Mg 9PO4, Cl', silicate N0 3 9S O 4
9
9NH 4 9Phenols Drinking water 9As, Ba, Cd, Cr, Pb, Hg, Se, Ag, rE"
Remaining Organic parameters
500
L
T, G
1000 50
L L
T, P T,P
500
L
24 hrs
T = Teflon; S = Stainless Steel; P = PVC, G = Borosilicate Glass. F = Field; L = Laboratory.
178
SITE INVESTIGATION STEP 5
Storage and transport: The storage and transport of groundwater samples often are the most neglected elements of the sampling protocol. Care should be taken to respect the maximum holding periods shown in Table 7.11 for various parameters. The documentation of actual sample storage and treatment should be kept in order.
Table 7.12: Percentage of variance due to laboratory error, field error and natural variability by chemical and site (Barcelona et al., 1989) Sand Ridge Beards Town Beards Town (up-gradient) (down-gradient) Parameter L F N L F N L F N Water quality 9 N O 3" 98042
9SiO2 9 O - P O 4 2-
9T-PO42" 9CI 9Ca 9Mg 9Na 9K
0 0 0 1.2 0 7.2 0 0 0 0
Geochemical 9NH 3 0 9NO 2 NA 9 S 2NA 9Fe 2+ NA 9Fe-r 0 9Mn-r 0
0 0 NA 1.2 NA NA 45.7 20 0 0
100 100 100 97.6 100 92.8 54.3 80 100 100
0.1 0.2 0 0 2.8 0 0 0 0 33.9
NA NA 20 0 NA 3.3 2.3 2.2 0.3 NA
99.9 99.8 80 100 97.8 96.7 97.7 97.8 99.7 66.1
0.2 1.4 0 0 0.9 0 0 0 0 87.1
NA 99.8 0.1 98.6 6.8 93.2 0 100 NA 99.1 17.2 82.8 3.6 96.4 2.8 97.2 7.1 92.9 NA 12.9
0 NA NA NA NA NA
100 NA NA NA 100 100
0 0.1 NA 0 0 0
0 NA NA 0.1 0 40.1
100 99.9 NA 99.9 100 59.9
0 0.3 NA 0 0 0
0 100 NA 99.7 NA NA 5.9 94.1 NA 100 73.6 26.4
Pollutant indicator
L+F
L+F
L+F
9TOC 15.4 84.6 29.9 70.1 40.6 59.5 9TOX 0 100 12.5 87.5 24.6 75.4 NA indicates that the number of observations on which the estimated variance was less than 5 or the estimated variance was negative, L = laboratory, F = field, and N = natural.
GEOCHEMICAL DATA COLLECTION 7.8
179
GEOCHEMICAL DATA COLLECTION
Many measured environmental parameters generally exhibit various degrees of uncertainty or randomness in the results obtained. Random errors result from slight differences in the execution of the same sampling procedures. Systematic errors result from procedures that alter the properties of the sample. Random error is unavoidable, but must be evaluated to determine its effect on accuracy. 7.8.1
Sources of Errors
Possible sources of error, in groundwater sampling, can be found in: (1) site selection, (2) sampling, (3) measurement methods, (4) reference samples for calibration, and (5) data handling. Both random and systematic errors may be involved in each stage. Errors at each stage are cumulative, but are not of equal significance or magnitude. Total variance in geochemical data results from the combination of natural geochemical variability and the cumulative error. The percentage of variance attributable to natural variability may often be greater than either field or laboratory error. Natural variance cannot be reduced. However, variance resulting from field and laboratory error can be reduced so that the actual variance closely approximates the natural variance. Estimates of the relative contribution of natural variability, field error, and laboratory error to total variance at three sites of groundwater investigations are shown in Table 7.12 (Barcelona et al., 1989). For most chemical constituents, natural variability accounted for more than 90% of the variance. For most inorganic constituents, where field and laboratory could be estimated, field error contributed a large percentage of total variance. Organic pollutant indicators showed much higher percentage of variance due to field and laboratory error than inorganic pollutant indicators.
Table 7.13: Sources of errors in groundwater sampling (Barcelona et al., 1985) Step 9 Establishing a sampling point 9 Field measurements 9 Sample collection 9 Sample transfer 9 Field blanks/standards 9 Field determinations 9 Preservation/storage 9 Transportation
Sources of error
~" c,t
Improper well construction and material selection. Instrument malfunction; operator error. Sampling mechanism bias; operator error. Sample exposure; de-gassing; oxygenation. Operator error; matrix interferences. Instrument malfunction; operator error. Matrix interferences; handling/labelling errors. Delay; sample loss.
Field Error Specific possible sources of error at various steps in groundwater sampling are shown in Table 7.13 (Barcelona et al., 1985). The largest sources of error are unrepresentative sample
180
SITE INVESTIGATION
locations and disturbances caused by drilling and well construction. Sample collection is the next largest source of error. Major sources of systematic sampling error include: (1) well construction and improper screen design, (2) improper purging, (3) well casing materials, (4) sampling mechanisms and grouting/sealing, (5) sampling tubing can result in uncertainty in VOC measurements, (6) changing sampling procedures and personnel without a strictly defined sampling protocol, and (7) failure to document unavoidable deviations from sampling protocols. Potential contributions of sampling methods and materials to uncertainty in geochemical results are shown in Table 7.14 (Barcelona et al., 1988). Well purging procedures can result in large variations in pH, TOC, Fe (II), and VOC. The next largest source of error is well casing, followed by sampling mechanisms and grouting. Poorly grouted or cemented wells can greatly alter the pH of the water.
Table 7.14: Potential contributions of sampling methods and materials to uncertainty in geochemical results (Barcelona et al., 1988) Sampling method and material
pH
Chemical parameters TOC (mg/l) Fe(II)(ms/1) VOC (~tg/1)
Concentration range Drilling muds Grouts, seals Well purging Well casing
5- 9 .... +, 4- 5 +, 0.1- 5 ....
0.5- 25 +, 300% . . . . . +, 500% +, 200%
0.01-10 0.15- 8000 . . . . . . . . ,500% .... -, 1000% +, 10- 1000% +, 1000% +, 200%
Sampling 9Mechanism 9Tubing
gas lift, bailer, gas left, +, 0.1 - 3 +, 150% -, 500% . . . . . . . . . . . . .
suction, -, 1- 15% ,10- 75%
(+) indicates an increase; (-) indicates a decrease.
Analytical Error Geochemical analysis, including measurement methods and water sample analysis, is subject to the most stringent quality assurance/quality control procedures (QA/QC) procedures, and consequently analytical uncertainty tend to be a relatively minor component in relation to the total uncertainty of groundwater monitoring program. Failure to analyse blanks, standards, and samples by exactly the same procedures may result in either a biased blank correction or a biased calibration (Kirchmer, 1983). Indirect Measurement Indirect measurement of the desired property occurs when the desired property is not amenable to direct measurement, thereby requiring the use of another measurable property of known correlation. A good example of this is the batch equilibrium test which is used to evaluate the
GEOCHEMICAL DATA COLLECTION
181
distribution coefficient.
Data Handling Faulty recording of observations in field or laboratory note books or indirect coding for computer analysis are examples of data handling uncertainty. It is conceptually important to distinguish between the sources of uncertainty in order to devise the appropriate analytical model and to be able to analyse and combine data sets from different sources.
7.8.2
Sampling Methods and Types
To develop the basis for the analysis, it is noted that in the sampling and geotechnical testing of subsoil, distinction must be made between block sampling and bore hole sampling. In block sampling, where one or more blocks of reasonable size are recovered from the site (from a known physical location) by an appropriate technique, the blocks are trimmed into several smaller samples, which are subsequently tested in the laboratory to furnish a cluster of values of the measured property, corresponding to the physical location of the block in the subsoil. Therefore, block sampling is useful for evaluating the mean and the inherent scatter of the soluble and exchangeable chemical constituents at a particular point in the subsoil. On the other hand, borehole sampling furnishes single values of the measured property at various points along a borehole. In the general practical situation, several boreholes are irregularly located within the zone of interest and tests along the borehole are also irregularly spaced, resulting in a collection of values of the measured property, with each value corresponding to a particular point of known coordinates. Thus borehole sampling is mainly useful in establishing the spatial trend of the measured property. In actual situations, borehole sampling could be used in conjunction with block sampling to assess subsoil geochemical conditions. The types of water samples that may be taken during investigation are: (1) FieM blank: It is a sample of distilled or de-ionized water taken from the laboratory out into the field, poured into a sampling vial at the site, closed, and returned as if it were a sample. The level of contamination of the field blank is the zero analyte signal for determining the limit of detection. (2) A rinse or cleaning blank: It is a sample of the final rinse before it is put in a new cell. This type of sample is used to evaluate whether a sample may have been contaminated from material taken in the previous sample. (3) Field samples: They are those samples that are taken in the field as representative of conditions at the site and analysed in the laboratory for constituents of interest. If sampling points or locations are unrepresentative, or biased sampling procedures are used, no amount of care in QA/QC in subsequent stages will salvage an accurate picture of actual field conditions. (4) Duplicate samples: They are collected and not analysed unless it is later determined that they contain additional useful information. Soil samples are commonly duplicated. (5) Split samples: They are field samples that are split between two storage vessels or cut in half in the field. One sub-sample may be analysed by one laboratory and the other sub-sample may be archived or given to another laboratory. (6) Spiked samples: They are field samples that may be split with one sample receiving a spiked volume of a reference standard to estimate the recovery of the analyte in the laboratory.
182
(7)
(8)
7.9
SITE INVESTIGATION Spiked samples allow estimates of accuracy and detect possible matrix interference problems. Laboratory blanks: They are similar to field blanks except that the distilled de-ionized water is used in the laboratory at the time each batch of samples is tested. This type of sampling may detect contamination that occurs in the laboratory. Standard reference samples: They have analyzed previously in outside laboratories. These samples are available from reference laboratories to detect either instrument calibration error or the use of inappropriate laboratory analytical methods.
GEOCHEMICAL DATA ANALYSIS
It is reported that chemical constituent concentrations of soil and groundwater sampling are neither normally distributed nor independent (Mann, 1987; Kite, 1989). This in turn creates special challenges for statistical analysis of geochemical sampling data because many of the traditional statistical techniques for analysing sampling data, such as linear regression and t-testing, assume that the population sampled has the symmetric, bell-shaped Gaussian (normal) distribution. Linear regression is probably the most frequently misused statistical technique in this context. The first step in analysing geochemical data is to determine whether they are normally distributed. If they are, traditional techniques in standard text books on statistics can be used. If not, one of the following methods, which are discussed in Chapters 8 and 9, may be used (Boulding and Barcelona, 1991): (1) Data transformation such as logarithmic conversion to create data sets that are normally distributed and, hence amenable for analysis by conventional methods. With this transformation, bias will be introduced and can be evaluated (Wilson et al., 1990); (2) Geostatisticaltechniques that facilitate differentiation of correlated and non-correlated data sets and interpolation of values between sample points. The fuzzy linear regression technique may be useful in hydrologic situations where the relationship between variables is imprecise, data are inaccurate, and/or sample sizes are insufficient (Bardossy et al., 1990). Subsurface pollution investigations typically involve measurements of concentration changes in geochemical parameters over time. Consequently, statistical techniques designed specifically for analysis of trends in time series are important (Harris et al., 1987; Montgomery et al., 1987); and (3) Pollutant transport modelling.
7.10
SUMMARY AND CONCLUDING REMARKS
Field investigation at polluted sites presents unique challenges to the investigator. Often conducted in non-homogeneous anisotropic environments which may be polluted by a complex diversity of pollutants. Not only the findings of the investigation must be accurate, but also they must be conducted in the most cost and time effective manner possible. The highly variable nature of subsurface conditions makes it impossible to define specific investigation strategies that will be appropriate in all cases. The selection of appropriate geophysical, drilling and sampling techniques should be made by experienced geophysicists and hydrogeologists.
SUMMARY AND CONCLUDING REMARKS
183
Although several techniques may be applicable, careful consideration of site- specific factors will usually dictate a certain methodology for cost effective field activities. When designing monitoring networks and choosing monitoring well types and configurations, the field hydrogeologist should confer with the contaminant hydrogeologist, the geophysicist, and to some degree, the mathematical modeller to be sure adequate, efficient monitoring well network is constructed for all end-users. No proper site remediation plan can be prepared or implemented unless the findings of the site investigation are used in concert with close field supervision during the remediation process. The concept of phased site investigation will provide the basis for this approach. Without these planned provisions, most of the remedial actions will likely experience cost overrtms in excess of those which may have been actually budgeted for.
This Page Intentionally Left Blank
CHAPTER
EIGHT
GEOSTATISTICS
8.1
INTRODUCTION
Statistics is generally concerned with the analysis and interpretation of uncertainty resulting from limited sampling of the variable of interest. It typically employs a population distribution model that assumes that samples are normally distributed and uncorrelated. Experimental data in environmental, geotechnical, or earth sciences tend to be specially correlated. Clearly, it is to be expected that pollutant concentrations in a subsurface, for example, will be more far apart. This observation is true of most earth science data since closely spaced locations, hence samples would have subjected to similar physical, chemical and biological processes. Such correlated data cannot, therefore, be effectively manipulated with the classical statistical technique. Geostatistics is a statistical method that is particularly useful in situations where a sample value is affected by its location and its relationship with its neighbours, i.e., spatially correlated data. Historically, geostatistical application originated in the mining and petroleum exploration industries, through the pioneering work of D.C. Krige in South Africa in the 1950s, and its mathematical formalization by G. Matheron in the early 1960s (Matheron, 1971). It is now routinely employed in the environmental field, for example, to estimate chemical and/or physical parameters at locations where data is unavailable. Geostatistics are potentially powerful tool for the analysis of pollutant concentrations in a subsurface. Many restrictions have been introduced, such as the need of a large number of samples, but the method remains very efficient. The geostatistical process is a two step procedure. The first is the calculation of the experimental variogram and fitting a model to it. The variogram merely describes the special relationship between the data points. This is perhaps the most important step in kriging process. Kriging is a moving average technique which uses the variogram parameters to obtain the relationship between the data points. Estimation techniques based on the theory of regionalised variables can be used to obtain the best (in the sense that the estimation error is minimized), linear (i.e., estimated concentrations are given by linear combinations of measured concentrations), unbiased estimates of pollutant concentrations at unmeasured points in the plume based on a limited number of groundwater samples. Then, these estimates may be used to map pollutant concentrations within the plume. Furthermore, because the estimation errors at each point are also computed, it is potentially possible to evaluate the "information content" of a groundwater sampling scheme, i.e., to determine the value of individual measurements in defining the areal extent of groundwater pollution. The estimation errors may also be useful in developing procedures for optimizing the design and cost of future groundwater sampling and monitoring programs.
185
186
GEOSTATISTICS
8.2
DATA ANALYSIS C O N C E P T S
8.2.1
Histogram
The histogram is a common way to represent the distribution of experimental data. Consider a set of n measurements that are sorted in increasing order, z~ < z2 < " < z,. The interval between the largest and the smallest value is divided into m intervals by the points a o, a~, ..., am.~, am. The intervals are usually of equal length, and are selected in such away that the histogram is relatively free from abrupt ups and downs. A measurement z belongs to the k-th interval if ak.t -< z < ak. Let nk be the number of measurements that belong to the k-th interval. The ratio n J n represents the frequency of occurrences in the k-th interval. Plotting the number nk or the frequency of occurrences nk/n as a bar diagram results in a histogram, such as depicted in Figure 8.1.
Figure 8.1. Histogram of pollutant concentration data.
8.2.2
Ogive
The ogive is the cumulative frequency distribution of the data set. For the sorted data, z 1 < z: < ... < z,, we compute p~ -
i - 0.5
,
for
i = 1, . . . , n
[8.1]
n
where p, is a number that increases from 0 to 1, and represents the ratio of number of data values smaller than or equal to z~. A plot of p, against z, yields an ogive, such as shown in Figure 8.2.
DATA ANALYSIS CONCEPTS
187
1.0 0.8 er 0.6
.~
0.4
E (~
0.2
I
I
I
I
I
I
I
I
0
100
200
300
400
500
600
700
800
Concentration (ppm)
Figure 8.2. Ogive of pollutant concentration data.
8.2.3
Summary Statistics Arithmetic Mean The arithmetic mean represents the average value of the data set, and is given by:
m =
Z l + Z2+
"'" + Zn
n
-
l
2.., zt n i=2
[8.2] Ik.
_1
Median The median, Zm, is defined as the data value that is larger than half of the data set and smaller than the other half. It is defined as follows:
Zm
-
zt, (zt+zt+l)/2,
l = (n + 1)/2, l =n/2,
/f n = odd if n = even
Mean Square Difference The mean square difference is a measure of the spread of the data set. It is given by:
[8.3]
188
GEOSTATISTICS
112 = (Zl- m)2+ "'" + (an- m)2
1 s (z,- m) 2 n i=l
[8.4]
where 112is known as the variance, and its square root, 11, is the standard deviation.
Skewness Coefficient The skewness coefficient, ks, is a measure of the degree of symmetry about the central value. It is expressed as:
1 s (Z~- m)3 / :
n
~:l
[8.S]
1i3 It is a dimensionless number. A symmetric distribution has a k,. value of zero. If the data contains many values slightly smaller than the mean and a few values much larger than the mean, the coefficient of skewness is positive. If there are many values slightly larger and a few values much smaller than the mean, the coefficient of skewness is negative.
8.2.4
Normal Distribution
)
The normal probability density function (PDF) is given by:
f (z) -
1 exp V/2rt:o2
(Z - m)2 2o 2
[8.6]
For data set which is normally distributed (bell-shaped), the mean and standard deviation provide enough information to reconstruct the histogram with acceptable accuracy. Figure 8.3 shows a histogram with mean m and variance o 2 and the normal probability density function distribution with the same mean and variance.
8.3
SEMIVARIOGRAM
The semivariogram is a graph that is most commonly used in applied geostatistics to explore spatial interdependence. It contains information about the scale of fluctuation of a variable. Consider the case of n measurements, Z(Xl), z(x2) .... , z(x,), where the argument x denotes the array of coordinates of the point where these measurements were taken. A plot of the square difference 1/2 [z(x,) - z ( x ' ,)] 2 against the separation distance I x, - x ~1 for n(n- 1)/2 measurement pairs, yields a scatter plot, as shown in Figure 8.4. A semivariogram is the smooth line through this scatter plot.
DATA ANALYSIS CONCEPTS
189
Figure 8.3. Histogram and theoretical distribution of normal data.
m,,
v
r>-
2
--
1
-
0 0
o"
',,
i",
I
I
1
2
.,'
~/
",
I 3
4
Figure 8.4. Semivariogram.
In the standard method of plotting the semivariogram, the axis of separation distance is divided into consecutive intervals. The k-th interval is [hlk, hUk] and contains Ark pairs of measurements [z(x,), z(x ~)]. The semivariogram function, Y(hk) is given by:
v(h k) = ~
1
Nk
~,=, [z(x~) - z(x'~)] ~
[8.7]
190
GEOSTATISTICS
The index i refers to each pair of measurements z(x3 and z(x ~) for which i
[8.8]
h tk -< IIxi - x ill < h uk
This interval is represented by a single point hk equal to the average value: Nk
k
=
1 ~ N k i:l
[[xi _ x,i[
[8.9]
A connected plot of Y(hk) against hk results in a semivariogram, as shown in Figure 8.4.
Nonstationary -..
Range ~>-
Sill
2 Stationary
0
1
2
3
h
Figure 8.5. Semivariogram illustrating stationary and nonstationary behaviour.
The behaviour of the semivariogram at distances comparable to the size of the domain determines whether the function is stationary or not. A function is considered stationary if it consists of small-scale fluctuations (compared to the size of the domain) about some well-defined mean value. For such a function, the semivariogram should stabilize around a value, called the sill, as shown in Figure 8.5. The sill is approximately equal to the variance of the data. For a stationary function, the length scale at which the sill is obtained describes the scale at which two measurements of the variable become practically uncorrelated. This length scale is known as range, correlation length or radius o f neighbourhood.
INTRINSIC MODELLING 8.4
191
INTRINSIC MODELLING
Consider that pollutant concentration, or any other spatially variable quantity, is a function of the spatial coordinates and may be represented as z(x~), z(x~, x2), or Z(Xl, x2, x3), depending on whether the quantity varies in one, two, or three dimensions. The notation z(x) will be used to include all three cases, where x is the location index (a vector with one, two, or three components). Thus,
Xl/
X1 X
=
" X1 ,
X
=
X2
;
X
=
X2 X3
[8.10]
The function z(x), known as a regionalised orfield variable, is not known everywhere, and needs to be estimated, at unknown locations, from available observations. The actual unknown z(x) is one out of a collection (or ensemble) of possibilities of z(x; 1), z(x; 2) .... This ensemble defines all possible solutions to the problem at hand. The members of the ensemble are known as realization or sample functions. The ensemble of realizations defines what is known as a random function. Since we are interested in calculating averages over all possible realizations, we use statistical moments. In linear estimation, we use the first two statistical moments of the random field, namely: (1) The mean function (m) (first moment), which gives the expected value (E) at any point x:
m(x) - E[z(x)] (2)
[8.11]
The covariance function (second moment), which is the covariance for any pair x and x':
R(h) = E[(z(x)- m(x)) (z(x')- m(x'))]
[8.12]
where h = II x - x' II = x/(x,- X ' l ) 2 -at- ( x 2 - x 2) 2 + ( x 3 - x ; ) 2 . Eqs. [8.11] and [8.12] constitute a stationary model. This model is also isotropic because it uses only the length and not the orientation of the linear segment that connects the two points. The covariance at h = 0 is known as the variance or sill of the stationary function. The sill should be finite. R(h) vanishes or tends to vanish when h exceeds a certain value, known as the range. In order to apply this model in interpolation, we first need to find the mean m, and select an expression for the covariance function and determine the optimum values of its parameters. In most cases, the mean is not known beforehand, and needs to be inferred from the data. To avoid this, it is often convenient to work with a semivariogram: 1 y(h) = -~ E[(z(x)- z(x')) 2]
[8.13]
192
GEOSTATISTICS For a stationary function, the relation between the semivariogram and the covariance function
is: 1 y(h) : ~1 E[(z(x)- z(x')) 2] : -~ E[((z(x)- m(x))- (z(x')- m(x'))) 21 1
: - E[(z(x)- m(x)) (z(x')- m(x'))] + -~ E[(z(x)- m(x)) 2]
[8.14]
+ 1 E[(z(x')- m(x')) 2] : - R(h) + R(O) 2
8.5
STRUCTURAL ANALYSIS
Structural analysis is the selection and fitting of mathematical expressions for the required first two moments of the regionalised variable. The form of these expressions comprises the model. A list of such mathematical models is given below. 8.5.1
Stationary Models
Gaussian Model
The Gaussian model is defined by the covariance and semivariogram functions: R(h) = o2 exp(- Lh--~~ } [8.15]
where o is the variance, and L is the length parameter, with values greater than zero. The covariance function decays asymptotically, as shown in Figure 8.6. Exponential Model
For the exponential model, the covariance and semivariogram functions are given by: R(h) = 02exp( - h ) [8.16] Y(h) = 0 2 ( 1 - e x p ( -
hi)
where o is the variance, and L is the length parameter. It should be noted that both O 2 and L are greater than zero. This model is quite popular, particularly for hydrogeological application. A schematic of the semivariogram and covariance function is shown in Figure 8.7.
STRUCTURAL ANALYSIS
193
1.0 0.8
v(h)
0.6 0.4
----~11~ .........................
(h) , I
0.2
I
0 0.1 ~_ range (a) ~
0.2 h
0.3
0.4
0.3
0.4
Figure 8.6. Gaussian semivariogram and covariance functions.
1.0 0.8 "
0.6
y(h)
I
0.4
R(h) 9
0.2
0
0.1 range (a )
I
0.2 h
}~
Figure 8.7. Exponential semivariogram and covariance functions.
194
GEOSTATISTICS
Spherical Model For the spherical model, the covariance and semivariogram functions are given by: \
R(h) =
1 --
3h 2a
-t-
1 h3]o 2 2a )
O<_h<_a
3
h>a
0,
[8.17] l( 3 h V(h) =
2 a
1 h31 02 2a 3
O<_h<_a h>a
122 ,
where the parameters are the variance 02 > 0 and the range a > 0. A schematic of the semivariogram and covariance function is shown in Figure 8.8. 1.0
i
i
y(h)
0.8 0.6 0.4
R(h)
0.2
0
0.2
0.1
Range (a)
-1~
0.3
0.4
h
Figure 8.8. Spherical semivariogram and covariance functions.
Nugget Effect Model The covariance and semivariogram functions for the nugget effect model are given by: 0, R(h) = Co6(h) = Co ,
h>O h =0
y(h) = C o(1 - 6 ( h ) ) = {Co0~
h>0 h=0
[8.18]
STRUCTURAL ANALYSIS
195
where the parameters are the nugget variance Co > 0 and the Kronecker delta, 6(h), which equals one if h = 0 and 0 for all other cases. A schematic of the semivariogram and covariance functions is shown in Figure 8.9. The semivariogram jumps from 0 (at h = 0) to o 2 (at h > 0). The covariance function drops off from o 2 (at h -- 0) to 0 ( at h > 0).
1.0
, I
0.8
.
0.6
.
.
.
.
.
.
.
.
.
.
.
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.
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.
.
.
.
.
Z
q- .
.
.
.
.
.
T-
.
.
.
.
.
.
I 1
I- .
.
.
.
.
.
_L
.
.
.
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T
.
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I .
.
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.
.
, 0.4
I
i - -
i . . . . . . . . . . . . . . . . . . . .
1 . . . . . . . . . . . . . . . . . . . . .
i
.....................
0.2
~. . . . . . .
[ IP 0
0.1
R(h)
l ,]
i
0.2
0.3
} ......
',"
_
0.4
Figure 8.9. Nugget effect semivariogram and covariance functions.
8.5.2 Nonstationary Models Power Model The semivariogram of the power model is given by: y(h) - ~h ~
[8.19]
where ~ and 13 are the model parameters, with tt > 0 and 0 < 13 < 2. A schematic of the semivariogram is shown in Figure 8.10.
Linear Model The semivariogram of the linear model is given by: y(h) - t~h
[8.20]
where the only parameter is the slope tx > 0 of the semivariogram. A schematic of the semivariograrn is shown in Figure 8.11.
196
GEOSTATISTICS 1.0 0.8 0.6 0.4
I
0.2
. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
J.
I
. . . . . . . . . . . . . . . . . . .
~t
. . . . . . . . . . . . . . .
I
9
I
0
0.25
0.5
0.75
1.0
0.5
0.75
1.0
Figure 8.10. Power function semivariogram.
1.0 0.8
..C:
0.6 0.4 0.2
0
0.25
Figure 8.11. Linear function semivariogram.
8.5.3
Model Superposition
Linear with nugget effect The semivariogram for the linear with nugget effect model is given by:
STRUCTURAL ANALYSIS
?(h)
= ICo+ ah, t 'Ymean'
197
h <_ a
[8.21]
h > a
where ? is the sample variance, a is the linear coefficient, a is the range, Co is the variability due to nugget effect. A schematic of the semivariogram is shown in Figure 8.12. ....
1.0 h
0.8
L .........................
I
0.6
I
I
0.4
,
I
0.2
I
'
I
Co (VariabOity due to nugget effect) 0
0.25
0.5
0.75
1.0
Figure 8.12. Linear with nugget effect semivariogram.
1.0
I
P
o.e Y I ~
I
0.4 ~ir 0.2
I
"
r .
/
. . . . . . .
.
.
.
.
1 t
(~,artance of _C~ t~e transition model)
.
, . . . . . . .
~.......
!
-
Co (Variab ity due to nu~rget effect) 0
0.25
I I
'
, I
0.5
0.75
Figure 8.13. Spherical with nugget effect semivariogram.
1.0
198
GEOSTATISTICS
Spherical with nugget effect The semivariogram for the spherical with nugget effect model is given by:
l ( -23h_ 21ah 33) '
h<_a
[Co+ C 1,
h>a
y(h) = C~
C1
a
[8.22]
where C~ is variance of the transition model. A schematic of the semivariogram is shown in Figure 8.13.
8.6
KRIGING
Kriging is merely a weighted moving average technique. It can be used to assign an estimated value to a particular location (point kriging) or to a block (block kriging). Like any other estimate, a Kriged estimate is a weighted combination of the sample values around the point to be estimated. Other linear unbiased estimators exist, such as polygons, triangles, and inverse distance methods. However, Kriging is the best linear unbiased estimator. Kriging is the process of estimating the value of specially distributed variable from adjacent values while considering the interdependence expressed in the semivariogram. The Kriging process involves the construction of a weighted moving average equation which is used to estimate the true value of a regionalised variable at specific locations. This equation is designed to minimize the effect of the relatively high variance of the sample values by including knowledge of the variance between the estimated point and other sample points within the range. The main aspects to consider in understanding the theory of Kriging are the estimation error and the weighting coefficients. Given n measurements of z at locations with spatial coordinates xl, x2,..., x n, the value of z at an arbitrary point Xo is estimated by assuming that it is a linear combination of the measurements, i.e.,
Zo* - ~
~'i z(x i)
[8.23]
i=l
Thus, the problem is reduced to selecting a set of coefficients ~1, ..., ~.,. These coefficients are the weights which are assigned to the n sample values. The difference between the estimate zo~and actual value Z(Xo) is the estimation error:
Zo* - z(x o) = ~
)~i z(xi) - z(x o)
i=l
The estimator should meet the following specifications:
[8.24]
KRIGING (1)
199
U n b i a s e d n e s s : On the average, i.e., over all possible solutions or realizations, the estimation error must be zero. That is:
ZoZ Xo , mm( l)mO1 l
[8.25]
The numerical value of the mean, m, is not specified. Therefore, for the estimator to be unbiased for any value of the mean, it is required that:
~
.~ = 1
[8.26]
i=l
Imposing the unbiasedness constraint eliminates the unknown parameter m. (2)
M i n i m u m Variance: The mean square estimation error must be minimum. The mean square error in terms of the semivariogram is given by (Kitanidis, 1997)"
.E~Zo*-Z~.o~l:
-
:s s ~,~,~,-,--, ~+~ :s ~,,~-,--o /=1 j=l
[8.27]
i:1
Thus, the problem of best (minimum mean square error) unbiased estimation of the )~ coefficients may be reduced to a constrained optimization problem. Values of)~l, ..., )~nthat minimize the objective function given by Eq. [8.27] while satisfying the constraint given by Eq. [8.26] are sought. This problem can be solved easily by using Lagrange multipliers, a standard optimization method. The necessary conditions for the minimization are given by the linear kriging system of n+ 1 equations with n+ 1 unknowns:
- ~2 ~ (
x ~ - xjll~ + ~ - - v ( i x , -
Xoll~,
i = 1, 2 ..... , n
[8.28]
j=l
Xj = 1 j=l
where g is the Lagrange multiplier.
[8.29]
200
GEOSTATISTICS It is common practice to write the kriging system in matrix notation, in the form:
[8.30]
Ax=b where x is the vector of the unknowns
X
[8.31]
=
kt
b is the vector -Y(llx I - Xoll) -Y(llx 2 - Xoll) b
[8.32]
__.
-Y(llx. 1
xoll)
and A is the matrix of coefficients 0 -y(Ix= - xlll)
-y(llx~ - x= I) 0
-Y(llx I - x , ll) -Y(llx 2 - x , ll) 9
a
-y(Ix.
- x~l)
1
-y(Ix.
- x 2 I)
1
0 1
~
[8.33]
1 0
Therefore, the problem is reduced to the solution of a linear system 9Solving this system, we obtain ~,~, ..., Xn, and ~t. In this manner, the linear estimator of Eq. [8.23] is fully specified. In addition, we can quantify the accuracy of the estimate through the mean square estimation error (MSE), which can be obtained by substituting into Eq. [ 8.27] the values of ~.~, ..., Xn obtained from the solution of the kriging system.
SOLUTION METHODOLOGY 8.7
(1)
(2)
(3) (4)
201
SOLUTION M E T H O D O L O G Y A geostatistical analysis of pollutant concentrations, for example, consists of four steps: Determine if the measured pollutant concentrations are additive and normally distributed. If they are not normally distributed, appropriate transformations are required (e.g., a natural log transformation often improves the fit of the data to a normal distribution); Estimate the spatial correlation of pairs of measured concentrations as a function of the distance and the direction of their separation, i.e., determine the experimental semivariogram; Perform a structural analysis, i.e., fit a theoretical model to the experimental semivariogram; and Perform kriging. The fitted model is used in an interpolation procedure, known as kriging, to estimate expected values of pollutant concentrations at points within the plume where measurements were not taken.
Step 1: Additivity and normal distribution The use of linear geostatistics requires that the representative volume (ReV) be additive (Journel and Huijbregts, 1978). Ion concentrations expressed in units of mass (or weight or volume) of solute per unit volume of groundwater do not represent true accumulations and are not additive. Thus, the pollutant concentrations obtained from groundwater should be multiplied by the porosity of the soil at the point where the sample was taken. The need to convert pollutant concentrations to additive variables means that the porosity of the soil also must be measured (or estimated) for each sample location. A second requirement of geostatistical analysis is that the ReV be normally distributed. Although geostatistical techniques have been proposed for cases where the probability distribution of the ReV deviates from normality, they are not usually necessary since deviations from normality can often be corrected by an appropriate transformation or by the elimination of erroneous measurements in the sample data. Often, a log-transformation will improve the fit of the ReV to a normal distribution. This transformation is given by:
y(x) = log [z(x) + A]
[8.34]
where y(x) is the natural log-transformed pollutant concentrations, z(x) is the pollutant concentrations, and A is a constant added to z(x) to improve the fit. The mean and variance of z(x) are given by (Rendu, 1978):
m =exp(, + o 2 = m 2 [exp (132) - 1]
[8.35]
[8.36]
202
GEOSTATISTICS
where rn is the mean of z(x), o 2 the variance of z(x), a is the mean ofy(x), and
B2
is the variance of
y(x). Step 2: Calculation of experimental semivariogram Given a normally distributed set of data z(x) (it may be the natural concentration or the logtransformed concentration), an estimation of the two first moments is required for a linear geostatistical analysis. The first order moment is simply the expectation or mean. There are three second-order moments in geostatistical analysis, namely, variance, covariance, and variogram. These are, usually, functions of the position, (x). However, under the assumption of stationarity, the covariance and semivariogram do not depend on the positions of the pairs of measurement points employed, but only on the vector h separating them (i.e., h = xi - xm). Most of the time, it is not possible to know, a priori, the variance and covariance of z(x) as required for second-order stationarity. In this case, a less strict hypothesis of stationarity can be used. Since only one realization of pollutant concentrations is available in the sample data z(x), only an estimate of the semivariogram, given by Eq. [8.7], is possible. There are two sources of errors associated with the estimate of the semivariogram. Since a realization of the pollutant concentration is not known for every point in the plume, the first source of error is due to the limited number of measurements that are available. This error is called the variance of estimation, and it depends on the number of samples N(h) used to estimate ),(h), for each vector h. The error decreases as N(h) increases. The second source of error is from the fluctuations of the local mean (defined from point to point in the plume) about the assumed mean for the entire plume. These fluctuations cause estimates of the semivariogram to also vary from point to point in the plume. A theoretical analysis of these two sources of estimation errors can be used to determine the minimum number of sample pairs and the maximum magnitude of the vector h for which the semivariogram can be approximated. The analysis requires that the fourth-order moments of z(x) be known. For few samples, these moments cannot be computed accurately. However, practical rules exist for estimating the semivariogram from a set of sample values (Journel and Huijbregts, 1978): Rule #1:
N(h) > 30 - 50
[8.38]
Rule #2:
Ithll< L/2
[8.39]
where IIhllis the magnitude of the separation vector h, and L is the longest dimension (in direction of h) of the pollutant plume. The restriction on Ithll given by Eq. [8.39] has important implications for the estimation procedure (kriging) to be described in step 4. Since the experimental semivariogram cannot be computed for I[hllgreater than L/2, any calculation made with the fitted semivariogram models are restricted to I[hllless than L/2. One important consequence of this restriction is that kriging should not be used to estimate pollutant concentrations for points that are at a distance greater than L/2 away from the nearest sample. The restriction on the number of sample pairs given by Eq. [8.38] also affects the estimation of the true semivariogram. In the general case, an adequate number of sample pairs can be obtained by grouping sample pairs into distance classes for each specified vector h o. Sample pairs can be
APPLICATION
203
grouped by defining tolerance intervals for Itholl,i.e., A Itholl.If a sample pair falls within the tolerance intervals for the specified vector h o, that sample pair is retained and used in the calculation of T(ho). Sample pairs that do not meet these criteria are discarded.
Step 3: Structural analysis Structural analysis is the fitting of a mathematical model to the experimental semivariogram. Any mathematical function may be used to model an experimental semivariogram as long as it is positive-definite, i.e., increasing or constant with increasing h and non-negative. In practice, only a few types of functions are typically used. Models are of two types depending on whether or not the experimental semivariogram exhibits a sill. The most commonly used models are discussed in section 8.5. The procedure for fitting a mathematical model to the experimental semivariogram consists of several steps. Preliminary estimates of the model parameters (e.g., a and Co) can be obtained by visual inspection or by the method of weighted least squares (weighted because the number of sample pairs used to compute each point of the semivariogram is, usually, different). At this stage, the fitting procedure can be guided by knowledge of the physical properties of the pollutant, plume, subsurface, and local groundwater flow pattern.
Step 4: Estimation of the unknown contaminant concentrations (Kriging) When we know only the two first moments of the distribution, the only way to estimate the unknown contaminant concentration is to assume that it is a linear combination of the known values, i.e.,
z *(x) : ~ I., z(xi)
[8.40]
i=1
where ~'i are the weights assigned to the n sample values and z*(x) is estimated value of the contaminant concentration at location x. The best linear unbiased estimate may be obtained by using the Lagrangian method, as discussed previously.
8.8
APPLICATION
The data used in this example, reproduced in Table 8.1, came from Chem-Dyne waste site, located in Hamilton, Ohio, near the confluence of Great Miami River and the Fort Hydraulic Canal (Cooper and Istok, 1988). The Chem-Dyne Corporation began operation in 1975 as a chemical waste transfer, disposal, and storage facility. Wastes that have been present at the site include pesticides, PCBs, polychloronated biphenyls, lab packs, acids, resins, solvents, heavy metals, and cyanides (CH2M Hill, 1984). As of 1979, the waste disposal and storage activities at the site were terminated, and a cleanup program was implemented under the management of the US Army Corps of Engineers (US EPA, 1984). The site has the following geologic characteristics: the aquifer underlying the site consists of unconsolidated alluvial deposits extending to depth greater than 45 m. The deposits consist
204
GEOSTATISTICS
primarily of sandy gravel or gravelly sand. Overlying these deposits, but sometimes extending below the water table, are finer deposits of silts, silty sands, and sandy silty clays. On top of these finer materials, a cap of fill and rubble has been placed. This fill is extremely variable in composition but does not appear to extend below the water table (CH2M Hill, 1984).
Table 8.1" Well locations (x, y, z) and barium concentrations Well x y z Conc. Well x Num. ( m ) ( m ) (m) (mg/m 2) Num. (m)
y (m)
z (m)
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18
547 541 752 759 350 310 451 470 538 538 412 688 562 534 496 421 245 207
9.89 2.69 3.08 11.95 11.95 1.37 1.05 11.20 2.35 10.88 2.16 19.14 2.65 19.14 10.03 1.95 9.88 2.4
795 802 769 678 623 573 573 536 517 519 451 472 477 551 496 370 115 335
320 331 543 348 412 288 260 194 292 374 310 442 455 656 611 83 237 489
10.31 2.49 2.23 0.59 1.50 0.36 0.99 1.45 0.81 1.88 1.07 8.35 0.51 0.00 2.35 2.19 1.63 1.63
1400 1280 901 4330 3460 7360 0 1700 2680 4630 707 13500 1950 0 1410 0 1100 0
19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36
151 162 309 327 566 519 534 534 536 547 586 376 453 175 346 406 397 395
Conc. (mg/m 2) 832 790 0 4940 2680 1680 3080 9220 4520 755 887 0 7560 1530 0 4440 700 1690
With the geostatistical procedures described in the previous sections, the mean and the variance of the concentration in the soil were calculated. Then, histograms and the cumulative distributions were obtained, as shown in Figures 8.14 to 8.16 for barium. From the cumulative distribution plot, it is possible to determine whether the concentrations are normally or In-normally distributed. To get the In-transformed data, we apply the transformation C' = In (C). Most of the time, these operations give a smooth curve that better approximates a normal distribution. The test of normality (or In-normality) is usually done by visual inspection. Quantitatively the probability that the ReV is normally distributed may be determined from any of several statistics, e.g., the Chisquared statistic, the Kolmogorov and Smirnov statistic, or the Shapiro-Wilk Statistic. From Figures 8.14 to 8.16, it is possible to see that this distribution is most likely a Innormal distribution.
APPLICATION
205
Figure 8.15. Cumulative distribution of barium concentration.
To analyse barium distribution in the polluted soil, the semivariogram shown in Figure 8.17 is constructed. A theoretical semivariogram that approximates the experimental semivariogram, shown in Figure 8.17, is then thought. Various theoretical semivariogram models, some of which are detailed in section 8.5, are evaluated, and that with the minimum mean squared error (MSE), given by Eq. [8.27], is selected.
206
GEOSTATISTICS
Figure 8.17. Experimental semivariogram and the best fit model of semivariogram (linear with nugget effect).
In the present example, a linear model with a nugget effect provided the best fit, of the models attempted, to the experimental data. The main coefficients are: (1) range = 300 (m), (2) slope of the linear part - 0.002, and (3) nugget effect = 0.95 (In (semivariogram)). The calculated MSE
SUMMARY AND CONCLUDING REMARKS
207
is 0.1. From Figure 8.17 we see, however, the linear with a nugget effect model is not in good accord with the experimental semivariogram. Table 8.2 shows results of calculations of barium concentrations and variance at different locations, specifically at z = 2 m, x = 400 m, and y varying from 200 m to 600 m at 25 m intervals.
Table 8.2: Calculated barium concentrations and variance Result x y z Kriging estimates No. m m m * 103 (m$/m 2) 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17
8.9
400 400 400 400 400 400 400 400 400 400 400 400 400 400 400 400 400
200 225 250 275 300 325 350 375 400 425 450 475 500 525 550 575 600
2 2 2 2 2 2 2 2 2 2 2 2 2 2 2 2 2
4 3.7 0.7 5.4 1.9 4.2 4.4 2.7 4.5 3.8 2.2 5.3 3.8 1.8 2.8 1.2 3.8
Kriging variance * 106 (mg 2/m 4) 2.8 2.6 2.8 2.6 2.8 2.6 2.8 3.2 2.8 2.8 2.8 2.8 3.0 2.8 2.8 2.8 3.0
SUMMARY AND CONCLUDING REMARKS
In summary, the geostatistical analysis gives a good estimation of the pollution trends for subsurface soil. Then, from the calculated results, we can estimate the area of high pollutant concentrations and determine the locations of the next sampling points and groundwater monitoring well locations.
This Page Intentionally Left Blank
CHAPTER
NINE
SUBSURFACE POLLUTANT TRANSPORT
9.1
INTRODUCTION
When a pollutant is released from a waste storage facility, it migrates downward through the unsaturated zone (i.e., vadose zone) to the water table and then laterally in the direction of the hydraulic gradient in the saturated zone. Throughout this process its fate is controlled by a myriad of physical, chemical, and biotic processes. These include the physical processes of advection, diffusion, dispersion, and capillary, and the biotic and abiotic processes of bioaccumulation, biodegradation, immobilization, retardation, and volatilization. Quantification of these various processes at a field site is very difficult. Assessment of subsurface fate and transport must address questions of source characterization (what is released, where, when, how much, etc.), vadose zone transport and processes, groundwater transport, and exposure and dose assessment.
9.2
MODELLING PROCESS
The hydrogeological modelling process starts with identifying and defining the problem and setting the modelling objectives. In polluted site assessment, the objectives often focus on determining the chances of a pollutant plume reaching a critical point, or on a comparative evaluation of various remediation strategies. The key steps are: (1) Formulation of a conceptual model: Because the real hydrological system is complex, we create an idealization of the system. The information consists of the extent, configuration, and properties of the hydrogeological units as well as the processes relevant to the main objectives. The coupling and interaction among these processes are particularly important, as are the assumptions that justify excluding certain processes and interactions. We must consider the spatial dimensions that are necessary to properly represent the selected processes in the system; (2) Model Design: On the basis of the conceptual model, we can design the physically-based mathematical model. This step involves: (a) identifying the governing physical principles and the corresponding equations that express these principles, (b) defining the boundary and initial conditions and material properties, and (c) selecting and implementing the solution method; (3) Calibration: In the cases where model parameters are unknown, calibration can be used to estimate these parameters by fitting a simulation to an observed response under controlled conditions; (4) Sensitivity: One of the most important uses of models is to determine the sensitivity of the 209
210
(5)
(6)
9.3
SUBSURFACE POLLUTANT TRANSPORT system with respect to the various parameters. A sensitivity analysis is usually done by defining a base case, and then varying the parameters (one at a time) within their respective ranges. These ranges are generally known to an experienced modeler. Sensitivity analysis provides an insight to the behaviour of the system under various conditions, and about the importance of the individual parameters and processes; Prediction: When the sensitivity of the model is sufficiently understood, the model can be used for predictive purposes. We have to bear in mind that the model is not the real system. It is only an idealization of the system and it is subject to many assumptions and simplifications; and Decision: Model prediction forms one component in the decision making process related to the problem under investigation. The decision process also involves regulatory, economic, environmental, and political components. The integration of all these aspects requires a broad understanding of the problem. A sound model can contribute to this understanding.
TRANSPORT PROCESSES IN SOILS
There are three basic physical mechanisms by which miscible and immiscible pollutants are transported in the subsurface environment: advection, diffusion, and dispersion. Emphasis in this section is placed on soluble species, and only single-phase flow is considered. 9.3.1
Advection
Advection is the process by which pollutants are transported along with the flowing fluid or solvent in response to a hydraulic gradient. Due to advection, non-reactive solutes are transported at an average rate equal to the seepage velocity of the fluid. For saturated flow, the seepage and Darcy velocity are related by
v -
v
[9.1]
n
where v, is seepage velocity, n is soil porosity, v is Darcy's water flux which is given by: Q - - k Oh = kwi~' A WOx
[9.2]
where Q is volumetric water flow rate, A is cross sectional area perpendicular to the flow direction, kw is hydraulic conductivity, h is total hydraulic head, x is the flow direction, and ih is hydraulic gradient. The advective mass flux of a particular chemical species is given by: dadvection
= VC = k i h C
=
nvsc
[9.3]
where Jauvect,o,,is the advective mass flow (mass flowing through a unit cross sectional area per unit time), and c is the concentration of the solute in the liquid phase of the porous material. Thus the
TRANSPORT PROCESSES IN SOILS
211
advective transport is proportional to the medium hydraulic conductivity, the hydraulic gradient, and the local concentration. 9.3.2
Diffusion
While advection is associated with the bulk macroscopic groundwater movement, diffusion is a molecular-based phenomenon. If we could see the individual molecules, we could note the continual movement of each molecule and of one molecule relative to the other. Hence, diffusion is caused by random thermal motion, as is the Brownian movement of colloidal particles observable under the microscope. In unstirred liquids, all molecules and ions of the solvent and solute have this random movement. As a result of this motion, which occurs randomly in all directions, irregularities in the concentration of a solution eventually disappear. Measurements of diffusion usually entail measuring the rate at which the irregularities disappear. Thus, if neighbouring volumes of solution have different concentrations, more solute molecules move into the volume of lower concentration than move out until the concentrations become equal. The rate of change of concentration depends on the initial difference of concentration between the two volumes, and on the mean distance between them. Expressed in other words, the rate of transfer by diffusion between two volume elements varies directly with the difference of concentration and inversely with the distance between them. This can be represented by Fick's first law of diffusion as: Jdiffusion
=
-
D
0__c_c 0x
[9.4]
where Jdiffusion is the rate of flow, or flux, and is the amount of solute diffusing per unit time across unit cross sectional area, D is diffusion coefficient, and x is the direction of transport. The two commonly used definitions of diffusion coefficient are the self-diffusion coefficient and the bulk diffusion coefficient. They can be understood by referring to the methods by which they are measured: (1) Self-diffusion coefficients (Do) are measured by adding a small amount of radioactive isotope to a system containing a uniform concentration of unlabeled ion or molecule throughout. The isotope equilibrates with the unlabeled ions or molecules, and the rate at which this occurs is used to measure the self-diffusion coefficient. Typical values are shown in Table 9.1 (Robinson and Stokes, 1965). (2) Bulk diffusion coefficients (D~) are measured where there is movement of a solute within a soil mass due to a concentration gradient. Solute is therefore transferred to the volume of solution at the lower concentration. The coefficient is also sometimes known as a salt diffusion coefficient. The bulk diffusion coefficient is usually much less than the selfdiffusion coefficient.
Effects of Soil Properties on the Magnitude of D~ The two soil properties that affect the diffusion coefficient in saturated soil are discussed below.
212
SUBSURFACE POLLUTANT TRANSPORT
Table 9.1: Self-diffusion coefficients for representative ions at infinite dilution in water D o * 10-1~m2/sec D o * 10 1~ m2/sec Anion 0~ 18~ 25~ Cation 0~ 18~ 25~ H+ Li + Na § K§ Mg 2+ Ca 2§ Fe z+ CO 2+ Ni 2+ Cu 2+ Zn 2+ Cd z+ Pb 2+ Cr 3+ Fe 3+ A13+
56.1 4.72 6.27 9.86 3.56 3.73 3.41 3.41 3.11 3.41 3.35 3.41 4.56
2.36
81.7 8.69 11.3 16.7 5.94 6.73 5.82 5.72 5.88 5.88 6.13 6.03 7.95 3.90 5.28 3.46
93.1 10.3 13.3 19.6 7.05 7.93 7.19 6.99 6.79 7.33 7.15 7.17 9.45 5.94 6.07 5.59
OH F CI Br I SO42 NO2
25.6
N O 3"
9.78 4.39
CO32
10.1 10.5 10.3 5.0
44.9 12.1 17.1 17.6 17.2 8.9 15.3 16.1 7.8
52.7 14.6 20.3 20.1 20.0 10.7 19.1 19.0 9.55
Soil Porosity Because we are considering diffusion in soil pore solution, the need to consider the porosity of the soil is self-evident. A cross a unit cross-sectional area of soil, the diffusion coefficient is proportional to the cross-sectional area of the liquid contained in the soil. We can write: D
= nD ~
[9.5]
Tortuosity of the Pathway of Diffusion Ions diffusing through the soil solution do not pass along straight tubes but follow irregular and winding paths. Diffusion is along these irregular paths because of the changes in soil pore size distribution (Porter et al., 1960; Olsen and kemper, 1968; Bear, 1972). The complexity of the pathways is too great to calculate or measure directly. So tortuosity is regarded as an empirical factor. We can now write: Ds " r,n D o
[9.6]
where 1: is the tortuosity factor, which needs to be measured indirectly. It varies with the moisture content of the soil because as soil becomes drier the diffusive pathway becomes more tortuous (Rowell et al., 1967). In free liquid the value of z = 1, in saturated soil 1: = 0.4 at field capacity of 0.2. Tortuosity is introduced into the diffusion equation to adjust x, the distance between two points in the soil, in order to give the true concentration gradient, d c / d x . Generally, tortuosity is expressed as:
TRANSPORT PROCESSES IN SOILS
213 [9.7]
: l/l e
where l is the macroscopic, straight line distance between two points defining the path, and l e is the actual, microscopic or effective distance of transport between the same two points. Since le is greater than l, it follows that z is less than 1.0. Typical reported values of 1: are in the range 0.01 to 0.67 (Perkins and Johnston, 1963; Freeze and Cherry, 1979; Daniel and Shackelford, 1988; Shackelford, 1989; Shackelford and Daniel, 1991). Fick's law for diffusion in soil can now be modified to include tortuosity and porosity: Oc
Jdiffusion = VoY'n -~x
[9.81
where c is the concentration of the diffusing ions in bulk soil solution.
9.3.3
Dispersion
The d i s p e r s i o n mechanism of pollutant transport is associated with bulk fluid movement in the porous medium. Fluid particles that are at one time close together tend to move apart or spread. The spreading nature of pollutant is attributed to variations in seepage velocity, which occur during pollutant transport in soils. These variations are related to the following mechanisms (Fried, 1975; Bear, 1979; Freeze and Cherry, 1979): (1) The flow velocity across any cross section within the soil will be greater in the middle than near the walls of the pore channels where there is greater frictional resistance; (2) From continuity principle within the soil, flow velocity across a smaller pore opening is greater than that across a larger pore opening; (3) Flow occurs along paths of varying tortuosity, giving different path lengths and therefore different rates of flow per unit length of soil; and (4) Due to the heterogenous nature of soils, which result in variations in hydraulic conductivities, flow velocity changes. This spreading mechanism is known as dispersion. In a very definite sense, dispersion occurs because of our inability to follow the details of groundwater movement from one pore scale to another. Statistically, advection refers to the average rate of movement while dispersion refers to the deviation from the mean. Also, statistically, dispersion is scale-dependent. The further a particle moves in the subsurface, the greater the range of heterogeneity of hydraulic conductivity it will experience. For example, consider a subsurface containing sand and clay layers, a particle may either start off in sand or in a clay. For short distances of movement, it will remain in the same type of material it started off in and the dispersion coefficient will be characteristic of that material. However, as it moves further from its initial point it may move from sand to clay to sand, etc., with each unit having its own characteristic velocity. Considering two particles, it is apparent that the expected deviation of their locations from the mean position will increase more through the actual heterogeneous system than it would through an idealized homogeneous system.
214
SUBSURFACE POLLUTANT TRANSPORT The dispersion mass flux is usually modelled as a Fickian-type process: OC Jdispersion = -Od ~
[9.91
where Jdispers,onis the dispersive flux, and Dd is the dispersion coefficient which is assumed to be a function of seepage velocity and longitudinal dispersivity and given by: [9.10]
D d = nOt,lVsf~
where ff is the longitudinal dispersivity of the porous medium in the transport direction, and 13is a constant ranging between 1.0 and 2.0 (Freeze and Cherry, 1979). In most applications, it is assumed to be 1.0. The dispersivity, trt, is scale-dependent. In laboratory experiments, a / is found to vary from 0.1 to 10 mm, as shown in Table 9.2. In the field, the dispersivities are sometimes measured through single and multiple well tracer tests. More often, however, what is usually done is that measured field concentrations are simulated with mathematical models and the coefficients adjusted to get an adequate match. The values found in this fashion are usually much larger than laboratory values. Recent literature has shown field values for at to vary from 1 to 100 m or larger. These values are larger than laboratory values by a factor of up to 105, suggesting that dispersion plays a different role in the field than in the laboratory. In practice one usually combines the coefficients of diffusion and dispersion into a single hydrodynamic dispersion coefficient.
Table 9.2: Longitudinal dispersivity values (Gillham and Cherry, 1982) Test type
a t (m)
Laboratory tests Natural gradient tracer tests Single well tests Radial and 2 well tests
9.4
0.0001- 0.01 0.01-2.0 0.03- 3.0 0.5- 15.0
TRANSPORT EQUATION
The fundamental equation of pollutant transport is the conservation o f mass equation. It states that for an arbitrary volume the net rate of mass increase within the volume is equal to the net mass flux into the volume plus any increase in mass due to reactions within the volume. The mass increase term represents the total mass per bulk volume, including both the sorbed mass and that in solution. For a saturated medium the mass density, m, which is the mass per bulk volume, may be represented as: m = nc + (1 - n)psq
[9.11]
TRANSPORT EQUATION
215
where m is the bulk concentration, 9s is the soil density, and q is the sorbed concentration (units of mass or activity sorbed per mass of soil). The product (1- n) p, is the bulk density of the soil, Pb. The net flux includes advective, diffusive, and dispersive mass transport. The mass flux vector, J, is the mass crossing a unit area per unit time. The reaction term includes radioactive decay, biodegradation of organic pollutants, precipitation and redox chemical reactions that may mobilize a pollutant, and others, and may be represented with the symbol S t . The source strength S ~ has units of mass per unit volume per unit time. For an arbitrary control volume, the conservation of mass equation takes the form
dmact - JfJf Jf m d V = -
f f J . n dA + f f f s * av
[9.12]
where the first term represents the time rate of increase in the total mass within the control volume, the second term is the net flux of mass into the volume across the control surface with n the outward unit normal vector to the control surface, and the last term is the mass increase due to sources located within the volume. Eq. [9.12] is an integral form of the continuity equation. Since the control volume is arbitrary, we may also write the continuity equation in the form
Om + ~7. J - - S + Ot
[9.13]
This is the general form of the continuity equation and serves as the starting point for most further investigations of subsurface fate and transport. For most applications we work with a form of the general continuity equation that is simplified in one fashion or another. These simplifications involve making further assumptions as to how to model the various processes that are of interest. Non-polar organic compounds in groundwater are found to be sorbed by soil organic matter present in the porous medium. This sorption is due primarily to hydrophobic interactions resulting in weak, non-specific sorption forces, as discussed in Chapter 5. When the organic compounds are present in trace concentrations, linear adsorption isotherms are often observed, as discussed in Chapter 5: q = Kac
[9.14]
where K d is the distribution or partition coefficient for the chemical species (1/kg). The distribution coefficient is found to be a function of the hydrophobic character of the organic compound and the amount of organic matter present and may be written as:
Kct = K f o c
[9.15]
where Koc is the organic carbon partition coefficient, and foe is the fraction of organic carbon within the soil matrix. Sorption partition coefficients indexed to organic carbon can be estimated from physico-chemical properties of pollutants such as octanol-water partition coefficient and solubility, as discussed in Chapters 5 and 6. Eq. [9.15] is valid only for foc greater than 0.001. Otherwise, sorption of organic compounds on non-organic solids (clays and mineral surfaces) can become
216
SUBSURFACE POLLUTANT TRANSPORT
significant. Also, the linear isotherm model is valid only if the solute concentration remains below the solubility limit of the compound. Typical calculated values for Kd are shown in Table 9.3 (Acar and Haider, 1990).
Table 9.3: Partition coefficient (Ka) and retardation coefficient (R) of selected organic pollutants (Acar and Haider, 1990). Pollutant Location and soil type Kd R foc % (cm3/~a) Acetone Benzene N-Butyl alcchol N-Butyl benzene Carbon Tetrachloride Chlorobenzene Chloroform 3,5-Dichlorobenzene O-Dichlorobenzene 1,2-Dichlorobenzene Dichloroethane 1,4-Dimethylbenzene Ethylbenzene Napthalene Nitrobenzene Quinoline 1,2,3,4-Tetrachlorobenzene Tetrachloroethylene 1,2,4-Trichlorobenzene Trichloroethylene
NA 0.0018 ...... Woodburn 0.34 --NA 0.018 ...... Glatt Valley, Switzerland 3.69 ----0.42 ...... Woodburn 0.50 --NA --1.2 Haggerstown silty loam 2.50 --Woodburn 3.45 --Willanette silty loam 3.30 --Catlin silty loam 4.10 --Glatt Valley, Switzerland 0.50 --Woodbum 0.91 --Haggerstown silty loam 8.00 --NA 0.12 1.4 Haggerstown silty loam 4.30 --Glatt Valley, Switzerland 10.48 --Catlin silty loam 17.30 --NA --7.00 Baton Rough high plasticity 8.30 4.00 Clay Toluene Glatt Valley, Switzerland 0.37 --O-Xylene Catlin silty loam 10.40 --foc = organic carbon content; NA = not applicable; a = organic matter content (%).
1.10 NA 1.10 NA 1.94 1.10 1.60 a 4.04 NA 1.10 1.94 NA 1.10 NA 4.04 NA 1.0 NA 4.04
Using the linear sorption isotherm and assuming local equilibrium, the expression for the bulk concentration can be written as:
m = ngbKaC
[9.16]
Losses due to biotic processes are often modelled as either first or zero order decay. Assuming that the source term is actually represented as a first order equation with an apparent or effective rate constant, A, then:
TRANSPORT EQUATION S + = - ~.m
217 [9.17]
The relationship given by Eq. [9.17] states that the loss rate is proportional to the total mass present. As long as linear partitioning relationships hold, then Eq. [9.17] remains true as an effective rate constant. This does not, however, actually represent the loss rate from the aqueous phase. The mass flux vector, J, in terms of advection, diffusion and dispersion, is given by: J = Jdvection + Jdiffusion + Jdispersion
[9.18]
J = nVsC - nDhd" Vc
[9.19]
where Dhd is hydrodynamic dispersion coefficient which is given by: [9.20]
D hd = T,D O+ a lYs ~
Substituting Eqs. [9.16], [9.17] and [9.19] into Eq. [9.13] gives: (n + PbKd) OC + 7 " ( n v s c ) + ~ (n + PbKd) c = V " (nDhd" 7C) Ot
[9.21]
If the porosity is constant then we introduce the retardation f a c t o r as: R=I+--K
Pb
[9.22]
d
n
The physical significance of the retardation f a c t o r is that it measures how much slower a solute migrates than water. Thus a retardation factor of ten means that the average speed of the solute is ten times slower than that of water. In addition, if the flow is steady and there are no volumetric sinks (leakage, infiltration, evaporation, etc.) then Eq. [9.21] may be written as: R Oc + v s 9Vc + )~Rc - V ' ( D h d " Vc) Ot
[9.23]
Eq. [9.23] is the form of the pollutant transport equation that is used for most analytical models in one, two, and three dimensions with the added assumption that the velocity field is uniform. In particular, the one-dimensional form of the pollutant transport equation is: OC Ot
-
Dhd 02C R
Ox 2
-
Vs OC R
Ox
-,~c
[9.24]
218
SUBSURFACE POLLUTANT TRANSPORT
The first term gives the rate of change of concentration at a given location. The second term accounts for the effect of hydrodynamic dispersion. The third term accounts for the advection effect and the last term accounts for the sink term which is modelled as first order decay.
9.5
SOLUTE TRANSPORT MODELS
Analytical and numerical models are used to simulate the subsurface transport of chemicals. The numerical models are the most general since they may be tailored to address site specific conditions. However, these numerical models require a significant data base, which may not be available in the initial phases of a site investigation. Analytical models require simplifying assumptions, but are much more computationally efficient and need less specific data. They are specially useful for screening calculations. This section focuses on analytical models for column experiments, chemical spills, and chemical plumes from continuous releases of pollutants. 9.5.1
Conservative Tracer
A conservative tracer is a substance that moves through a porous media without interacting with the matrix or undergoing chemical transformations. Understanding the transport of conservative species is the first step toward understanding the fate and transport of hazardous and radioactive chemicals, which may be influenced by a wide range of processes, namely advection, diffusion, and dispersion. Diffusion
For cases where the seepage velocity is very low, diffusion will govern the transport process. The transport equation is given by:
de Ot
- Ds
02C OX 2
[9.25]
This equation represents the rate of change of solute concentration. The solution for this equation will depend on the initial and boundary conditions. Given a constant source concentration and the following initial and boundary conditions: at
t = O:
at x = O atx=l at x = 0
C = Co > 0 c=O
at
t>0."
c=co
the solution to Eq. [9.25] is given by Crank (1956) as:
erC/co
[9.26]
SOLUTE TRANSPORT MODELS
219
where Ds is soil diffusion coefficient.
Advection- Dispersion The advection- dispersion equation for conservative species takes the form: 0r
02e
Ot
- o hd OX 2
Table 9.4" Complementary error function
0 0.05 0.1 0.15 0.2 0.25 0.3 0.35 0.4 0.45 0.5 0.55 0.6 0.65 0.7 0.75 0.8 0.85 0.9 0.96 1.0
erf({)
erfc(~)
0 0.056372 0.112463 0.167996 0.222703 0.276326 0.328627 0.379382 0.428392 0.475482 0.520500 0.563323 0.603856 0.642029 0.677801 0.711156 0.742101 0.770668 0.796908 0.820891 0.842701
1.0 0.943628 0.887537 0.832004 0.777297 0.723674 0.671373 0.620618 0.571608 0.524518 0.479500 0.436677 0.396144 0.357971 0.322199 0.288844 0.257899 0.229332 0.203092 0.179109 0.157299
OC -
Vs
[9.27]
OX
(erfc(~))
1.1 1.2 1.3 1.4 1.5 1.6 1.7 1.8 1.9 2.0 2.1 2.2 2.3 2.4 2.5 2.6 2.7 2.8 2.9 3.0
erf(~)
erfc({)
0.880205 0.910314 0.934008 0.952285 0.966105 0.976348 0.983790 0.989091 0.992970 0.995322 0.997021 0.998137 0.998857 0.999311 0.999593 0.999764 0.999866 0.999925 0.999959 0.999978
0.119795 0.089686 0.065992 0.047715 0.033895 0.023652 0.016210 0.010909 0.007210 0.004678 0.002979 0.001863 0.001143 0.000689 0.000407 0.000236 0.000134 0.000075 0.000410 0.000022
The solution to Eq. [9.27] subject to the above stated initial and boundary conditions is given by Crank (1956) as"
Co
2
2 D ~h~)
exp ~
,erfc
2D~J)
[9.28]
220
SUBSURFACE POLLUTANT TRANSPORT
The erfc appearing in the analytical solution is the complementary error function which, for any argument ~, is given by: erfc ( - ~) = 2 - erfc (~)
erfc (~) = 1 - e r f (~);
[9.29]
where erf(~) is the error function of the argument ~ and is given by:
err (~)
= 2
,(-, (-1)" ~2,+1
,/-~ ~_- n((Zn + 1)
[9.30]
y--
Values of erf(~) and erfc (~) are shown in Table 9.4 (Freeze and Cherry, 1979).
Sample problem 1: A site investigation study has revealed that the site contains mixed chemicals buried beneath the ground surface and extending to a depth of 1 m, as shown in Figure 9.1. A thick layer of clay measuring 3 m in depth was found under the buried waste. The groundwater level was found at a depth of 0.5 m from the ground surface. At a depth of 4 m from the ground surface, the groundwater analysis has shown that one of the dominant pollutants has a concentration, Ce, of one half of the initial concentration, Co, at a depth of 1 m from the ground surface. The clay soil has a porosity of 0.5 and hydraulic conductivity of 5 x 10.9 m/sec. It is required to calculate the transient times by considering advection only, diffusion only, and advectiondispersion processes.
Figure 9.1. Schematic diagram showing site characteristics.
SOLUTE TRANSPORT MODELS Solution: hydraulic head, h, = soil layer depth, L, = hydraulic gradient, i, = Soil diffusion coefficient, D, = Dispersion coefficient, ~ =
221
0.5 m 3m
(h+L)/L = (0.5 + 3)/3 = 1.167 ~ v,. 1m
(a) Advection process
v~=ki/n t = L/vs
5• -9 x 1.167/0.5 = 1.167 • 10 .8 m/sec = (3/1.167 • 104) • (1/365 • 24 • 60 • 60) = 8.15 years =
(b) Diffusion process Using Eq. [9.26]"
c/c o X
D,.
=0.5 =3m = 1 x 1.167 x 1 0 4 = 1.167 • 10 .8 m2/sec =x/2VDst=3/(2 v / 1 . 1 6 7 x 1 0 .8 x 365 x 24 x 60 x 60 t) = 2.4726/ V t
The transient time calculations are shown in Table 9.5. By plotting the relationship between
CJCo and t, the transient time at CJCo =0.5 found to be approximately 28 years.
Table 9.5" Transient time calculations for diffusion process t (years)
~
erfc (~)
c~/c~
5 10 20 30
1.105 0.78 0.55 0.451
0.1198 0.2889 0.4366 0.5245
0.1198 0.2889 0.4366 0.5245
(c) Advection-dispersion processes By using Eq.[9.28]
[9.28]
222
SUBSURFACE POLLUTANT TRANSPORT x vst ~
v~ t/D~ ~2
=3m = 1.167 • 10 .8 • 365 • 24 • 60 • 60 t = 0.368 t = (x- v, t) / 2 v/ D,. t = (3- 0.368 t )/(2 v/ 0.368 t) = ( 3 - 0.368 t) / 1.2133 ~/t = 1.167 x 10 .8 • 3/1.167 x 10-8= 3 = (x + vs t) / 2 ~/Ds t = (3 + 0.368 t )/(2 v/ 0.368 t) = ( 3 + 0.368 t) / 1.2133 v/t
The transient time calculations for advection- dispersion process are shown in Table 9.6. By plotting the relationship between C/Co and t, the transient time at c/co =0.5 found to be approximately 7 years.
Table 9.6: Transient time calculations for advection- dispersion process
9.5.2
t (years)
~l
erfc (~l)
~z
erfc (~z)
c,/c o
5 7 8 10
0.427 0.424 0.0163 -0.177
0.5716 0.5716 0.943 1.168
1.784 1.737 1.732 1.741
0.0109 0.0162 0.0162 0.0162
0.395 0.448 0.634 0.747
Reactive Chemical Species
For reactive chemical species with retardation factor, R, the equivalent forms of Eqs. [9.26] and [9.28] are:
Co and Co
er cl I 2V/(
/R)t
l(e lX Vsl tl+ IvJ
2
2~/(D~a/R)---~tj
exp ~ h a
,erfc
[9.311
-2v/(Dhd/R)t) -
I
[9.32]
respectively.
Sample P r o b l e m 2" For the same information given in sample problem 1, what are the transient times if R - 2.
SOLUTE TRANSPORT MODELS
223
Solution: By following the same procedures as described previously, the transient times will be doubled. For diffusion only, t = 56 years, and for advection dispersion, t = 14 years. 9.5.3
Spill of Pollutants
Pollutant releases to the subsurface may occur over short periods of time, or there may be a slow continual release over a long time period. The first of these cases may be modelled by assuming that the release is instantaneous, while the second leads to plume models. If a mass of pollutant is released to the water table over a short duration, the chemical slug is subsequently transported downgradient as it spreads in the direction of flow, transverse horizontally and transverse across the thickness of the aquifer. If the flow is uniform in the x-direction at a seepage velocity, vs, then the concentrations in one, two, and three dimensions are given by (Hunt, 1978; Wilson and Miller, 1978):
(x-(v]R)t)2) 4(Dx/R)t
Mexp(-M) exp[-
c(x,t) =
[9.33]
nR~(4nD~~)
(x -(v]R)t) 2 M exp(-M) exp
c(x,y,t)
-
4~Dx~
ye
)
- 4(/~-)R)t
=
[9.34]
nR
~(4~~-~)(4r~_Dff_[)
and M c(x,y,z,t)
( exp(-M) exp [ -
-
(x
y2
(v]R)t) 2
4(D]R)t
-
_
4(D~/R)t
z 2
)
4(D=/R)t [9.35]
nR
d(4~-~)(4~-D~
f ) ( 4 n ( D ; t)
where c is the concentration (M/L3), M is the mass of spilled pollutant which varies with the number of dimensions: (a) in one dimension, it is an instantaneous plane source (M/L2), (b) in two dimensions, it is an instantaneous line source (M/L), and (c) in three dimensions, it is an instantaneouspoint source (M), x is distance measured in the direction of down-gradient from the point of spill (L), D~, D~, and D= are hydrodynamic dispersion coefficients in x, y, and z directions respectively (L2/T), n is the porosity (L3/L3), t is time (T), R is retardation parameter, and v~, is
224
SUBSURFACE POLLUTANT TRANSPORT
seepage velocity (L/T). According to Eqs. [9.33] to [9.35] the maximum concentration occurs at the location x = vs t/R and y = z = 0, and is given by: M exp(-~t) C m a X --
nR
~(4~-Dff f ) ( 4 ~ _ D ~ ) ( 4 ~
(Dzt)
[9.36]
The main feature of interest in Eq. [9.36] is that near the source, the maximum concentration decreases as:
Cmax
N
1
t3/2
or since the distance of migration of the centroid of the polluted mass is given by concentration decreases a s L -3/2.
[9.37]
L = vs t/R, the
Sample Problem 3: An instantaneous release of 5 kg of a pollutant due to an accident took place on a surface of a soil deposit. The soil was characterized by having a porosity of 0.3 and seepage velocity of 0.2 m/day. Assume a dispersion coefficient of 1 m and decay constant of 0 day -1 (no decay). It is required to calculate the following, considering one-dimensional analysis: (1) The concentration at a distance of 1 m from the source after 1 day, and (2) The maximum concentration and it is location.
Solution: We use Eq. [9.33 ]" Mexp(-~,t) exp (-(x-(vs/R)t)2)4-(D--~
c(x,t)
=
[9.33]
nRI(4~D~~)
where: M
D~ R A n x
t
=5kg = 0.2 m/day = err v,.= 1 • 0.2 = 0.2 m2/day =1 = 0 day l =0.3 =lm =lday
SOLUTE TRANSPORT MODELS
225
yielding the concentration at x = 1 m and t = 1 day: c(1,1) = { 5 e x p [ - ( 1 - 0 . 2 x l ) 2/(4x0.20]}/{0.3xl~/4x• = 4.72 kg/m 3 The maximum concentration is located at a distance x = Cmax = {5}/ {0.3 X 1 X/4 X • 0.2 t} = 10.513 kg/m 3
9.5.4
Vs t/R
= 0.2 m.
Pollutant Plume
If a pollutant is released at a constant rate from the source for a long enough period of time, a concentration distribution in the shape of a pollutant plume develops. The steady plume models are useful in conservative predictions because they provide the maximum likely concentration that may be observed at a given location. The concentration distribution in one, two, and three dimensions are given by (Hunt, 1978; Wilson and Miller, 1978):
c(x,t)
= Am exp (-~,t)
2n(vJR)
erfc ( x - (Vs/R)t I ( -V/4(D-~t )
[9.381
where A m is continuous plane source (ML2T~). The concentration in two dimensions is given by:
c(x,y,t)
-
Am exp (-~,t)
exp
(x
2n X/4X(Vs/R)(D~/R)r
r)vs 2D
erfc - - ~ ~/4(Dxx/R)t )
[9.39]
where A m is continuous line source (MLITI), and r=
X2 +
D,y
y
The concentration in three dimensions is given by:
c(x,y,z,t) -
Am exp (-Xt)
exp
8nxr ~/(D /R)(DJR)
rvsI i r vs ,l
219=
erfc - V/4(Dxx/R)t )
where d m is continuous point source (MT~), and r = I x 2 + D yy xx
2 +D= Z 2
Dzz
[9.40]
226
SUBSURFACE POLLUTANT TRANSPORT
Sample P r o b l e m 4: A continuous release of 5 mg/day of a pollutant due to leakage is taking place on a surface of a soil deposit. The soil was characterized by a porosity of 0.3 and seepage velocity of 0.2 m/day. Assume a dispersion coefficient of 1 m and a decay constant of 0 day 1. It is required to calculate the concentration of pollutant as a function of distance after one year, using onedimensional analysis. Solution: We use Eq. [9.38]
c(x,t) =
Am exp (-~,t)
erfc
2n(vs/R)
x - (v/R)t]
[9.38]
(4(D J R ) t )
where: Am
= 5 mg/day = 0.2 m/day = a'r v,.= 1 • 0.2 = 0.2 m2/day =1 = 0 day l =0.3 = 365 days
Ys
D~x R 2 n t
and obtain:
c(x,
365 days) = {5 / (2 • 0.3 • 0.2)}
erfc
[(x- 0.2 • 365)/( v/4 • 0.2 • 365)]
The concentration profile is tabulated in Table 9.7.
Table 9.7: Concentration variations with distance after 1 year for a pollutant source with a constant rate of release x (m)
~
erfc (~)
c (mg/m 3)
1 10 20 30 40 50 60 70 80 90 100
-4.21 -3.686 -3.10 -2.516 -1.931 -1.346 -0.761 -0.176 0.409 0.995 1.58
2 2 2 1.999 1.993 1.934 1.711 1.223 0.572 0.157 0.024
83.33 83.33 83.33 83.29 83.03 80.58 71.29 50.96 23.83 6.55 0.987
METHODS FOR CALCULATING TRANSPORT PARAMETERS
9.6
METHODS FOR CALCULATING TRANSPORT PARAMETERS
9.6.1
Laboratory Methods for Hydraulic Conductivity Testing
227
Notable differences are observed between hydraulic conductivities measured in the laboratory and in the field. The ratio between field and laboratory hydraulic conductivities may be as large as 1000. The difference may be attributed to the fact that the laboratory apparatus is capable of testing only a small volume of soil. Therefore, the sample may be too homogeneous, and not representative of actual conditions. In the field, soils tend to be heterogeneous, with cracks, fissures, roots, and animal burrows. These factors will all affect soil hydraulic conductivity. Also, construction procedures may not be highly controlled, creating variation in water content and compaction. For these reasons, extreme care must be taken when testing soils for hydraulic conductivity in the laboratory. Due to time constraint and expense of field testing, laboratory measurements will almost always be required. There are at least two basic types of apparatus used for laboratory testing: the rigid wall permeameter and the flexible wall permeameter. Considerable discussion on the relative merits of these permeameters exists (Daniel et al., 1985; Bowders et al., 1986; Madsen and Mitchell, 1989). Each has its advantages and disadvantages and a discussion of these considerations is presented here.
(a) Compaction mould permeameter
Figure 9.2. Compaction mould permeameters.
(b) Double-ring compaction mould permeameter
228
SUBSURFACE POLLUTANT TRANSPORT
Rigid Wall Permeameter The are two principal types of rigid or fixed wall permeameters: the modified compaction mould, and the consolidation cell. The compaction mould permeameter is the same as that used in ASTM D698 for determining moisture-density relations for soils. It consists of two end plates, a cylindrical compaction mould, and a collar, as shown in Figure 9.2(a). The soil is compacted directly into the mould, trimmed, and permeated with the liquid stored in the collar place above the soil specimen. Back pressure saturation is not applied, and the hydraulic pressure gradient is achieved by pressurizing the permeant with compressed air. The inflow and outflow quantities are measured to determine the rate of flow for a period of time. Darcy's Law is used to calculate the hydraulic conductivity (Bowders et al., 1986). The main advantages of this cell, as reported by Madsen and Mitchell (1989), are the low cost of test sampling, and the simplicity of operation. Other strong points include the ease in testing compacted samples, and the use of reasonable confining pressures. However, the compaction mould has several serious drawbacks. The first is that since no back pressure is applied, complete saturation cannot be achieved. Also, there is no means of controlling the applied stresses, since the applied vertical stress is zero. This is inconsistent with field conditions where an overburden creates vertical stress. The greatest disadvantage is the potential for sidewall leakage. This is a major concem when permeating the soil with organic chemicals, which have a tendency to reduce the diffuse double layer and cause shrinkage. Daniel et al. (1985) reported tests where visible gaps were seen between the soil and the cell wall following permeation of kaolinite with heptane. A double-ring compaction mould consists of a compaction mould and a ring built into the base plate, as shown in Figure 9.2(b). The primary function of the ring is to separate the outflow that occurs through the central portion of the soil from the outflow that occurs near or along the sidewall. If there is significant sidewall leakage, the rate of flow into the outer collection ring will be much greater than the rate of flow into the inner ring. If the rates of flow, adjusted to take into account the differences in cross-sectional area of the test specimen that is intercepted, are unequal, one could reject the test and setup a new one. Although at present there is relatively little experience with this device, the cell has worked well in a few tests and shows excellent promise. The disadvantages are the same as those indicated previously for the compaction mould permeameter, except that one has some indication of the magnitude of sidewall leakage. For hydraulic conductivity testing, the standard consolidation procedure is followed, whereby the sample is trimmed into a ring (typical height of 13 to 25 mm and typical diameter of 50 to 80 mm) and then clamped into the base. The reservoir surrounding the ring is filled with water, and the soil is consolidated by applying the desired vertical stress, with deformation measured by a dial gauge. Permeant is then introduced at the base of the sample, with the leachate flowing upward. A schematic presentation of the consolidation cell is shown in Figure 9.3. A modification may be made whereby a thicker walled cell is used, so that compaction takes place directly in the ring. Consolidation and permeation follow the same procedures as for the standard consolidation cell. Since the sample is consolidated, the applied vertical stress can be of the same magnitude as that found in the field. The applied stress tends to push the soil against the sidewalls, helping to minimize leakage. For the modified cell, compacting directly in the mould ensures a seal between the material and the cell walls. The dial gauge provides for the determination of vertical deformation. Other advantages include cost efficiency and the fact that the tests do not take as long to conduct since the samples are relatively thin. The upward flow of the permeant helps saturation. Although the potential for sidewall leakage is reduced, it can never be eliminated for a rigid wall permeameter.
METHODS FOR CALCULATING TRANSPORT PARAMETERS
229
Another drawback is that complete saturation cannot be ensured. The application of back pressure may expand the ring, thereby creating sidewall leakage. Finally, the thin samples, though they shorten testing time, may not be representative of field conditions.
Figure 9.3. A schematic of the consolidation cell.
Flexible Wall Permeameter
Triaxial cells are used to perform flexible wall hydraulic conductivity tests. The specimen is confined in a membrane which is pressurized to keep it in contact with the soil specimen. This virtually eliminates sidewall leakage. A variety of specimen heights and diameters may be used, depending on the size of the apparatus. A pressure transducer measures pressure drop across the sample, and double drainage lines help to clear air bubbles. Back pressure is generally applied to fully saturate the soil. All stresses, horizontal and vertical, may be controlled, and specimen deformation may be measured. A schematic presentation of the flexible-wall permeameter is shown in Figure 9.4. A drawback of this device is the possibility of chemical pollution of the membrane. Some chemicals react with latex, thereby destroying the membrane. To alleviate this problem, the specimen may be wrapped with Teflon tape prior to placement of the membrane, or, for short term tests, the membrane could be placed in the chemical solution before starting the test to determine if chemical attack will occur. The high confining pressure applied to the specimen tends to close any cracks in the soil. Since a field soil is not homogeneous, this situation does not replicate actual conditions. The laboratory hydraulic conductivity in this case will be lower than the true value. Other disadvantages include higher costs and increased testing time due to the use of larger samples.
230
SUBSURFACE POLLUTANT TRANSPORT
Figure 9.4. A schematic presentation of the flexible-wall permeameter.
9.6.2
Laboratory Methods for Adsorption Characteristics
In order to study the adsorption characteristics of soils, two experimental techniques are generally used in the laboratory: (1) batch equilibrium test, and (2) soil column leaching test. These techniques are described below:
Batch Equilibrium Test The main objectives of this test are to: (I) study soil attenuation of pollutants at equilibrium, (2) estimate the number of pore volumes required to achieve breakthrough of a constituent into the effluent liquid, and (3) calculate the retardation parameter required in the pollutant transport equation. Test procedures The clay is air-dried and ground with a mortar and pestle until a uniform powdery texture is obtained. A fixed amount of soil, say 4 g, is placed in glass sample bottles. Various concentrations of leachate are prepared, say 60, 250, 500, 1000, 1500, 2000 ppm. A constant volume of each leachate concentrations, say 40 ml, is then added to each bottle which is then capped tightly. A solution with a soil ratio of 10:1 was recommended by the United States Environmental Protection Agency (EPA, 1987) for estimating soil attenuation of chemicals from batch adsorption tests. Triplicates are prepared for each concentration to ensure accuracy. The bottles are then placed in a rotary shaker and kept at a constant temperature of 20 ~C. The bottles are shaken for at least 24 hours to ensure equilibrium. At the end of the shaking period, the samples are centrifuged to separate the clay from the liquid. One of the bottles, however, contains just the leachate and no clay (blank). The supernatant liquid from the bottles is filtered, and the equilibrium concentration in the liquid phase
METHODS FOR CALCULATING TRANSPORT PARAMETERS
231
of a constituent of interest (c, expressed in units of mass of constituent per unit volume of liquid) is measured using the appropriate analytical method. The concentration of the constituent in the leachate itself is determined from the bottle with no soil and is denoted Co. Analysis The adsorption mass ratio, q, is computed for each bottle as follows: q :
(Co- C) V
M
[9.41]
where V is the volume of liquid in a bottle (40 m/), and M is the mass of soil in the bottle (4 g). The numerator in Eq. [9.41 ] represents the mass of constituent adsorbed onto the solid phase, and it is divided by the mass of the soil to obtain a measure of the relative mass of the constituent adsorbed on the solid phase. The values of q are plotted as a function of the equilibrium concentration. For constituents at low or moderate concentrations, the relationship between q and c can be expressed as: q = kdc b
[9.42]
where kd and b are coefficients that depend on the constituents, nature of the porous material and the interaction mechanism between it and the constituents. Eq. [9.42] is known as the Freundlich isotherm. If b = 1, then q versus c data will be a straight line. Such an isotherm is termed linear, and Eq. [9.42] with b = 1 reduces to: dq _ kd dc
[9.43]
where kd, known as the distribution coefficient, is used for pollutant partitioning between liquid and solids only if the reactions that cause the partitioning are fast and reversible and only if the isotherm is linear. For cases where the partitioning of the contaminants can be adequately described by the distribution coefficient (i.e. fast and reversible adsorption, with linear isotherm), the retardation factor, R, can be expressed as: R = 1 + Pd kd
[9.44]
n
where Pd is the dry mass density (mass of dry solids divided by the total volume of the soil) of the test specimen, and n is the porosity of the test specimen. The retardation parameter also can be expressed as the ratio of the breakthrough times of an adsorbed chemical relative to that of a nonadsorbed tracer.
232
SUBSURFACE POLLUTANT TRANSPORT
Leaching Column Test The main objectives of leaching column test are to: (1) study pollutant migration and attenuation by soils, and (2) estimate the transport parameters which control the migration of pollutants through soils.
Test procedures In the leaching column test, rigid-wall or flexible-wall permeameter is used. The soil is compacted into four replicate soil columns, and the leaching cells are then assembled. First, steadystate fluid flow is established through the soil specimen. After steady state has been established, the fluid in the effluent reservoir (usually water) is changed to a solution with known and constant concentration (Co) of particular chemical species. The concentration, c e, of a chemical species appearing in the effluent reservoir is measured over time and the results are plotted in the form of solute breakthrough curves, or relative concentration, CJCo, versus time or pore volumes of flow (PV). A pore volume of flow for a saturated soil is the cumulative volume of flow through the soil divided by the volume of the void space in the soil. Chemical Analysis Soluble ions: A known amount of soils, 30 g, is taken from each layer. The soil is placed in two plastic containers, to which distilled water is added in two stages. The soil/liquid ratio is 1:10, which is the same ratio used for the adsorption isotherm. The mixture is shaken for approximately 3 hours, poured into small tubes, and centrifuged. The supernatant is separated from the clay and stored in two glass bottles. This washing procedure removes pore liquid, which may contain some of the pollutants, from the clay. The soil remaining after washing is placed in an oven and allowed to dry overnight. The supernatant, washes, and the leachate are analysed with an atomic absorption spectrophotometer. The reading from the machine minus the concentration found in the background solution is the true quantity of the required cation in ppm. The conversion to meq/100g from ppm requires the following formula meq/lOOg = ppm x 2.5 EWC
[9.45]
where EWC is equivalent weight of cation. The amount of specific constituent in the pore solution is determined as follows: m =
[C] x WDS • 400 ml WDS., x 1000 ml
[9.46]
where mc is mass of constituents (g), [C] is the concentration of specific constituent in the wash in (g/l), WDS is the weight of dry soil for each layer (g), 400 ml is the volume of distilled water used in washing, WDSw = 30 g/moisture content at end of test, and 1000 ml is the amount of ml in one litre. The total mass of specific constituent is the sum of the mass of constituents per layer.
METHODS FOR CALCULATING TRANSPORT PARAMETERS
233
Exchangeable ions: Exchangeable ions can be determined from the Silver-thiourea method (Chhabra et al., 1975) method in which ammonium acetate is prepared under a fume hood by adding 57 ml of concentrated acetic acid and 68 ml of concentrated ammonium hydroxide to 700 ml of distilled water. The pH of the solution can be adjusted as required. Once the required pH is reached, distilled water is added to the flask to make one litre. Four grams of the dry soil are carefully weighed into a plastic centrifuge tube. Triplicates are desired to ensure accuracy, so a total of 12 g of dry soil is needed. To each of the three centrifuge tubes, 33 ml of ammonium acetate is added. The tubes are then placed in a mechanical shaker for approximately 2 hours. Following shaking, the tubes are placed in a high speed centrifuge for approximately 10 minutes. The supernatant is then decanted into a 100 ml container. The whole process is repeated two more times by adding 33 ml and 34 ml of ammonium acetate, respectively. The supernatant is, then, analysed using double beam atomic absorption spectrophotometer. The analysis involves measuring various ions in the supernatant, as well as any background levels in the ammonium acetate, with the final concentration equal to the difference between the two. To convert from p p m to meq/100g, the following formula is used m e q / l OOg -
[C] • 100 ml • 100 g x 1000 1000 ml • 4 g • E W C
[9.47]
where 100 ml is the amount of ammonium acetate, 1000 ml is the amount of ml in one litre, 4 g is the amount of dry soil, and 1000 is used to convert to meq. The amount of specific constituent in the soil is found using the expression: _ [C] x W D S x 100 ml 4 g x 1000 ml
m -
[9.48]
where [C] is the average concentration of specific constituent in soil in (g/I), WDS is the weight of dry soil for the layer, 100 ml is the amount of ammonium acetate added, 1000 ml is the amount of ml in one litre, and 4 g is the weight of the dry soil used.
Mass Balance: To determine the total volume input, the total number of pore volumes collected is multiplied by the size of one pore volume, in ml. The final output component is the leachate. For each pore volume, the exact volume of solution which passed through the sample is determined. The amount of specific constituents in g/ml is multiplied by the volume of leachate, thereby giving the amount of specific constituent in a particular pore volume in grams. The procedure is repeated for all the pore volumes. The final mass balance equation for specific constituent is M e : M L + Msl + MEt 4-sink~source
[9.491
where M I is the mass introduced, M E is the mass leached, MsI is the mass of soluble ions (in pores), and MEI is the mass of exchangeable ions. The sink term is due to microbial action while the source term is due to desorption of ions.
234
SUBSURFACE POLLUTANT TRANSPORT
Data Extracted From Leaching Column Tests Four kinds of data can be obtained from leaching column test. These are: (1) breakthrough curve (effluent concentration versus pore volume), (2) migration profiles (concentration versus depth) of soluble ions, (3) exchangeable profiles (concentration versus depth) of adsorbed constituents by soils, and (4) adsorption isotherm (q versus c). To highlight these kinds of data, a specific example is discussed (Mohamed et al., 1994). Material and Methods: The natural micaceous soil used in this investigation consists of illite, phlogopite, hydrobiotite, and vermiculite as basic elements. Specific surface area determined by using the Ethylene Glycol Monoethyl Ether (EGME) adsorption method (Carter et al., 1965) = 206 m2/g. Cation exchange capacity of soil was determined by using two methods" (1) batch equilibrium test - ASTM D 4319 (1984), CEC = 14.89 meq/100g, and (2) the silver-thiourea method (Chhabra et al. 1975), CEC = 13.2 meq/100g. The low CEC values indicated that the soil is mainly illitic and mica in composition. The engineering properties of this soil are" (1) maximum dry density = 1.81 Mg/m 3, (2) optimum moisture content = 16.1%, and (3) hydraulic conductivity, using rigid wall permeameter = 2.3 x 10.9 m/sec.
Test Procedure: A leaching column test was conducted as described above. Soil was compacted at its optimum moisture content and maximum dry density using the static compaction method. Leaching was carried out using municipal solid waste leachate spiked with heavy metals (Pb 2+, Zn2+) and cations (Na +, K § Mg 2+, Ca 2+) in the form of chlorides. The pH of the reconstituted leachate was adjusted to pH of 1.33 by adding concentrated hydrochloric acid. The chemical composition of the reconstituted leachate is: (1) heavy metal concentrations (ppm): Pb 2+= 1372.2, and Zn 2§ = 1141.6, and (2) cation concentrations (ppm): Na 2+ =346, K + = 164.8, Mg 2§ = 43.8, and Ca 2+= 94.5.
Figure 9.5(a). Variation of heavy metal relative concentration with number of pore volumes.
METHODS FOR CALCULATING TRANSPORT PARAMETERS
235
Leaching was carried out under a constant pressure of 103.5 kPa, resulting in a hydraulic gradient of 87.2. During the leaching process, the effluent was collected as a function of time and analysed to determine its chemical composition. At the end of the test, the specimen was extruded, cut into 10 mm thick slices, which were, then, analysed for pore fluid contents (soluble ions) and exchangeable ions.
Migration of Heavy Metals: In a clay soil system, heavy metal may (1) occur in ion exchange sites, (2) be incorporated into or on the surface of crystalline or non-crystalline compounds, or (3) be in the soil pore solution. Most investigations have recognized that heavy metals occur predominantly in a sorbed state. Because of their low solubility, movement of heavy metals in soil has generally been considered to be minimal. Figure 9.5(a) shows the effluent relative concentration, C/Co, in the leachate, for lead, Pb 2+, and Zinc, Zn 2+, as a function of pore volume passage. Ceis the concentration of the ion concerned in the effluent, and Co is the original concentration of the ion concerned in the influent. From the diagram, it is evident, therefore, that most of the Pb > and Zn > are retained in the soils. The results show that a significant amount of heavy metals was retained in the top portion of the soil samples, as seen in the concentration profiles depicted in Figure 9.5(b). Due to lower pH in the influent leachate, it is expected that the retention capacity of the soil in the top part of the column is reduced. However, this depends on the buffering capacity of the soil to any change in pH. It is known that heavy metals would generally precipitate out of the solutions if the solution's pH is high (e.g. Pb > precipitates at pH >5). Since soil pH was initially about 6.5, Pb> precipitates in the soil at the start of leaching. Further leaching decreases soil pH and enhances the mobility ofPb > in solution. After 5 pore volumes, the top 25% of the soil column has pH values ranging from 1.33 to 5, thus enhances the mobility of Pb 2§ in this part of the soil column. Pb > was retained by cation exchange replaceability, as shown in Figure 9.5(c). For the rest of the soil column, which has pH greater than 5, Pb 2+retention mechanism will be due to precipitation in various forms.
Figure 9.5(b). Variation of heavy metal pore fluid concentration with distance.
236
SUBSURFACE POLLUTANT TRANSPORT
Figure 9.5(d). Heavy metal adsorption isotherm; Y(1) and Y(2) are the left and the right vertical axes, respectively.
It can be seen also from Figure 9.5(b) that the amount of Zn 2§ retained by the soil column is less than the amount of Pb 2§ retained. This can be explained by the ease of exchange or the strength with which cations of equal charge are held, which is generally inversely proportion to the hydrated radii or proportional to the unhydrated radii. Therefore, the predicted order of soil retention based on: (1) unhydrated radii is Pb 2§ (0.120 nm) > Zn 2+ (0.074 nm) which agrees with the experimental
METHODS FOR CALCULATING TRANSPORT PARAMETERS
237
data, and (2) metal ion softness, which is a function of ionization potential, change of metal ion and ionic radius, is Pb 2+ (3.58) > Zn 2+ (2.34). In order to investigate the adsorption isotherm of Pb 2+ and Zn 2+, chemical test data are presented in terms of exchangeable cations and equilibrium solution concentration in the pore fluid, as shown in Figure 9.5(d). It can be seen from Figure 9.5(d) that exchangeable Pb 2§ and Zn 2§ increased for lower pHs up to pH = 5, hence the exchangeable cations decreased with further decrease in pH.
Migration of Cations: Figure 9.6(a) shows the effluent relative concentration, C/Co, in the leachate for cations (Na+, K § Mg 2+, and Ca2+). With increasing number of pore volumes permeated, the relative concentration of cations increased. For Na § Mg 2§ and Ca 2§ cations, the relative concentration exceeds 1.0. This can be attributed to the elution of cations from the solid particles. The high relative concentrations of Ca2+and Mg 2+in the effluent collected can be attributed to cation exchange or replacement by Pb 2§ and Zn 2+. Low concentration of K § in the effluent leachate is due to the fact that K § is often adsorbed and incorporated into the interlayer lattice of micaceous soils. Since the soil tested is basically micaceous, there is greater affinity for K § hence adsorption and incorporation into the inter-layer lattice of the soils of the K § ions occurred during the leaching process. Also, it can be noticed that Na + relative concentration reached a steady state approximately after 4 pore volumes while for K § Mg 2+, and Ca 2+, steady state conditions arrived after approximately 2 pore volumes.
Figure 9.6(a). Variation of metal ion relative concentration with number of pore volumes.
The migration profiles of pore fluid cation Concentration versus depth of the soil column are shown in Figure 9.6(b). The migration profiles depict how a particular cationic species migrate or move through the soil column with increasing permeation by the leachate. The initial concentration ofNa + in the influent leachate was 345 ppm while the measured concentration in the pore fluid was
238
SUBSURFACE POLLUTANT TRANSPORT
greater than 500 ppm, indicating Na § desorption. Similar results are obtained for K § Mg 2+ and Ca 2+. As discussed, the increase in concentration of the pore fluid as a function of time is attributed to cation exchange or replacement by A13+, Pb 2+, and Zn 2§ in the top part of the soil column while in the bottom part, Na § and K § are exchanged by Ca 2+ and Mg z+. This is due to the higher valence of Ca 2+ and Mg 2+ compared to Na § and K § and, hence higher replacing power, as shown in Figure 9.6(c).
Figure 9.6(c). Variation of adsorbed metal ion concentration with distance.
METHODS FOR CALCULATING TRANSPORT PARAMETERS
239
The adsorption isotherms of various cations are shown in Figure 9.6(d). Generally, it can be seen that maximum exchange capacity of various cations were obtained at lower pH values due to replacement by heavy metals. For higher pH values, cation exchange decreased.
Figure 9.6(d). Metal ion adsorption isotherm.
9.6.3
Estimation of Transport Parameters
In order to predict the transport and the fate of various pollutant species in the subsurface environment, the transport parameters involved in the goveming set of equations that describe the transport process need to be accurately defined. The laboratory methods which can be used to estimate the transport parameters of chemical species diffusing through waste containment barriers are discussed. In general, methods used to calculate the transport parameters fall into two broad categories-- steady and transient states. This section describes some of the more common procedures which have been used to calculate the transport parameters.
Steady State Methods Decreasing Source Concentration: A schematic diagram illustrating the steady state method is shown in Figure 9.7. The soil is contained between two reservoirs -- a source reservoir containing the solution of interest and a collection reservoir from which samples are drawn for specified chemical analysis. The concentration of the chemical species of interest is higher in the source reservoir than it is in the collection reservoir, hence a concentration gradient is established across the soil sample. Once the steady state condition has been reached, Fick's first law for diffusion can be applied and the diffusion parameter can be calculated as follows:
240
SUBSURFACE POLLUTANT TRANSPORT J=-D
a___c_c
[9.50]
8x The diffusion coefficient can then be calculated (Shackelford, 1991; Yong, Mohamed, and Warkentin, 1992) using: [9.51]
where Js is the mass flux, D is the diffusion coefficient, L and A are the length and cross sectional area of the soil sample respectively, and Ac is the change in mass of the chemical species in an increment of time, At. Since the quantity (L/A Ac) in Eq. [9.51] can be measured or set independently of the test, only the change in mass with respect to time, (Am~At), is measured during the test.
Figure 9.7. Steady state method with decreasing source concentration.
At steady state,
Ami At
-
Am2 At
-
Am At
[9.52]
where Am~ is the decrease in mass of the chemical species in the source reservoir, and Am2 is the increase in mass of the chemical species in the collection reservoir. The use of the difference operator in Eq. [9.52] implies that the concentration gradient across the sample is linear. However,
METHODS FOR CALCULATING TRANSPORT PARAMETERS
241
due to coupled flow processes, the concentration gradient within the soil sample is non-linear. As a result, the calculated diffusion parameter using the external (across the clay) concentration gradient may not be the same as that determined using the internal (within the clay) distribution of concentration. In order to apply this technique, the following conditions have to be satisfied: (1) The concentration of the chemical species of interest should be higher in the source reservoir than in the collection reservoir; (2) The chemical species diffusing from the source reservoir must be continuously replenished while the mass of the chemical species diffusing into the collection reservoir needs to be continuously removed in order to maintain constant concentration difference Ac across the sample; (3) The concentration of the chemical species of interest at steady state in the soil at x = 0 is higher than it is at x = L. In other words, the concentration profile within the soil should be inward concave, as shown in Figure 9.7(c); and (4) The solution pH in the source reservoir as well as in the soil column should be greater than 5. This is due to the fact that when pH < 5, heavy metals interact with clay through exchangeable cations, hence most cations (Na +, K +, Mg 2+, Ca 2+) are desorbed from clay resulting in outward concave profiles in the soil. Time-Lag Method: This method is commonly used to obtain the diffusion coefficient through porous membrane (Jost, 1960; Crank, 1975). A schematic diagram illustrating the time- lag method is shown in Figure 9.8. The soil is contained between two reservoirs -- a source reservoir containing the solution of interest and a collection reservoir. This soil is initially at zero concentration and the concentration at the interface with the collection reservoir is maintained effectively at zero concentration. In this case, the total amount of diffusing substance per crosssectional area, Q,, which has passed through the soil approaches a steady state value as t -~oogiven by (Jost, 1960; Crank, 1975):
[9.53]
where Qt = f0tJ~ 9 Cl is the concentration in the source reservoir, which is maintained at a constant value with time. Eq. [9.53] is the equation of a straight line on a plot of Qt versus t, as shown in Figure 9.8(c). The intercept on the time axis is the time lag, TL, which is given by: L2
TL- 6D
[9.54]
Therefore, D can be calculated using Eq. [9.54] by plotting Qt versus time and determining the value for the intercept TL. The time lag method requires less control of the test conditions than the decreasing source concentration method since steady state condition has to be established, not maintained. Furthermore, the time required to establish steady state condition can be excessive, especially for relative thick samples.
242
SUBSURFACE POLLUTANT TRANSPORT
Figure 9.8. Time lag method
Root Time Method: This method was developed by Mohamed and Yong (1991). The approach is based on an analytical solution of the differential equation of solute transport in clay barrier. The equation is first cast in a non-dimensional form and the Fourier Series is used to solve the differential equation for specified initial an boundary conditions. The concentration, in a nondimensional form, is given by:
c*(~,1:) = (1-~)- 1 exp(_nzz) sin(g~)
[9.55]
and the relative change in concentration is given by: CRC
= exp (-11:21:)
[9.56]
where: Dt
c - c2
x
L 2'
c1 -
L
c2
and 1: is non-dimensional time factor, { is non-dimensional distance, c" is non-dimensional concentration, c is the concentration at specific time and distance, Cl is the concentration at x = 0, c 2 is the concentration at x = L, and L is the length of soil specimen. The theoretical relationship between the non-dimensional relative concentration Cec* and the root time factor, z, is shown in Figure 9.9(c). The theoretical curve is linear up to relative
METHODS FOR CALCULATING TRANSPORT PARAMETERS
243
concentration of 0.2 (80% equilibrium). At a relative concentration of 0.1 (90% equilibrium), the abscissa (AC) is 1.055 times the abscissa (AB). This characteristic is used to determine the point on the experimental curve corresponding to relative concentration of 0.1 (i.e. 90% of the steady state equilibrium time).
Figure 9.9. Root time method
The experimental data reduced in terms of relative change in concentration of specific ion in the collected effluent versus root time generally consists of a short curve representing initial increase in concentration (in the effluent), a linear part and a second curve as shown in Figure 9.9(d). The point (D) shown in Figure 9.9(d) corresponding to the initial condition is obtained by projecting the linear part of the curve until it meets the vertical axis at zero time. A straight line(DE) is then drawn having abscissa 1.055 times the corresponding abscissa on the linear part of the experimental part. The intersection of the line (DE) with the experimental curve locates the point (ago) corresponding to relative concentration of 0.1. The corresponding value t9o can then be obtained. The value of 1: corresponding to CRC*is 0.2436 and the diffusion coefficient, D, is given by: 0.2436 L 2 O
__
%
[9.57]
The method is applicable for the case of adsorption and desorption. It is also applicable for various values of pH in the influent solution (i.e., acid and alkaline conditions). It should be noted that in the case of low pH in the influent, some of the cations are desorbed from the clay, yielding
244
SUBSURFACE POLLUTANT TRANSPORT
c 2 > Cl, hence the experimental relative change in concentration is negative. Transient M e t h o d s
Several different transient methods have been used to calculate transport parameters. The column method is described. The soil column test, traditionally known as leaching column test, has been used to study adsorption and migration of contaminants through clay barriers. First, steady state flow is established through the soil sample by using distilled water in the source reservoir. After steady-state fluid has been established, the fluid in the influent reservoir is changed to a solution with known and constant concentration (Co'S) of particular chemical constituents. The concentration, c e, in the effluent reservoir is measured as a function of time. The data are reduced in the form of breakthrough curves, of relative concentration, C/Co, versus time or pore volumes of flow. Breakthrough curves are modelled using either one of the analytical solutions given by Eqs. [9.26], [9.28],[ 9.31 ], and [9.32]. The diffusion coefficient can then be calculated once, ce, c o, v, L, and t are known.
9.7
S U M M A R Y AND CONCLUDING REMARKS
The manner in which the transport parameters are determined for use in modelling is important. At least five options are available: (1) Using values and relationships reported in various publications for similar situations and circumstances; (2) Using the infinite solution diffusion coefficient as the starting point for the determination of the diffusion-dispersion coefficients and also using adsorption isotherms and adsorption characteristics, with modifications for tortuosity and advective velocity effects in regard to diffusion-dispersion phenomenon; (3) Using experimentally determined values from laboratory experiments with soil from the site and representative leachates; (4) Using information from monitoring wells and chemical analysis from the samples recovered, to back-calculate the transport and adsorption parameters; and (5) A combination of any of the preceding options. Where possible, it should be standard practice to match numerical solutions with established analytical solutions, to have full confidence of the model developed. In the final analysis, a combination of field, laboratory, and analytical studies is the best approach for evaluating pollutant transport in subsurface.
CHAPTER
TEN
RISK A S S E S S M E N T
10.1
INTRODUCTION
Risk may be defined as the probability of injury or harm from exposure to a hazard. Risk assessment is the process of determining the nature and magnitude of adverse effects posed by a given hazard. Hazardous situations to which the risk assessment process may be advantageously applied include contaminant releases into the environment, hazardous waste generation and disposal, emission of air pollutants, and contaminated sites. These hazards, and many others, may pose significant risks to human health as well as to the environment. The objective of risk assessment is essentially to generate the information necessary to make the best possible decision concerning a potentially hazardous situation. Such information will typically permit an evaluation of the acceptability of the hazard, as well as highlight and provide valuable insight into those features and exposure pathways with the greatest risk potential, and upon which the remedial effort should be focussed. In general, risk assessment offers an estimate of the likelihood of occurrence of adverse effects from exposure of humans and ecological receptors to chemical, biological or other hazardous agents present in the environment. It typically requires input from practitioners in the various fields and disciplines related to the hazardous situation under investigation, notably engineering, chemistry, biology, ecology and statistics. A major challenge of the risk assessment process is that the calculated risk estimates are often based on incomplete information and, hence, are characterized by uncertainty. Risk assessment is actually the first of a two-phased approach to handling a hazardous situation. If the baseline risk assessment, which is conducted under the assumption of "no-action" to control or mitigate the contamination or reduce the risk, yields an unacceptable level of risk, a program of action to reduce the risk (e.g., cleanup, regulation, education) is embarked upon in a risk management phase. In effect, the risk assessment phase provides a numerical estimate of the probability of injury or harm from a hazardous situation while the risk management phase combines the risk assessment [the scientific input] with the directives of [the enabling] regulatory legislation, together with socio-economic, technical, political, and other considerations, to reach a decision as to whether or how much to control future exposure of suspected toxic agents [substances] (US EPA, 1986b). Thus, risk management can be construed as the selection of an "acceptable" level of a hazardous substance. Risk assessment is motivated by consideration for human health and/or the ecology. A human health risk assessment is concerned principally with the health risks posed to potentially exposed human populations whereas ecological risk assessment focuses on the adverse effects on the environment or ecosystem. Although both types of assessment involve different processes, they do share a common philosophy and some data requirements. Thus, a carefully planned and coordinated 245
246
RISK ASSESSMENT
data collection phase could provide information necessary to assess both risk types. This chapter will be devoted to the human health risk assessment process.
10.2
BASIC ELEMENTS OF HUMAN H E A L T H RISK ASSESSMENT
A human health risk assessment characterizes the adverse health effects due to human exposure to hazardous substances. Estimates of the extent to which humans have been, or could be, exposed to the hazard and the toxicity of the substance are utilized in the risk assessment process to determine the present or future health risks to the population. Human health risk assessment can be performed for various environmental hazards, such as chemical contaminants, ultraviolet radiation, and electromagnetic field. One of the more common applications is the site risk assessment, which is the determination of the risks to human health from exposure to hazardous substances at a contaminated site. Although specific cases of risk assessment may differ considerably in scope, they do share a common philosophy and methodology. In general, a human health risk assessment consists of the following basic steps: (1)
Hazard Identification This involves the identification of the nature and extent of contamination, and a determination whether or not exposure to the contaminant can potentially increase the incidence of an adverse health effect;
(2)
Exposure Assessment Identifies the pathways or routes through which the contaminants of concern might reach humans or ecological receptors, and quantifies the frequency, magnitude and duration of such exposures;
(3)
Toxicity Assessment Evaluates the potential adverse effects of each contaminant of concern, and characterizes the relationship between the amount of exposure (dose) and the magnitude of the adverse health effect (response) in humans;
(4)
Risk Characterization Characterizes the nature and magnitude of the risk posed by integrating the findings of the exposure and toxicity assessments; also evaluates the attendant uncertainty, so that proper judgement can be made regarding the acceptability, or otherwise, of the predicted risk estimates.
These elements of the human risk assessment process are outlined in Figure 10.1, and discussed in more detail in the rest of this chapter. Additional and complementary information can be found elsewhere in the literature (e.g., Asante-Duah, 1993, 1996; US EPA, 1989).
HAZARD IDENTIFICATION
247
Data C o l l e c t i o n and E v a l u a t i o n Ill Gather and analyse relevant site data a Identify potential chemicals of concern
Toxicity Assessment
Exposure Assessment 0 Analyse contaminant release [] Identify exposed populations g Identify potential exposure pathways [] Estimate exposure concentrations for pathways 0 Estimate contaminant intakes for pathways
! I Collect qualitative and
quantitative toxicity informatior
IN Determine appropriate toxicity values
Risk C h a r a c t e r i z a t i o n IN Characterize potential for adverse health effects to occur 9 Estimate cancer risks 9 Estimate nonc~ncer hazard
quotients
11 Evaluate uncertainty i
Figure 10.1. Components of human health risk assessment (US EPA, 1989).
10.3
HAZARD IDENTIFICATION
The hazard identification step of the health risk assessment process entails the identification of the contaminants present at a site, and a determination whether or not exposure to the contaminants would produce adverse health effects in humans. A review is conducted of all available site environmental monitoring data for the purpose of identifying potential chemicals or contaminants of concern upon which the assessment will focus. The selection of chemicals with potential to impair human health, for consideration in the risk assessment process, may be guided by such factors as mobility, persistence, bioaccumulation, concentration and toxicity, and fate and transport properties. Historical information reliably associating certain chemicals with the site is quite pertinent.
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Generally, such chemicals should be selected for inclusion in the risk assessment process, even if other considerations, such as noted above, suggest otherwise. In addition to the identification and characterization of potentially hazardous contaminants, a determination of their extent in their primary sources and in each of the environmental media (e.g., air, soil, sediments, surface and groundwater) is an essential part of the process. Also of interest are the characteristics of (1) the sources of the contaminants, particularly information related to their release potential, and (2) the environmental setting, notably its possible impact on the fate, transport and persistence of the contaminants. Details regarding data collection and evaluation, i.e., the gathering and analysing of relevant site data and identification of chemicals with potential adverse health effects on exposed populations, can be found in the literature (e.g., US EPA, 1989). If a contaminant is judged to be potentially harmful to humans, it is further investigated in the exposure and toxicity phases of the risk assessment process.
10.4
EXPOSURE ASSESSMENT
Exposure assessment may be defined as the contact of a chemical or physical agent (i.e., contaminant) with humans or ecological receptors. Exposure assessment is the (qualitative or quantitative) determination of the magnitude, frequency and duration of exposure, the nature and size of population potentially at risk and the pathways by which the contaminants may reach the risk population. A major aspect of the exposure assessment process involves the identification of the exposure scenarios, and will typically include a determination of: (1) Environmental transport media (e.g., soil, sediment, groundwater, surface water, air); (2) Contaminant sources, fate and transport, including intermedia transfers (e.g., leaching of contaminants from soil into groundwater, production of fugitive dust during excavation); (3) Exposure points (e.g., on-site, off-site, sediments, groundwater, surface water); Exposed pathways (e.g., dermal absorption, ingestion, inhalation); (4) (5) Exposed population (e.g., sensitive sub-populations such as children, the elderly and pregnant women; population size); and (6) Exposed durations (e.g., acute or short term, chronic or long term). The exposure assessment per se begins after relevant data on the site contaminants have been gathered and evaluated, and those contaminants of potential concem, which be the focus of the risk assessment, have been selected. The exposure assessment process consists of three major steps (US EPA, 1989): (1) Characterization of exposure setting; (2) Identification of exposure pathways; and (3) Quantification of exposure; Step 1: Characterization of exposure setting This step involves the identification of (1) the physical characteristics of the site, and (2) the population potentially at risk, and subgroups with increased risk potential. Important physical
EXPOSURE ASSESSMENT
249
characteristics of the site include: Climate (e.g., temperature, precipitation); Meteorology (e.g., wind speed and direction); Geology (e.g., location and characteristics of underlying strata); Vegetation (e.g., unvegetated, forested, grassy); Soil type (e.g., sandy, organic, acid, basic); Groundwater hydrology (e.g., depth, direction and type of flow); and Surface water location and description (e.g., type, flow rates, salinity). Common sources of this type of information include county soil surveys, wetland maps, aerial photographs, and study reports (e.g., preliminary assessment (PA), site investigation (SI), and remedial investigation (RI)). It may also be useful to consult technical experts, such as hydrogeologists. The populations potentially at risk may be characterized by factors that influence exposure, notably, proximity to the site, behaviour and activity patterns, and by the presence of sensitive subpopulations. In general, the following populations will be potentially at risk: People residing on or near the site and, therefore, have the greatest risk potential; Distant and local consumers of water supply, sea food (e.g., fish) or agricultural products (e.g., fruits and vegetables) originating from the contaminated site region; People who may in the future be exposed to chemicals that have migrated from the site; People whose behaviour patterns or activities put them at increased risk, e.g., children due to increased likelihood of soil contamination on the playground, construction and other out-door workers, individuals exposed chemicals during occupational activities, heavy consumers of home-grown vegetables and locally caught fish; and People with increased sensitivity to chemical exposures, e.g., infants and children, pregnant and nursing women, the elderly, and those with chronic illnesses. These increased risk groups will typically be found in schools, day care centres, hospitals, retirement centres and communities, nursing homes, construction sites, near-by fisheries and vegetable gardens.
Step 2" Exposure pathway identification Exposure pathways relate to the routes that a contaminant follows, or would follow, to reach the population at risk. Exposure pathways are identified and characterized by integrating available information on the sources, locations, mechanism(s) and type of contaminant releases, and the locations and activity pattems of populations potentially at risk. An exposure pathway typically involves the following four elements: (1) source and mechanism of contaminant release, (2) medium for contaminant retention, transport or migration, (3) exposure point, i.e., location of potential contact between contaminant and humans, and (4) exposure route, i.e., the mechanism by which a contaminant comes into contact with humans (e.g., ingestion, inhalation, dermal contact).
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Table 10.1: Common chemical release sources (US EPA, 1989) Receiving Medium
Release Mechanism
Air
II Volatilization
Release source
O Fugitive dust generation
or. Surface wastes-- lagoons, ponds, pits, spills o:. Contaminated surface water o:~Contaminated surface soil or~Contaminated wetlands ~ Leaking drums o:. Contaminated surface soil ~ Waste piles
Surface water
O Surface runoff tl Episodic overland flow O Groundwater seepage
o~- Contaminated surface soil or. Lagoon overflow ~. 9 Spills, leaking containers :~ 9 Contaminated groundwater
Groundwater
tl Leaching
.t. Surface or buried wastes ol. Contaminated soil
Soil
0 Leaching II Surface runoff 0 Episodic overland flow 0 Fugitive dust generation/deposition n Tracking
:. 9 Surface or buried wastes :. 9 Contaminated surface soil ~ Lagoon overflow ~9 Spills, leaking containers ~ Contaminated surface soil ~ Waste piles ~ Contaminated surface soil
Sediment
El Surface runoff
.t. Surface wastes-- lagoons, ponds, pits, spills or. Contaminated groundwater .t. Surface or buried wastes ~176 Contaminated soil
O Groundwater seepage O Leaching
Biota
(1) (2) (3)
n Uptake (direct contact, ingestion, inhalation)
t. 9 Contaminated soil, surface water, :.9 Sediment, groundwater or air
An exposure pathway identification process embodies the following steps (US EPA, 1989): Identification of contaminant sources and the receiving media; Evaluation of contaminant fate and transport; Identification of exposure points and exposure routes; and
EXPOSURE ASSESSMENT (4)
251
Identification of complete exposure pathways.
Identification of contaminant sources and the receiving media The sources of contaminants at a site, the release mechanism(s), and the receiving media may be determined from information in reports (e.g., PA, SI, RI), monitoring data, and information on source locations. Some information may be obtained deductively. For example, the presence of contaminated soil in the vicinity of a tank would identify the tanks as a likely source of the contamination, leakage or rupture as the release mechanism, and the ground as the receiving medium. A listing of common contaminant sources, release mechanisms and receiving media is shown in Table 10.1. Evaluation of contaminant fate and transport The fate and transport of contaminants are evaluated for the purpose of predicting future exposures, and identifying likely sources of currently polluted media. This step is essentially qualitative, and is aimed at identifying current and/or future receiving media of the site contaminants. In making this determination, it is important to note that a chemical, after release into the environment, may be (1) Transported (e.g., by convection in water or on suspended sediment, or through the atmosphere); (2) Transformed (a) biologically (e.g., biodegradation); (b) chemically (e.g., photolysis, hydrolysis, oxidation, reduction); (c) physically (e.g., precipitation, volatilization); (3) Retained or accumulated (e.g., at the release source, in the receiving media); The physical/chemical and environmental fate parameters of potentially hazardous chemicals may be used to predict their fate at a given site. Some commonly used fate parameters include organic carbon partition coefficient (Koc), distribution coefficient (Kd), octanol/water partition coefficient (Kow), solubility, Henry's law constant, vapour pressure, diffusivity, bioconcentration factor (BCF), and half-life, as discussed in Chapter 6. Site-specific characteristics that may influence a chemical's fate and transport include soil moisture content, organic carbon content, and carbon exchange capacity. The water table location is also an important characteristic in as much as a high water table increases the likelihood of leaching of chemicals in soil into the groundwater. Using chemical, site-specific and monitoring data, contaminant fate and transport within and between each medium may be evaluated. Thus, contaminated and/or potentially contaminated areas are identified.
Identification of exposure points and exposure routes Exposure points, where the identified contaminants may potentially come in contact with humans, are determined by considering population locations and activity patterns in the vicinity of the contaminated or potentially contaminated areas. Contaminated media and sources are obvious potential on-site exposure points, particularly if the site has current or future use, or if access to the site is not restricted or limited, for example, by distance. Potential off-site exposure points are generally down gradient or downwind of the contaminated site. Exposure routes, through which the identified contaminants can enter the human body, depend on the exposure media (e.g., groundwater, surface water, sediment, soil/dust, air, and food)
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and the activity patterns at the exposure points. The major exposure routes of interest are (1) ingestion, (2) inhalation, and (3) dermal contact. Ingestion is the route involving eating and/or drinking. Contaminants present in food or drinks are carried directly into the digestive system, and from there to target areas or organs in the body where they could cause adverse health effects. The solubility of ingested contaminants is an important factor that influences the contaminant's transport and metabolism in the human body. Lipophilic (i.e., fat-soluble) chemicals, such as benzene or DDT, are rapidly absorbed into the body on ingestion. Hydrophilic (i.e., water-soluble) chemicals may be absorbed throughout the body since the chemistry of human metabolism is water-based. Consumption of fish polluted with methyl alcohol, a chemical known to cause damage to the central nervous system and even death, is a classic example of this exposure route. Inhalation involves absorption of airborne chemicals during breathing. On inhaling, the chemical in question passes through the lungs to the blood stream, and from there to the brain and the rest of the body. An important factor that influences the severity of the resulting adverse health effect is the solubility of the contaminant in the blood. A common example of this route of exposure is the inhalation of carbon dioxide from, for example, smoking, automobile exhaust, and industrial processes. Carbon monoxide has a much stronger binding affinity for hemoglobin than does oxygen. Consequently, it can irreversibly displace oxygen from hemoglobin (the iron-based organic compound by means of which oxygen is transported through the blood), and curtail the amount of oxygen in the blood, leading to oxygen starvation of cells. The inhalation of airborne particles, possibly carrying toxic chemicals, is very much dependent on the particle size. The smaller-sized particles are more likely to penetrate deeper and, hence, potentially cause more harm. Dermal or skin absorption involves contact of the skin with contaminants, and is enhanced by breaks (e.g., cuts, scratches, abrasions) in the skin, increased concentrations of contaminants, and decreased particle size. The fat or oil solubility of the contaminant is also an important absorption enhancement factor. Oil or solvent-based chemicals, being fat or oil soluble, are readily absorbed into the lipophilic layers of the skin. It is for this reason that gasoline, for example, would remain on and be detectable on the skin for a long period, even after repeated soap and water washing. It should be noted that the existence of an exposure point is not necessarily indicative of the presence of an exposure route. For example, if contaminated soil (the exposure point) is touched while wearing gloves, the potential exposure route (dermal contact) is blocked.
Identification of complete exposure pathways
(1) (2) (3) (4)
An exposure pathway is considered complete if the following elements are present: Contaminant sources; Mechanism of contaminant release into the environment; Exposure point where contact can occur; and Exposure route by which contact can occur.
An exposure pathway will be incomplete if any of the elements above is absent. Contact with contaminated media while in protective clothing is an example of an incomplete (dermal contact) exposure pathway. Another example involves a situation in which a contaminant is released into the air in an unpopulated area. Complete pathways are identified by integrating information on contaminant sources, release mechanisms, fate and transport, exposure points, and exposure routes. Human monitoring data, if available, may be used to support conclusions on the completeness of an
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253
exposure pathway, particularly if the data indicates chemical accumulation or related effects. However, data lacking confirmatory evidence, i.e., negative data, does not imply that a pathway is incomplete.
Step 3: Quantification of Exposure A determination of the magnitude, frequency, and duration of human exposure to a contaminant may be made based on the identified complete pathways and information on the populations potentially at risk. Generally, this is accomplished in two stages: (1) estimation of exposure concentration, and (2) calculation of chemical intakes.
Estimation of exposure concentration Exposure concentrations may be estimated by using monitoring data alone or in combination with environmental fate and transport models. Use of monitoring data alone is adequate in cases where there is direct contact with a chemical in a contaminated medium, or where monitoring is/was conducted directly at an exposure point, such as a drinking water well. Monitoring data yield good estimates of current exposure concentrations. Use of monitoring data alone may not be adequate in a number of instances, namely (1) exposure points are remote from the contaminant sources, and there are mechanisms for the release and transport of contaminants to the exposure points (e.g., air dispersion, surface water transport), (2) in the absence of temporal distribution of data, with which long term future exposure concentrations may be predicted, and (3) no reading is registered because the concentrations of the contaminants are below the monitoring equipment's detection limit. In these instances, monitoring data are used in combination with contaminant fate and transport models to predict exposure concentrations. A variety of models for estimating exposure concentrations in various environmental media (e.g., groundwater, surface water, air, soil, sediment) are available and well documented in the literature (e.g., US EPA, 1988; US EPA, 1989b). A useful discussion of important elements of exposure concentration determination in various environmental media is given in US EPA (1989). Calculation of chemical intakes Chemical intake is defined as the amount a chemical in contact with the receptor's exchange boundary (e.g., skin, lungs), and that is available for absorption. It is different from the absorbed dose, which is the amount of chemical absorbed into the blood stream. Chemical intake, hence dose, is clearly dependent on the exposure point concentrations in the relevant media. The dose may be obtained from the intake the application of an absorption factor-- a function of physiological parameters such the gastrointestinal absorption rates. If the absorption is unknown or cannot be reliably estimated, a 100% absorption can be assumed, resulting in the equality of the dose and the chemical intake. This is obviously a conservative estimate of the actual dose. It is also often assumed, for sake of simplicity, that the potential receptor remains at the same location and is exposed to the same ambient concentration. This assumption yields a conservative estimate since, in general, some time will be spent away from the exposure point(s), and lower or near-zero chemical exposures will be endured.
254
RISK ASSESSMENT A general equation for calculating chemical intakes is given as:
I = C x
CR x EFD 1 x BW AT
[10.1]
where I is the intake; the amount of chemical at the exchange boundary (mg/kg body weight-day), C is the concentration of pollutant; the average concentration contacted over the exposure period (mg/litre of water), CR is the contact rate; the amount of contaminated medium contacted per unit time or event (litres/day), EFD is the exposure frequency and duration; describes how long and how often exposure occurs. Often calculated using two terms (EF and ED), EF is the exposure frequency (days/year), ED is the exposure duration (years), BW is the body weight; the average body weight over the exposure period (kg), and A T is the average time; period over which exposure is averaged (days). Three categories of variables can be identified in the chemical intake equation: (1) chemical-related, i.e., exposure concentration (C), (2) population-related, i.e., contact rate (CR), exposure frequency (EF) and duration (ED), and body weight (BW), and (3) assessment process-based, i.e., averaging time (AT). Specialization of the general intake equation (Eq. [ 10.1 ]) to the various exposure media and exposure routes can be found in US EPA (1989) and Asante-Duah (1996).
10.5
TOXICITY ASSESSMENT
10.5.1 Introduction
Toxicity assessment evaluates the potential for selected contaminants to cause adverse health effects in exposed individuals. Toxicity assessment generally involves two steps: (1) hazard identification, and (2) dose-response assessment. Hazard assessment is aimed at determining whether exposure to contaminants can cause an increase in the incidence of an adverse health effect (e.g., cancer, birth defects, mental retardation). The nature and strength of the evidence of causation are also characterized in the hazard assessment step. The second step, dose-response assessment, evaluates (quantitatively) the toxicity information and characterizes the relationship between the dose of contaminant administered or received and the incidence of adverse health effects in the exposed population. From this quantitative dose-response relationship, the potency of the contaminant is estimated. Specifically, toxicity values, which can be used to estimate the incidence or potential for adverse health effects in terms of human exposure to the contaminant, are derived. Toxicity values (e.g., reference doses and slope factors) are used in the risk characterization step to estimate, at varying human exposure levels, the likelihood of occurrence of adverse effects. A schematic representation of typical dose-response curves is given in Figure 10.2. In general, response increases with dose. The smaller the dose required to cause a given effect, the more potent (toxic) is the substance deemed to be. The dose-response concept, upon which the toxicity assessments are based, assumes that noncarcinogenic effects have a threshold dose below which no observable adverse effect occurs while carcinogenic effects are nonthreshoM substances, i.e., produce a response at even the smallest dose, as illustrated in Figure 10.2.
TOXICITY ASSESSMENT
255
Toxicity assessments usually employ one of two approaches depending on whether carcinogenic or noncarcinogenic effects are involved.
(b 0 Q.
Nonthreshold (carcinogenic) /
o~ (b
~
(carcinogenic) r
Dose(Exposure) Figure 10.2. A schematic representation of typical dose-response relationships.
10.5.2 Sources of Toxicity Information
Toxicological information, for use in assessing the potential for a substance to cause adverse health effects (carcinogenic and noncarcinogenic) in humans, are obtained primarily from epidemiological (human) studies and experimental animal studies. Secondary or supporting sources of such information include in vitro tests and comparison of structure-activity relationships. Epidemiological studies Epidemiological data indicating a strong clear-cut relationship between a chemical and an adverse health effect are considered to be the most convincing evidence of the risk potential of the chemical. Such positive data are, however, not often available due to a variety of factors. For example, since humans are normally exposed to toxic chemicals inadvertently or by accident, data may not have been generated for the specific chemical of concern. Where data is available, some important characteristics of the exposure (e.g., concentration) may be lacking or poorly defined. Also, the exposed population tend to be quite limited (in size) and rather varied (e.g., by age, sex, activity patterns, genetics, and other factors that may influence susceptibility). Thus, available human exposure-response data may not be sufficient for a quantitative assessment. In these situations, the limited data is better used in a supporting role, to support conclusions drawn from animal studies. Clearly, if adequate high quality data are available, they should be given priority over data obtained from animal or any other studies, since our primary interest is with humans. In this instance, animal studies data may play a supporting role.
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Animal studies Because of the inherent difficulties in obtaining adequate data on human dose-response to all chemicals of interest, as discussed above, toxicity information gathered in animal studies are commonly used to infer the potential of a chemical to cause an adverse effect in humans. It is assumed that humans and animals are similar in their response to chemicals, i.e., if a chemical causes an adverse effect in an animal so will it in humans. However, extrapolating animal data to humans is complicated by a variety of factors that differ among species and may potentially affect the response to the chemical of interest. These factors include sex, strains, body size, lifespan, genetic homogeneity, metabolism and excretion patterns. In general, the likelihood that a chemical will have an adverse effect in humans is considered to increase as similar results are observed across sexes, strains, and routes of exposure in animal studies. Commonly used animals are rat, mouse, rabbit, guinea pig, dog and monkey. Animal species that are most similar to humans in terms of physiology, metabolism and pharmacokinetics is normally given priority. A common approach for making interspecies comparisons is to use standardized scaling factors, such as mg/kg body weight, ppm in the diet or water, mg/m 2 body surface area per day, and mg/kg body weight per lifetime. Supporting studies There are a number of other studies that may be used to assess a chemical's potential health effect on humans. These studies, which are used primarily to support conclusions drawn from epidemiological and animal studies, include (1) metabolic and other pharmacokinetic studies, (2) cell cultures or microorganism studies, and (3) structure-activity studies. Metabolic studies seek evidence of a chemical's toxicity potential by comparing the metabolism of a chemical which exhibits toxic effects in animals with the corresponding metabolism in humans. Cell cultures or microorganism studies investigate a chemical's potential for biological activity, hence potential for and possible mechanisms of carcinogenicty, by testing for point mutations, numerical and structural chromosome aberrations, DNA damage/repair, and cell transformation. Structure-activity studies estimate the toxicologic activity of a chemical from known toxicologic activities of structurally related chemicals.
10.5.3 Toxicological Parameters Human health risk assessment usually evaluates the toxic effect of chemicals differently depending on whether the chemical is a carcinogen (cancer-causing) or a noncarcinogen.
Noncarcinogenic Effects Noncarcinogenic chemicals give rise to toxic endpoints known as noncancer or systemic toxicity, i.e., which do not include cancer or gene mutations. It should be noted that carcinogens also commonly exhibit noncancer effects, i.e., systemic toxicity. In general, chemicals with noncancer toxicity have a different effect on different organs of the body and, hence, are often referred to as systemic toxicants. Factors which determine which organs are most impacted include dose, route of exposure, oil or fat solubility, and the chemical's effect on enzyme activity. Common systemic toxicants and their target organs include (Clayton and Clayton, 1986; Griffin, 1986):
TOXICITY ASSESSMENT (1) (2) (3) (4)
(5)
257
Hepatotoxic chemicals, which mostly affect the liver (e.g., carbon tetrachloride, tetrachloroethane); Nephrotoxic chemicals, which affect the kidneys (e.g., halogenated hydrocarbons); Hematopoietic toxins, which affect the blood or blood cells (e.g., aromatic compounds such as benzene, phenols, aniline, nitrobenzene and toluidine); Neurotoxic substances, which affect the nerve system (e.g., methyl alcohol, carbon monoxide, heavy metals and organometallics); and Anesthetics or narcotic chemicals, which affect consciouness (e.g., acetylene hydrocarbons, olefins, aliphatic ketones, aliphatic alcohols, esthers, paraffin hydrocarbons, ethy and isopropy ether).
Noncarcinogenic effects are often characterized by a threshold dose below which no effect occurs. The concept is similar to that of a drug which yields no beneficial effect if administered in too small a dose. The dose at which no effects are observed in humans or experimental animals is known as the No Observed Effect Level (NOEL). In situations where data that definitively identifies a NOEL are not available, a safe threshold level may be determined based on the Lowest Observed Effect Level (LOEL). NOEL is the exposure level at which no effect at all has been observed; usually, effects, albeit of no toxicological importance, are observed. The No Observed Adverse Effect Level (NOAEL) is used to characterize situations involving doses that do not produce an adverse effect. It is the highest dose at which no significant adverse effect is observed. Where NOAEL has not been specifically identified, the Lowest Observed Adverse Effect Level may be used instead. The reference dose (RfD) is the toxicity value most commonly used to evaluate noncarcinogenic effects. The RfD is an estimate of the noncarcinogenic chemical which produces no appreciable risk of adverse effects in the general human population, including any sensitive subgroups. It is defined as the maximum amount of a toxic substance that an individual can absorb with no adverse health effects. A commonly employed unit of measure is mg/kg body weight/day. Various types of RfD are available depending on (1) exposure duration (chronic, subchronic or single event), (2) exposure route (oral or inhalation), and (3) critical effect (developmental or other). Chronic RfD relates to long term exposures, and is defined as an estimate of the daily exposure level that will not produce any appreciable risk of deleterious effects during the lifetimes of the exposed populations, including sensitive subpopulations. Exposure periods between 7 years (approximately 10% of a lifetime) and a lifetime are typically considered in this class. Subchronic RfD is used for estimating noncarcinogenic effects associated with short term exposures, spanning approximately between 2 weeks to 7 years. Developmental RfD is used to estimate potential adverse effects on a developing organism as a result of a single exposure event. Additional information relating to the estimation of the various RfDs can be found in US EPA (1989). RfDs are generally considered to have an uncertainty as much as an order of magnitude or more. A RfD is typically derived from the experimental dose value (e.g., NOAEL) by an application of uncertainty (i.e., safety) factors. These factors relate to uncertainty in such areas as (1) variations in the general population, particularly sensitive subpopulations, and (2) interspecies variability between humans and animals (e.g., extrapolating from animals to humans). Details on the application of uncertainty factors can be found in Asante-Duah (1996) and US EPA (1989).
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Table 10.2: weight-of-evidence classification system for carcinogenicity Group
Description
A B1 or B2
Human carcinogen Probable human carcinogen B 1 indicates that limited human data are available B2 indicates sufficient evidence in animals and inadequate or no evidence in humans Possible human carcinogen Not classifiable as to human carcinogenicity Evidence of noncarcinogenicity for humans
Carcinogenic Effects Carcinogenic effects, unlike noncarcinogenic effects, are often based on the nonthreshold model, which assumes that even one molecule of a cancer-causing agent can lead to the disease. This is the so-called "one-hit" model. The toxicity value most often used to quantify carcinogenic effects is the slope factor and the associated weight-of-evidence classification. Weight-of-evidence classification The chemical contaminant is assigned a class that reflects the strength of the evidence pointing to a carcinogenic effect in humans. Based on (a) human studies, and (b) animal studies, the evidence is considered (1) sufficient, (2) limited, (3) inadequate, (4) no data, or (5) evidence of no effect. Depending on the extent to which the chemical has been shown to have carcinogenic effects in humans and/or experimental animals, a weight-of-evidence is assigned according to the classification system depicted in Table 10.2. The assigned class may be revised upward or downward based on any available supporting studies (e.g., structure-activity). Slope factor calculation The slope factor (SF) may be defined in several ways. It is essentially a measure of the (potential) carcinogenic effect of a chemical. It may be considered an upper-bound estimate of the probability of developing cancer from a unit intake of a potential carcinogen over a lifetime. It is the risk of cancer (i.e., proportion of exposed population that develop cancer) per unit dose. It is a toxicity value that quantitatively defines the relationship between dose and response, and is usually, but not always, regarded as the 95% confidence limit of the slope of the dose-response curve, with units of (mg/kg-day) ~. Toxicity values for carcinogenic effects may also be expressed in terms of risk per unit concentration. These measures, known as unit risks, may be defined for inhalation as well as for oral exposure routes, and are related to the slope factor as follows: Unit risk for air (inhalation) - slope factor x inhalation rate / 70 kg
[10.2]
RISK CHARACTERIZATION Unit risk for water (oral) = slope factor x consumption rate / 70 kg
259 [10.3]
Typical values commonly used are: inhalation rate = 20 m3/day, and consumption rate = 2 litres/day. Unit risk for air and drinking water are usually expressed in units of micrograms per cubic metre (gg/m 3) and micrograms per litre (gg/l), respectively. 10.5.4 Exposure Route Considerations
Toxicity values (e.g., RfD, SF) are, in general, dependent on the exposure routte (oral or inhalation). In the absence of RfDs or SFs for the dermal route of exposure, it may, in some instances, be appropriate to use oral RfD and oral SF to estimate noncarcinogenic and carcinogenic effects, respectively, associated with dermal exposures. A pre-condition is that the adverse effect does not occur directly at the point of application; skin carcinogens, for example, should not be evaluated using the oral slope factor. It is worth noting that direct use of oral SF or oral RfD for a dermal exposure route is not correct for differences in absorption and metabolism between the dermal and oral routes. Thus absorption factors, typically 10% and 1% for organic and inorganic chemicals respectively, are usually applied. Depending on the specific chemical involved, the oral SF or RfD for dermal exposure may over- or underestimate the risk of adverse health effect, thus increasing the uncertainty of the calculated risks. However, this is not expected to significantly underestimate the total risk, considering the other routes of exposure embodied in the risk assessment process. Oral SF or RfD may also be used to characterize inhalation exposure routes where inhalation SF or RfD is not available. Similarly, inhalation SF or RfD may be used as a surrogate for both ingestion and dermal exposures in cases where oral SF or RfD is lacking.
10.6
RISK CHARACTERIZATION
Risk assessment is the final step in the risk assessment process. It integrates exposure and toxicity assessments to provide qualitative and quantitative estimates of risk to exposed or potentially exposed population. The risks to potentially exposed populations are characterized by noncarcinogenic hazard quotients and hazard indices and/or carcinogenic risks. As the final step in the risk assessment process and the first input in the risk management process, risk characterization serves as a bridge between the two elements. It provides risk information to the risk manager for integration with other nontechnical factors (e.g., economics, regulation, technical feasibility) to arrive at an appropriate course of action for mitigating the estimated risk. The first step in the risk characterization process is to gather and organize available exposure and toxicity information. Then a check is conducted to verify the consistency and validity of the key assumptions common to the both the exposure and toxicity outputs for each contaminant and exposure pathway of concern. The major assumptions relate to the averaging period for exposure, the exposure route, and the absorption adjustments. The goal of the verification phase is to ensure that the exposure estimates are in concert with the assumptions used in developing the toxicity values. For example, absorption adjustment verifies that the exposure estimates and the toxicity
260
RISK ASSESSMENT
values are either both expressed as absorbed doses or as intakes (i.e., administered doses). Following the verification step, risk and hazard quotients for carcinogenic and noncarcinogenic effects are calculated for each pathway analysed. 10.6.1 Calculation of Carcinogenic Risks The risk of carcinogenic effects is defined as the incremental probability of an individual developing cancer over a lifetime from exposure to a cancer-causing agent (carcinogen). Carcinogenic risk can be estimated by integrating data on the potency and the level of exposure to the carcinogen. More specifically, the risks are calculated by multiplying the route-specific cancer slope factor by the estimated intakes, to obtain the excess or incremental individual lifetime cancer risk. Typically, carcinogenic effects are calculated using the linear low-dose and one-hit models, defined by (US EPA, 1989b; Asante-Duah, 1996).
Linear low-dose model CR = CDI x SF
[10.4]
CR = 1 - exp (- CDI• SF)
[10.5]
One-hit model
where CR is the probability of an individual developing cancer (nondimensional), CDI is the chronic daily intake for long term exposure (i.e., averaged over a 70-year lifetime) (mg/kg-day), SF is the slope factor ([mg/kg-day]~). The linear-dose model is generally valid only at low risk levels (i.e., less than 0.01). Where chemical intakes are high, and potential risks are estimated to be greater than 0.01, the one-hit model is considered more appropriate. Computation of cumulative health risks from exposure to chemical mixtures or multiple carcinogens is usually based on the assumption of additivity of the carcinogenic effects. The aggregate cancer risk for all contaminants and relevant exposure routes/pathways may be estimated using the following equations:
For the linear low-dose model at low risk levels: Total cancer risk= ~
~
j=l
[ CDIj• SFj]
[10.6]
i=1
For the one-hit model used at high carcinogenic risk levels:
o,ol co,, r ri ,,
Elexp j=l
i=1
[10.7]
RISK CHARACTERIZATION
261
where CDI0 is the chronic daily intake for the ith contaminant andj ~hpathway, SF0 is the slope factor for the ith contaminant andjth pathway/exposure route, n is the total number of carcinogens, and p is the total number of pathways or exposure routes. Incremental risks ranging between 10.4 and 10.7 are normally considered acceptable. However, since populations may be exposed to additional contaminants from other sources not accounted for in the study, cancer risks well below 10.6 should be targeted, to allow for a reasonable margin of safety. Cancer risk is often expressed in terms of an individual risk and/or population risk. The latter is often taken to be the product of the individual risk and the size of the potentially exposed population. 10.6.2 Calculation of Noncarcinogenic Hazards Noncarcinogenic effects are normally expressed by the hazard quotient (HQ) and/or the hazard index (H1). HQ is defined as the ratio of a single chemical exposure level, over a specified period of time, to the reference dose for the chemical obtained from a similar exposure duration (US EPA, 1989):
HQ = E / RfD
[10.8]
where HQ is the hazard quotient, E is the chemical exposure, and RfD is the reference dose (mg/kg-day). The aggregate noncancer risk due to multiple contaminants and exposure routes/pathways may be estimated by the hazard index ((H/) as:
Total hazard index= ~ ~ E~ y:l ,:l RiD0 j=l
[10.9]
i=1
where E 0 is the exposure level (or intake) for the i th contaminant and fh pathway, RfD0 is the acceptable intake level (or reference dose) for the ith contaminant andfh pathway/exposure rome, n is the total number of chemicals presenting noncarcinogenic effects, and p is the total number of pathways or exposure routes. HI refers to the sum of more than one hazard quotient for multiple hazard chemicals and/or exposures. Certain assumptions are inherent in the hazard index computation specified in Eq. [ 10.9], namely (1) a receptor will experience the same reasonable maximum exposure from each of the multiple pathways, and (2) similar toxicological endpoints are involved. The benchmark hazard index is unity (1). The likelihood of an adverse effect increases as HI increases. Below the benchmark value of 1, adverse effects are deemed unlikely; above 1, adverse health effects are considered very likely to occur. Since humans may be exposed to contaminants from sources unrelated to those accounted for in the risk assessment process, it is recommended that HI well below 1 be considered the safe level, thus extending the margin of safety. The hazard index is usually calculated separately for chronic (long term) and subchronic
262
RISK ASSESSMENT
(short term) exposure periods, by a simple modification ofEq. [ 10.9]. The total chronic hazard index and the total subchronic hazard index are obtained by replacing the exposure level (Ev) by the chronic daily intake (CDlv) and the subchronic daily intake (SDlv), respectively. Finally, it should be remembered that carcinogenic substances may also induce noncarcinogenic effects. Such effects should be included in the hazard index calculation.
10.7
S U M M A R Y AND CONCLUDING R E M A R K S
The overall objective of the risk assessment process is the protection of human health and the environment. Risk assessment provides the necessary information upon which an effective corrective action and management program can be based. A human health risk assessment consists of the following major steps (1) hazard identification, (2) exposure assessment, (3) toxicity assessment, and (4) risk characterization, including attendant uncertainty. The following tasks will typically be involved in the risk assessment effort: (1) (2) (3) (4)
(5) (6)
Determine and characterize the contaminants present at the site; Determine the chemicals of potential concern and their toxicological profiles; Identify all probable contaminant migration pathways, those of potential concern; Identify populations potentially at risk, especially sensitive subgroups such as children, the elderly, and pregnant women; Develop exposure scenarios, by integrating information on the populations potentially at risk with the probable migration/exposure pathways, and Calculate carcinogenic risks and noncarcinogenic hazard indices for each population deemed to be potentially at risk.
Ecological risk assessments follow the same general approach employed in assessing noncarcinogenic health effects in humans. The objective here is to identify habitats, plants and/or wildlife that may potentially be exposed to harmful chemicals, select target species, identify probable pathways and routes of exposure, predict likely exposure doses, and characterize/estimate the potential risk.
CHAPTER
ELEVEN
RISK M A N A G E M E N T
11.1
INTRODUCTION
In Chapter 2, we discussed the concept of sustainable development and its relation to environmental protection. It will be recalled that the 1992 UN Conference on Environment and Development (UNCED) Principle # 4 characterized sustainable development as: "In order to achieve sustainable development, environmental protection shall constitute an integral part o f the development process and cannot be considered in isolation from it."
Therefore, in developing risk management strategies we should be aiming for environmental protection. This can only achieved when the developed risk management strategies are ecologically viable, economically feasible, and socially desirable. In doing so, the solution will be environmentally sound and politically acceptable. In many aspects the foundation underlying the need to establish risk management strategies are the same as those behind the existing spectrutn of regulations and standards governing pollutants in the environment, notably drinking water standards, ambient air and water quality criteria, air emissions from incinerators, discharges to surface waters from wastewater treatment plants and land spreading of sewage sludges. Establishing risk management strategies for polluted sites is extremely complex due to such factors as: (1) the heterogenous nature of the soil, (2) difficulties in characterizing the occurrence and predicting the transport and fate of pollutants in soils, (3) the unknown and mixed variety of substances in polluted soils, (4) the multiple pathways by which pollutants may reach humans and other receptors, and (5) the uncertainty and highly variable exposure conditions. In this chapter, the basic elements of a sound risk management strategy are discussed.
11.2
E L E M E N T S OF A RISK M A N A G E M E N T P R O G R A M
An effective risk management program should provide responsible administrative control, logical application of the best possible technical analyses, and the involvement of the public. A successful risk management plan must be credible, organized, and thorough, i.e., address the concerns of the public, be relevant, doable and economical, and based on existing technology with flexibility to adapt to later advances.
263
264
RISK MANAGEMENT PROBLEM
IDENTIFICATION
(1) Define areas of concern. (;2~) I d e n t i f y / p o l l u t a n t s o u r c e s . HAZARD (1) (2)
T
CHARACTERIZATION
Characterize pollutant sources. Map pollutant boundaries.
II
L~
EXPOSURE (1) (2) (3) (4) (5)
ASSESSMENT
Identify potential pathways. Identify potential receptors. Identify exposure pathways. Develop a conceptual model. Determine the extent of pollutant migration.
T
CONSEQUENCE (1) (2) (3)
ASSESSMENT
Assess pollutant toxicity. Evaluate human health risks. Evaluate socio-economic impacts.
Implement monitoring programs
RISK (1) (2)
Implement monitoring programs
No
MITIGATION Develop Develop
risk profile for site. strategy for risk mitigation.
Is r e m e d i a t i o n
require Yes
RISK (1) (2) (3) (4)
MITIGATION
Develop remedial action options. Identify regulatory criteria. Determine cleanup goals. Select appropriate remedial option.
IMPLEMENTATION OPTION (1) (2)
PLAN
Evaluate Evaluate
OF
SELECTED
option performance. socio-economic impacts.
Figure 11.1. Key elements of risk management program for polluted sites.
ELEMENTS OF A RISK MANAGEMENT PROGRAM
265
In addition, a thorough program identifies potential scenarios for accidental releases, potential consequences of such releases, and the actions planned to alleviate a problem when it occurs. The key elements of the risk management program are shown in Figure 11.1 and discussed in the following sections.
11.2.1 Hazards Identification Program The basic objectives of this phase are to define the system and to identify in broad terms the potential hazards. The review process during this stage should provide a summary of information that indicates potential health effects that may be associated with exposure to particular chemicals and the level at which exposures present risk to human populations, as discussed in Chapters 7 to 10. The basic steps in the hazard identification program are: Step 1." Identify the hazard(s). Is it a toxic release, an explosion, a fire ...? Step 2." Identify the sub-components of the facility, i.e., soil, groundwater, surface water, air and receptors, which give rise to the hazard(s). Step 3." Bound the study. Will it include detailed study of risks from physical and environmental factors, earthquakes .....? Frequently, the study will involve more than a preliminary identification of the subcomponents of the facility or set of events that lead to hazards. If the analysis is extended to include consideration of the event sequences which transform a hazard into an accident, as well as corrective measures and consequences of the accident, the study is called a preliminary hazards analysis. For example, at waste disposal sites, a preliminary analysis determines whether waste materials present at the site can be transported off-site, as discussed in Chapter 9. The analysis identifies the available pathways, determines whether any human or wildlife receptors can be reached by the waste, and identifies a set of transport and exposure scenarios, as discussed in Chapter 10. Each scenario encompasses a complete exposure pathway. The potential consequence of each scenario is the transmission of risk to receptors in the environment. A common hazard ranking scheme is:
Class 1Hazards: Class II Hazards: Class 111Hazards: Class IV Hazards:
Negligible effects Marginal effects Critical effects Catastrophic effects
11.2.2 Consequence Analysis The preliminary hazards analysis discussed earlier in this section represents a first attempt to identify the gross facility and the set of events which can lead to hazards while the facility is still in a preliminary investigation stage. This phase begins with the task of identifying potential pathways in which a release might occur. The study of hazards identification indicated the pathways which initiate the risks, i.e., soil, groundwater, surface water, air and receptors. Hence, the risk study begins by following the potential course of events. The first question that needs to be posed is "what is the possibility of transporting hazardous waste constituents with groundwater?" Groundwater is considered to be the pathway with the greatest long term human health consequences because aquifers supplying drinking water can be contaminated by chemical wastes for an indefinite time following a release. Wastes not present in
266
RISK MANAGEMENT
the liquid phase can be dissolved and leached by the action of surface or groundwater and transported off-site. Two dominant factors control the extent and rate at which infiltrating water or groundwater can affect the transport of mobile waste components: (1) the depth of the bottom of the waste material relative to the height of the upper most layer of groundwater, and (2) the hydraulic conductivity of the waste material and adjoining soil. Surface water can furnish a rapid and direct transport route for particulate matter and soluble chemical species. For the surface water transport to be complete, waste materials must be within the influence of runoff, and the surface elevation and slope must be able to support the movement of the runoff at a velocity capable of carrying the wastes in suspension. Once moved by surface water, the pollutants can be deposited atop soils, yielding a source for direct contact pathway. Waste components can be transported through the air if exposed to and affected by air currents. The waste may be transported either in vapour phase, if their vapour pressure favours their volatilization, or adhered to soil particles, if the soil is in a friable form. Wastes are also transported in the air when fugitive dusts are created. Heavy machinery, automobiles, or even foot traffic in an area of exposed waste material can generate dust upon which waste can be adsorbed. If captured by the wind, these dusts can be transported off-site. Another method of airborne transport of waste is the movement of gases generated by the digestion of landfill materials. Methane is commonly produced in landfills, and as it seeps to the surface, it can carry volatile waste components through the soil and into the atmosphere. To examine the use of the land and other resources proximate to the waste site, it is necessary to evaluate the possibility of transport of waste materials by various receptors, as discussed in Chapter 10. The basic steps in the consequence analysis phase are: (1) Calculate the amount of toxic material, as discussed in Chapters 7 and 8. Conceptually, this part of the analysis, which represents an assessment of the exposure/release sequence, is a key input into the consequence analysis. All the other inputs such as type of waste, site location, population density and prevailing weather patterns are site specific; (2) Follow the trajectory of the lethal toxins in the environment, i.e., exposure assessment, as discussed in Chapter 10; and (3) Assess the health effects, i.e., dose-response analysis, as discussed in Chapters 6 and 10.
11.2.3 Risk Mitigation Specific means used to control potential releases or their consequences to the environment must be identified. Figure 11.2 shows the possible remedial action techniques for controlling potential release of contaminants at a contaminated waste site. Also, to mitigate the release of toxic elements to the air, provision should be given for the use of: (1) scrubbing systems to neutralize or remove hazardous components, (2) flare systems to destroy hazardous compounds, and (3) secondary containment devices to temporarily hold the hazardous material until it is further processed before release. After the potential consequence analysis has been considered, a decision can be made regarding the acceptable level of risk.
QUANTIFIED RISK ASSESSMENT 11.3
267
QUANTIFIED RISK ASSESSMENT
Quantitative risk assessment estimates the degree of adverse effect due to release of specific wastes from a particular site. It is a tool to calculate a numerical value for existing and future health risks associated with exposures by means of identified complete pathways. To accomplish this end, the steps in a quantitative risk assessment are: (1) Specify the identity and rate of release of contaminants from the source, as discussed in Chapters 4 and 7; (2) Quantify the rate of transport, as discussed in Chapters 8 and 9; (3) Quantify the rates of exposure and uptake by receptors, as discussed in Chapter 10; and (4) Quantify the degree of toxicological effects, as discussed in Chapters 6 and 10.
I
Sediment
I
i
Hydraulic dredging Mechanical excavation
I PollutedI site I Surface water and soils
I I
Clay capping Regrading slopes Diversion structures Revegation Physical, chemical electrical and biological treatments
OPTIONS
I
I
Air
I I
Synthetic liners Clay capping Gas vents Gas treatment
I
Groundwater
I
I
Passive barriers Reactive walls Pump and treatment Physical and chemical treatments
Figure 11.2. Possible remedial action techniques for a contaminated site.
Data are developed on chemical residues in the various transport media. Assumptions and estimates based on available information regarding the transport, fate, and toxic effects of the compounds at hand are also applied. A typical scenario developed to describe exposure from an aquifer utilized as drinking water would include leaching of wastes in contact with groundwater, transport of those leached materials with the movement of groundwater toward domestic wells, and human exposure by ingestion. A sample of groundwater from a receptor's drinking water source is
268
RISK MANAGEMENT
necessary to verify the concentration and identify the pollutants in water, and a profile of the receptor's water use is needed. The exposure and uptake data are correlated with known toxicological effects of similar dosages. Finally, the quantitative assessment furnishes numerical values for the leaching and transport rates and duration over which the residues are expected to remain in the groundwater.
11.4
ROLE OF REGULATORY AGENCIES
Regulatory criteria that influence risk management include national, state or provincial, and/or local statutes. Pertinent regulation may include those specific to risk management of polluted sites as well as those of a more general or environmental nature. In general, regulations can be categorized as location-specific, pollutant-specific, or action/technology-specific (Holm, 1993). Location-specific regulations include those related to the site where the pollution and remediation occur. Pollutant-specific regulations include those relating to type, quantity, and/or character of the waste. Action~technology-specific regulations include those pertaining to the means and methods of risk management. An important aspect of regulation categories relates to when in the risk management process each category of regulation can reasonably be considered. While locationspecific regulation can be brought to bear on the remedial considerations relatively early (when site identification has been made), consideration of pollutant-specific regulations may await site characterization information. Likewise, action/technology-specific regulations can only be considered after the concepts, alternatives, and technologies of the risk management implementation are formulated. As an example of the regulatory framework for remedial planning and implementation, the US EPA (1988) has articulated regulatory compliance criteria for hazardous waste remediation in terms of applicable or relevant and appropriate requirements (ARARs). The term applicable refers to promulgated, legally enforceable laws and statutes that specifically address waste substances or pollution while the term relevant and appropriate refers to promulgated laws and statutes that relate to situations sufficiently similar to the particular waste situation and that are well suited to the situation. In addition, public participation in remedial planning is mandated by CERCLA (1980), and Superfund Amendments and Reauthorization Act (SARA) (1986).
11.5
REGULATORY APPROACHES
One of the first issues faced by agencies responsible for assessing the suitability of soils or other conditions at polluted sites is whether or not numerical criteria should be established. Numerical criteria, which are adopted, for example, by Canada, France, Germany, Netherlands, Norway, and United Kingdom, offer several advantages -- they are relatively easy to use and administer, they facilitate communication between interested parties, and reduce confusion. However, numerical criteria that are intended for broad application are insensitive to site specific conditions and often imply a level of understanding or confidence in the underlying science that may or may not exist. These limitations have contributed to decisions made by some regulatory agencies to refrain from setting numerical criteria. As an alternative, these agencies have chosen to establish procedures that intend to determine site specific objectives only. The procedures typically involve
REGULATORY APPROACHES
269
risk assessment in which the health risks that site users, visitors or neighbours can experience are
estimated. Examples of agencies that have selected this approach include US EPA, California Department of Health Services, and the New York Department of Environmental Conservation. While the latter approach is capable of considering site specific factors, it also imposes burdens upon all parties to apply the procedures correctly and defend the results. The methodology and equations are the subject of considerable debate and their use requires that issues such as inherent uncertainties in interpreting toxicological information and assigning a definition to acceptable risk levels be addressed. 11.5.1 Risk-Based Mitigation Criteria The US EPA approach for remediation involves a site specific risk assessment, as discussed in Chapter 10, conducted according to procedures which are comprehensively presented in the Superfund Public Health Evaluation Manual.
(1) (2) (3) (4)
(5)
The steps involved in remediation goal setting are: Selection of indicator compounds for a given site; Estimation of both short term concentration (STC) and long term concentrations (LTC) of the indicators in media at points of maximum human exposures; The STCs and LTCs are compared with applicable or relevant and appropriate standards (e.g., drinking water standards); Estimation of human sub-chronic and chronic daily intakes (SDI and CDI). SDI is the projected human intake averaged over short time and it is defined as the peak short term concentration multiplied by the human intake factor times body weight factor. CDI is the projected human intake averaged over 70 years and it is defined as the peak long term concentration multiplied by the human intake factor times the body weight factor. The estimation of daily intake is made assuming human exposure can occur from different media (air, groundwater, surface water, soil and fish) and intakes (ingestion, inhalation, skin absorption). The intake is estimated separately for each indicator compound, route of exposure, duration of exposure and population exposed. Total human intake for each route of exposure is a sum of daily intakes from all media by the same route. Additivity only applies to the same population, same time and same duration (i.e., sub-chronic versus chronic). For carcinogens, CDI values are used to calculate lifetime cancer risk where lifetime risk is equal to the CDI multiplied by the carcinogenic potency factor. In addition, additivity principles are used for assessing multiple chemicals; and The hazard indices for sub-chronic and chronic exposures (HIS and HIC) are then computed. HIS is defined as the ratio between SDI and AIS, where AIS is the acceptable daily intake for sub-chronic daily exposure which does not cause adverse effects during short term exposure. HIC is defined as the ratio between CDI and AIC, where AIC is the acceptable daily intake for chronic exposure which does not cause adverse effects during long term exposure.
The target levels for remediation are determined differently for indicator compounds with standards versus those without. If standards exist, that sets the upper limit on target levels. For those without standards, the compounds are divided into two groups: (a) chemicals with non-carcinogenic
270
RISK MANAGEMENT
toxic effects, and (b) potential carcinogens. For non-carcinogens, the daily intake for each compound must be maintained equal to or less than the acceptable daily intake. In addition, for multiple substances and routes, the overall hazard index must be maintained less than or equal to one. For carcinogens, remediation is intended to maintain the cancer risk in the range of 104 to 107, with 10~ established as a target point. The target concentration is that concentration that will produce a chronic daily intake associated with this range of risks. New remediation approaches by Superfund Amendments and Reauthorization Act (SARA), 1986, require that superfund remedies must be: (1) protective to human health and the environment, (2) cost effective, (3) utilize permanent solutions, (4) alternative treatment technologies and resource recovery to the maximum extent practicable, and (5) on-site remedies must meet applicable or relevant and appropriate regulations such as Resource Conservation and Recovery Act, Toxic Substance Control Act, Safe Drinking Water Act, Clean Water Act, and Clean Air Act. In 1985, the US Office of Technology Assessment (USOTA), acting in behalf of the US Congress, evaluated past practices associated with setting remediation goals at Superfund hazardous waste sites. The USOTA identified seven alternative approaches for establishing remediation goals at Superfund sites: (1) ad hoc, (2) site-specific risk assessment, (3) national goals for residual pollution, (4) clean to background or pristine levels, (5) best available technology or best engineering judgement, (6) cost-benefit approach, and (7) site classification. In addition, the USOTA drew the following important conclusions: (1) It is no longer acceptable to continue remediation under the current ad hoc approach. Dealing with each site as a unique case is inefficient and there is increasing likelihood that sites with similar problems will not be cleaned to comparable level of protection; (2) Pursuing remediation to background or pristine levels does not make environmental, technical or economic sense; (3) Though seemingly attractive and extensively used, best available technology or engineering judgement approaches do not offer environmental protection comparable to the likely high costs of implementation; (4) Though use of existing standards, risk assessment and cost benefit pose considerable problems, they may be used; (5) Remediation strategy based on site classification could be the most beneficial approach to be used as shown in Table 11.1. For this strategy to be successful, the decision regarding land use must be made at the local level; (6) There is a need to raise the issue of remediation goals to the highest levels of policy making with an open debate. The success of the Superfund program and private and state remediation depends on equitable and technically sound resolution of this issue; and (7) What is ultimately important and realistically achievable is consistency in the process of determining remediation goals, rather than necessarily making all remediation the same.
11.5.2 Numerically-Based Mitigation Criteria Over the past 20 years, the factors considered in establishing soil remediation criteria have changed both in terms of the number considered and in the relative importance assigned to individual factors. While some of the earlier efforts considered only one or two factors (such as background concentrations and analytical detection limit), some of the current initiatives by various agencies
REGULATORY APPROACHES ting remediation Dals (USOTA, 1985) Classes of national priority list sites
Remediation goals set by
Likely course of action
EPA classes of groundwater
Known or likely exposures to people or sensitive ecological elements requiring restoration of site, including groundwater
Site risk assessment
1. High priority initial response to recontrol site; 2. Obtain necessary data and perform risk assessment; 3. High priority full scale permanent cleanup when technology is available to meet remediation goals.
1. Groundwater is
vulnerable to pollution; 2. Irreplaceable source of drinking water to substantial populations.
Known or likely exposures exist, but limited number of people and sensitive environments. Clear alternatives to site remediation such as relocation and use of alternative water supply; site restoration or reuse not critical.
Cost-benefit analysis
1. Initial response. 2. After cost-benefit analysis choose risk management analysis.
Current and potential sources of drinking water or have other uses.
Site not likely to lead to exposures to people and not situated near sensitive environment. No site restoration or reuse anticipated.
Applicable and relevant environmental standards
1. Low priority initial
Not potential source of drinking water and of limited use.
response. 2. Reevaluation every 5 years to assess need for remediation.
272
RISK MANAGEMENT
consider several factors and utilize techniques such as environmental fate modelling and health risk assessment. In establishing numerically based-criteria for remediation, the following factors are generally considered: (1) background or ambient concentrations of pollutants, (2) environmental mobility of pollutants, (3) relationship between soil and water quality, (4) health of terrestrial plants and animals, (5) human health, (6) aesthetics, (7) limits of analytical capability, and (8) land use.
Table 11.2" Mean elemental content (mg/kg) of soil and crustal rocks, and the soil enrichment factor (SEE) Elem. Li Be B C N O F Na Mg A1 Si P S CI K Ca Sc Ti V Cr Mn Fe Co Ni Cu
Soil 2.4E+ 1 9.2E -1 3.3E+1 2.5E+4 2.0E+3 4.9E+5 9.5E+2 1.2E+4 9.0E+3 7.2E+4 3.1E+5 4.3E+2 1.6E+3 1.0E+2 1.5E+4 2.4E+4 8.9E 0 2.9E+3 8.0E+l 5.4E+1 5.5E+2 2.6E+4 9.1E 0 1.9E+l 2.5E+1
Crust 2.0E+ 1 2.6E 0 1.0E+I 4.8E+2 2.5E+1 4.7E+5 4.3E+2 2.3E+4 2.3E+4 8.2E+4 2.8E+5 1.0E+3 2.6E+2 1.3E+2 2.1E+4 4.1E+4 1.6E+ 1 5.6E+3 1.6E+2 1.0E+2 9.5E+2 4.1E+4 2.0E+l 8.0E+l 5.0E+l
SEF 1.2E 0 3.5E - 1 3.3E 0 5.2E+1 8.0E+l 1.0E 0 2.2E 0 5.2E -1 3.9E -1 8.8E -1 1.1E 0 4.3E -1 6.2E -1 7.7E -1 7.1E -1 5.9E -1 5.6E - 1 5.2E -1 5.0E -1 5.4E -1 5.8E -1 6.3E -1 4.6E-1 2.4E -1 5.0E -1
Elem. Zn Ga Ge As Se Br Rb Sr Y Zr Nb Mo Ag Cd Sn Sb I Cs Ba La Hg Pb Nd Th U
Soil 6.0E+I 1.7E+l 1.2E 0 7.2E 0 3.9E -1 8.5E -1 6.7E+1 2.4E+2 2.5E+1 2.3E+2 1.1E+I 9.7E -1 5.0E-2 3.5E -1 1.3E 0 6.6E -1 1.2E 0 4.0E 0 5.8E+2 3.7E+1 9.0E-2 1.9E+ 1 4.6E+1 9.4E 0 2.7E 0
Crust 7.5E+1 1.8E+l 1.8E 0 1.5E 0 5.0E -2 3.7E -1 9.0E+l 3.7E+2 3.0E+l 1.9E+2 2.0E+l 1.5E 0 7.0E-2 1.1E-1 2.2E 0 2.0E -1 1.4E-1 3.0E 0 5.0E+2 3.2E+1 5.0E-2 1.4E+ 1 3.8E+ 1 1.2E+l 2.4E 0
SEF 8.0E -1 9.4E -1 6.7E-1 4.8E 0 7.8E 0 2.3E 0 7.4E -1 6.5E -1 8.3E -1 1.2E 0 5.5E -1 6.5E -1 7.1E-1 3.2E 0 5.9E-1 3.3E 0 8.6E 0 1.3E 0 1.2E 0 1.2E 0 1.8E 0 1.4E 0 1.2E 0 7.8E -1 1.1E 0
Background Conditions Typical background concentrations naturally existing in the environment, is one of the most frequently used factors in setting criteria. Background concentrations are generally assumed to represent environmentally sound and acceptable conditions and establish the ultimate conditions that remedial actions can achieve. Information about background concentrations provides little guidance
REGULATORY APPROACHES
273
for compounds that are solely anthropogenic since the background concentration is zero. Another limitation of this factor is defining what "background" means and whether it is necessary to differentiate between background concentrations in various types of areas such as urban and rural areas. Soil Background Concentrations: The relationship of soil to crustal rock can be obtained from data shown in Table 11.2, which lists average mass concentration of 50 chemical elements in soils and surficial rocks (Sposito, 1989). The soil concentrations refer to samples taken approximately 0.2 m beneath the land surface from unpolluted soils across the United States. These concentration data are quite comparable with elemental composition of soils sampled worldwide. The 10 most abundant elements in soils are O > Si > A1 > Fe > C > Ca > K > Na > Mg > Ti, whereas in crustal rocks they are O > Si > A1 > Fe > Ca > Mg = Na > K > Ti > P. The elements in the list from carbon to calcium, with the exception of F, A1, and P, are quantified as macro-elements. The remaining 40 elements, therefore, are micro-elements. The soil enrichment factor (SEF), defined as the ratio of soil to crustal rock concentrations, is shown in Table 11.2. The SEF is a quantitative measure of relative enrichment (or depletion) of a chemical element in soil as compared with rock. Given the variability of both soil and rock composition, depletion or enrichment can be classified as: (1) SEF is less than 0.5 would indicate significant depletion; (2) SEF is greater than 0.5 and less than 2.0 would indicate no significant depletion or enrichment; (3) SEF is greater than 2.0 and less than 10.0 would indicate some enrichment; and (4) SEF is greater than 10.0 would indicate strong enrichment. For most of the micro-elements, there is a close correspondence between soil and crustal rock concentrations. Hence, it would be safe to say that by taking SEF = 1, the background concentration of various elements will correspond to that of the bedrock shown in Table 11.2.
Environmental Mobility Environmental mobility refers to the ability and/or ways with which a substance can move in the environment, as discussed in Chapter 6. Relatively mobile substances include those that are relatively soluble in water or volatile. Mobility is also influenced by environmental conditions such as soil properties and the characteristics of the groundwater regime. Mobile substances are likely to move off-site and/or come into contact with various types of receptors. Mobility is not explicitly considered in most remediation criteria. Factors such as soil pH (Alberta Criteria), soil characteristics (Ontario Criteria) and soil organic matter (The Netherlands Criteria) are considered. As environmental fate modelling techniques come into wider use, it is probable that environmental mobility will increase in importance. Environmental fate modelling techniques can also be expected to become an integral part of efforts to develop site specific remediation criteria.
Relationship Between Soil and Water Quality The relationship between soil conditions and those of local groundwater is an obvious one and several agencies that have issued remediation criteria for soil have also issued complimentary criteria for groundwater. In many cases the groundwater criteria are derived from drinking water guidelines and are based on the assumption that the groundwater is used directly as drinking water (British Columbia, Quebec, and the Canadian Council of Ministers of the Environment, CCME).
274
RISK MANAGEMENT
Environmental partitioning of pollutants between soil and groundwater has been used to set the reference values in The Netherlands.
Health of Terrestrial Plants and Animals Information concerning the health of terrestrial plants has been used by several agencies in setting remediation criteria to avoid phototoxic or other adverse effects on grazing animals. With regard to concentrations of substances in soil which are capable of adversely affecting vegetation, the scientific literature provides enough information for selected pollutants. Most of the available data pertains to agricultural crops and the pollutants of concern typically include boron, copper, nickel, and zinc. The data available with regard to adverse effects on grazing animals are a result of conditions observed in agricultural animals. For example, disorders associated with excessive amounts of molybdenum, selenium, and copper in diets or soils have been observed in cattle and sheep. There is very little information of this type for organic compounds.
Human Health Human health considerations, usually in the form of assessments of health risks, have been used increasingly over the past few years to develop remediation criteria. In some of the recently developed methodologies, human health considerations are the primary factor in setting remediation criteria. This philosophy is often predicated on the assumption that criteria that are sufficiently protective of human health will be sufficiently protective of the environment. For some pollutants this is known or suspected not to be the case. Examples include zinc and some esters which are capable of causing phototoxic effects (environmental degradation) before being of concern to human health. An approach based only on human health may be capable of establishing soil criteria for areas with potential to produce odours or taint locally grown produce. Risk assessment requires numerous assumptions to be made regarding the people being exposed, the pathways of exposure, the relationship between dose and response, and the environments in which exposure can occur. For carcinogens, it is assumed that any dose poses some level of risk and, therefore, there are the additional prerequisites of defining acceptable risk. Each of these aspects is laden with uncertainties and in some instances there is considerable debate as to proper procedures.
Aesthetics Pollutants in the environment can be sources of odours, staining of soil, discolouration, film or foams on water, and impart disagreeable tastes to water, plants, and the animals that live in such environments. Many odours or tastes can be detected at concentrations lower than those neecied to cause other types of adverse effects. Criteria based on aesthetic considerations are needed to avoid such effects. While those pollutants most likely to cause aesthetic concerns are well known, the concentration in the soil at which those effects occur are not well documented. Consequently, this factor is not often considered when establishing remediation criteria.
Analytical Capabilities Analytical detection limits have been used by several agencies in establishing remediation criteria. For some anthropogenic substances, it has been assumed that any measurable concentration is unacceptable for relatively sensitive land uses, such as residential or agricultural, and that a maximum acceptable concentration for less sensitive land use could be defined as a multiple ( a
REGULATORY APPROACHES
275
factor of 10 or 100) of the detection limit. The use of analytical capabilities in setting criteria likely will diminish as other factors rise in importance. Weaknesses of the analytical capabilities are: (a) detection of a substance does not automatically mean that an adverse effect will occur, (b) analytical detection limits have steadily improved over the past several decades, thus providing a constantly changing target, (c) the detection limit achieved on soil samples is a function of interfering compounds or conditions that may be present in samples, and (d) analytical detection limits are sensitive to the procedures followed during sample collection, transportation to the laboratory, handling, and transportation.
Land Use Land use is a frequently used factor in establishing remediation criteria. The types of land use most often addressed are residential/parkland (RfP), agricultural (AG), and commercial/industrial (C/I). Virtually all agencies that differentiate according to land use, advocate the use of lower criteria for AG _
276 11.6
RISK MANAGEMENT MITIGATION TECHNOLOGIES FOR POLLUTED SOILS
Mitigation technologies for polluted soils can be grouped, as illustrated in Table 11.3, into: (1) natural attenuation, (2) containment, (3) removal for subsequent on-site or off-site treatment, (4) in-situ treatment, and (5) land disposal 11.6.1 Natural Attenuation
Natural attenuation requires that the natural processes currently existing in the subsurface continue to provide adequate environmental protection. For example, due to existence of a thick clay layer beneath the disposal site, dissolved pollutants may be attenuated before an unacceptable risk to the environment occurs. Another example is the use of naturally occurring wetlands for pollutant attenuation. For this technique to be implemented, subsurface monitoring is required to ensure that the natural attenuation processes continues to provide the anticipated protection to human health and the environment.
Table 11.3: Mitigation technologies for polluted soils. Mitigation type
Available technology
Natural attenuation
@ Clay deposits @ Wetlands
Containment
O Physical (!) Covering systems | Slurry walls | Sheet piling O Chemical | Organic-based | Inorganic-based @ Thermal | Vitrification
Removal and treatment
O Product recovery 9 Groundwater extraction 9 Soil vapour extraction
In-situ treatment
O Soil flushing | Surfactant-based | Solvent-based O Electrochemical @ Biological
MITIGATION TECHNOLOGIES FOR POLLUTED SOILS
277
11.6.2 Containment
Containment techniques generally involve creating barriers or immobilizing pollutants for preventing unacceptable pollutant migration. Techniques such as physical, chemical, and thermal are used~
Physical Covering systems, pumping wells, steel casing, slurry walls of cement and bentonite are typical physical barriers for vapour and groundwater migration. These systems are discussed in Chapters 12 to 15. Chemical This technology attempts to encapsulate the pollutants directly, hence reducing the potential leachability. In micro-encapsulation technology, a liquid monomer is mixed with the polluted soil and, then, a catalyst is injected to polymerize the monomer and encase the polluted soil. In inorganicbased encapsulation processes, cement is used to solidify the polluted soil, as discussed in Chapter 20. Critical issues such as (1) compatibility of the waste material with the solidifying agents, (2) effect of the environmental conditions (e.g., pH and oxidation-reduction), (3) potential biological transformation, and (4) effect of freeze/thaw and wet/dry conditions are of major concerns in evaluating the technology performance. Thermal Thermal technique such as vitrification transforms the polluted soil into a solid mass similar to rock. This technology may create harmful gases, hence air pollution monitoring is required during its application. 11.6.3 Removal and Treatment
Recovery of non-aqueous phase liquids, groundwater, soil vapour and polluted soils can be removed and treated on-site or off-site.
Product Recovery Non-aqueous phase liquid (NAPL) products are recovered with the use of extraction wells and interceptor trenches, as discussed in Chapter 12. Subsurface monitoring is required to evaluate the removal efficiency of pollutants. Soil Vapour Extraction Soil vapour extraction is used for recovering pollutants from the soil gas phase, as discussed in Chapter 16. The technique is applicable to organic pollutants with high vapour pressure. Monitoring wells are required to assess the technology. The extracted pollutants will be treated onsite or off-site. Groundwater Extraction Pump and treat is usually used to recover dissolved pollutants in groundwater, as discussed in Chapter 12. It is usually used as a containment technique if the pollutants are highly adsorbed by
278
RISK MANAGEMENT
soils. The extracted pollutants may be treated on-site or off-site. 11.6.4 In-Situ Treatment
In-Situ Soil Flushing Solubilizing agents can be injected into the subsurface to dissolve pollutants and flush them from the soil. These agents are divided into two main categories: (1) surfactant-based techniques, as discussed in Chapter 18, and (2) solvent-based techniques, as discussed in Chapter 17. Extraction wells located below the water table can be used to collect the polluted water for treatment. Electrochemical Treatment This technology is used to treat in-situ polluted soils with low hydraulic conductivity. The technology relies on the use of a direct current (dc) and a set of anodes and cathodes to mobilize pollutants from the anode to the cathode, as discussed in Chapter 19. Extraction wells are used at the cathode to extract pollutants for treatment. Biological Treatment In-situ bioremediation is usually used for organic polluted soils. Current technology enhances aerobic biodegradation by adding oxygen and nutrients into the subsurface. Biodegradation of some organic pollutants can create new toxic compounds. For example, anaerobic biodegradation of trichloroethylene can produce vinyl chloride, a more hazardous pollutant. A detailed discussion of this technology is given in Chapter 21.
11.7
SELECTION OF MITIGATION OPTIONS
The selection of an appropriate mitigation option depends on a careful assessment of both short and long term risks posed by the polluted site. The selection process ensures that the selected mitigation option aims at achieving the protection of human health and the environment, and full compliance with mitigation criteria specified by regulatory agencies. Mitigation alternatives considered for potentially polluted sites should be evaluated based on sustainable development concept, as discussed in Chapter 2. The basic criteria in the sustainable development concept are: (1) Environmental: This criterion is used to evaluate wether a mitigation option is capable of eliminating the potential risk posed by the polluted site; (2) Social: This criterion is used to encourage public participation and account for the social costs associated with the mitigation option; (3) Technical: This criterion is used to evaluate the performance of a specific technology in achieving the requirements posed by the environmental criteria; (4) Economic: This criterion is used to evaluate the total costs associated with a mitigation plan; and (5) Land: This criterion is used to ensure land sustainability, which is achieved by integrating the above four criteria. In Chapter 22, a hypothetical site condition is used to demonstrate how these criteria can be
SUMMARY AND CONCLUDING REMARKS
279
used for evaluating various mitigation options and choosing the most appropriate solution.
11.8
SUMMARY AND CONCLUDING REMARKS
Any successful approach to risk management must address a multiplicity of technical and regulatory issues. Such complexities demand the detailed attention of a coordinated, multidisciplinary project team and integration of issues in a way perhaps unprecedented in other engineering and scientific endeavours. Inherent uncertainties in site conditions, geochemical data, regulatory criteria, and technology performance must be recognized in formulating a risk management approach. Regardless of the level of efforts expended, uncertainty cannot be eliminated and can only be managed through an approach that recognizes this fact, strives to understand the certainties, and establishes and implements an adaptive methodology. The following chapters, from 12 to 21, deal with detailed examination of risk mitigation technologies for polluted soils. These chapters cover a wide range of technologies such as subsurface control systems, ex-situ, and in-situ treatment techniques.
This Page Intentionally Left Blank
CHAPTER
TWELVE
SUBSURFACE CONTROL SYSTEMS --- G R O U N D W A T E R
12.1
EXTRACTION
---
INTRODUCTION
Subsurface control systems can be considered as means of modifying an unacceptable set of environmental conditions, e.g., groundwater pollution from an uncontrolled waste disposal site. The subsurface control systems involve the following: (1) Removal of a plume after measures have been taken to halt the source of pollution; (2) Diversion of groundwater from flowing through a source of pollution and from contacting a drinking water supply; and (3) Prevention of leachate formation by lowering the water table beneath a source of pollution. Subsurface control systems include groundwater pumping, subsurface drains, low hydraulic conductivity barriers and in-situ treatment methods. Groundwater pumping involves extraction of water from or injection of water into wells to capture a plume or alter the direction of groundwater movement. Subsurface drains consisting of gravity collection systems are designed to intercept groundwater. Low hydraulic conductivity barriers consist of a vertical wall of low hydraulic conductivity materials constructed underground to divert groundwater flow or minimize leachate generation and plume movement. In-situ treatment methods involve biological or chemical treatment methods. These control systems can be used singularly or in combination to control groundwater pollution. The materials, techniques and procedures currently available cannot provide complete and/or permanent isolation of the pollutants from the environment. It should be realized that subsurface control systems, except treatment methods, do not influence the nature of the physico-chemical characteristics of the pollutants. System failure due to reactions between the pollutants and the subsurface control materials may result in the release to the groundwater of either the original target pollutants or additional pollutants released as a result of these reactions. Care must also be taken to ensure that pollutants are not released to the groundwater as a consequence of the construction activities. Monitoring of subsurface control systems should be included as an integral part of the remedial strategy. It is essential that the effectiveness of the system be recorded during its entire lifetime. Monitoring parameters should include changes in hydraulic conductivity, physico-chemical characteristics of the pollutants, and variations in pollutants concentration. Monitoring of the system may assist in sensing if any reactions are occurring and, therefore, allow correcting actions to be initiated before the system breaks down entirely.
281
282 12.2
GROUNDWATER EXTRACTION W E L L HYDRAULICS
As withdrawal of water from an aquifer begins, the water level in the well declines as water is removed from storage in the well bore. This withdrawal lowers the head in the well below the level in the surrounding aquifer. As a result, water will flow from the aquifer into the well. As the water level in the well continues to decline, the rate of flow into the well from the aquifer will continue to increase until the rate of inflow equals the rate of withdrawal. The difference between the initial pumping rate and inflow to the well is due to borehole storage which can become an important factor in the analysis of short duration aquifer (pumping) tests and in large diameter wells. As pumping continues, the original water table is depressed resulting in the formation of a cone o f depression. The surface of the cone is the draw-down curve and the outer limit is the radius o f influence, as shown in Figure 12.1. The draw-down becomes negligible beyond the radius of influence, r o. The draw-down at a specific point near a pumping well is a function of the distance from the well, the draw-down in the pumping well, the properties of the aquifer, the effect of other pumping wells, and the aquifer boundaries. A draw-down curve produced from pumping data indicates the variation in the draw-down based on distance from the well and the draw-down at a specific point as a function of time.
Figure 12.1. Radial flow to a well in an unconfined aquifer.
In an idealized aquifer, flow to the well is radial as water converges on the well from all directions. The area through which the flow occurs decreases near the well. Therefore, the hydraulic gradient also increases toward the well to accommodate the same volume of flow. Mathematically, radial flow equations describe flow into the well are based on Darcy's law and the equation of continuity of mass. There are several important differences between the cone of depression in a confined aquifer
WELL HYDRAULICS
283
and that in an unconfined aquifer. Withdrawals from an unconfined aquifer result in the drainage of water from the pores in the soil or rocks, hence the water table declines and the cone of depression forms, as shown in Figure 12.1. Lowering of the water table as a result of de-watering of the aquifer results in a decrease in the saturated thickness and, therefore, in the transmissivity of the aquifer which is defined as the measure of the ability of an aquifer to deliver water to a well. The radial flow to the well bore is accommodated by an increase in draw-down in the well and in the aquifer. In the case of a confined aquifer, the water in the aquifer is under an artesian pressure due to the presence of the upper and lower confining stratums, as shown in Figure 12.2. When water is withdrawn due to pumping, a draw-down in the artesian pressure occurs. Since the storage coefficient of a confined aquifer is low because it consists of sandy soils, the cone of depression will be established rapidly.
Figure 12.2. Radial flow to a well in a confined aquifer.
12.2.1 Steady State Equations At steady state, the principle of conservation of mass dictates that the rate at which water crosses an imaginary cylindrical boundary at a known distance from the well must be the same as the rate at which water is pumped from the well. The area of the cylindrical boundary, A, is equal to the product of its height, b, which is equal to the saturated thickness of the aquifer, and the circumference, 2~r. The specific discharge, v, at any point is equal to the product of the hydraulic conductivity, Kw, and the hydraulic gradient, dh/dr. Hence, the rate, Qw, at which water is pumped from the well is given by:
284
GROUNDWATER EXTRACTION Q., = vA = - Kw -~r dh 2 rcr b
[12.1]
where Kw is the hydraulic conductivity, h is hydraulic head, r is radial distance, and b is aquifer thickness from which the water is drawn. For convenience, changes in water table height, h, can also be expressed in terms of the distance the water table lies below the height it had in the absence of pumping (the static condition). This distance is defined as the draw-down, s, as shown in Figure 12.2. Therefore, Eq. [ 12.1 ] can be written as:
Q~ =
[12.2a]
K ds 2rcrb ~ dr
or
as
Qw
m
dr
[12.2b]
2 =Kwbr
In designing a remediation well for an aquifer, it is important to be able to predict the total draw-down in the aquifer. Theoretically, in an idealized aquifer having unlimited extent and no recharge, the cone of depression advances outward to infinity. Thus, steady state analysis is approximate and useful only after pumping has occurred for some time. Steady state draw-down at any given radius r I from the well, relative to draw-down at another radius r2, can be determined by integrating Eq. [ 12.2b]:
s I - s2 =
r2
Qw 2rcK b
In - -
rI
[12.3]
where Sl is the draw-down at a radius r~ from the well, and s 2 is the draw-down at a distance r 2 from the well. In practice, absolute draw-down may be estimated if it is possible to define a radius of influence, r o, which represents the horizontal distance beyond which pumping of well has little influence on the aquifer. In other words, beyond ro, no significant draw-down due to pumping is assumed to exist. Then, Eq. [ 12.3] can be written as:
s -
Qw 2nKwb
r In __s
[12.4]
r
where s is the draw-down at a radius r from the well, and ro is the radius of influence. For an unconfined aquifer in which draw-down is a significant fraction of the saturated thickness, Eq. [ 12.3] may be expressed in terms of head rather than draw-down:
WELL HYDRAULICS ho2
-
Ow
h;2 -
285
r In o
~K w
[12.5]
r i
where ho is the hydraulic head at a distance ro from the-well, and h; is the hydraulic head at a distance r; from the well. Eqs. [12.3] to [12.5] are applicable to steady state or equilibrium conditions. Therefore, pumping needs to be continued at a uniform rate for a sufficient period of time to reach steady state condition in which the draw-down, s, at the point of measurement is changing negligibly with time. Other assumptions inherent in the development of the Thiem equations (i.e., Eqs. [ 12.4] and [ 12.5]) include: (1) The observation wells are located close to the pumping well; (2) The aquifer is homogeneous and isotropic, and has uniform thickness and infinite areal extent; and (3) The well completely penetrates the thickness of the aquifer.
12.2.2 Unsteady State Equations The first unsteady state radial flow equation was developed for confined aquifer situation by Theis (1935), and was subsequently extended by Wenzel (1942). The equation was developed on the basis of the following assumptions: (1) The aquifer is homogeneous and isotropic; (2) The aquifer is of infinite areal extent and constant thickness; (3) The discharge well has an infinitesimal diameter and completely penetrates the thickness of the aquifer; (4) Water obtained from storage in the aquifer is discharged instantaneously as the head declines; and (5) Discharge from the well is constant and the head distribution around the well is changing with time. In an idealized aquifer which fulfills the above assumptions, the general equations that define radial flow toward a pumped well are as follows:
s(r,t) -
Ow
[12.6]
W(u)
4rcK b
where W(u) - -0.577216- In(u)+ u-
and
U 2
2 + 2!
+
U3 -
U4
3 + 3!
4 + 4!
~
+
(-1)"u " n + n!
286
GROUNDWATER EXTRACTION
u -
0~s/"
2
[12.7]
4Kbt
where s(r, t) is the draw-down at time t and radius r, t is time since start of pumping, r is distance from discharge well, and a~ is storage coefficient (dimensionless). The solution of Eq. [12.6] is too tedious to be of practical application. However, Wenzel (1942) provided a simplified solution through a table of values of W(u) for a wide range of (u), as shown in Table 12.1. For small values ofu less than 0.01, Eq. [12.6] can be written as (Cooper and Jacob, 1946):
( 0.183Q w
s(r,t) = I
log( 2.25K bt)~ 2r
[12.8]
Kwb
Table 12.1" The well function W(u) u
1
•
0.219
• • • • •
13.24 15.54 8 17.84 -9 20.15 1~ 22.45
• •
•
-7
3
4
0.049 0.0130.0038
-1 1.82 1.22 .2 4.04 3.35 .3 6.33 5.64 .4 8.63 7.94 .5 10.94 10.24
x l 0 "6 •
2
0.91 2.96 5.23 7.53 9.84
12.55 12.14 14.85 14.44 17.15 16.74 19.45 19.05 21.76 21.35
• ll 24.75 24.06 x l0 12 27.05 26.36 • "13 29.36 28.66 • -14 31.66 30.97 • 15 33.96 33.27
23.65 25.96 28.26 30.56 32.86
5
6
7
8
9
0.0011 0.00036
0.00012
0.000038 0.000012
0.70 2.68 4.95 7.25 9.55
0.56 2.47 4.73 7.02 9.33
0.45 2.30 4.54 6.84 9.14
0.37 2.15 4.39 6.69 8.99
0.31 2.03 4.26 6.55 8.86
0.26 1.92 4.14 6.44 8.74
11.85 14.15 16.46 18.76 21.06
11.63 13.93 16.23 18.54 20.84
11.45 13.75 16.05 18.35 20.66
11.29 13.60 15.90 18.20 20.50
11.16 13.46 15.76 18.07 20.37
11.04 13.34 15.65 17.95 20.25
23.36 23.14 25.67 25.44 27.97 27.75 30.27 30.05 32.58 32.35
22.96 25.26 27.56 29.87 32.17
22.81 25.11 27.41 29.71 32.02
22.67 24.97 27.28 29.58 31.88
22.55 24.86 27.16 29.46 31.76
Note that each entry in the table represents the value of W(u) for a value of u equal to the product of the top row and the left column of the table.
WELL HYDRAULICS
287
In Eq. [ 12.8], the draw-down is proportional to log (t), indicating that pumping data may be plotted on semi-log paper as straight lines, as shown in Figure 12.3. The aquifer transmissivity, T &Kwb) can be determined by noting As~0, the difference in draw-down between any two values of time differing by a factor of 10: 0.183Q w
T-- Kwb -
[12.9]
Aslo
The storage coefficient, as, can be calculated by using the following equation: 2.25 Tt ~ as =
[12.10]
2
r
where T is transmissivity, to is the intercept of the straight line slope with the zero draw-down axis, and r is distance from the pumped well to the observation well, where the draw-down measurements were taken.
0 0.5 E "-" t.-
O "O
1.5 20
'
2.5
'-
3.0
D
A-
1.0 .................
............................
~l[
"
_[.
.
.
i
"
.
Ik
....................
!
3.5
-
]
9
.....
!
i
i
4.0 0.1
1.0
10
100
1000
T i m e (minutes~
Figure 12.3. Draw-down plot as a function of time.
Exact and approximate methods for calculating the radius of influence, r~ for steady state and unsteady state are shown in Table 12.2.
288
GROUNDWATER EXTRACTION
Table 12.2: Methods of calculating radius of influence, r e. Pumping condition
Method
Reference
Steady state Exact, unconfined
In r o = [~zK~(H2 - h,2)/Qw] + In r,
Thiem Eq.
Exact, confined
In ro = 2 ~ b K w [ ( H - h,)/Q~] + In r,
Thiem Eq.
Semi-empirical
ro = H ( K J 2N) m
(m)
Bear (1979)
Empirical
ro = 3000 s (Kw) !/2
(m)
Bear (1979)
Empirical
ro = 575 s (HK~) 1/2
(m)
Bear (1979)
Unsteady state Exact Semi-empirical (confined)
Draw-down versus log distance plots ro = C,e ( H Kw t / s)l/2 (m)
Bear (1979)
Semi-empirical R = Qe (H Kw t / G) ~/2 (m) Bear (1979) (unconfined) Kw is hydraulic conductivity; s is draw-down; H is initial head; h, is head at distance r, from the pumped well; N is rate of recharge; t is time since pumping starts; C.~eis a constant, varying between 1.9 and 2.45.
12.2.3 Capture Zone Analysis and Optimization Modelling Capture zone is defined as the portion of the aquifer which is affected by pumping. For relatively simple hydrogeologic settings (homogeneous isotropic aquifers), analytical solutions, presented in the previous sections, may be adequate to determine the capture zone. For more complex sites, numerical computer models may be required. These models provide insight to flow patterns generated by alternative pump and treat approaches and to the selection of monitoring points and frequency. The WHPA model (Blandford and Huyakorn, 1991), and Capture Zone Analytic Element Model (CZAEM) developed by EPA (Haitjema et al., 1994; Strack et al., 1994) are examples of relatively simple computer softwares that perform capture zone and groundwater pathline analysis. The numerical MODFLOW and MODPATH models developed by US Geological Survey are commonly used to model more complex hydrogeological settings. Cohen et al., (1994) identify a number of computer codes of potential value for capture zone analysis. Optimization programming methods are being used increasingly to improve pump and treat system design (Gorelick et al., 1993). As applied to the design of pumping systems, optimization involves defining an objective function, such as minimizing the sum of pumping rates from a number of wells.
GROUNDWATER MANAGEMENT BY HYDRODYNAMIC CONTROLS 12.3
289
G R O U N D W A T E R M A N A G E M E N T BY H Y D R O D Y N A M I C C O N T R O L S
Groundwater modification systems can be employed to: (1) lower groundwater level in the vicinity of the polluted sites, (2) increase the separation between the pollutants and groundwater, (3) supplement a barrier system, and (4) extract polluted water for treatment. The components of a hydraulic management system are extraction wells, infiltration or injection wells, and infiltration trenches. The extraction wells are generally installed as single or in groups to withdraw groundwater from the aquifer. Infiltration or injection wells can also be installed as single or in groups to recharge the subsurface, hence preventing ground subsidence. Hydrodynamic controls are the components that make up hydraulic barriers. They can be grouped into the following categories: extraction, injection, and a combined extraction-injection system. Extraction systems create an artificial depression in the groundwater table while removing water from the ground. These include: (1) gravity drains as trenches, (2) shallow wells and well point systems, and (3) deep wells. Injection systems have the reverse effect. By inducing water to flow into the ground, they create an artificial impression or mound, often called a pressure ridge, in the groundwater table. These include: (1) recharge basins, and (2) injection wells. These systems can be employed independently or combined to achieve the desired effect. Pumping systems are very often combined with injection systems in order to modify the flow of groundwater locally without affecting the surrounding groundwater level. In this system, polluted groundwater can be recovered at the surface, treated, and re-injected into the aquifer, in effect cleansing the groundwater. 12.3.1 Extraction systems
Gravity Drains Drains or trenches are usually excavated to a depth of a few metres below the ground surface. They are effective in lowering the water table by a few metres in unconsolidated materials and can be used to separate a superficial polluted zone from the groundwater table. Depending on their depth and the degree of extraction required, drains are the least costly of the extraction systems. Unfortunately, they are rather ineffective in deep aquifers. Shallow Wells and Well-Point Systems Shallow wells are generally between 5-10 metres deep and can be operated with a suction pump. They are an effective method of extraction and can also control the lateral and vertical movement of a polluted plume. Shallow wells can be used to lower a groundwater table that is near the surface while collecting shallow leachate. They can also be employed to intercept a polluted plume which is near the surface. They are relatively inexpensive to install. Furthermore, they use the least costly pumping equipment. Well-point systems are simply groupings of closely spaced shallow wells, normally connected together at the surface by a suction header which leads to a central suction pump. Well point systems have been successfully employed for de-watering construction sites, and many highly experienced people who can give sound advice on the practice are available. This type of system lends itself very well to lowering the groundwater table at large polluted sites such as landfills.
290
GROUNDWATER EXTRACTION
Deep Wells Deep wells are used where an aquifer or polluted plume does not permit the use of shallow well systems. Each well must be equipped with a pump. A variety of pumps, all of which have relatively high maintenance costs, may be considered. These include submersible pumps, and jet injectors.
12.3.2 Injection Systems
Recharge Basins Recharge basins are simply ponds of water created at or above the groundwater surface from which water will naturally infiltrate downward into an aquifer. Use of basins is generally limited to shallow, unconfined aquifers. Soil below the basin must be permeable enough to allow adequate infiltration rates. The bottom of the basin must be inspected and cleaned periodically to remove algal growths and less permeable sediments which may reduce the rate of infiltration. These basins do not need to be very deep, since the flow downward is principally dependent on the surface area.
Injection Wells Injection wells are used to inject water back into the ground either after treatment or as a method of controlling the movement of a contaminated plume. Wells may be shallow or deep but the small pipe used to apply water must penetrate at least below the water level of the targeted aquifer. This is important in order to avoid free fall of water into the well bottom which could lead to air entrainment in the aquifer. Dissolved air in the recharge water could cause "air binding" in the aquifer and seriously impair pumping-recharge process. Recharge water should be clean and free of a particulate matter, but even so, injection wells in unconsolidated aquifers eventually clog up the interface between the well and the aquifer. Therefore, periodic pumping and/or redevelopment (surging, jetting) of the well is required. The specific capacity of an injection well in an unconsolidated aquifer is only about half that of a pumping well due to potential clogging. In fractured rock or limestone with solution channels, however, rates may approach those of a pumping well. Although injection wells are more expensive to operate than recharge basins, they have several advantages: (1) recharge rates may be controlled as desired in order to follow an adjacent pumping rate, and (2) specific depths and aquifers can be targeted for recharge, including confined aquifers.
12.4
HYDRAULIC BARRIER SYSTEMS
12.4.1 Barriers Using Extraction Systems
Gravity Drains (1)
Gravity drains are used to: Reduce the flow from unpolluted sources as shown in Figures 12.4(a) and 12.4(b). Figure 12.4(a) illustrates the conventional subsurface drain that receives recharge from the stream as well as the leachate plume, resulting in a larger collection and treatment. Figure 12.4(b) illustrates the use of subsurface drain with clay or plastic barrier to reduce the flow to the drain.
HYDRAULIC BARRIER SYSTEMS (2)
(3)
291
Lower groundwater out of contact with a polluted plume. Here the water table is lowered by gravity provided the drains have a clear discharge. A sump equipped with a pump may be required on flat terrain. A schematic representation of this case is shown in Figure 12.5. Control plume movement below a waste disposal facility and to collect pollutants from the subsurface. A schematic representation is shown in Figure 12.6. It should be noted that the presence of viscous or reactive chemicals could lead to clogging of the drains, i.e., formation of iron manganese or calcium carbonate deposits.
Figure 12.4(b). Subsurface drain with clay or plastic barrier for reducing flow from unpolluted sources.
292
GROUNDWATER EXTRACTION
Figure 12.6. The use of subsurface drains to completely encapsulate the site.
Pumping Wells Pumping wells can be used for either short term or long term measures. Short term measures are generally established by lowering the water table through pumping. The following are the short term measures that can be undertaken to control migration of pollutants. Surface Water Bodies: Discharge of pollutants to surface water bodies can be prevented by changing the location of the discharge zones and moving it beyond bodies of surface water. Figure
HYDRAULIC BARRIER SYSTEMS
293
12.7 illustrates a hypothetical scenario of discharging the underground water to an adjacent stream of water body. The pollutant plume is suspended and a greater separation between the pollutant plume and the groundwater is achieved. In addition, the pollutant travel distance through the soil is increased, hence allowing the soil to attenuate the pollutants.
Figure 12.8. Management of direct contact between the waste and groundwater table.
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GROUNDWATER EXTRACTION
Direct Contact: Prevention of direct contact between the pollutants and the groundwater can be achieved by modifying the level of the groundwater table and creating a separation between the pollutant source and the top of the saturated zone. Figure 12.8 illustrates two conditions, one before the installation of the pumping well and the other after the implementation of the remedial technique. It can be observed that the pumping well is anchored in the confining strata to achieve the maximum separation distance between the pollutant and the new established water table. It should be noted that monitoring is required to ensure that the established separation continues to exist in the long term performance. Aquifer Pollution: Prevention of pollution of an underlying aquifer can be achieved by creating a localised upward hydraulic gradient. Construction of extraction wells around the pollutant source permits the development of a local upward hydraulic gradient, thus limiting the migration of pollutants. Figure 12.9 illustrates the hydraulic conditions before and after installation of the extraction wells. In order to ensure that the zone of influence is covered, a well- point system can be designed to overlap the individual cones of depression.
Figure 12.9. Management of a polluted aquifer by creating a localised upward hydraulic gradient.
Long term measures are considered for cases where a plume of polluted groundwater cannot be eliminated in a short term measure. The following long term measures may be employed: (1) Extraction at a specific designed rate of pumping with no subsequent recharge to the aquifer; (2) Use of a series of extraction and injection wells that will allow water within the plume to be pumped, treated, and pumped back into the aquifer; and (3) Extraction via a series of pumping wells and recharge through a basin or gallery. This technique can be adopted with or without treatment.
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Extraction-Discharge to Surface: Pumping without subsequent recharge may be an acceptable approach when a small flow of groundwater is involved. However, when large groundwater flows are involved or when residents are dependent on groundwater as a source of drinking water, recharge will probably be necessary. The main objectives of this system are: (1) plume containment to prevent polluting water supply systems, and (2) prevent pollution of streams or confined aquifers. Figure 12.10 illustrates the plume condition before and after adopting the extraction-discharge to surface technique of plume contaminant.
Figure 12.11. Management of salt water intrusion via extraction system.
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GROUNDWATER EXTRACTION
This system of extraction-well barrier can also be used to prevent salt water intrusion as shown in Figure 12.11. A line of wells is installed parallel to the coast and pumped continuously. This creates a thorough in the water table which draws in salt water from the coastal side and fresh water from the inland aquifer. Extracted brine is returned to the sea. The barrier formed by this system protects portable water wells located inland. Due to the volume of water that must be drawn from the aquifer, there is a considerable waste of fresh water using this method.
12.4.2 Barriers using Injection System For this specific application, injection well is used to control salt water intrusion. A line of injection wells is installed parallel to the coast through which fresh water is injected into the aquifer. This has the effect of forming a pressure ridge which pushes the salt water back towards the ocean. The fresh water injected must be of high quality and must often be transported to the site from inland. A schematic representation of this system is shown in Figure 12.12.
Figure 12.12. Management of salt intrusion via injection system.
12.4.3 Barriers Using Both Extraction and Injection Systems When designing a system using both extraction and injection wells, it is preferable to locate the two types of wells far enough apart so that they will not interfere with their respective zones of influence. This will give maximum flexibility to the system. Figure 12.13 illustrates the relationship between extraction and injection wells and the effect an injection well will have on the cone of depression. The designer must have knowledge of the impact of withdrawal upon recharge because if there is an increase in the area of interface between the two zones of influence, the behaviour of both components is slowed until hydraulic equilibrium is achieved. When this occurs, there is a net
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297
reduction in the effectiveness of the extraction well, which can lead to a change in the zone of influence of the system. Here it is important to have adequate monitoring in order to ensure that the resultant zone of influence is large enough to contain the polluted plume. Pumping wells and an injection well can be used, as shown in Figure 12.14, for plume containment with or without treatment of polluted fluids.
Figure 12.14. Management of polluted plume via groundwater extraction and injection systems.
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GROUNDWATER EXTRACTION
12.5 DESIGN AND CONSTRUCTION CONSIDERATIONS The design of effective hydraulic barriers, whatever the purpose may be, is entirely dependent on the geotechnical and hydrogeological characteristics of the location in question. The design of a groundwater extraction system must take into consideration a wide range of issues. A pump and treat groundwater remediation program can generally be categorized into four basic components: (1) comprehensive characterization of the site and assessment of constraints, (2) selection of an appropriate extraction and treatment system, (3) monitoring of system performance, and (4) establishment of cleanup goals. It must be understood that each of these components is complex and site-specific. The intent of the following discussion is to present an overview of these issues. 12.5.1 Site Characterization
The sources and characteristics of the groundwater pollution must first be established, as discussed in Chapter 7. A detailed assessment of the site history is necessary. The effort expended on collecting an inventory of the potential sources of pollution, the quantity of pollutants in the subsurface and the timing of the impact will be vital to the selection of the monitoring locations. The establishment of a monitoring-well field is necessary in order to define both the extent of pollution and the surface stratigraphy. The determination of the hydraulic conductivity of the various stratigraphic units, the hydraulic gradients and the direction of pollutant migration is critical. The areal and vertical limits of the pollutant plume must be clearly defined prior to considering remedial action. The selection of the appropriate drilling method is also important, particularly when drilling near the source area of pollution. Indiscriminately drilling through a source area containing separate phase product may creme vertical pathways for the migration of denser than water pollutants (i.e., DNAPL). If this occurs, the remediation of the site becomes much more difficult. The determination of the location of the source of the pollutants is obviously critical. The farther pollutants are transported from the point of release the more difficult the cleanup problem becomes. An efficient extraction and treat system will remove pollutants from areas as close to the source as possible. This may involve the excavation of surficial pollution. However, the identification of the source of separate phase chlorinated solvent pollutants (DNAPL) is often very difficult. The separate phase product will migrate vertically until it is physically confined by an impermeable strata, thereby creating its own source region. In these cases, the identification of the location of the source may, therefore, be impossible. The investigation of the vadose zone is an important aspect of site remediation. Pollutants within unsaturated formation materials act as potential sources, and may eventually reach the groundwater table at concentrations that may require cleanup. A frequently encountered practical engineering problem is the non-conventional recording of subsurface information. A monitoring well network may be established with the exclusive goal of collecting chemical data. However, it is important that the engineering properties of the soil and bedrock are conventionally defined. It is critical during the site characterization stage to anticipate questions which may arise during the subsequent development of the remedial action plan. These questions can often be answered if an attempt is made to maximize the information obtained during the installation of monitoring wells.
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12.5.2 Assessment of Property Constraints During the site characterization process, the impact of human practices on the objectives of the project will become evident. Groundwater pollution problems are often situated in industrialized areas, where pollution migrates across property boundaries. Issues such as the denial of property access, vandalism and unknown water extraction rates may play a role in the development of a pumping strategy. Long term security of the groundwater extraction system must be considered prior to addressing technical issues. A common scenario is for pollution originating from a property designated (or zoned) as industrial site and to migrate onto relatively densely populated residential property. When this occurs, a wide range of complex public health and safety issues will develop.
12.5.3 Remedial System Design Following the characterization of the site, a decision must be made as to the nature and extent of remediation required. The principal motivating forces which drive remediation include concern over public health and safety, and the requirement to achieve compliance with environmental regulations. The design of an extraction system should focus on two issues: (1) arresting the down gradient movement of the plume by placing a barrier to the leading edge of the plume, thereby facilitating containment, and (2) extraction of the greatest pollutant mass possible by extracting groundwater at the source areas, where the pollutant concentrations are the highest. In general, a staged approach to source removal and groundwater extraction has been proven to be most effective. A small number of wells should initially be constructed to achieve the goals of both pollutant mass removal and hydraulic control. The system should then be operated until equilibrium conditions or trends are established through the collection of water level and chemical data from the monitoring well field. The new information can then be used to adjust groundwater extraction rates or possibly modify system design.
Groundwater Interceptor Trench The construction of a trench-type active barrier is generally considered the preferred option for hydraulic control. The advantages of trench construction are: (1) a continuous zone of permeable material from which groundwater can be extracted will provide an effective hydraulic control, and (2) assuming low influent seepage rates, the trench will have a storage capacity, which will minimize the cycling of pumps and facilitate the efficient extraction of groundwater. The disadvantages of trench construction are: (1) construction of the trench would disrupt the residential neighbourhood. Public perception is of paramount concern, (2) disposal of excavated materials would present problems, and (3) health and safety issues related to trench construction.
(1)
(2)
The principal engineering design issues that must be considered are: The seepage into the trench must be determined. This can be performed by using either steady state or transient state groundwater flow models, as discussed previously. The calculation will determine the groundwater levels in the trench at various pumping rates, the potential cycling intervals of the pumping system, and the capture zone of the trench; Evaluating the number of extraction points, the backfill (filter) requirements, and the sizing of the treatment unit. The performance of a subsurface drain over the long term is critically dependent upon the selection of an appropriate backfill material. If an unsuitable material is
300
(3) (4)
(1)
(2)
(3) (4)
(5) (6)
GROUNDWATER EXTRACTION selected, there is the probability that the drain may clog with silt and clay. The selection of an appropriate drain material must consider both the estimated seepage into the trench and the trench wall material; A construction methodology must be established for the purpose of assessing the practical construction methods of the design, and determining the schedule, the quantities of materials to be used and the associated costs; and Health and safety considerations during trench construction. These concerns focus both on the regional and confined-space monitoring of air quality throughout the entire construction phase. Groundwater Extraction Well The principal engineering design issues related to extraction well are: Identification of the required extent and configuration of the extraction well network. The well locations are generally selected to provide plume capture based upon the available information regarding potential well discharge rates and radius of influence; Extraction well drilling and preliminary installation and testing methodology. The measurement of aquifer parameters may be carried out using either piezometer (response) tests or pumping tests. Response tests are generally used for the estimation of bulk hydraulic conductivity whereas pump tests are used to verify the capacity of an aquifer to yield water. In-situ hydraulic conductivity may be determined by response tests carried out on a single well. A detailed analysis o f the hydraulic conductivity estimation is discussed by Hvorslev (1951); Evaluation of the potential pore conductivity of the extraction well using conventional pump test draw-down analysis. Assessment of the aquifer yield is generally carried out by estimating the aquifer transmission properties; Recommendation of the optimal extraction system operating conditions in order to maximize capture of the leading down-gradient portion of the plume. The concept of capture zone analysis (Keely and Tsang, 1983) describes the portion of the aquifer affected by pumping which actually yields water to the well. The capture zone is generally much smaller than the zone of pressure influence. Capture zone analysis requires the pump test data for each well. Additional wells are added until sufficient pumping capacity is provided to create a capture zone of an appropriate size. The design input parameters include local ambient hydraulic gradient, hydraulic conductivity, aquifer thickness and available draw-down. To maintain draw-downs in the operating wells at stable levels, pumping rates may need to vary. The pumping system may require adjustment depending upon the observed performance. Such adjustment may include increasing or decreasing the pumping rates of selected wells, or selecting fewer or additional pumping wells to operate; Final design and construction of the groundwater extraction system; and Demonstration that the capture of the plume has been achieved based on water level data collected during system operation.
Pulsed pumping Pulsed pumping technique is a promising new technology (US EPA, 1989). The technique relies on the cycling of extraction or injection wells on and off in active and resting phases. In the resting phase of a pulsed pumping operation, sufficient time is allowed for pollutants to diffuse out
DESIGN AND CONSTRUCTION CONSIDERATIONS
301
of low hydraulic conductivity zones and into adjacent high hydraulic conductivity zones until maximum concentrations are achieved in the high hydraulic conductivity zones. For sorbed pollutants and NAPL residuals, sufficient time is allowed for equilibrium concentrations to be reached in the groundwater. Subsequent to each resting phase, the active phase of the pulsed pumping cycle removes the minimum volume of polluted groundwater, and the maximum possible concentrations, for the most efficient treatment. The flow chart shown in Figure 12.15 outlines a recommended active (trench and extraction systems) design procedure. The chart assumes that the decision has been made to proceed to construction of an active barrier system. It is intended to illustrate the decision process, by outlining a rational methodology for selecting either a trench or well-type active barrier system.
Define the aerial and vertical extent of plume
No
" construction of " a trench-type b rri ~feasible?
Ground water flow velocitv
Site sensitivity High
High
Design trench-type barrier , 9, - " " " l ~ t ~ esign acceptable?
No
Low
,.._l Design well-type barrier
Yes
P roceed with construction
Figure 12.15. A recommended active barrier design procedure.
12.5.4 Monitoring of System Performance The establishment of a monitoring well network in the vicinity of the pumping well is necessary in order to assess the hydraulic and chemical effect of the entire pumping well field. This
jj--J
302
GROUNDWATER EXTRACTION
is normally carried out using selected monitoring wells from the site characterization well network. Water level measurements are used to determine the actual flow field imposed by pumping wells. Understanding the flow field will improve the understanding of the pathways for pollutant transport, and will aid in the location of additional extraction wells. Monitoring of concentration with time will facilitate the tracking of pollutant mass removal in the vicinity of the well and serve to verify the active remediation. 12.5.5 Establishment of Cleanup Goals At present, all remediation cleanup goals are set by the local regulatory agencies, as discussed in Chapter 11. At most industrial sites, these cleanup goals have been set to meet the drinking water standards. The requirement for aquifer cleanup to drinking water standards in industrial areas or areas where the groundwater resources have no reasonable foreseeable use is being questioned. Many industries argue that the establishment of risk-based cleanup levels, as discussed in Chapters 10 and 1 l, can be fully protective of the human health and the environment. The debate on this issue is likely to continue for many years to come.
12.6
CRITICAL EVALUATION OF GROUNDWATER MANAGEMENT
The current groundwater management technique relies on pump and treat which can effectively contain widespread, deep groundwater pollution. Pump and treat systems for groundwater remediation came into wide use in the early to mid 1980s. By the early 1990s, failure of the pump and treat approach was identified as its inability to achieve restoration in 5 to 10 years. Pump and treat systems were criticized more pointedly by Travis and Doty (1990), who asserted as a simple fact that polluted aquifers cannot be restored through pumping and treating. The drawbacks of the use of this method are numerous. When the objective of the system is to cleanup groundwater to drinking water standards, the results have been disappointing. The reason for the lack of success of pump and treat remediation can be attributed, in part, to the fundamental lack of understanding of the soil-pollutant-groundwater systems. Two principal issues include the inadequate knowledge of the flow of groundwater in heterogeneous media and lack of appreciation for the time required for the dissolution of separate phase pollutants (tailing and rebound phenomena). These problems combined with the improper use of predictive computer modelling have contributed to the current level of scepticism surrounding pump and treat system. The phenomena of tailing and rebound can best be described with the use of Figure 12.16. Tailing refers to the progressively slower rate of decline in dissolved pollutant concentration with continued operation of pump and treat system while Rebound refers to the rapid increase in pollutant concentration that can occur after pumping has been discontinued. This increase may be followed by stabilization of the pollutant concentration at a somewhat lower level (Cohen et al., 1994). 12.6.1 Tailing and Rebound Impacts on Remediation Tailing presents two main difficulties for groundwater restoration: (1)
Longer treatment times: Without tailing, and from a theoretical viewpoint, pollutants can be removed by pumping a volume of water equivalent to the volume of the pollutant plume.
CRITICAL EVALUATIONS OF GROUNDWATER MANAGEMENT
(2)
303
However, the tailing effect increases the pump and treat system operating time by hundreds of years; and Residual concentrations in excess of the cleanup standards: As discussed in Chapters 3, 5, and 6, soils have a high capacity for pollutant adsorption. During pumping, the concentration in soil pore fluid is decreased at a faster rate than that in the adsorbed phase, hence a residual concentration is retained by the soil constituents. For soils with high clay content, residual concentrations could be in excess of cleanup standards. When pumping is halted, for enough time to allow equilibrium between the liquid and solid phases, pollutant concentrations in soil pore fluid will be increased (rebound).
Figure 12.16. Concentration versus pumping duration showing tailing and rebound effects.
12.6.2 Contributing Factors to Tailing and Rebound The degree to which tailing and rebound complicate remediation efforts at a site is a function of the physical and chemical characteristics of the pollutant being treated, the subsurface material, and groundwater flow. Major factors and processes that contribute to tailing and rebound are discussed below. Groundwater Velocity Variations Tailing and rebound occurs as a result of the variable travel times associated with different flow paths taken by pollutants to an extraction well. For example, at the edge of a capture zone which is created by a pumping well, groundwater travels at low hydraulic gradient whereas adjacent to the well, it travels with high hydraulic gradient. Between the two extremes, groundwater travels with variable velocities.
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GROUNDWATER EXTRACTION Matrix Diffusion
As pollutants advance through relatively permeable pathways in heterogeneous media, concentration gradients cause diffusion of pollutant mass into less permeable media (Gillham et al., 1984). Matrix diffusion is most likely to occur with dissolved pollutants that are not strongly sorbed, such as inorganic anions and some organic chemicals. During pump and treat operation, dissolved pollutant concentrations in the relatively permeable zones are reduced by advective flushing, causing a reversal in the initial concentration gradient and slow diffusion of pollutants from the low to high hydraulic conductivity media. The significance of matrix diffusion increases as the length of time between pollution and cleanup increases. In heterogeneous aquifers, matrix diffusion contribution to tailing and rebound can be expected, as long as pollutants have been diffusing into less permeable material. Pollutant Desorption
In pump and treat system, groundwater flow velocity is initially high, causing an initial decrease in pollutant concentrations. The decline in concentrations will later tail off until the rate of pollutant dissolution is in equilibrium with the velocity of the pumped groundwater. If pumping stops, the groundwater velocity slows and concentration rebound occurs due to desorption process, rapidly at first and then gradually reaching equilibrium unless pumping is resumed. For example, non-aqueous phase liquids (NAPLs) tend to be relatively insoluble in water. However, with desorption, NAPL concentrations are low enough below the solubility limits, causing an increase in groundwater concentrations above the regulatory limits. Therefore, residual and pooled free products will continue to pollute groundwater. Precipitation-Dissolution
As with adsorption-desorption processes, precipitation-dissolution reactions are reversible. Thus, large quantities of inorganic pollutants may be found with crystalline or amorphous precipitates in the subsurface, as discussed in Chapter 5. Palmer and Fish (1992) studied the effect of solid phase pollutant precipitate on dissolved pollutant concentrations during pumping from a recovery well. The results, shown in Figure 12.16, indicate a tailing curve where the pollutant concentration is controlled by solubility. In this situation, if pumping stops before the solid phase precipitates are depleted, rebound can occur due to desorption and dissolution.
12.7
SUMMARY AND CONCLUDING REMARKS
Large expenditures are made each year to prepare for and operate remedial measures for groundwater pollution. Regulatory responsibilities require that adequate oversight of these remedial measures be made possible by structuring appropriate compliance criteria for monitoring wells. The oversight efforts are nominally directed at answering the question, "what can be done to show whether or not the remedial measures are achieving the desired pollution control?" Recently, other questions have developed because of the realization that pump and treat technology does not function as it has been presumed. Such questions include: "what can be done to determine whether the remediation will meet its intended goals?" and "what can be done to determine whether the remediation will be within the budget limit?"
Obviously, these questions can be answered by the use of sophisticated data analysis tools,
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such as computerized mathematical models of groundwater flow and pollutant transport. Computer models can indeed be used to make predictions about future performance, but such predictions are highly dependent on the quality and completeness of the field and laboratory data. The latter is just as true for models evaluating pump and treat technology, in contrast to the common belief that an accurate performance evaluation can be made simply by comparing data obtained form monitoring wells during cleanup to the data generated prior to the onset of remediation. Historical trends of pollution levels at local monitoring wells were rendered useless by the extraction and injection wells used in pump and treat. This is a consequence of the fact that the extraction and injection wells produced complex flow patterns locally, where previously there were comparatively simple flow patterns. Complex groundwater flow pattems present great technical challenges in terms of characterization and management of the associated pollutant transport pathways. For example, water moving along a flow line that proceeds directly into a pumping well from up-gradient is moving the most rapidly whereas those waters lying at the lateral lifts of the capture zone move much more slowly. This in turn results in certain parts of the aquifer being flushed quite well while others are remediated relatively poorly. In addition, the previously unpolluted portions of the aquifer, that form the peripheral bounds of the pollution plume, may become polluted by the operation of an extraction well that is located too close to the plume boundary. It is not possible to determine precisely where the various flow lines generated by a pump and treat system are located unless detailed field evaluations are made during remediation. Neither pollutant nor velocity distribution are constant throughout the zone of action. Consequently, more data must be generated during remediation. Decisions regarding the frequency and density of chemical sampling must consider the detailed flow paths generated by the remedial action.
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CHAPTER THIRTEEN
SUBSURFACE C O N T R O L SYSTEMS --- A L T E R N A T I V E S
13.1
TO GROUNDWATER
EXTRACTION
---
INTRODUCTION
Although groundwater extraction is currently the preferred option with regard to pollutant plume containment, alternatives such as passive barriers may require consideration. Passive barriers can be broadly grouped into either impermeable cutoff walls or permeable reactive walls, as shown in Table 13.1. The fundamental difference in philosophy is that the objective of the cutoff wall is to prevent all advective flow of groundwater whereas the reactive wall is intended to function as an insitu treatment system. The design of passive barriers has been traditionally criticized for two main reasons: (1) long term performance has been questioned. If the barrier begins to leak, further remedial work will be required. This may be extremely expensive if the property was redeveloped following cleanup, and (2) since the objective of passive barriers is to contain the contaminated plume, these barriers do not offer the possibility of long term plume and source removal. The principal advantage of passive barriers is obviously the low maintenance costs. Such barriers, unlike pump and treat systems, do not require continuous operation. The failure of pump and treat systems to offer a remedial solution in a reasonable time frame has lead to a resurgence of interest in passive barriers. This chapter discusses the principles of passive barrier systems and the various environmental problems that may impact on their performance.
Table 13.1: Types of passive barrier systems Type 9 Impermeable cutoff walls
9 Permeable reactive walls
Available technology @ Slurry walls c~" 0) Soil-bentonite @ Cement-bentonite | Diaphragm c~" 03 Rock grouting O Grouting | Grout curtains O Sheet piling O Ground freezing O Electrokinetic barriers @ Funnel-and-gate system O Continuous curtains
307
308 13.2
ALTERNATIVES TO GROUNDWATER EXTRACTION SLURRY WALLS
Slurry walls are the most common subsurface barriers because they are inexpensive. During construction of a slurry wall, the trench should be maintained full of a clay-based slurry while native material is being excavated. This will allow the excavation near vertical trench walls to remain stable during construction. When excavation is complete, the slurries are allowed to set, thereby forming a hydraulic barrier. 13.2.1 Structure Formation
Slurries or trench excavations are clay colloidal systems. Such a system is obtained when a small amount of clay powder, e.g., bentonite, is stirred in a large volume of water. Colloidal particles interact with each other via the diffuse ion layers of exchangeable cations, as discussed in Chapter 5. The total energy of interaction between colloidal particles can be calculated by the addition of repulsion and attraction forces. Repulsion results from inter-penetration of ion layers of adjacent particles, and from adsorption of water on surfaces of adjacent particles. Repulsion is manifested in swelling of colloidal particles. Water molecules adsorbed on the surface will force adjacent particles apart. This accounts for swelling at low water contents, at which adsorbed water is strongly held. When two colloidal particles are less than 15 A apart, the exchangeable ions are uniformly distributed in the inter-particle space and do not separate into two-diffuse layers, one associated with each surface. Under these conditions there is a net attraction between particles. However, when the inter-particle distance exceeds about 15 ~, diffuse ion layers form, with a resulting net repulsion. This repulsion can again be visualized as being due to water attracted between the particles forcing them apart. In this case, water movement is due to the osmotic activity of the ions between particles rather than to the properties of the surface. The concentration of ions is higher in the plane mid-way between two parallel particles than in the outside solution. Water moves in response to this concentration gradient. The concentration difference depends upon the distance between particles and upon how far the diffuse ion layers extend, i.e., upon the valence and concentration of ions. Repulsion will be greatest with monovalent exchangeable ions and with distilled water as pore water. In describing attraction between colloidal particles, several different forces must be considered. First, there is the attraction between molecules and atoms described by the London-van der Waals theory. These are short range forces, and are inversely proportional to the seventh power of the distance between atoms. They, therefore, decrease very rapidly with increasing distance of separation. The magnitude of the force depends upon the properties of the surfaces. At inter-particle distance of less than 15 A there is a net force of attraction between clay particles when their exchangeable cations are in the inter-particle space. The magnitude of the forces of attraction and repulsion in a colloidal system vary, with the maximum attraction lower than the maximum repulsion. The forces of attraction can be manifested only if the conditions do not favour repulsion. In soil-bentonite slurry, repulsion is dominant. The inter-particle forces of attraction and repulsion determine colloidal particle arrangement or fabric. The above concepts have immediate practical implications in slurry-wall construction. A flocculated slurry is unstable and exhibits marked variation in its flow characteristics. Flocculation is a frequent occurrence in a slurry system and leads to particle association which is govemed by
SLURRY WALLS
309
three distinct modes: face-to-face, edge-to-face, and edge-to-edge, as shown in Figure 13.1. Since the energy of interaction is governed by each combination, a different geometry must be considered when summing the interaction of approaching plates (Yong, Mohamed, and Warkentin, 1992). The net of interaction energy varies with each mode of association, and the three forms may not occur simultaneously or to the same degree whenever a clay suspension is flocculated. The physical results from each mode have different effects and can change the properties of flocculated slurries in different ways.
Figure 13.1. The three principal modes of particle configuration.
Face-to-face association merely leads to thicker and probably larger flakes while edge-toedge and edge-to-face association cause the formation of a three dimensional structure. Among the three basic forms of flocculation, only the edge-to-edge and edge-to-face lead to aggregation. The thicker particles resulting from face-to-face association are essentially aggregated forms. Aggregation and flocculation do not proceed simultaneously or to the same degree. Thus, a colloidal system may be well dispersed but not flocculated, or it may be de-flocculated but not well dispersed.
13.2.2 Factors Affecting Structure Formation
The various parameters that could contribute to the formation of different particle configurations are salt concentration, pH, anionic species, and clay mineralogy. Salt Concentration The rheograms of homoionic sodium kaolinite (Hydrite PX) saturated with NaC1 and dispersed in 0.01 to 3000 meq/1 of NaC1 with a 9.1% (w/w) solids concentration and 10 meq/1 of
310
ALTERNATIVES TO GROUNDWATER EXTRACTION
NaHCO 3 with 4.85% (w/w) solids concentration are shown in Figure 13.2 (Yong et al., 1977). The region to the left of the curve identifies a dispersed clay particle system while the region to the right signifies that the clay system is flocculated.
Figure 13.3. Differential viscosity for kaolinite dispersed in various concentration of sodium chloride (NaC1) and sodium bicarbonate (NaHCO3).
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Clay with solutions of sodium chloride (NaC1) concentration greater than 1 meq/1 will tend to flocculate. If the NaC1 concentrations are increased, a distinct and separate family of curves will be obtained. These are characteristic of increased flocculation activity. The differential viscosity (shown in Figure 13.3) shows a significant increase in value when the NaC1 concentration begins to exceed 1 meq/1. As the differential viscosity increases, the flocculation of the initially dispersed kaolinite suspension can be initiated with NaC1 salt. Thus, when the NaCl concentration is in excess of 1 meq/1, the initially dispersed kaolinite system shows evidence of flocculation. In contrast, noting the lack of increase in the differential viscosity for the kaolinite in solution of sodium bicarbonate (NaHCO3), it can be deduced that the kaolinite in NaHCO3 system remains dispersed up to a concentration of 100 meq/1. The is attributed to the effect of potential determining ions (HCO3) on surface charge characteristics, as discussed in Chapter 4. The adsorption-desorption of the sodium saturated kaolinite system, shown in Figure 13.4, reveals that adsorption starts only when the NaC1 concentration of the initial solution is greater than 1 meq/1 while desorption is observed when the concentration is less than 1 meq/1. In view of Figures 13.3 and 13.4, it is evident that desorption occurs when the clay is dispersed and adsorption occurs when the clay is flocculated.
Figure 13.4. Adsorption-desorption of sodium kaolinite dispersed in various concentrations of sodium chloride (NaC1) and sodium bicarbonate (NaHCO3); Y1 and Y2 refer to the left and the right vertical axes, respectively.
The maximum energies developed in inter-particle action and the corresponding inter-particle separation distances for kaolinite are shown in Figure 13.5. The calculations have been performed using the actual zeta potential measurements and DLVO theory (Verwey and Overbeek, 1948) for the case of similar charged surfaces of clay particles. At low NaC1 concentrations, i.e., less than 1 meq/1, the double layer repulsion exceeds van der Waals attraction, resulting in a dispersed system.
312
ALTERNATIVES TO GROUNDWATER EXTRACTION
At higher NaC1 concentrations, i.e., greater than 1 meq/1, the double layers present at both edge and face are compressed, thereby reducing inter-particle repulsion at a given inter-particle distance. Since the van der Waal attractive force for a given mode of particle configuration (edge-to-edge, edge-toface, and face-to-face) remains constant at a given inter-particle distance regardless of NaC1 concentration, a reduction in the net inter-particle repulsion occurs.
Figure 13.5(a). Inter-particle separation distance at various sodium chloride concentrations.
In view of these considerations and results shown in Figure 13.3, we would deduce that the probable particle configuration of the flocculated clay system would seem to be edge-to-face flocculation for the kaolinite clay with NaC1 concentrations in excess of 1 meq/1. For NaC1 concentration less than 1 meq/1, the probable particle configuration is face-to-face.
Anionic Species In the presence of potential determining anions such as bicarbonates, phosphates and polyphosphates, the influence of anionic species becomes indeed very important, as discussed in Chapters 3 and 5. From Figure 13.3, it can be seen that a 10 meq/1 solution of NaHCO 3 yields a suspension which produces a lower viscosity than a 0.1 meq/l solution of NaCI. From the results shown in Figures 13.1 and 13.3, it is observed that if the kaolinite is dispersed in NaC1 with a concentration greater than 1 meq/1, the clay is essentially non-dispersive or flocculated. This observation has also been made by van Olphen (1977) for illite and montmorillonite where the initial concentrations of NaCl at which flocculation starts are obviously somewhat higher, i.e., 2-6 meq/1 NaC1. The observation could suggest that for those cases where the bedrock near a slurry wall consists of limestone or some other carbonate containing minerals, the criterion used to evaluate the potential dispersivity of the slurry wall may differ from that used for polluted groundwater containing no carbonates, bicarbonates or phosphates.
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Figure 13.5(b). Maximum interaction energies at various sodium chloride concentrations; Y 1 and Y2 refer to the left and the right vertical axes, respectively.
Figure 13.6(a). Rheograms of kaolinite, illite and their mixtures.
Mineralogy The rheograms of kaolinite, illite, kaolinite plus illite in a mixed portion of 60% kaolinite and 40% illite dispersed in 15 meq/1 NaOH or NaHCO3 are shown in Figure 13.6(a). A mixture of
314
ALTERNATIVES TO GROUNDWATER EXTRACTION
kaolinite and illite in either NaOH or NaHCO3 yields a greater Bingham yield stress than kaolinite or illite dispersed separately at the same solids content. This can be attributed to the edge-to-face flocculation of the clay particles where the negatively charged illite faces are attached to the positively charged edges of kaolinite by mutual attraction. The effect of montmorillonite on the structure formation can be seen from the results shown in Figure 13.6(b). In these experiments, varying amounts of montmorillonite (0-20% of mineral solids) were added to the 60% kaolinite and 40% illite mixture and dispersed in 15 meq/1 NaOH or N a I - I C O 3. The addition of 5% montmorillonite to the kaolinite-illite system in NaOH causes a further increase in both Bingham yield stress and differential viscosity. This can be attributed to interaction of montmorillonite with kaolinite edges resulting in an increase in inter-particle bonding. When the montmorillonite content is increased to 10% of the total mineral content, Bingham yield stress and differential viscosity are reduced. This is probably due to the fact that the initial addition of approximately 5% montmorillonite has covered all the edges of kaolinite and illite, with the excess montmorillonite creating a situation that induces face-to-face repulsion.
Figure 13.6(b). Effect of montmorillonite and anions contents on shear stress and viscosity of 60% kaolinite and 40% illite mixture; Y 1 and Y2 refer to the left and the right vertical axes, respectively.
A further increase in montmorillonite content results in an increase in Bingham yield stress. However, the differential viscosity remains relatively constant. In a N a H C O 3 solution, the Bingham yield stress increases with increasing amount of montmorillonite while the differential viscosity remains relatively constant. This would indicate an increase in the inter-particle bonding which could be due to increased edge-to-face clay particle flocculation with increasing amounts of montmorillonite. The results again demonstrate the importance of anions in the control of interparticle forces.
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13.2.3 Influence of Pollutants on Slurry Wall Hydraulic Conductivity
The effect of various pollutants on slurry wall hydraulic conductivity can be grouped under the categories of inorganic and organic chemicals.
Inorganic Chemical Effect The experimental results reported by Alther et al., (1985) are the basis for the following discussion on the effect of inorganic chemicals on slurry wall hydraulic conductivity. The reported experimental results are used to calculate the relative change in hydraulic conductivity, Kp/Kn, where Kp is the hydraulic conductivity of polluted bentonite with various salts, and Kn is the hydraulic conductivity of unpolluted (natural) bentonite. The calculated results are shown in Figures 13.7 to 13.9 for various cations in different associations with CI-, CO32, and SO42- anions, respectively.
Figure 13.7. Relative change in hydraulic conductivity as a function of cation type and concentration, and anion type (CI).
For the same anion, the relative change in hydraulic conductivity increases with increasing cation concentration. The increase in hydraulic conductivity can be largely attributed to the significant decrease in the repulsive inter-particle forces due to the increase in cations (i.e., Na § K +, Mg 2+, Ca 2§ in the pore fluid. Another possible consequence of the introduction of higher concentrations of various cations could be the disruptive effects of different cations on water structure (Lutz and Kemper, 1959). At high solution concentrations, cations would compete with the original water structure, which could in turn cause a reduction of its thickness around the clay particles, and a subsequent increase in hydraulic conductivity.
316
ALTERNATIVES TO GROUNDWATER EXTRACTION
Figure 13.8. Relative change in hydraulic conductivity as a function of cation type and concentration and anion type (CO32).
The effect of cation valence on the resultant hydraulic conductivity of bentonite slurry has considerable consequence in the specification of the type of bentonite to be used. At low concentrations, hydraulic conductivity showed a 4 to 5 fold increase for divalent cations relative to monovalent cations. The increase in hydraulic conductivity could be attributed to the reduction in the diffuse ion layer thickness, as discussed in Chapter 5. As the concentration of cations increases, the thickness of diffuse ion layer decreases and, therefore, hydraulic conductivity increases. In the case of divalent cations, a relatively lower concentration is required to balance the negative charges on the surface of bentonite resulting in smaller thickness of the diffuse ion layer. With an increase in concentration, the thickness of the diffuse ion layer remains relatively unchanged, hence the hydraulic conductivity does not show any significant change. However, for monovalent cations, with an increase in concentration, the diffuse ion layer thickness continues to decrease and the structure favours edge-to-edge or edge-to-face formation. The other important aspect that can be observed from Figures 13.7 to 13.9 is the role of anions on the resultant changes in hydraulic conductivity. Before embarking on any discussion related to changes in hydraulic conductivity, we well discuss the mechanisms by which anions can be adsorbed to clay surfaces. An anion approaching a charged surface is subject to attraction by positively charged sites on the surface, or repulsion by negative charges. Layer silicates in the clay fraction of soils are normally negatively charged, as discussed in Chapter 4, so that anions tend to be repelled from the mineral surfaces. An anion approaching soil solids may thus be subjected to simultaneous repulsion by negatively charged surfaces and attraction to positively charged sites on clay edges. Factors affecting anion repulsion, as discussed in Chapter 5, include: (a) anion charge and concentration, (b) species of exchangeable cations, (c) pH, (d) presence of other anions, and (e) nature and charge of the colloid surface. If the negative charge on a soil colloid surface remains constant, anions of higher charge are
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repelled more than anions of lower charge (i.e., CO32"> SO42 > CI'). Also, increasing the anions concentration increases anions repulsion. The volume of the diffuse double layer from which the anions are excluded is highly dependent on the cations valence. For example, the C1- exclusion volume of layer silicate suspensions decreases in the order Na § > K § > Mg z+ > Ca 2+. The more tightly adsorbed cations produce a more condense double layer, so that a smaller volume is affected by anion exclusion.
Figure 13.9. Relative change in hydraulic conductivity as a function of cation type and concentration and anion type (SO42).
In summary, it is evident that for the same concentration of Ca 2+ ions, the order of hydraulic conductivity variations with respect to anion association is: CaCO3 < CaSO4 < CaC1. This order can be explained by considering anion exclusion and ion-pair stability constants. From anion exclusion viewpoint, it is known that anions of higher charge are repelled more than anions of lower charge. Therefore, inter-particle repulsion energies would be less for CI than CO32 ion associations resulting in the formation of flocculated structure in the case of CI and dispersed structure in the case of CO32 ions. Based on ion-pair stability constants, it is known that anions of higher charge are more stable than anions of lower charge. Therefore, the stability of the formed structure would be less in the case of CI than that of CO32 ions.
Organic Chemical Effect The effect of organic chemicals on the hydraulic conductivity of soil-bentonite slurry mixtures was experimentally investigated by Anderson et al., (1985). Soil-slurry mixture composed of calcareous smectitic clay soil mixed with a 9% solution of bentonite was placed in 150 mm diameter double-ring permeameters, as discussed in Chapter 5. The soil-slurry mixture was flushed with water containing 0.01 N C a S O 4 for one month, and then it was allowed to consolidate over a
318
ALTERNATIVES TO GROUNDWATER EXTRACTION
period of four weeks. After consolidation, initial hydraulic conductivity values were measured. Next, specimens were permeated by either xylene or methanol. The reported experimental results by Anderson et al., (1985) were recalculated and presented in Figure 13.10 in terms of relative changes in hydraulic conductivity, Kp/Kw,where, Kpis the hydraulic conductivity of the passing pollutants (i.e., xylene or methanol), and Kw is the hydraulic conductivity of water containing 0.01 N C a S O 4. Xylene is characterized by having molecular weight of 106, dielectric constant of 2.4, and dipole moment of 0.4 while methanol is characterized by molecular weight of 32, dielectric constant of 31.2, and dipole moment of 1.66.
Figure 13.10. Relative hydraulic conductivity of soil-bentonite slurry mixture permeated by xylene and methanol.
The hydraulic conductivity of soil-bentonite slurry mixture increases by approximately 170 fold after passing less than 0.5 pore volume of xylene. However, for the case of methanol, the hydraulic conductivity increased by 180 fold after collecting less than 2.0 pore volumes. Such changes in hydraulic conductivity indicate differences in the initial structure formation of soilbentonite slurry mixture. During the initial structure formation, soil particles tend to be dispersed. When flushed with water containing 0.1 N CaSO4, the thickness of the diffuse ion layer tends to decrease due to the effect of divalent cation and SO42 anion exclusion on the diffuse double layer. This in turn tends to shift the structure towards flocculation, hence an increase in the hydraulic conductivity of soil-bentonite slurry. Further leaching with xylene or methanol reduces the thickness of the diffuse ion layer due to the decrease in pore fluid dielectric constant and the repulsive energy (Yong, Mohamed, and Warkentin, 1992). Consequently, an increase in volume change, or decrease in swelling, is obtained. The results can also be explained on the basis of the relationship between liquid limit and pore fluid dielectric constant. As the dielectric constant increases, the liquid limit increases. The decrease in liquid limit with a decrease in dielectric constant can be attributed to the
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319
aggregation of soil particles, hence an increase in the size of pore channels and, consequently, hydraulic conductivity.
13.2.4 Slurry Trenching Quality Control Parameters The quality control parameters, which are of major concem when constructing a slurry trench, are those related to: (1) slurry quality, (2) trench dimensions and continuity, and (3) cutoff wall composition and placement. These parameters are discussed below.
Slurry Quality The slurry material utilized in both soil-bentonite and cement-bentonite methods is sodium bentonite. This highly plastic clay material has the ability to adsorb nearly 5 times its weight in water and to swell to 10 to 13 times its dry size upon complete hydration. When a sufficient quantity of water is added to bentonite and mixed vigorously, the result is a stable thixotropic gel that has ideal viscosity, density, and filter loss properties for slurry trenching. Thixotropy is defined as an isothermal, reversible, time-dependent process occurring under conditions of constant composition and volume, whereby a material stiffens while at rest and softens or liquefies upon remoulding. As previously stated, the primary function of the slurry is to maintain trench wall stability. The two primary factors cited for trench wall stability are the hydrostaticforce exerted by the slurry and thefilter cakeformation on the walls. Actually, these two factors work together. The weight of the slurry must exert its force not only on the soil pore fluid but also on the soil particles if it is to support the trench walls. The filter cake, which forms when water is squeezed out of the slurry through the trench walls, allows the hydrostatic force of the slurry to be more directly transferred to the walls of the trench. This is why unstable wall conditions are frequently associated with poor cake formation. Filter cake formation is a function of many parameters including clay type, clay concentration, makeup water, formation time, level of bentonite, pollution, quality and type of chemical additives, and hydrostatic pressure. The hydraulic conductivities of bentonite and cementbentonite filter cakes as low as 2.3x 10~1 m/sec and 1.0• 10.9 m/sec, respectively, have been reported (D'Appolonia, 1980; Nash and Jones, 1963). The higher filtrate losses associated with cementbentonite cakes is due to the calcium in the cement, which causes a partial flocculation and aggregation of the bentonite. It may be noted that filtrate loss is, indeed, a measure of slurry stability because the rate of filter loss is directly proportional to the level of bentonite pollution. High filtrate loss can be controlled in a number of different ways (Jepsen and Place, 1985): (1) utilize specially treated pollutant resistant bentonite, which have proven very effective in a number of slurry trenching projects, (2) utilize chemical treatments as dispersants or bulking agents in controlling filter loss, and (3) increase the bentonite concentration in the slurry. In the soil-bentonite method, the slurry may contain between 4 and 7% bentonite by weight. An ideal makeup water should be low in total dissolved salts and have a pH in the range of 5.6 to 10. A poor quality makeup water consumes more bentonite than a relatively fresh water to achieve the same viscosity, density, and fluid loss properties. When introduced to the trench, the slurry should be fully hydrated with an apparent viscosity of 0.015 to 0.020 Pa.s (15 to 20 cP) and have a density of approximately 1.05 Mg/m 3 (La Russo, 1963). In the cement-bentonite method, the self-hardening slurry typically contains 4 to 7% bentonite (by weight), 8 to 25% Portland cement, and 65 to 88% water (Jefferis, 1981). If too little
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ALTERNATIVES TO GROUNDWATER EXTRACTION
bentonite is used, the cement will settle out, causing excessive water bleeding. If too much bentonite is used, the slurry will be too difficult to work with and the wall may be too weak. If too little cement is used, no settling of the slurry will be achieved. The cement, which acts as a binder for the wall, is always added to a fully hydrated bentonite-water slurry just prior to being introduced in the trench. This requirement, called for in most specifications, minimizes the viscosity of the slurry in the trench during excavation so that excessive suspended solids from the soil will not be entrained in the slurry when it eventually hardens. When the cement-bentonite slurry is introduced in the trench, it should have a density of approximately 1.09 Mg/m 3 and a Marsh funnel viscosity ranging from 30 to 50 s (Jefferis, 1981; Ryan, 1980).
Cutoff Wall Composition Soil-Bentonite Cutoff Walls: Most soil-bentonite cutoff walls are formed with the soil excavated from the trench. However, borrowed soil materials are used in the following situations: (a) if the soil is too polluted to be mixed with bentonite, (b) if the soil is too coarse to meet low hydraulic conductivity requirements at economical bentonite application rates, and (c) if the soil is too fine to meet the necessary strength requirements of the wall. In trenches excavated through soil highly polluted with volatile organic constituents, there are also air quality considerations that may eliminate the excavated soil as a viable constituent in the backfill mix. The typical soil-bentonite backfill consists of 2 to 4% bentonite (by weight of total mixture), 25 to 35% water, with the balance being soil (D'Appolonia, 1980). The hydraulic conductivity of soil-bentonite cutoff walls ranges between 10 -9 and 10~~ m/s. The hydraulic conductivity of the backfill is dependent on the gradation of the soils being used, the bentonite content, and the type of pollutants that interact with the backfill. The greater the percentage of fines (soil particles smaller than 7.5x 10.5 m) in the backfill and the more plastic those fines are, the lower the wall hydraulic conductivity will be. The greater the percentage of bentonite in the backfill, the lower the wall hydraulic conductivity will be. In addition, the effect of pollutants on the resultant wall hydraulic conductivity is of great importance, as discussed in the previous sections. Cement-Bentonite Cutoff Walls: The quantity of suspended solids that are entrained in the cement-bentonite slurry should be kept to a minimum. Excessive suspended particles (particularly sand) tend to increase the wall hydraulic conductivity and increase wall strength, which can make it more brittle and susceptible to cracking. Tests have shown, however, that suspended solids are not an important factor in the hydraulic conductivity of cement-bentonite cutoff wall until the concentration is high enough to permit inter-particle contact. This can be reached at a percentage (by volume) in the range of 10% (Ryan, 1980). The initial mix is very plastic and can withstand compressive strains of several percent without cracking. Low strength, high plasticity, and low hydraulic conductivity (1.0• 108 m/s) are all qualities of cement- bentonite cutoff walls that are associated with the rather large quantities of bentonite in the slurry mix, 4 to 7% by weight. The soil-bentonite method has several advantages, aside from the lower cost, over cementbentonite. These include: (a) the resultant wall is generally of lower hydraulic conductivity than cement-bentonite walls, (b) the backfill can have various materials blended in to suit design conditions, (c) soil-bentonite backfill is generally more resistant to degradation by most pollutants, and (d) the excavated material can be mixed as backfill and placed back into the trench. Given the relative advantages of the two systems, the project requirements should be
GROUTING
321
evaluated to determine the best method to be selected.
13.3
GROUTING
Grouting is a commonly employed ground modification technique used to strengthen the ground, reduce hydraulic conductivity of fill voids (Hausmann, 1990). In geoenvironmental engineering applications, grouts can be used to: (a) consolidate polluted soils via shallow low pressure injection, (b) provide for solidification via injection into the waste, and (c) form a barrier to lateral or vertical pollutant migration via injection to seal the soil around the site. The critical issues in grouting are: (a) the grout must set or harden in contact with waste components, and (b) the grout must not deteriorate in the presence of waste during normal temperature or moisture cycles within the expected life time of the grouted structure. 13.3.1 Types of Grouts In general, grouts can be divided into two main categories, namely, chemical grouts and particulate grouts. Chemical grouts include bitumen, silicates, lignochromes, phenolic resins, acrylic resins, aminoplasts and urethanes. They react to produce a gel or polymer that fills the pore space. It has low initial viscosity that increases rapidly during grouting. Particulate grouts, on the other hand, are typically composed of cement, lime, clay, fly ash or some combination of these materials. Setting of particulate grouts is provided via chemical reactions and flocculation of the dispersed solids. Micro-fine cements are being increasingly used, in lieu of ordinary cement grouts, because of their ability to penetrate fine materials. The selection of the grout to be used for a specific task will depend on the compatibility of the grout with the chemicals that are present in the ground, the ease with which it may be injected, and its ability to adequately seal the soils. The chemicals present can affect the setting time of the grout and its long term hydraulic conductivity.
Chemical Grouts Silicate Grouts: They are composed of a sodium silicate base, a reactant, an accelerator, and water. The reactant is typically an amide, an acid, or some polyvalent cation. A salt, such as calcium chloride or sodium aluminate, is used to accelerate the set or gel of the grout. Sodium silicate, nSiO2.Na20, is commercially available as an aqueous solution. The silica/alkali ratio, n, is important in that ranges of 3 to 4 yield gels with adhesive properties particularly suitable for grouting. If n is greater than 4, the silicate becomes unstable. The concentration of the grout used depends on the material into which it is to be injected. Soils with 20% sand are considered to be the lower limit of injection (Bowen, 1975). When sodium silicate solution and a concentrated solution of appropriate salt are mixed, the gel-forming reaction, as indicated by Eq. [ 13.1 ], is virtually instantaneous.
Si02.Na 20 + CaCl 2+ H 20 -. Ca(09)2
+
Si02.2NaCl
[13.1]
Acrylamide Grouts: They are a mixture of organic monomers which could be polymerized at ambient temperature (Bowen, 1975). The basic composition of acrylamide-based grouts is: (1) Acrylamide (or methanol acrylamide, methacrylamide, etc.) which is generally 95% of the
322
(2)
ALTERNATIVES TO GROUNDWATER EXTRACTION mixture and will polymerize into long molecular chains; and 5% cross-linking agent, such as methylenebis-acrylamide, which binds the acrylamide chains together.
Gel consists of 8% to 12% solids, in the form of long molecular chains randomly cross linked to each other, and 88% to 92% by weight of water molecules which are mechanically trapped in the "brush-heap" structure of the gel. In gel placement under water, no volume change is expected. However, gel placed in an arid environment will lose water and can shrink to the volume occupied by its solids, about 10% of its original volume. The critical issue is the migration of ions through the water in the gel. Soluble salts that are added to a grout solution may be expected to leach out with time. Also, gels made with pure water can be expected to absorb salts from a saline environment.
Lignoehrome Grouts: They consist of lignosulfonate or lignosulfite and hexavelant chromium compound. Lignosulfonates and lignosulfites are by-products of the wood pulp processing industry and, depending on the extraction process used, can be salts of calcium, ammonium, or sodium, as discussed in Chapter 4. For grout use, sodium lignosulfonates are unstable while calcium lignosulfonates provide the best water proofing and the most stability (Tallard and Caron, 1977). Lignochrome grouts set up in an acid medium (i.e., lignosulfonic acid) where the highly toxic hexavelant chromium is reduced to the less toxic trivalent form. The set time of the grout is controlled by varying the concentrations of both the chromate and the accelerator (usually a metallic salt). Phenoplasts: They are chemical grouts formed by the polycondensation of a phenol and analdehyde (Tallard and Caron, 1977). The reaction of these two compounds to form the set grout is achieved by an increase in the pH, with a minimum set time obtained at a pH of 9.3. Set time can also be controlled by varying the dilution of the grout, with more dilute solutions taking longer to set. Aminoplasts: The major grout in this class is urea-formaldehyde. If forms a resin by a condensation reaction caused by heat in an acid medium (Tallard and Caron, 1977). Because of the requirement of heat and a low pH, and their high viscosity, urea-formaldehyde grouts are not often used. Particulate Grouts
Portland Cement: It is an extremely popular construction material, and one of the first to be used as a grout. When mixed with water, it sets up into a crystal lattice in less than two hours. For grouting, a water-cement ratio of 0.6 or less is more effective (Bowen, 1975). The smallest voids that can be effectively grouted are no smaller than three times the cement grain size. A more finely grouted cement makes a more water tight grout (Bowen, 1975). Portland cement is often used with a variety of additives that modify its behaviour. Among these are clays, sands, fly ash, and chemical grouts. Bentonite Grouts: Due to its extremely small particle size, bentonite is the most injectable, and thus the best suited for grouting (Tallard and Caron, 1977). Bentonite grouts can be injected into
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323
materials with lower hydraulic conductivities than can other suspension grouts. Medium to finetextured sand, with hydraulic conductivities of around 10.5 to 10.6 m/sec, can be sealed with a bentonite grout. Dry bentonite is mixed with water on site at a composition of 5 to 25% by dry weight (Bowen, 1975). In these ratios, bentonite will adsorb large amounts of water and, with time, form a gel. This gel, although it imparts little if any structural strength, is an extremely effective water barrier. Chemical compatibility of bentonite with organic and inorganic chemicals were discussed in previous sections. 13.3.2 Compatibility of Chemical Grouts with Pollutants
Grouted specimens with sodium silicate were subjected to hydraulic gradients as high as 500 for periods of up to 60 days and the amount of leached grout was measured (Caron, 1965). From these results, it was concluded that grout is more likely to be retained in fine soils due to the fact that small voids support grout better than large voids. The amount of leached sodium was also found to decrease with higher gradients due to the reduced contact time, and the apparent hydraulic conductivity increased initially and then decreased. Also, Baker (1982) reported that high pH leachate can be detrimental to sodium silicate-based grouts, either by preventing gelation or by promoting deterioration with time. Low pH leachate may accelerate the gelling of sodium silicatebased grouts. However, alkaline water normally contains large amounts of calcium carbonate and this may result in calcium ions reacting with the silicate to form a hard calcium silicate compound. Hydraulic conductivity under these conditions may increase slightly, but the strength of the grout matrix will remain intact (Krizek and Madden, 1985). It should be emphasized that silicate based grouts cannot be injected into formations containing sea water unless a buffer solution is added, because the magnesium ions in the water will react with the silicate and give undesirable results. When silicate grouts were mixed with: (a) ethyl acetate and formamide, (b) Hardener 600, (c) Terraset, and (d) Geloc 4 as reagents, all specimens exhibited large increase in hydraulic conductivity (Krizek and Madden, 1985). The rate and magnitude of the increase was influenced by the curing time of the specimens and the magnitude of the applied hydraulic gradient. Specimens with short curing times exhibited complete leaching of the grout due to inadequate gel strength. Longer curing times resulted in a gradual decrease in hydraulic conductivity without complete leaching of the grout. The effect ofpH and hardness on the hydraulic conductivity of sodium silicate-ethyl acetateformamide grout was studied by Siwula and Krizek (1992). It was found that the hydraulic conductivity decreased as the pH of the permeant decreased, particularly in the case of specimens cured for periods of 1 hour or less. The authors have attributed the effect of pH on the resultant changes in hydraulic conductivity to gel dissolution. Water hardness affects hydraulic conductivity by altering the solubility of the silicate. Calcium and magnesium tend to form insoluble calcium and magnesium silicates. For curing times greater than one hour, water hardness does not cause a significant difference in hydraulic conductivity. In general, the effect of the curing time has a much greater effect (up to about four orders of magnitude) on the hydraulic conductivity of the sodium silicate-ethyl acetate-formamide grout than either the water pH or hardness. In the studies reported by Petrovsky (1982), various percolating fluids were used to evaluate the potential leachability of cement grout. These fluids were characterized as soft water (carbonate hardness between 12 and 75 ppm) with pH values between 6.8 and 8.6, and a maximum average
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ALTERNATIVES TO GROUNDWATER EXTRACTION
sulfate concentration of the water of 25 ppm. Based on the experimental results, three states of leaching were identified: (a) The rapid dissolution of the free calcium hydroxide (CaOH2), which resulted in the highest concentration of calcium in the leachate during the leaching process; (b) The hydrolytic action of the percolating water which released lime and caused the calcium concentration of the seeping water to be reduced; and (c) The dissolution of calcium hydroxide and the decomposition of hydrated grouts. The effect of various waste components on the grout setting time is shown in Table 13.2 (Karol, 1983). In order to further evaluate various chemical grouts, a relative ranking scheme developed by Karol (1985) is shown in Table 13.3. In summary, the following three issues are the basic criteria for indicating which chemical grout is the best under a set of environmental conditions: (1) have a set time that can be regulated, (2) be reasonably non-corrosive to mixers and pumps, and (3) be formulated from materials that are low in toxicity.
13.3.3 Grouting Methods Penetration, displacement, and jet grouting are different grouting methods that are useful at landfills and hazardous waste sites. These methods are described as follows.
Penetration Grouting Penetration grouting involves the injection of a particular slurry or a fluid into the ground to fill the inter-particle voids and fissures. One use of penetration grouting is to make grout curtains or to seal the bottom of the site (US EPA, 1983). The biggest problem with using penetration grouting to reduce the hydraulic conductivity is that the placement of the grout cannot be controlled well enough to ensure a complete seal. When penetration grouting is used to reduce the hydraulic conductivity of the rock or soil mass then it is important to have adequate overlap of the grout bulbs to prevent holes in the barrier. The success of the grouting program will depend on the initial site characterization and on the selection of a suitable grout, drilling equipment and procedure (Mitchell, 1981). Penetration grouting has been useful for stabilizing landfills and mine tailings, and for providing in-situ remediation. Use of fly ash as a grout has been successful in stabilizing landfills (Moulton et al., 1977) and for stabilizing waste piles from coal mining (Bowders and Almes, 1990). Lime and fly ash injections have also been used to stabilize landfills. Dredged materials and sewage sludge can also be injected into landfill for stabilization or to enhance the biological activity and methane generation.
Displacement Grouting Displacement grouting is a process which uses a highly viscous material for compressing the surrounding fill or soil. This grouting technique is useful for filling void space in municipal waste landfills to improve the utilization of the landfill capacity, stabilize the fill, and enhance the biological activity in the wastes. The void space may be injected with a variety of waste materials.
Vl
GROUTING
P J
m
Table 13.2: Effect of pollutant on the grout setting time.
I
Character of pollutant
Conc.
Acrylate
30% silicate
50% Silicate
Urathane
Portland cement
. .
Strong oxidizer
10%
Inhibited
OK
OK
Inhibited
OK
Inorganic acid
10%
Inhibited
OK
OK
Inhibited
OK
Base
10%
Inhibited
OK
OK
Inhibited
OK
. ..
Base
10%
Inhibited
Inhibited
Slow
Inhibited
OK
Salt
10%
Inhibited
Flash set
Flash set
Inhibited
Slow
Salt CHC HCM HCM SB SB Halogenated
10% Sat. Sat. Sat. Sat. Sat. Sat.
Inhibited Inhibited Inhibited Inhibited Inhibited Inhibited Inhibited
Flash set OK OK OK OK OK OK
Flash set OK OK OK OK OK OK
Inhibited Inhibited Inhibited Inhibited Inhibited Inhibited Inhibited
Flash set OK OK OK OK OK OK
Pollutant components
I
Type of grouts
Characteristics of pollutant test solution used
.
Potassium chromate Hydrochloric acid Ammonium hydroxide Sodium hydroxide Ammonium chloride Copper sulfate Benzene Gasoline Oil Phenol Toluene Trichlorotethlene
I
I I
CHC= cyclic hydrocarbon; HCM= hydroci on mix.;
I= substitute(
lenzene; Sat.=saturated.
ALTERNATIVES TO GROUNDWATER EXTRACTION Table 13.3: Relative ranking of solution grouts as to their toxicity, viscosity and strength (Karol, 1985).
Grouts
. . . . . .. .
Silicates
Lignosulfonates
Phenoplasts Aminoplasts Acrylamides
Polyacrylamide Acrylate
Polyurethane
Corrosivity or toxicity
Viscosity
Strength
Low Medium Low High High Medium Medium Medium Medium High High High Low Low Low Low Low High High High High High
High Medium Medium Medium Medium Medium Medium Medium Medium Low Low Low High Low Low Low Low High High High High High
High Medium to high Low Low Low Low Low High High Low Low Low Low Low Low Low Low High High High High High
~~
0 Joosten process 0 Siroc 0 Silicate-bicarbonate 0 Terra Firma 0 Blox -all 0 Terranier 0 Geoseal 0 HerculOx 0 Cyanalog 0 AV- 100 8 Rocagel BT 0 Nitto-SS 0 Injectite 80 0 AC-400 0 Terragel 0 Flexigel 0 Durigel 0 CR-250 0 CR-260 0 TACSS 0 CG5610 @ AV220
SHEET PILING
327
A higher landfill capacity is obtained because more of the available volume would be filled with waste rather than air.
Jet Grouting Jet grouting is the use of a very high pressure water jet to erode soil for in-situ mixing. It can be used to build hydraulic barriers for containment of polluted plume. The vertical barriers are built as a line of overlapping columns. Construction of horizontal barriers for sealing the bottom of the site is also possible using this technique, allowing for complete isolation of the polluted area.
13.4
SHEET PILING
Sheet piling can be used to form a groundwater barrier. Sheet piles are made of wood, precast concrete, or steel. Wood is ineffective as a water barrier. Concrete sheet piles are used primarily where high strength is required. Steel is the most effective in terms of groundwater cutoff and cost. In the construction of a sheet piling cutoff, the piles are assembled at their edge interlocks before they are driven into the ground. This is to ensure that earth materials and added pressures will not prevent a good lock between piles. The piles are then driven a known distance at a time over the entire length of the wall. This process is repeated until the piles are all driven to the desired depth. Starr et al., (1992) present the recent development in specialized steel sheet piling with sealed joints for containment of polluted plumes. A sealable joint provides a cavity which can be filled with a sealant after driving. The joint may be modified such that a double capacity configuration may be used for additional effectiveness. Work is currently progressing on a tool to log the size of the cavity. Field tests have been carried out using both single and double walls for the containment of polluted plumes. The results indicate that the hydraulic conductivity ranged from 10.9 to 10-12m/see across the barrier. A principal advantage of this technique is the readily available installation equipment and the relative ease of installation. The disadvantages of sheet piling are that it is limited to areas where the subsurface conditions are relatively free of boulder and bedrock. In addition, the cost of materials for sheet pile walls is higher than for other types of walls. Steel corrodes because of an inherent tendency to revert to its natural state. Almost all metals exhibit this characteristic. The amount and extent of corrosion depends on the environment in which the sheet piling wall is functioning. The overall performance of sheet pile structure is judged by its weight loss and/or by a corresponding reduction in thickness. Extending the life of steel sheet piling can be done by applying protective coatings such as coal-tar epoxy. Another proven method for protecting steel from corrosion is cathodic protection, which uses direct current.
13.5
GROUND FREEZING
Ground freezing can be used as a temporary measure to contain polluted groundwater. In saturated soils, ground freezing forms a semi-permeable wall as ice crystals form and fill the pore spaces (Iskander and Jenkins, 1985). If the soil is not saturated, then the injection of additional water may be necessary to form ice barrier. The freezing pipes are installed around the plume to be contained and angled toward the centre to get under the plume for complete containment (ASCE
328
ALTERNATIVES TO GROUNDWATER EXTRACTION
News, 1991). The freezing front will spread out from the pipes until a complete wall is formed. The front may eventually spread all the way to the centre of the site and form a solid, frozen mass.
13.6
ELECTROKINETICS
When electrical charges are generated at the solid liquid interface, a diffuse layer of electric charges is formed on the liquid-phase side of the interface, as discussed in Chapter 5. If an external electric field is applied to this system, a relative movement of the two phases will take place with respect to each other. This is referred to as electrokinetic phenomena. Electroosmosis, electrophoresis, and electrolysis are examples of these phenomena, which are discussed in Chapter 19. A brief description follows.
13.6.1 Electroosmosis Electroosmosis refers to the movement of soil moisture or groundwater from the anode to the cathode. Electroosmotic transport in a soil medium depends on the following factors: (1) mobility of the ions and charged particles within the soil moisture or groundwater, (2) hydration of the ions and the charged particles, (3) charge and direction of the cations and charged particles, (4) ion concentration, (5) viscosity, (6) dielectric constant, and (7) temperature.
13.6.2 Electrophoresis Electrophoresis refers to the movement of soil particles within the soil moisture or groundwater under the influence of an electrical field. This definition includes all electrically charged particles like colloids, clay particles floating in the pore solution, organic particles, droplets, etc. The mobility of these particles is similar to that of ions. Within the pore solution, these particles transfer electrical charges and affect electrical conductivity and the electroosmotic flow. Clay particle as such have two electrical polarities. One consists of the structure-based dipole moment, which depends on the atomic mass and has an orientation parallel to the longest axis of the clay particle. The second polarity stands at right angle to the first and is caused by external electrical field. It depends on the manner of polarization of the electrical double layer. The mobility of clay particles is an interplay between these two moments and is, therefore, less than the electroosmotic mobility which varies between 10.9 and 101~ m2/V.s.
13.6.3 Electrolysis Electrolysis refers to the movement of ions and ion complexes within the soil moisture or groundwater. In electroosmosis and electrophoresis, one considers only water transport and particle transport, respectively. In electrolysis, the movement of ions and ion complexes is taken into consideration. The average mobility of ions lies around 10.8 m2/V.s, which is one order of magnitude greater than that of electroosmotic mobility. Therefore, the energy necessary to move all ions over an average distance of 1 m through a cross sectional area of 1 m 2 of soil is one order of magnitude less than that required in electroosmosis.
REACTIVE BARRIERS
329
Figure 13.11. Concept of electrokinetic containment barrier.
Electrokinetics could potentially be used for waste management in the following ways: (1) de-watering of waste sludge slimes, dredged sediments by first concentrating the solid particle using electrophoresis and then consolidating it by electroosmosis, (2) electroosmotic flow barriers, (3) leak detection systems for disposal facilities, (4) injection of grouts to form barriers, (5) provide nutrients for biodegradation process, (6) in-situ generation of reactants, such as hydrogen peroxide, for cleanup and/or electrolysis of pollutants, and (7) decontamination of polluted soils and groundwater. As indicated before, electrokinetics can be used for subsurface containment of pollutants. Figure 13.11 illustrates the concept of electrokinetic hydraulic flow barrier (Mitchell and Yeung, 1991). In practice, the vertical barrier would be made by installing a wire mesh and periodically applying a small electric gradient across the soil medium to generate an electroosmotic flow in the direction opposite the flow of polluted plume
13.7
REACTIVE BARRIERS
The concept of using permeable in-situ reactive barriers to treat polluted plume as it moves through an aquifer under natural hydraulic gradient, as shown in Figure 13.12, was first suggested by McMurty and Elton (1985). It has received significant attention from the research community (Blowes and Ptacek, 1992 and 1994; Blowes et al., 1994; Starr and Cherry, 1994). The great promise of in-situ reactive barriers is that they require little or no energy input once installed. The main engineering challenges involve determination of suitable type and amounts of reactive materials in a permeable wall and proper placement techniques.
330
ALTERNATIVES TO GROUNDWATER EXTRACTION
Figure 13.12. Schematic diagram showing porous reactive wall for treating polluted groundwater.
13.7.1 In-Situ Reactive Zones
In situ reactive zones are installed by excavating a portion of the aquifer material and replacing it with a porous reactive mixture over the desired depth of treatment. Above the zone of treatment, aquifer material or other backfill is used to fill the excavation to ground surface. The reactive medium is installed in a zone down-gradient from the pollutant source. The distance between the pollutant source and the reactive zone can vary depending on site conditions and the composition of the reactive media in the wall. Altematively, reactive zones can be installed as impermeable barriers containing permeable zones or windows, as in funnels-and-gates system (Starr and Cherry, 1994). Preliminary results of laboratory and small-scale field experiments (Blowes and Ptacek, 1992 and 1994; Blowes et al., 1994) suggest that, at some sites, in-situ reactive media have the potential to greatly reduce the expense associated with plume control. Because reactive zones are installed down-gradient from the source of groundwater pollution, natural hydraulic gradients carry the polluted groundwater through the treatment zone. This in tum avoids the need for pumping the polluted groundwater to ground surface, where problems and costs are incurred. For the in-situ reactive system to be effective, three basic requirements must be met (Blowes et al., 1995): (1) The pollutant must be sufficiently reactive such that suitable transformations take place during the time the polluted groundwater flows through the treatment zone. Without sufficient reaction rates, only partial transformation and removal of the pollutant will occur, leading to incomplete treatment or, in some cases, formation of by-products that are less desirable than the original pollutants; (2) The reactive media in the treatment zone should be sufficiently abundant and have persistent reactivity so that it performs well for an economically viable period of time; and (3) The reactive media itself should not release additional pollutants that would be unacceptable in groundwater on the down-gradient side of the treatment zone.
REACTIVE BARRIERS
331
13.7.2 Types of Reactive Media Zero-valence metals Zero-valence metals are those which do not have any charges on the surface, hence they are in their elemental form. They are used for: (1) Enhancing the degradation of chlorinated organic compounds (Senzaki and Kumagai, 1988; Senzaki, 1991; Gillham and O'Hannesin, 1992, 1994; Lipezynska-Kochany et al., 1994; Matheson and Tratnyek, 1994); and (2) Removal ofhexavalent chromium (Cr (VI)) (Blowes and Ptacek, 1992, 1994; Gillham et al., 1994).
To enhance the degradation of chlorinated organic compounds, a permeable reactive wall, consisting of a mixture of 22% (w/w) granular iron (zero-valence iron) and 78% (w/w) sand, was installed to treat a polluted plume containing 250 mg/1 trichloroethylene (TCE) and 43 mg/1 of tetrachloroethylene (PCE) (Gillham et al., 1994). The experimental results showed that the concentrations of both TCE and PCE decreased rapidly as the plume entered the reactive wall, with the rate of decline decreasing with travel distance across the wall. Approximately 90% of the TCE and 86% of the PCE were removed within the reactive wall. The concentration in the effluent was well above drinking water limits. However, Gillham et al., (1994) have indicated that greater removal and possibly complete remediation could have been achieved if a higher percentage of iron had been used in the reactive mixture. About 10% of the initial TCE and PCE appeared as chlorinated degradation products (i.e., dichloroethane (DCE) isomers), and vinyl chloride was not detected as a by-product. There was no apparent change in the performance of the reactive wall over a 16 month period of monitoring, and core samples collected 24 months after installation indicated no substantial change in the physical characteristics of the reactive material. The experimental results suggest great potential for the use of zero-valence iron for in-situ remediation of groundwater polluted by chlorinated organic chemicals. Though the process of degradation is widely accepted as reductive dechlorination, details of the mechanism remain uncertain. More research studies are in progress. The use of zero-valence iron to remove hexavalent chromium (Cr(VI)) from waste stream was reported by Blowes and Ptacek (1992 and 1994). In their experiments, fine grained iron filings were mixed with a small percentage of calcite and coarse grained quartz sand. The calcite was used to maintain the moderate pH conditions, needed for the formation of the insoluble oxyhydroxide solid, while coarse grained quartz sand was added to maintain a high hydraulic conductivity. The mixture was used to treat more than 150 pore volumes of a 20 mg/1 Cr(VI) solution at a flow rate of 8 m/year. The principle behind the removal of hexavalent chromium (Cr(VI)) is its reduction, using solid phase zero-valence iron (Fe~ to Cr(III) which is subsequently precipitated to form a mixed Cr(III)-Fe(III) oxyhydoxide. These combined reactions lead to extremely low aqueous concentrations of Cr(VI). At pH greater than 5, concentration of Cr(III) are also very low. The resulting concentrations of the total Cr are less than the recommended drinking water limits established for Cr.
332
ALTERNATIVES TO GROUNDWATER EXTRACTION Other Reactive Materials
Other reactive materials are: (1) those which promote the precipitation of insoluble metal hydroxides and carbonates through a simple increase in pH; and (2) those in which ion exchange or adsorption reactions are promoted through the addition of additives. For example, McMurty and Elton (1985) proposed the use of reactive barriers containing pH-neutralizing materials to remediate acidic groundwater containing high concentrations of Cr(III).
13.7.3 Engineering Aspects of the Reactive Barrier System Installation Method
There are a number of ways in which a reactive wall can be installed at a field site. The wall can be installed by first inserting steel sheet piling into the formation, followed by excavation of the native aquifer materials, insertion of the reactive medium, and the withdrawal of the sheet piles to allow groundwater to flow through the reactive medium. The replacement material may consist of a single reactive medium, or a mixture of materials. If appropriate, local aquifer material can be blended together with the reactive materials to reduce installation costs. The primary physical constraint is that the hydraulic conductivity of the mixture be similar to or greater than that of the native aquifer materials. This requirement prevents groundwater from by-passing the reactive material.
Figure 13.13. Aerial view schematic of a funnel-and-gate system.
Alternative installation methods include slurry walls excavated using biodegradable materials. These slurries later biodegrade, allowing the hydraulic conductivity to be regained and normal flow of groundwater to return. Reactive media may be installed through the use of in-situ techniques that blend the reactive material with the aquifer material in the zone through which the polluted groundwater flows. This latter method is less likely to ensure the creation of a continuous
SUMMARY AND CONCLUDING REMARKS
333
zone in the aquifer, and may lead to incomplete treatment of the pollutants. Any installation technique that allows replacement of aquifer material with the reactive materials to form a relatively continuous treatment zone without destroying hydraulic conductivity is likely to be suitable.
Installation Design Reactive zones can be installed as a broad, continuous curtain that extends the entire width of a pollutant plume or it can be installed as one or more discrete zones joined to segments of impervious walls. This configuration is known as a funnel-and-gate system, as shown in Figure 13.13. The impervious wall segments direct the plume through the gaps where the permeable medium is placed (Starr and Cherry, 1994). In both cases, the wall can either be installed through the entire aquifer depth or it can be installed slightly deeper than the maximum depth of the polluted zone. The geochemical and physical conditions at a site will determine the stability of a given design.
Continuous Curtain: Continuous reactive curtain is desirable at sites where installation is relatively inexpensive, where the pollutant plume is relatively narrow in depth, and the treatment media is inexpensive. Continuous walls have the advantage of minimizing the groundwater flow velocity through the treatment zone, and maximizing contact with individual grains of reactant in the wall. With the funnel-and-gate system, the presence of low hydraulic conductivity sections leads to increased velocities within the treatment zone (i.e., the gates). Funnel-and-Gate System: There are a number of methods that can be used to install fiameland-gate systems. The impermeable zones can be constructed of impermeable materials, such as sheet piling, or low hydraulic conductivity materials, such as bentonite slurry. The primary advantage of the funnel-and-gate system is improved control of reaction zone. Figure 13.13 shows an example of a funnel-and-gate system where sheet piles are installed to form impermeable sections of the wall (funnels), and openings containing reactive mixtures are constructed within the wall (gates) (Starr and Cherry, 1994). The openings are unique in that they are constructed by inserting sheet piling perpendicular to the funnels. The reaction gates can be filled with a single reactive media, or, if needed, different media can be installed in series (Figure 13.13). The funnel-and-gate length ratio can be adjusted to optimize residence times within the gate sections. To increase the residence time within the gate, the length of the gate in a direction parallel to groundwater flow should be increased. This in turn will increase the contact time with the reactive mixture. The final design of the system requires characterization of the physical hydrogeology and chemistry of the site, and an understanding of the rates and nature of reactions taking place in the treatment zone.
13.8
SUMMARY AND CONCLUDING REMARKS
Subsurface barrier systems are essentially concerned with the control of movements of polluted groundwater. Water samples taken from appropriate locations provide an apparently obvious means of monitoring performance. Depending on the purpose of the barrier, water samples may need to be taken from only one or both sides of the barrier. An example of the latter is when the aim is a controlled net flow of water into the isolated area in order to determine whether any unwanted reversal of flow is occurring. Water sampling from inside or outside the barrier is not as straight forward as it might appear. Identification of the proper parameters to be analysed and
334
ALTERNATIVES TO GROUNDWATER EXTRACTION
determination of appropriate locations for sampling are difficult tasks. The large number of samples to be analysed may also be a problem. The main difficulty, however, is ascertaining whether the analysed pollutants came from within the encapsulated area or was already present outside the barrier. This will greatly impact the decision to repair a failed section of a barrier system. In assessing the long term performance of subsurface barriers, it is necessary to address the following issues: (1) will there be any inherent and unavoidable changes in material properties?, (2) to what extent are such changes affected by external factors such as interaction with pollutants?, (3) at what rate will the changes occur, and does the rate matter?, and (4) to what extent is the system sensitive to any physical damage?
CHAPTER
FOURTEEN
COVERING SYSTEMS
14.1
INTRODUCTION
Pollutants generated from hazardous wastes can be transported through groundwater, surface water and air. The plume generated via groundwater transport could by contained by limiting the recharge of the groundwater. Generally, blocking the transport routes, hence permitting on-site containment of the wastes is principally achieved through three common techniques, namely, inplace containment, secure burial ceil, and neutralization. In-place containment is certainly the most acceptable and economical solution to remediating a hazardous waste site when appropriate soils are available and the substrata are sufficiently impermeable to the flow of groundwater. The movement of groundwater and surface water through the waste and the potential for erosion are thus controlled, ultimately preventing, the migration of waters from the site. Groundwater can be controlled by barrier systems and by leachate and groundwater extraction systems. Surface water is controlled by diverting runoff around the landfill, grading it to control drainage, and covering it to prevent percolation through the site or volatilization from it. Clearly, instituting a well-designed cover system is an effective means for controlling surface water and, ultimately, enhancing the protection of the waste disposal site.
14.2
FUNCTIONS OF COVERING SYSTEMS
The objective of covering systems is to reduce or eliminate the transport of fluids through the waste. Such fluid transport may produce a leachate seepage which poses a threat to water quality. A covering system is also intended to prevent wind dispersion of water particles and hence reduce the threat to water quality. In addition, depending on waste characteristics, the intended function of the covering system will be different. For example, the principal concern in the decommissioning of mine tailings dump sites is the potential for generation of acid leachate. The problem is common in metal and coal mining. Most base metal, precious metal, and uranium mines contain sulphide minerals either in the ore or the surrounding waste rock. The sulphide minerals are unstable and oxidize when exposed to oxygen and water, yielding sulphuric acid. Sulphuric acid increases the solubility of heavy metals and promotes their mobility. Therefore, protection of the contained mine tailings against water and oxygen entry calls for engineered soil covers or some other substitute materials which are impermeable to both oxygen and water. Generally, coveting involves the partial or complete isolation of the waste material from the surrounding environment. The key design elements of this technology include: (1) Covers which overlie the waste and extend to the underlying natural or engineered containment system; 335
336
(2)
COVERING SYSTEMS
(3)
Surface water handling systems, such as dikes and channels, to prevent runoff contact with waste; and Augmentation of the existing containment system by cutoff walls, as discussed in Chapter 13, constructed around the facility, or by providing surface drainage that prevents groundwater from contacting the waste.
14.3
TYPES OF COVERING SYSTEMS
Covering systems vary according to the site and waste specific conditions. Also of importance to the selection of the covering system is the project status. Coveting system design must be selected based on practicality and suitability of actual site condition. In such instances, data on the facility design, operation, or performance may be incomplete. It may be necessary to undertake a field investigation, as discussed in Chapter 7, prior to the design of the covering system. On the other hand, active sites generally have sufficient data available to facilitate selection of appropriate covering system technologies. Adequate information on design, construction, operation, waste characteristics, and facility performance is usually available. New disposal facilities offer the best opportunities to develop a cost-effective closure strategy during the planning phase, and to integrate this plan with the design of the containment system and other operational waste management considerations. The important factors that emerge from examination of the status of the site are the design elements of the cover system. Regarding the design elements, there is a significant difference in the covering system design requirements for waste management units in arid climates and those in humid climates. In arid climates there is little potential for leachate formation. Runoff from waste management facility is usually not an issue, and no significant threat is posed to the ground or surface water. Closure requirement can usually be met by limiting access and leaving waste piles inplace. However, if the waste generates dust, stabilization may be required. In humid climates and where the waste has the potential to generate a leachate that poses a threat to the ground or surface water, more comprehensive closure measures are required. Cover and/or revegetation is usually required for erosion protection and infiltration control. In extreme cases, waste piles may require multi-layer covers to limit infiltration. In regions where substantial freezing occurs during the winter periods, the effects of freezethaw cycles on the performance of high clay content materials can be destructive. Because of the interactive forces in the clay material, pore space readjustment occurs after a few cycles of freezethaw, creating large pore spaces with a high degree of continuity. This, in essence, destroys the design requirement of a low permeable material (Mohamed, 1997c, Mohamed et al., 1993 a, b, c). Slurried tailings impoundments offer a different kind of problem because of the amount of unconsolidated (soft) material left behind. Methods to improve the surface conditions, as well as the shape of the surface, include managed deposition to create a mounded surface, and selectively depositing the coarser fraction of the tailings in the hollow areas.
SOIL-BASED COVERING SYSTEMS 14.4
337
SOIL-BASED COVERING SYSTEMS
As discussed in Chapter 3, any landfill undergoing closure must be covered with a final cover that minimizes long term migration of liquids through the closed landfill (US EPA, 1989). In addition, it must function with minimum maintenance, promote drainage, minimize erosion of the cover, accommodate settling, and have hydraulic conductivity less than or equal to that of any bottom liner system or natural soil present. Coveting system design components, for hazardous waste site closure, are shown in Figure 14.1 (US EPA, 1989). Many of the layers shown in the figure are composed of soils or have soil components. Each layer has a distinct purpose. The materials must be selected and the layers designed to preform the intended functions.
Figure 14.1. Covering system design components for hazardous waste site closure.
(1) (2)
(3) (4)
(5)
The basic components of multi-layered covers are (US EPA, 1989): A top layer, consisting of a vegetated soil layer to minimize erosion and oxygen transport, A drainage layer that minimizes water infiltration into the waste or the infiltration barrier located under it; A one- or two-component infiltration barrier layer (low hydraulic conductivity flexible membrane liner, FML/soil), that limits water infiltration into the underlying wastes; Biotic barriers to prevent damage to the infiltration barrier by burrowing animals and plant roots; and Foundation layers to support a cover and/or provide the cover shape necessary for control of surface runoff and intemal drainage.
The selection and design of the above components must meet the closure objectives and incorporate the site-specific conditions, i.e.,
338 (1) (2)
(3) (4)
COVER/NG SYSTEMS
Climatic conditions that determine the amount of infiltration and the surface water erosion potential; Waste characteristics, such as the physical and chemical properties which influence the potential quantity and quality of leachate; The existing containment system and its performance history that determines if there is a need for additional containment and also can dictate the amount of infiltration control that the cover needs to provide; and The characteristics of the underlying geological units and the distances to the ground and surface water bodies that affect the potential risks to water quality.
The need for different cover components and their specifications, such as material type, thickness, aerial extent, slope, and method of construction, should be based on the most costeffective combination of the components that meets the closure objectives and performance goals. It should be emphasized once again that the above US EPA-recommended design may not be suitable for other types of waste management practices, discussed in section 14.3. The following sections describe in more detail the design considerations of the various covering system components.
14.5
TOP LAYER
The top layer consists of a vegetative or armouring of gravel sized material. In some areas, the prevailing climate may inhibit the establishment and maintenance of vegetation, or a planned alternative use of the site may preclude vegetation. The vegetation component of the top layer should meet the following general specification: (1) locally adopted perennial plants, (2) resistant to drought and temperature extremes, (3) roots that will not disrupt the infiltration barrier, (4) capable of thriving in low nutrient soil with minimum nutrient addition, (5) sufficient plant density to minimize cover erosion, and (6) capable of surviving and functioning with little or no maintenance. In addition, attention should be paid to the selection of plant species. The use of shrubs and trees is usually inappropriate because of the root system extension to a depth that normally would invade either the drainage layer, the infiltration barrier, or the waste. A surface armour component of very coarse material promotes infiltration rather than runoff. Such a layer may be more applicable in arid areas or used in combination with an infiltration barrier layer. Where an armouring layer is used, it is recommended that the material possesses the following characteristics: (1) capable of remaining in place and minimizing erosion of itself and the underlying soils component during extreme weather events of rainfall and/or wind, (2) contains durable materials that are not likely to weather significantly over an extended period, and (3) capable of accommodating settlement of the underlying material without compromising its performance. For both vegetated and armoured top layers, the shape of the top surface should ideally be slightly convex, and uniformly sloped. To prevent pounding of rain water due to irregularities of the surface, the cover should be sloped to achieve the recommended US EPA (1989) values. Slopes greater than 5 to 10 percent will regularly require special controls such as armouring or vegetation to prevent erosion. The erosion potential of coveting surfaces can be evaluated with the use of the universal soil loss equation (Williams, 1975). The rate of soil removal by erosion can be simply expressed by:
TOP LAYER A
=
339 [14.1]
RxK•
where A is average soil loss (tons/acre/year), R is the rainfall and runoff erosivity index by geographic location (dimensionless), K is soil erodibility (tons/acre), L is the slope-length factor (dimensionless), S is the slope-steepness factor (dimensionless). LS is generally referred to as topographic factor, C is the cover management factor, and P is support practice factor. Rainfall and Runoff Factor (R): The rainfall and runoff erosivity index, R, takes into consideration the total, the intensity and the seasonal distribution of the rainfall. It also takes into account the erosive effects of storms. Computed rainfall indices vary from less than 20 in dry climate to more than 600 for humid subtropical regions, as shown in Table 14.1.
Table 14.1: Approximate rainfall and runoff factor, R, by climate condition Rainfall and runoff factor Climate condition (R) Mediterranean climate (dry summer, mild wet winter) Arid climate (hot, dry) Humid subtropical (mild winter, hot and wet summer) Humid continental (short winter, hot summer) Humid continental (long winter, warm summer)
20 - 180 10-80 300 - 600 100 - 200 20- 100
Table 14.2: Soil erodibility factor, K, by soil texture Erodibility factor Texture class (K) Texture class
Erodibility factor
Fine sand Very fine sand Loamy sand Loamy very fine sand Sandy loam
O.16 0.42 0.12 0.44 0.27
Very fine sandy loam Silt loam Clay loam Silty clay loam Silty clay
(/Q 0.47 0.48 0.28 0.37 0.25
Soil Erodibility Factor (K): Soil Erodibility, K, is a measure of a soil's inherent susceptibility to erosion. The two most significant and closely related soil characteristics influencing erosion are infiltration capacity and structural stability. In turn, these are affected by soil properties such as organic matter content, soil texture, the kind and amount of swelling clays, soil depth, tendency to crust, and the presence of impervious soil layers. Approximate K values are shown in Table 14.2 (Wischmeier and Smith, 1978) for different soil types. Note that K factor normally varies
340
COVERING SYSTEMS
from near zero to about 0.6. It is low for soils into which water readily infiltrates, such as well drained sandy soils or friable tropical clays high in hydrous oxides of iron and aluminum or kaolinite. Erodibility indices of less than 0.2 are normal of these readily infiltrated soils. Soils with intermediate infiltration capacities and moderate soil structural stability generally have a K factor of 0.2- 0.3 while the more easily eroded soils with low infiltration capacities will have a K factor of 0.4 or higher.
Topographic Factor (LS): It reflects the influence of length and steepness of slope on soil erosion. The longer the slope, the greater is the opportunity for concentration of the runoff water. However, soil loss is generally less sensitive to slope length than to steepness of slope. Also, the complexity of the slope (its unevenness) influences soil erosion. Values for the topographical factor are shown in Table 14.3 (USDA, 1995).
Table 14.3: Values for the topographical factor (LS) Slope length (m) Slope 25 50 75 100 Slope
100
(%)
(%) 2 4 8
Slope length (m) 25 50 75
0.21 0.36 0.64
0.23 0.43 0.79
0.25 0.46 0.90
0.26 0.50 0.99
12 16
1.01 1.38
1.31 1 . 5 2 1.69 1 . 8 5 2.18 2.46
Cover Management Factor (C): The cover management factor, C, indicates the influence of vegetation systems and management variables on soil loss. The C value is the ratio of the soil loss under the conditions found in the field in question to that which occur under clean-tilled, continuously fallow conditions. The reported values by the USDA (1980) are derived from experiments on agricultural crop lands and have less value for waste management practices. When there is no vegetation cover, the management factor, C, has a value of 1. Clearly, a low value of C is desirable. Mulch may be used to reduce the value of C to an acceptable level. Chemical binders and rocks can be considered as mulch. Application of organic mulches composed of residues from agricultural crops or industrial products are usually part of the revegetation process. The direct effect of mulch on the C factor can be visualized with the aid of Figure 14.2 (Lyle, 1987). A total cover reduces erosion by as much as 30 fold relative to a system with no cover at all. Support Practice Factor (P): Support practices include tillage on the contour, terrace systems, and grassed waterways, all of which will tend to reduce the support practice factor, P. If there are no support practices, the P factor is 1.0. In the following sections, the physical, environmental and chemical parameters that affect the performance of the top vegetative layer are discussed.
TOP LAYER
341
r 0.8 0
0.6
E 0.4
E 9 0
0.2
(D 0
0
20
40
60
80
100
Mulch c o v e r (%)
Figure 14.2. Effect of mulch on cover management factor, C.
14.5.1 Physical Parameters Soil Water
Plant growth is possible only in unsaturated soils where both water and air are present. Growth is very limited on dry soils and increases as soil becomes wetter until moisture reaches a level at which oxygen is driven from the soil. Basically, soil can be classified as being in one of the following three states, indicated in Figure 14.3: (1) water saturated where all the voids are filled with water, (2) at field capacity which corresponds to the optimum soil moisture content, and (3) at wilting point where the remaining moisture cannot be adsorbed by the plant. Particle Size Distribution or Texture
The relative percentage of sand, silt and clay largely determines to which extent water is available to plants. Coarse soils, e.g., soil with high sand content, have high hydraulic conductivity. They allow water to enter freely and pass through without retaining enough for plant use. The water that moves downward without being held by the soil is called gravitational water. On the contrary, fine textured soils with high silt and/or clay content have low hydraulic conductivity. They have the capacity to hold much water, but so strongly that its movement and availability to roots is considerably curtailed. A typical water holding capability of various soils is shown in Figure 14.4. Soils with intermediate texture, such as loams (approximately 40% sand, 40% silt and 20% clay), are highly desirable, notwithstanding their low water holding capacity. The ranges of plant-available water for various soil texture are shown in Figure 14.5 (Lyles, 197).
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COVERING SYSTEMS
Figure 14.3. Phase diagrams and soil water potential curve showing the status of water in soils.
Organic Matter Organic matter in soils results from the decomposition of plants and animals, as discussed in Chapter 4. It has many great impact on soil water as well as the soil nutrient cycle, namely: (1) On the soil surface, fresh organic materials reduce the impact of falling raindrops and the speed of surface running water, thereby reducing erosion. They are also valuable since they reduce the adverse effects of wind and temperature, and allow the surface to remain moist. Mulch is often added to make up for the lack of surface natural organic matter; (2) Fresh organic matter is a source of food for animals, such as insects or worms, which in turn helps supply the soil with water and oxygen by burrowing into it; and
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343
Decomposed organic matter is the most important factor promoting the formation and stabilization of soil structure. It can also absorb and hold water and nutrients for plant use, and is particularly valuable for compensating for the disadvantages of sandy soils.
Figure 14.4. Soil water holding capacity for various types of soil.
Figure 14.5. Effect of texture on soil's ability to hold water.
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COVERING SYSTEMS
Soil Structure Soil structure is related to the aggregation of soil particles into clusters to produce a characteristic solid and void distribution. The state of water in a soil depends on the number, size and shape of the pores. These properties control the amount of water and air that enter the soil, and are just as important as the solid portion. Basically, we can distinguish between macro- and micropores. Macro-pores are open spaces between aggregates of soil particles and are needed to get water and air into the soil and distribute them throughout the root zone. Micro-pores are mainly needed to hold some of the water by capillarity for use when the soil becomes drier. Poor soil aeration is a common characteristic of a compacted soil with few or no large pores or a badly drained soil after a rainfall. The optimum configuration for plant growth is obtained when the pore spaces are equally divided between the micro- and macro-pores with 50% of the soil material being solid material and the other 50 percent pore space.
Soil Depth Soil depth can be defined as the thickness of the soil layer from which plants can adsorb water and nutrients. The lower limit of the soil layer can be considered to be the depth at which roots cannot penetrate deeper into the soil. Shallow soil cannot store much water. For landfills, it is critical that the top soil be thicker than the evaporation path. Indeed, if evaporation occurred within the underlying clay liner, the resulting desiccation of the clay will affect its imperviousness. 14.5.2 Environmental Parameters
Sunlight Under high luminosity and temperature, the photosynthesis of plants almost completely stops. This characteristic is particularly important since most accumulations of mill tailings are light in colour due to the type of metals they contain, and may reflect excessive radiation to plants on the surfaces, thus intensifying the physiological stress. As the sloping sides of waste piles receive greatly varying amount of solar radiation, depending on their direction of exposure, vegetation that may be effective on northern or eastern exposures may not be suitable for southern or western ones. Dark coloured coal mining wastes can provide some protection from heat reflection and offer a favourable environment for plant development. Where there is a mixture of plant species or very dense vegetation, the taller plants may shade the others and significantly reduce their development.
Temperature Because of the intensity of solar radiations, temperature, though very difficult to control, is quite important. It has a direct effect on the basic plant function as well as on the adsorption of water and nutrients by roots. At low temperatures when the soil water approaches freezing, water becomes more viscous and therefore more difficult to absorb by plants. The uptake of nutrients is also significantly retarded. From 0~ to 65 ~C, water becomes increasingly easier to adsorb. However, high temperatures promote the evaporation of soil water near the surface where additional water has to be supplied (for irrigation) and can slow down, or even stop, the photosynthesis process. The adverse effects of temperature, including temperature extremes caused by solar heating by day and radiation cooling at night, can be significantly reduced by using mulch to prevent the rise in temperature and subsequent loss of soil water. It can be particularly useful after seeding during the germination period. The amount of annual precipitation, the temperature range, and the intensity of
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sunlight radiation are the most important factors to be considered in choosing the soil type and selecting the right species
Living species Animals can have a positive effect on plant growth whenever they contribute to the aeration of the soil structure or, like some bacteria, when they adsorb nutrients from the atmosphere and make them available for plants. However, animals often have adverse effects on plant growth. They can be diseased organisms, insects which feed on plant parts or higher animals such as rabbits and deer that destroy plant through grazing. Some plants can affect the growth of others by shading as mentioned earlier, as well as through competition for water and nutrient absorption. A suitable plan has to be established to prevent such adverse effects and ensure the growth of the desired species. Regular inspections have to be carried out and a remedial action taken as soon as the problem is discovered. 14.5.3 Chemical Parameters
The (1) (2) (3) (4)
Nutrients Each plant species requires some essential element nutrients to grow and develop properly. source of the 15 essential elements usually used by plants are: Carbon dioxide from the atmosphere: carbon (C); Soil water: hydrogen (H); Oxygen in atmosphere: oxygen (O); and Soil water and soil solids: nitrogen (N), phosphorus (P), potassium (K), calcium (Ca), magnesium (Mg), sulfur (S), iron (Fe), zinc (Z), copper (Cu), manganese (Mn), molybdenum (Mo), and Boron (B).
Dealing with nutrients in soils is particularly critical and rather difficult. The advice of an agricultural specialist should be sought. Nutrients can be present in the soil in different forms, some of which are available for plant root absorption while others are not. There is a constant movement of nutrients from available to unavailable forms, depending on soil conditions (e.g., pH and microorganisms). Nutrients can also be removed from the soil by natural processes such as soil erosion, action of microorganisms, and evaporation. Moreover, when plants are harvested by humans, the nutrients they contain are removed from the cycle. Likewise, nutrients can be added by precipitation, decomposition of organic matter or chemical fertilizers. Selected nutrients and their global effect on plants as well as the different processes they are involved in are discussed below.
Nitrogen: Nitrogen, which is one of the essential elements used in great quantities by plants, is a good example for illustrating the complexity of all the processes by which nitrogen moves from one form to the other. Nitrogen exists in waste, soil and the atmosphere, in several forms: (1) Organic N, such as alkyl and aromatic amines. This type of nitrogen is not available to plants; and (2) Inorganic N, such as ammonium (NH4+) which is held by cation exchange, ammonia (NH3) which exists in gaseous form, nitrite (NO2) which is highly mobile, nitrate (NO3) which is highly mobile and toxic to the environment when leached from the soil, and molecular nitrogen (N2) which exists in gaseous form. Ammonium (NH4+) and nitrite (NO2) are the
346
COVERING SYSTEMS only compounds in this group that are available to plants.
A brief description of the processes by which nitrogen moves from one form to the other is given below. For a detailed discussion, the reader should refer to Chapter 21. (1) Mineralisation involves the conversion of the plant-available organic form of nitrogen to an available inorganic form by microbial decomposition; (2) Fixation is the process by which atmospheric nitrogen (N2) is transformed to available inorganic nitrogen by bacteria. For instance, the Rhizobium bacteria which live in the root of leguminous plants can convert N 2 to NH4; (3) Nitrification is the conversion of ammonium (NH4 § to nitrate (NO3-) by two types of bacteria, under certain soil conditions (aerobic medium); and (4) Denitrification is the microbial process whereby NO3 is converted to gaseous nitrogen. It is largely governed by soil organic matter content, pH and temperature. Metals: Some metals are essential to plant development even if they are mostly needed in small quantities. Metallic compounds exist in a variety of forms, not all of which are available to plants. They cannot be adsorbed by plants and are leached with difficulty if they are insoluble at the pH of the soil or strongly sorbed or chelated on the soil constituent particles, as discussed in Chapter 6. But when they are in a soluble form or weakly adsorbed, they are available for plant uptake and easily transported by leaching and runoff. Metals may accumulate in a plant and have an adverse effect on its growth, a condition referred to as phototoxicity.
Fertilization To determine the exact amount of fertilizer to be added and thereby reduce the associated costs, a good knowledge of the soil characteristics is required. It is well known that adding more fertilizer than the optimum amount will just increase the costs without promoting further growth, or may even have adverse effect if the added nutrients are toxic to the plant. Consequently, laboratory soil tests, which consist of extracting, identifying and quantifying the soil nutrients, are conducted. This is usually done by leaching the soil with a mild acid solution. The amount of nutrient extracted is then correlated to the amount of plant growth that the nutrient will support. Testing is usually done for the most important nutrients except nitrogen since the available forms of nitrogen remain in the soil for only a short period of time. It is therefore assumed that all soils will need additional nitrogen, an assumption which is rather reasonable for poor soils such as mine soils. Finally, fertilization decisions should take into consideration the type of vegetation to be grown, moisture, temperature, seed quality, and the amount of plant growth desired.
Soil pH Both acid and alkaline soils cause problems that result in poor plant growth. These problems can be grouped in the three following categories: (1) decreased availability of nutrients, (2) increased availability of toxic elements, and (3) decreased activity of beneficial soil organisms. Most natural soils range in pH from 4.0 to 8.5 but most crops grow best between pH 6.0 and 7.0. However, more acidic soils may be suitable for some categories of plant (Brown et al., 1986). Alkaline soils are characterized by excessive concentration of soluble salts and exchangeable sodium. They usually induce soil moisture stress and phototoxicity due to high concentration of boron and other ions. Acid soils develop where soil materials contain minerals that produce acidic
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reactions upon oxidation (for instance pyritic minerals) or where bases have been largely leached from the parent material. However, an acid soil reaction does not directly inhibit plant growth but rather increases concentrations of toxic soluble aluminum and manganese. Soil pH significantly influences the state of nutrients in the soil. Therefore, it should be adjusted before any re-vegetation operation is done. Adjusting soil acidity is usually done by adding limestone to the soil. Calcium carbonate (CaCO3) is the most widely used material to raise the pH of the soil. To ensure a quick neutralization, it is important that limestone should be very finely ground.
14.6
DRAINAGE LAYER
When required, the drainage layer should be designed to minimize the amount and residence time of water coming into contact with the underlying waste or infiltration barrier, thereby decreasing the potential for leachate generation. Its construction materials and configurations should therefore facilitate rapid and efficient removal of water to an exit drain. The drainage layer also functions as a capillary barrier. The material used requires a grain size distribution sufficiently coarse to prevent the upward migration of fluids by capillary action, thus preventing waste constituents migration to the root zone of a vegetated top layer. Meeting the requirements for drainage usually will satisfy the capillary barrier requirements. The drainage layer should be designed, constructed, and operated to function without clogging. The physical clogging may be prevented by incorporating a filter layer of soil or geosynthetic material between the top layer and the drainage layer. The prevention of biological clogging may range from limiting vegetation to shallow rooted species to the installation of a biotic barrier (US EPA, 1980). In arid locations, the need for a drainage layer should be based on precipitation frequency, intensity, and the sorption capacity of other soil layers in the cover system. It may be possible, for example, to construct a top layer that will bear most of the precipitation that infiltrates into the layer, thereby eliminating the need for a drainage layer. The suggested 0.3 m minimum thickness of the drainage layer (US EPA, 1989) allows sufficient cross-sectional area to transport drainage in most situations. However, for relatively flat drainage slopes, drainage layer thickness should be greater than 0.3 m. Care should be taken in the design to control the velocity of the existing water, both within and beyond the exit drains, in order to prevent soil loss and destabilization. The water balance method is a kind of mathematical accounting process which considers precipitation, evapotranspiration, surface runoff, and soil moisture storage, all of which have a beating on the extent to which infiltration can be expected to occur after a rain. Since infiltration is the major contributor to leachate generation, knowing how much can be expected under a given set of site conditions will provide the designer with valuable information on which to base his/her recommendations. Such recommendations might specify the soil types, drainage grades, plant species, or cover thickness required to minimize or preclude leachate production. The factors of critical importance in a water balance calculations are soil moisture storage, evapotranspiration, and surface runoff. The first is critical because a soil cover that has exceeded its field capacity (the maximum amount of water a soil can retain in a gravitational field without downward percolation) becomes a source of infiltration to the waste, possibly resulting in leachate production. Ideally, efforts by the designing engineer should be directed to ensuring that the soil
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cover and other landfill features selected adequately to keep the soil moisture storage below field capacity. Evapotranspiration represents the amount of water present in the soil that is lost to the atmosphere from a given area through direct evaporation from the soil and transpiration from plant tissues. Evapotranspiration occurs when soil moisture is at or near field capacity. However, as soil moisture approaches the wilting point (the moisture content below which moisture is unavailable for withdrawal by plants), the amount of water available begins to restrict the rate of evapotranspiration, resulting in reduced actual water losses. Surface runoff is the portion of rainfall which runoff the site before entering the cover soil. Variables affecting runoff include intensity and duration of rainfall, existing soil moisture, soil hydraulic conductivity, slopes, and type of vegetative cover. Runoff can be calculated using empirical runoff coefficients commonly used to design surface water drainage systems. By multiplying the coefficients by the mean monthly precipitation, a "mean monthly surface runoff' can be calculated. The basic equation for determining the amount of percolation anticipated at a given site is as follows: PERC
- P
- R/O
- ST-
AET
[14.21
where P E R C is percolation, P is precipitation (the mean monthly values are typically used), R/O is the surface runoff, S T is soil moisture storage, and A E T is the actual evapotranspiration. Once the designer has estimated the quantity of moisture that will percolate, he/she is in a position to make a decision concerning the type and size of the drainage system. Other infiltration models, such as the Hydrologic Evaluation of Landfill Performance (HELP) developed by US EPA (1984), may aid in the design of drainage layers. Materials used to construct the drainage layer should be durable, clean, washed or screened prior to construction to remove fines that may promote clogging. To further prevent clogging, it may be necessary that a granular or geosynthetic filter be placed directly over the drainage layer to minimize the migration of fines from the overlying top soil. If a graded granular filter is used, care should be taken to ensure that the relationship between the grain size distribution in the filter and adjacent material meets specified criteria. Although there are numerous variations for filter criteria, those presented by Cedergren (1967) are usually recommended for water drain construction and are used (US EPA, 1987) in the design of hazardous waste management covers. The criteria can be stated as follows: (A)
Filter criteria for high flow gradients (10 to 50): (DIs filter / D85 base ) < 5
[14.3]
5 _<(D~5 filter / D15 base ) _ 20 (Dso filter / Ds0 base ) < 25
[14.4] [14.5]
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349
Filter Criteria for low flow gradient conditions (less than 1): (D6o / D~0 ) 2 4; for gravel-sized material
[14.6]
(D6o / Dlo ) > 6; for sand-sized material
[14.7]
1 _ [ (D20)2 / (Dl0 D60)] < 3; for either gravel or sand material
[14.7]
where D10, D15, D2o, D5o, 060, and D85 are measures of the percentage of soil particles smaller than the denoted sizes, are generally used to characterize the particle size of a given soil. The replacement of granular filter material by woven or non-woven synthetic materials (geotextiles) is becoming more widespread. The design criteria can be adopted from those used for designing filters adjacent to perforated pipes since they are applicable to a filter adjacent to a woven textile, i.e., (Ds5 for fine material / Dwr ) >_2
[14.8]
where Dwr is the diameter of the opening in woven textile. Previous studies (Marks, 1975; Demery, 1980) have indicated that the percentage of fines in soil layers adjacent to geotextiles should be limited to prevent clogging of the fabric and/or the formation of a filter cake adjacent to it. It is recommended, therefore, that geotextiles not be placed directly against waste that contains fine particles. A granular filter material should be used between the waste material and the geotextile. It is recommended that a granular drainage layer adjacent to a geotextile layer should not have more than 2 % passing the No. 200 sieve. Geosynthetic drainage materials are manufactured in a variety of configurations which continue to evolve with experience in manufacturing and use. Geonets and geogrids are drainage components designed to prevent rapid flow. They are manufactured as single components that usually must be separated from overlying soils that could clog them. The separating materials also are geosynthetics in the form of filter fabric. The geogrid, and top and bottom filters may all be factory-bonded into one unit. These bonded materials, also referred to a geocomposites, may be applied in one operation as the entire drainage layer. The various forms of geocomposites are well described by Koerner (1989).
14.7
INFILTRATION BARRIER LAYER
The infiltration barrier layer can be provided by either a geomembrane, a compacted clay, or a combination of the two. A geomembrane barrier layer should have the following characteristics: (1) It should have adequate thickness to prevent failure under either potential stress of the post closure care period, or during construction; (2) There should be no surface unevenness, local depressions, or mounding that creates depressions capable of containing or otherwise impeding the rapid flow and drainage of infiltrating water;
350
(3) (4) (5)
(1) (2) (3)
COVERING SYSTEMS The geomembrane should be protected by at least 0.60 m of overlying material; The geomembrane should be in direct contact with the underlying soil or waste, and should be installed on a smoothed surface; and Bridging or similar stressed conditions in the geomembrane should be avoided by providing slack allowances for temperature induced shrinkage of the geomembrane during installation and during the period prior to placement of the protective layer or drainage layer. Slack should not be excessive to the extent that folds are created that later may crack. The clay barrier layer should have the following characteristics: Clay material should be compacted in layers not exceeding 0.20 m in thickness; The compacted materials must be free of clods, rock, fractured stone, debris, cobbles, rubbish and roots, etc., that would increase the hydraulic conductivity or serve to promote preferential water flow paths; and The hydraulic conductivity of the compacted layer should be less than 1x 10.9 m/sec.
Figure 14.6. Three types of infiltration barriers.
Figure 14.6 highlights the basic difference between the three types of infiltration barrier in terms of their flow rates (US EPA, 1991). The flow rates are calculated for the following cases. First, consider a clay liner. If a 0.3 m head of water is pounded on a 0.9 m thick clay liner that has a hydraulic conductivity of 1x 10 -9 m / s e c , the calculated flow rate based on Darcy's law is 120 gal/acre/day. Second, consider a geomembrane liner. Assume that the geomembrane has one or more circular holes (defects) in the liner, that the holes are sufficiently widely spaced that leakage through each hole occurs independently from the other holes, that the head (h) of liquid pounded above the liner is constant, and that the soil that underlies the geomembrane has a very large hydraulic conductivity, i.e., the subsoil offers no resistance to flow through a hole in the geomembrane. Giroud and Bonaparte (1989a) recommended the following equation for estimating flow rates through holes in geomembranes:
INFILTRATION BARRIER LAYER q = C13 a v/~h
351 [14.101
where q is the flow rate (ma/sec), Cp is the flow coefficient with a value of approximately 0.6, a is the area of the circular hole (m2), g is the acceleration due to gravity (9.81 m/sec2), and h is the water head above the liner (m). For example, if there is a single hole with an area of 1 x 10 .4 m z and a head of 0.3 m, the calculated flow rate is 3300 gal/day. If there is one hole per acre, then the flow rate is 3300 gal/acre/day. It should be noted that with good quality control, one hole per acre is typical (Giroud and Bonaparte, 1989a). However with poor quality control, 30 holes per acre is typical with most of the holes having an area less than 1 • 10.5 m 2. For the case of a composite liner, Giroud and Bonaparte (1989b) and Giroud et al., (1989) recommended the following equation for computing seepage rates for cases in which the hydraulic seal between the geomembrane and soil is poor: q : 0.851 h 0"9 a ~
[14.11]
where K~ is soil hydraulic conductivity (m/sec). Eq. [ 14.11 ] assumes that the hydraulic gradient through the soil is 1. If there is a good hydraulic seal between the geomembrane liner and underlying soil, the flow rate is approximately one-fifth the value computed from Eq. [ 14.11 ] since the first term on the RHS of Eq. [ 14.11 ] is then 0.155 rather than 0.851. For example, suppose the geomembrane component of a composite liner has one hole/acre with an area of 1 x 10-4 m 2 per hole, the hydraulic conductivity of the subsoil is 1 x 10-9 m/sec, the head of water is 0.3 m and a poor seal exists between the geomembrane and soil. Based on that information, the calculated flow rate, using Eq. [ 14.11 ], is 0.8 gal/acre/day. In summary, for the above specified conditions, the flow rates are 120, 3300, and 0.8 gal/acre/day for soil liner, geomembrane, and composite liner, respectively. Therefore, a composite liner (even built by poor to mediocre standards) significantly outperforms a soil liner or a geomembrane liner alone. For this reason, a composite liner should be recommended when there is enough rainfall to warrant a very low hydraulic conductivity barrier in the cover systems. In the following subsections, the physical and chemical parameters that affect the performances of soil-based infiltration barrier layer are discussed.
14.7.1 Physical Parameters Materials Since the primary requirement of an infiltration barrier layer is compactibility to produce a suitably low hydraulic conductivity in the order of 1 • 10 .9 m/sec, the material used to construct the infiltration barrier must be chosen carefully. The soil should have: (1) At least 20% fines (fines are defined as the percentage, on a dry weight basis, of material passing the 0.075 mm opening of a No. 200 sieve); (2) A maximum amount of 10% gravel (gravel is defined as the material retained by the 4.76
352
(3) (4)
COVERING SYSTEMS mm opening of a No. 4 sieve); Plasticity index of at least 10% although some soils with a slightly lower plasticity index may be suitable (US EPA, 1988); and No stones or rocks larger than 25 to 50 mm in diameter should be present in the infiltration barrier layer.
The most important factor that controls such soil properties as hydraulic conductivity, plasticity, ion exchange, adsorption and flocculation/dispersion is the type of soil mineral, as discussed in Chapters 4 and 5. If the soil material does not contain enough clay or other fine-grained minerals to be compactible to the desired low hydraulic conductivity, commercially produced clay minerals, such as sodium bentonite, may be mixed with the soil (US EPA, 1991). Figure 14.7 shows, for a wellgraded silty soil, the relationship between the percentage of bentonite added to the soil and the hydraulic conductivity after compaction. The percentage of bentonite is defined as the dry weight of bentonite divided by the dry weight of soil to which the bentonite is added (Wb/Ws). For wellgraded soils containing a wide range of grain sizes, adding just a small amount of bentonite may lower the hydraulic conductivity of the soil to below 1x 10.9 rn/sec. For properly-graded soils, e.g., those with a uniform grain size, more bentonite is often needed (US EPA, 1991).
Figure 14.7. Effect of Bentonite addition upon the hydraulic conductivity of a bentonite-amended soil.
Structure
The placement conditions and the details of construction of the infiltration barrier determine the fabric and structure of the compacted layer which in turn determine its hydraulic conductivity. As noted by Yong, Mohamed and Warkentin (1992) and Mitchell and Madsen (1987), three levels
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of first-order fabric can be recognized as shown in Figure 14.8. These are categorized on the basis of the degree of magnification required for a proper observation of the fabric patterns or soil particles: (1) Macro-fabric: Macro-fabric units are distinguishable with the naked eye. Granular particles, for example, will in general constitute single fabric units. In clay soils, fabric units that can be identified visually with the naked eye will generally consist of an aggregation of clay particles. These units are defined as peds. Each ped consists of an aggregation of particles. Measured liquid flow rate occurs through the pore spaces between the peds. (2) Micro-fabric: Micro-fabric units are visually observed under light microscope. In the case of clays, single particles will not be distinguishable at this level of viewing. The fabric units identified in the microscopic range consist of several particles or groups of particles known as clusters. The composition of a cluster is somewhat similar to that of a ped except in regard to size. Clusters can be combined to form peds. The amount of flow throughout the pores of a cluster is very little. Ultra-macro-fabric:Ultra-macro-fabric units are visually observed in the ultra-microscopic (3) level using electron microscopy (e.g., transmission or scanning electron microscopy). Single or individual clay particles can be distinguished at this level. The small fabric units observed at this level of viewing are domains or single clay particles. Domain units consist of two or more particles acting as a unit. Several domains could combine to form a cluster.
Figure 14.8. Schematic diagram showing both fabric and pore classifications.
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The factors that control the structure of a compacted infiltration barrier and its properties are described below. Water content The water content of the soil at the time of its compaction is an important variable controlling the hydraulic conductivity of the infiltration barrier layer. Soils compacted at water contents less than optimum (dry of optimum) tend to have a relatively high hydraulic conductivity whereas soils compacted at water contents greater than optimum (wet of optimum) tend to have a lower hydraulic conductivity. It is usually preferable to compact the soil wet of optimum to achieve minimum hydraulic conductivity, as evident in Figure 14.9.
Figure 14.9. Influence of moisture content hydraulic conductivity.
Compaction Energy As shown in Figure 14.10, increasing the energy of compaction reduces hydraulic conductivity. The hydraulic conductivity of a soil that is compacted wet of optimum could be lowered by one to two orders of magnitude by increasing the energy of compaction. The compaction energy delivered to a soil depends on the weight of the roller, the number of passes of the roller over a given area, and the thickness of the soil layers. Increasing the weight and number of passes, and decreasing the layer thickness, can increase the compaction effort. The best combination of these factors to use when compacting low hydraulic conductivity infiltration barrier depends on the water content of the soil and the firmness of the sub-base. Clod Size The clod concept proposed by Olsen (1962) suggested that most of the flow of water in compacted clay occurs in relatively large pore spaces located between peds or clods of clay rather
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355
than between the particles of clay within the clods. According to the clod concept, soft, wet clods of soil are easier to remould than hard, dry clods. When a soil is compacted wet of optimum, the soft, wet clods of soil are remoulded. This in turn results in smaller inter-clod voids and lower hydraulic conductivity. This concept has been investigated extensively by Benson and Daniel (1990) who concluded that when soil is compacted dry of optimum, the hydraulic conductivity increases by approximately 2 orders of magnitude for an increase in clod size from 1.5 mm to 9.4 mm. It should be pointed out that high plasticity soils form large clods while low plasticity soils (plasticity index less than 10%) do not. For soils that form clods, the clods must be remoulded into a homogeneous mass that is free of large inter-clod pores if low hydraulic conductivity is to be achieved.
Figure 14.10. Influence of compaction effort on hydraulic conductivity.
Desiccation Desiccation causes infiltration barrier material to shrink and potentially to crack. Shrinkage characteristics vary with the nature of the soil. Total shrinkage increases with increasing initial water content. Shrinkage is a function of the percent of clay portion in the soil, the kind of clay minerals, particle arrangement, and overburden pressure. Shrinkage cracks will form at the surface of clay soils on drying. Cracks form where the cohesion of the soil is lowest. Where drying is not uniform, cracks will form in the wetter portions of the soil. A change in particle orientation occurs at the crack surface and, on re-drying after wetting, the cracks will appear in the same places if the soil has not been otherwise disturbed. The number of cracks per unit area depends upon the clay minerals and the particle arrangement. A large number of cracks will form in flocculated clays while in semi-oriented clays with high cohesion, a few relatively large cracks will result. The cracks in a surface soil with a well developed crumb structure may not be noticed because each crumb unit becomes separated by a small distance from the adjoining one. The cracks formed in the compacted clay layer have an important role in water infiltration and movement.
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The effect of desiccation on hydraulic conductivity is shown in Figure 14.11 (Boynton and Daniel, 1985). It can be seen that at low confining pressure, one order of magnitude change in the hydraulic conductivity can be obtained. At high confining pressure, the effect of desiccation on hydraulic conductivity is diminished due to the closure of the cracks that had formed before. In a covering system, the overburden stress on the compacted clay layer seldom exceeds 28 kPa. Based on Figure 14.11, if desiccation of the compacted soil layer occurs, it is not expected that all the damage done by desiccation would be self-healing unless there is a large percentage of sodium montmorillonite in the soil mixture. A compacted soil layer of low hydraulic conductivity must be protected from the damaging effects of desiccation during and after construction. During construction, the soil must not be allowed to dry significantly either during or after compaction of each lift. Frequent watering of the soil is usually the best way to prevent desiccation during construction. Similar results can be achieved by spraying bituminous sealers, use of a geomembrane cover, or by covering the compacted layer with a layer of another soil.
Figure 14.11. Effect of desiccation on the hydraulic conductivity of compacted clay.
Bonding between Lifts Lift interfaces have important implications in regard to water infiltration and the resultant hydraulic conductivity. US EPA (1983) has performed a large-scale laboratory test to investigate this phenomenon. It was concluded that lift interfaces provide connecting flow channels. Better overall performance (low hydraulic conductivity) is achieved if lifts are bonded together to eliminate high conductivity at lift interfaces.
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Freeze and Thaw When soil is first exposed to sub-freezing temperature, free water within the macro-pores will start to form ice buds. As the freezing process continues, the ice buds continue to grow as heat is extracted. Free water from adjacent macro-pores will migrate towards the ice buds caused by the differences in local temperature and matric potential. As the ice crystals grow in the pore spaces, continued pore expansion will cause the soil mass to undergo significant volume change. This will break some of the bonding between fabric units and eventually result in alteration of the original soil structure, and formation of new fabric units if new bonding is possible. In clay soils where the surface activity is relatively high, particle aggregation and/or orientation might occur as a result of reduced inter-particle spaces due to ice pressure and/or migration of adsorbed water.
Figure 14.12. Relationship between percent of soil less than 2 ~tm and freeze-thaw cycles.
Figure 14.12 illustrates the effect of cyclic freeze-thaw on particle aggregation (Yong et al., 1985). The test results show that when no dispersing agent is used, the amount of clay-sized particles present in the clay soil after the first four freeze-thaw cycles decreased by about 15% in comparison to the virgin unfrozen samples. At the end of the eighth freeze-thaw cycle, there seems to be no aggregation in the closed system and some small aggregation in the open system after which both undergo significant drops. The decrease in the clay size fabric units is more pronounced in the open system than in the closed system. In contrast, when a dispersing agent is used, the clay size fractions of the soil increase during the first four freeze-thaw cycles by about 20%. Subsequently, the change is small in the closed system while a reduction as high as 25% is experienced in the open system. The effect of aggregation due to cyclic freeze-thaw can be investigated from the characteristics of soil water potential curves of the cyclic freeze-thaw, as shown in Figures 14.13 and 14.14 (Yong and Mohamed, 1992).
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COVERING SYSTEMS
Figure 14.14. Soil water potential characteristics of natural clay (closed system).
Variations in the hydraulic conductivity of clay as a function of the number of freeze-thaw cycles are shown in Figure 14.15 (Mohamed et al., 1993). The first set of data obtained when the compacted clay layer was allowed to uptake distilled water during freezing. It was observed that the hydraulic conductivity increased 23-fold after 3 cycles of freeze-thaw. This can be attributed to: (1) formation of micro-cracks in both vertical and horizontal directions, and (2) clay aggregation during
INFILTRATION BARRIER LAYER
359
freezing, hence a decrease in surface area and an increase in pore size geometry.
Figure 14.15. Clay permeability variations with freeze-thaw cycles.
For the second set of data, with the clay allowed to uptake acidic leachate solution during freezing, it was observed that after the first cycle of freeze-thaw the hydraulic conductivity increased 625-fold. This can be attributed to: (1) formation of micro-cracks in both vertical and horizontal directions, (2) reduction in the thickness of the diffused double layer due to replaceability of major cations in the clay by the inducing heavy metals in the leachate, and (3) clay aggregation during freezing. Furthermore, after the second freeze-thaw cycle, the hydraulic conductivity decreased 2.9fold relative to the value noted after the first freeze-thaw cycle. This can be attributed to compression of clay particles in an aggregated form, hence a reduction in the water content and specific surface area. It is anticipated that the hydraulic conductivity will not show any further decrease after the second cycle and will level off. 14.7.2 Chemical Parameters
In the case of decommissioning uncontrolled hazardous waste sites where the groundwater level is at shallow depths, chemical compatibility between the compacted soil cover and the waste is important. During winter periods, water migrates from the waste in the upward direction and interacts with the compacted soil layer. The effect of chemicals on hydraulic conductivity of clays is discussed in Chapter 5. A summary of the various effects is presented below. (1) Inorganic acids: Inorganic acids can dissolve some of the constituents that comprise the clay soil structure. The solubility of clays in acids varies with the nature of the acid, the acid concentration, the acid to clay ratio, the temperature and the duration of treatment (Grim, 1968). The effect of an acid on clay is enhanced when the acid has an anion about the same
360
(2)
(3) (4)
(5)
14.8
COVERING SYSTEMS size and geometry as the clay component. Leaching with acids leads to approximately 5 to 12-fold increase in the hydraulic conductivity of the montmorillonite specimen in comparison to that determined using water as the permeant; Replacement of exchangeable sodium by calcium ions: This can alter the hydraulic conductivity of a bentonite considerably. Mesri and Olson (1971) showed a 28-fold increase in the hydraulic conductivity of montmorillonite clay while the results of Sridharan et al., (1986) indicated a 8-fold increase for a similar exchange reaction; Replacement of monovalent sodium ions by trivalentferric ions: This has a more pronounced effect on the hydraulic conductivity of bentonite, the increase being 33-fold; Replacement of monovalent sodium ions by monovalent potassium ions: This leads to approximately 5-fold increase in hydraulic conductivity of bentonite at a given void ratio; and Organic chemicals: Hydraulic conductivity to organic chemicals decreases as the log of its octanol-water partition coefficient increases. Because the octanol-water ratio is a measure of the tendency of organic chemicals to escape from water, those substances least compatible with water should move most slowly through the soil. The more positive the value of the octanol-water ratio, the smaller the value of hydraulic conductivity. There is a good correlation between the hydraulic conductivity of a clay soil permeated with an organic chemical and its dielectric constant. The greater the dielectric constant, the greater is the value of the hydraulic conductivity coefficient. As a general rule, as the molecular weight of the organic chemical increases, the clay soil hydraulic conductivity to organic chemical decreases. This is due to the fact that more water molecules will be displaced because the large molecules have more points of contact with the active clay surface.
BIOTIC BARRIERS
Biotic barriers are typically layers of cobble or coarse gravel that are located between the top layer and infiltration barrier layer. This biotic barrier layer would directly underlie the soil component of the top layer, perhaps separated from it by a geosynthetic filter layer. In the case where both a biotic barrier and a drainage layer are required, a single granular layer can be installed to serve both purposes.
14.9
MINING WASTE COVERING SYSTEMS
Acid mine drainage is a natural phenomenon caused by the oxidation of metallic sulphides. The sulphides of primary concern are pyrite (FeS2) and pyrrhotite (Fel.xS) as they are among the most common minerals found in mining and mineral processing waste, as discussed in Chapter 3. Underground metal sulphides, in their natural crystalline form, are relatively stable but when mined and milled, they become exposed to oxygen and water, causing the minerals to oxidize. In the absence of calcareous materials, the oxidation of sulphide minerals produces sulphuric acid and liberates heavy metals (Ferguson and Erickson, 1988; Scharer et al., 1991; Mohamed et al., 1993a and 1993b; Amourgis, 1994; Mohamed, 1997c). This phenomenon occurs through a series of chemical, electrochemical, and microbial processes according to the following reactions (Stumm and
MINING WASTE COVERING SYSTEMS
361
Morgan, 1981):
7
FeS2(s ) + -~ 02 + H20 ~ Fe 2+ + 28042- + 2 H §
1
Fe2+ + __1 02 + H + ~ Fe3+ + _ 9 2 0 4 2
FeS2(s ) + 14Fe 3+
+
8920 ~ 15Fe 2+
+
28042- + 16H +
[14.12]
[14.13]
[14.14]
When pyrite is exposed to both oxygen and water (Eq. [14.12]), sulphur (S) oxidizes to sulphate (SO42). This results in a release of acid ( 9 +) which in tum causes a drop in the pH of the water. When the pH of the water falls to between 2 and 4, and oxygen is readily available, subsequent oxidation of ferrous iron (Fe 2+) will occur producing ferric iron (Fe 3+) (Eq. [ 14.13]). Released Fe 3+ will further oxidize itself forming more Fe 2(Eq. [14.14]) which can be further oxidized to Fe 3+. The result is a rapid process of large quantity acid production which in turn causes the release of heavy metals into solution. If left to chemical processes alone, acid formation would be very slow. Bacteria play an important role in the formation of acidic drainage as they catalyse the oxidation of sulphur and iron, as discussed in Chapter 21. Therefore, acid generation, as well as acid consumption, is the resultant of a number of interrelated chemical reactions. The primary ingredients for acid generation are: (1) sulphide minerals, (2) water or humid atmosphere, and (3) an oxidant (e.g., O2 or Fe3+). Exclusion of absolutely all moisture or oxidant will stop acid generation.
14.9.1 Dry Barrier Concept The current practice of shallow land burial of toxic waste materials has one major problem, namely, rapid deterioration of liners in certain chemical environments. Extensive field investigations on pyritic uranium tailings have shown that the zone of active oxidation is confined to relatively shallow zones above the water table (Feenstra et al., 1981; Dubrovsky et al., 1985). These studies have shown that oxygen concentrations in the partially gas-filled pore spaces decrease rapidly below the tailings surface and that detectable oxygen concentrations extend to greater depths in coarsegrained tailings than in similar fine-grained zones. Typical zones of oxidation were on the order of 0.5-3.0 m from the tailing surface. The zone of oxidation was shown to be limited by the rate at which oxygen diffuses into the tailings and the rate at which it is consumed. Oxygen moves readily by molecular diffusion through geological materials that have a high gas-filled porosity. Materials at or near water saturation exhibit reduced oxygen fluxes owing to the relatively low solubility of oxygen in water and the low diffusion coefficient for oxygen in water relative to that in air. The concept of using covers as oxygen barriers has been investigated by many researchers, among them Brown (1970), Nicholson et al., (1989), Mohamed et al., (1994), and Mohamed (1997c). The cover design relies on the slow rate of diffusion of oxygen through water-filled pores, and is therefore based on the hydraulic principles and characteristics of partially-saturated porous
362
COVERING SYSTEMS
media. Figure 14.4 shows the soil water potential curves for three types of soils -- a kaoline clay, sandy clay, and fine sand. As the pressure head progresses in the negative direction from zero, soils remain saturated until a pressure head of ~a, the air-entry value, is attained. The air-entry value is the pressure head necessary to overcome the capillary forces that can be exerted by the largest pore in the medium. As the pressure head becomes more negative, smaller pores drain until the residual water content of the particular material is attained. The shape of the water content-pressure head relation depends largely upon grain size and structure. The double barrier concept is illustrated in Figure 14.16. Material A is of very high suction potential (e.g., natural clays or compacted sand/bentonite mixture) and is essentially a hydraulic barrier. Material B has to be of very low suction potential, the best natural material being uniform sand. Since no significant migration of moisture can occur across the boundary unless the material becomes saturated, material B impedes flow and is indeed a capillary barrier. The combination of materials A and B constitutes a double barrier. For material A to remain saturated, the minimum thickness of material A would be the difference between the pressure head at the air entry of material A (~A) and the pressure head at the residual water content of material B (~aB). In addition to both physical and chemical parameters, which were discussed previously with regard to the infiltration barrier layer, the following factors should be taken into consideration when a double barrier concept is to be implemented for decommissioning reactive mine tailings dump sites:
Figure 14.16. Double barrier concept design.
Oxygen Flux The minimum flux of oxygen is limited by the infiltration of oxygen-saturated water into reactive tailings. Figure 14.17, adapted from Nicholson et al., (1989), shows the relationship between oxygen flux and the oxygen diffusivity of a 1 m thick clay cover layer. The flux varies from 10, for
MINING WASTE COVERING SYSTEMS
363
a completely dry state, to 0.001, for a completely saturated state. However, if infiltration is assumed to be in the range of 0.01-0.10 m/year, and the water is saturated with oxygen (10 mg/1), then the oxygen flux due to infiltration will be in the range of 0.001-0.01 kg H2SO4/m2/year. This range represents the practical lower limit of acid production. The lower limit of infiltration, i.e. 0.01 m/year, corresponds to a hydraulic conductivity in the order of 3 x 10"lO m/see.
Figure 14.17. Oxygen flux resulting from infiltration and diffusion. The flux range that corresponds to infiltration (0.01-0.10 m/year) is shown by the shaded bar.
Film Moisture Action Although it is known that moisture can migrate from points of thicker films to points of thinner films, very little has been done to understand the mechanisms that control film migration. For example, it would be of greater value to understand how film and capillary moisture come to equilibrium. It has to be realized, therefore, that gradual transfer of moisture into the capillary barrier will occur and that may be considerable when the upper barrier is close to saturation. Treatment of capillary barrier with hydrophobic materials, e.g. paraffins and oils, is a possible way to control that effect. Another feasible measure is to install a multi-stage double barrier system instead of a singlestage system.
Migration offines Migration of fines into the lower barrier may also cause serious problems. Provision of a special filter zone between the barriers is therefore an essential requirement in most cases.
Water channels Inhomogeneties within the hydraulic barrier may result in significant flow through small
364
COVERING SYSTEMS
channels or cracks. Proper control during the placement of all hydraulic barriers is required and serious cracking has to be avoided, as discussed earlier. A self-healing material, e.g. a compacted sand-bentonite mixture, is certainly the most desirable. 14.9.2 Water Barrier Concept
In contrast to the dry barrier concept, the storage of mine tailings underwater produces different reaction conditions. With atmospheric oxygen as the ultimate crucial oxidant in the acid producing reaction, underwater disposal can drastically curtail reaction rates. First, the concentration of dissolved oxygen in water can attain a maximum of 8.6 mg/l at normal conditions of pressure and temperature compared to the 21% by volume (or 285 mg/1) concentration of oxygen in air. Secondly, the diffusion coefficient for oxygen in water (2x 10.9 m2/sec) is nearly five orders of magnitude less than in air (1.78x 10.5 m2/sec). Water cover on mine tailings provides three important control measures (Mohamed et al., 1996c; Dave and Vivyurka, 1994): (1) limiting available oxygen for the oxidation process and, hence, controlling acid generation, (2) eliminating surface erosion caused by wind and water action, and (3) creating a reducing environment, suitable for supporting sulphate and nitrate reducing microorganisms in sediments where soluble metals are precipitated as sulphide. Disposal of mine tailings under water includes: (1) subaqueous or underwater disposal in natural fresh water bodies such as lakes and marine disposal, (2) disposal into man-made impoundments or reservoirs, (3) disposal into flooded mine workings and open pits, and (4) building a water cover on existing waste management sites. Studies on water covers for mine tailings have been ongoing for a number of years. In the study by Noranda Technology Centre (1979) on copper and zinc tailings, high concentrations in the fresh tailings, placed under water in 45 gallon drums, were recorded. This was attributed to the dissolution of a manganese-bearing sphalerite. Dav6 and Lim (1983) measured transfer coefficients and transfer fluxes from un-oxidized tailings under continuous flow and no-flow conditions. Measured transfer fluxes for Ra-226 were found to be greater, by one order of magnitude, than theoretical fluxes calculated from using the free water diffusion coefficient. Dav6 (1991) has indicated that the dominant process when partially-oxidized waste mineral is submerged under water is the oxidation of sulphide minerals by ferric iron, as given by Eq. [ 14.14]. The author reasons that heavily pre-oxidized wastes might not be safe for under-water storage. Oxidative leaching of residual sulphide would be stopped only after soluble oxidation products, such as ferric iron, have either reacted or been flushed out. The various interactive physical and chemical processes that occur between tailings solids, tailings pore water, and a water cover may include the following: (1) oxidation of sulphide by dissolved oxygen diffusing through water, (2) oxidation of sulphide minerals by dissolved ferric iron, (3) suspension of tailings solids by wave action, (4) upward diffusion of dissolved metals from tailings pore water, (5) solid/solution reactions (including adsorption), and (6) vertical advection of pore water in tailings. A typical water cover system is shown in Figure 14.18. The major issues that need to be considered in the design are: (1) depth of water required over the tailings, (2) requirement for a protective solid layer at the tailings/water interface and composition and thickness of this layer, (3) requirement for treatment of the tailings surface, (4) resultant quality of the overlying and interstitial water, to determine an applicable flushing time, and (5) stability of the tailing dams.
MINING WASTE COVERING SYSTEMS
365
Figure 14.18. Components of a water-based covering system for mine tailings.
14.9.3 Organic Barrier
Concept
Wastewater sludges can be a valuable resource both as a cover material and a surface amelioration provided that their content of toxic substances is not too high. Wastewater sludges have been used in a variety of locations in the United Kingdom as a source of plant nutrients and organic barrier for acid-producing mineral wastes. To develop a self-sealing membrane from wastewater sludge, several events have to occur before the sediment formed from the settling solids can act as an impermeable membrane to both oxygen and water (Yong and Chen, 1990). These relate to: (a) the composition and properties of the suspended solids, (b) the nature of the interactions between the solids, (c) the need for treatment of the suspended solids to promote beneficial interaction, (d) the developed rate of sedimentation of the suspended solids, (e) the developed physico-chemical properties of the initial sediment, (f) the consolidation rate of the sediment, and (g) the stability of the final sediment membrane or sludge cake.
Induced Sludge-Sealing Process The different possible processes involved in the development of membrane seals from bleach Kraft mill effluent (BKME), an example ofwastewater sludges, are shown in Figure 14.19. The three basic categories are identified via source material, e.g. non-coagulated BKME, coagulated BKME, and fluid phase of BKME. In the last category, we rely on the interaction of the fluid phase chemistry with the surface layer to produce swelling such that void plugging occurs, hydraulic conductivity is reduced, and a proper lining is obtained. In regard to the other two categories, coagulated and non-coagulated BKME, we rely on the interaction and settlement of the suspended solids of the BKME (sludge or sludge cake) to form a coating layer that would be impermeable. The key elements in the systems relate to the capability
366
COVERING SYSTEMS
of the settled untreated or treated sediment to form an impermeable membrane. The procedure used by Yong and Chen (1990) to evaluate the capability of BKME selfsealing properties to produce the resultant cover requirements involved the use of a pervious sand blanket overlying a more impervious base. This situation permits gradual filling and blocking of the pore spaces (sludge penetration) in the sand blanket during settling of the sludge solids, followed by a build-up of a sludge layer atop the in-filled sand blanket. The resultant in-filled sludge formation is considered an in-depth cake. If a further sludge forms on top of the in-filled sand blanket, the process leading to the development of the combined sludge layer can be identified as an in-depth and cake filtration process. Figure 14.20 shows the calculated solid concentration profiles as a function of time. It can be seen that after 3150 secs (0.875 hours), the developed sludge cake above the sand surface is about 7 mm and the penetration of the settling solids into the sand voids is about 5 mm.
Non-coagulated BKME Untreated BKME
I
Non-biological solid plugging and sludge cake-induced sealing
Physical type sealing
B io-treated BKME
I
Biologically generated solid plugging and sludge cakeinduced sealing
Bio-physical type sealing
Non-coagulated solid system
Coagulated BKME Untreated BKME
I
Coagulated solid plugging and sludge cakeinduced sealing
Physicochemical type sealing
B io-tre ated BKME
I
Coagulated biologically solid plugging and sludge cakeinduced sealing
Bio-physicochemical type sealing
Coagulated solid system
Figure 14.19. Types of membrane seals from bleach Kraft mill effluent (BKME).
The calculated hydraulic conductivity of the developed sludge cake as a function of time is shown in Figure 14.21. The final hydraulic conductivity after 38400 secs (10.67 hours) is about 5.5• 10.9 m/sec which suggests that a self-sealing from the BKME can be obtained via physicochemical processes.
TYPES OF COVERING MATERIALS
367
Figure 14.21. Calculated sludge cake hydraulic conductivity variation with time.
14.10 TYPES OF COVERING MATERIALS The choice of covering materials is largely dependent upon specific site conditions, availability and cost. Covering materials used in soil-based covering systems can be broadly classified as (Parry and Bell, 1985): (1) natural materials, (2) modified soils, (3) synthetic material,
368
COVERING SYSTEMS
and (4) waste materials. 14.10.1
Natural Materials
As discussed previously, soil structure, mineralogical characteristics and degree of compaction are the principal factors affecting the performance of natural soils as covering materials. In general, well-graded fine grained soils are selected to control the movement of water, because of their relatively low hydraulic conductivity. Top soil quality depends on the inter-relationship between several physical and chemical properties. 14.10.2
Modified Soils
When appropriate well-graded fine-grained soils are not available, the grain size distribution can be broadened and percolation reduced by blending two or more soil types. The addition of Portland cement, though in small quantities, has proved very effective in modifying soil stability. Lime, usually added to assist the cementing process, can also be used to increase the stability of soils. Indeed, the addition of cement or lime, in as low a proportion as 1%, to a granular soils has shown to have great stabilizing effects (Mohamed et al., 1991). For fine-grained soils, lime has a strengthening effect over time because of its reaction with clay minerals, and the addition of 2-8% lime to clay soils can be used to enhance flocculation and base exchange. De-flocculation of finegrained soils can be achieved by the addition of soluble salts such as sodium chloride. It should, however, be noted that these may increase the swell and shrink behaviour of clay, and be subject to erosion and cracking. When montmorillonite, which is known to swell between 10 to 15 times of its original volume, is mixed with soils, the swelling that occurs fills the pores. The resultant effect is thus a decrease in hydraulic conductivity, provided proper compaction of the montmorillonite is achieved. However, since low hydraulic conductivity to water does not necessarily mean low hydraulic conductivity to other liquids, these additives should be tested under actual site conditions to ensure optimal effectiveness. 14.10.3
Synthetic materials
Synthetic materials are used in association with soil-based covering systems or as an alternative to soil-based covers. Flexible membranes should be resistant to tearing and thus be placed over a smooth buffer of soil. It should be noted that these membranes may be subject to uneven settlement, causing severe stress in the membrane cover. They are also liable to microbial and/or chemical attack and sunlight, thus necessitating their protection with another cover. 14.10.4
Waste Materials
Fulfilling the engineering and isolation functions of covers can be alternatively achieved with a larger number of waste products. Many, however, may include trace elements or potential pollutants that may be toxic or harmful. It is important to investigate and test these materials before use. The most common of these products used in soils cover systems are summarized in Table 14.4.
SUMMARY AND CONCLUDING REMARKS
369
Table 14.4: Common waste materials used as substitutes for soil-based covers. Remarks
Waste Materials Fly ash
9 Good engineering and pozzolanic properties. 9 Quality and composition depend on the coal burnt.
Slags from iron- and steelmaking
9Used with caution because they tend to be physically and chemically unstable.
Non-ferrous slags
9 Contain high levels of toxic elements. 9 Source of contamination rather than a useful material for site remediation.
Domestic refuse incinerator ash
9 High heavy metal contents. 9Used with caution.
Overburden materials
9Result from mines, quarries, coal and aggregate extraction areas. 9 Can be used as soil cover substitutes.
Dredged silts
9May be heavily contaminated with toxic elements and other substances, such as oil and grease. 9Restricted uses.
Construction rubble
9 Good alkaline properties and physical structure. 9Useful product for use as a real layer, provided it is free of potential contaminants.
14.11 SUMMARY AND CONCLUDING REMARKS Covering systems are generally used to prevent infiltration of surface water into polluted sites and flow of gas to the atmosphere. Flow of water and gas is associated with natural processes that are time-dependent and that generally become increasingly adverse with time. Problems arising from the generation of methane are among the exception since the quantity of methane will generally decline with time. This is not to say that the effectiveness necessarily and inherently decreases with time but rather that covering systems may not be fully tested until many years after installation. In addition, we have to keep in mind that covering systems require the use of either natural soils or man-made materials. The physico-chemical properties of these materials are time- dependent. In assessing the long term performance of covering systems, we have to ask the following set of questions: (1) what are the possible environmental conditions that may influence the effectiveness of the designed cover system?, (2) what are the rate of changes of the physico-chemical properties?, (3) what is the critical point at which adverse environmental conditions may occur?, (4) how can the decreasing performance of the designed cover system be measured?, and (5) what are the mitigation measures? The methods available for monitoring and evaluating the performance of covering systems depend upon the nature of the system and the function that is to be examined. The time-dependent
370
COVERING SYSTEMS
changes that may affect system performance are major complicating factors in making an evaluation of performance. Most coveting systems will be multi-functional and no single monitoring measure will provide information on performance in relation to all functions.
(1)
(2)
From a regulation standpoint, the regulatory agencies should: Allow the planning and design of waste management unit closure by either of the following two approaches: (a) Categorize the waste unit characteristics and then design the cover system in compliance with prescribed regulatory closure requirements, based on type of facility, site and operational factors, and engineered containment provided in the operational design. (b) Conduct a site-specific waste pollutant assessment and demonstrate that the waste management unit, with the application of the closure system or other controls, if necessary, does not pose a threat to the environment. Allow the applicant to select alternative materials as components for the prescribed cover, so long as it can be demonstrated that such alternative materials will not compromise the intended function. Regulations should also provide for engineered alternatives, so long as it can be demonstrated that such alternatives are consistent with the applicable performance goals for protecting human health and the environment.
CHAPTER
FIFTEEN
LINING SYSTEMS
15.1
INTRODUCTION
Lining systems are used primarily to prevent a potential pollutant in a waste from migrating to the surface water, groundwater and food chain. The durability and service life of a given lining system in a waste impoundment depends to a great extent upon the specific leachate which contacts the liner. The leachate generated from the waste can be highly complex, and can continually change in composition with time. Thus, an important consideration in the operation of a disposal facility is that measures should be taken to minimize the variation in the character of the waste, as there is no single lining material which can resist all waste leachates.
Table 15.1: Classification of liners for waste disposal facilities (adopted from US EPA, 1980) Construction-based
Structure-based
Material and method of application
9 On-site construction
9 Rigid
9 Compacted
0 Raw materials brought to the site and liner construction on the site. Compacted soil. @ Mixed on site. 9 Prefabricated
0 Drop-in polymeric membrane liner. 9 Partially prefabricated
O Panels brought to the site and assembled on prepared site.
soil
O Soil. @ Soil-cement.
9 Semi-rigid
O Asphaltconcrete. 9 Flexible
0 Polymeric membranes @ Sprayed-on membranes
9 Admixtures
0 Asphalt-concrete. @ Soil-cement. 9 Polymeric membranes
0 Rubber. 0 Plastic sheets. 9 Sprayed-on liners 9 Soil sealants 9 Chemisorptive
371
liners
372
LINING SYSTEMS
Liner materials could vary in chemical composition from compacted soils to highly crystalline polymeric materials which are chemically resistant and have very low hydraulic conductivity. Similarly, wastes can vary from highly polar solvents, such as water, to highly nonpolar materials, such as lubricating oils and hydrocarbon solvents, as discussed in Chapter 3. Most wastes contain water as the principal carrier. All compounds, inorganic as well as organic, are to a certain degree soluble in water. Consequently, pollutants can be carried in the water. Dissolved organic contents in the leachate can interact with organic-based liner materials and, over extended period of time, cause failure of the lining system. There are indications that some organic chemicals have major adverse effects on clays, as discussed in Chapter 5.
15.2
CLASSIFICATION OF LINING SYSTEMS
Table 15.1 shows a classification scheme of lining systems on the basis of construction method, structure, and material and method of application (US EPA, 1980).
15.3
CLAY LINERS
Fine-grained soils are widely accepted as major components of lining systems in waste containment barriers. Figure 15.1 illustrates a typical cross section of a clay-based liner system. The properties of the compacted clay layer and its susceptibility to changes with time due to chemical interaction are major concerns in design.
Figure 15.1. Typical cross section of lining systems.
DESIRABLE PROPERTIES OF CLAY LINERS 15.4
373
D E S I R A B L E P R O P E R T I E S OF C L A Y L I N E R S
In order to determine the set of desirable clay liner properties, we must first identify the potential problems that arise in a barrier system. Desirable properties are those that overcome each and every one of potential problems listed in Table 15.2.
Table 15.2: Potential problems for clay barrier systems POTENTIAL PROBLEMS FOR CLAY BARRIER SYSTEMS (1) C O N S T R U C T I O N R E L A T E D 9 Dessication cracking 9 Penetration from equipment 9 Freeze-thaw cracking 9 Penetration from rocks in adjacent materials 9 Deformation cracking 9 QA/QC on material uniformity
(2) I N - A C T I V I T Y P E R I O D R E L A T E D 9 Dessication cracking 9 Slope instability 9 Freeze-thaw cracking 9 Ultraviolet degradation 9 Thermal stress cracking 9 Animal attack 9 Erosion (3) O P E R A T I O N R E L A T E D 9 Deformation cracking 9 Hydraulic conductivity changes 9 Penetration from equipment 9 Erosion 9 Penetration from loading 9 Slope instability 9 Chemical degradation (4) L O N G T E R M R E L A T E D 9 Dessication cracking 9 Biological degradation 9 Chemical degradation
15.4.1 Low Hydraulic Conductivity Low hydraulic conductivity is a prime reason for choosing clay as a major component of waste containment barrier systems. Care must be taken in choosing the material and in constructing the barrier in a way that achieves and maintains a low hydraulic conductivity. This property may vary over several orders of magnitude for a given soil, and it can be affected greatly by details of placement conditions, construction methods, and post construction changes. Detailed discussions on the effect of chemical species on changes of clay hydraulic conductivity are given in Chapter 5 and will not be repeated here.
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LINING SYSTEMS
15.4.2 High Attenuation of Chemicals Soil liners attenuate chemicals, to varying degrees, by lowering pollutant concentrations as the fluid passes through the materials. The greater the distance of travel, the lower is the expected concentration. Chemical attenuation processes include adsorption, precipitation, biological processes, oxidation-reduction reactions, acid-base reactions, and complex formation and chelation (Dunn, 1983). These processes are discussed in Chapter 5. Soil buffering capacity, which describes the capability of soil materials to resist a change in soil pH with the addition of strong acids or bases, can be used as a measure of the ability of a soil to adsorb and retain pollutants. The ability of a soil to retain pollutants can be explained by data presented in Figure 15.2 (Yong, Mohamed, and Warkentin, 1992). The results indicate that illite soil possesses better buffering capability in comparison to montmorillonite, and that kaolinite soil shows very little effectiveness as a buffering agent. Also, it can be noted that silicate minerals, carbonates, and gibbsite exhibit strong buffering upon addition of strong acid under slightly alkaline to slightly acid conditions.
Figure 15.2. Effective buffer capacity range for the selected four clay soils.
It should be noted that adsorption process will continue to occur until equilibrium is reached in the soil, after which breakthrough of pollutants will occur. Figure 15.3 illustrates typical breakthrough curves for reactive and non-reactive solutes obtained from leaching experiments (Yong, Mohamed, and Warkentin, 1992). Other processes, such as precipitation and biological clogging, may lead to a decrease in pore size within the clay and therefore to a decrease in hydraulic conductivity with time.
DESIRABLE PROPERTIES OF CLAY LINERS
375
Figure 15.3. Typical breakthrough curves for reactive (Ca and Na) and non-reactive (C1) solutes obtained from leaching experiments.
15.4.3 Low Diffusivity It is known that pollutants migrate through clay barriers by advection, diffusion and hydrodynamic dispersion, as discussed in Chapter 9. In most landfills, the effect of advection is negligible and pollutants migrate essentially by diffusion. In addition, the diffusion coefficients obtained from laboratory or field measurements are of major importance in calculating the diffusion front in a compacted clay barrier. Figure 15.4 shows the effect of various chosen values of diffusion coefficients on the calculated chloride concentration profiles in the substrate after 25 years of field performance. The sensitivity of the pollutant concentration profile to a change in the value of diffusion coefficient is more pronounced in the latter years. It can be seen that for a one order of magnitude increase (or decrease) in the diffusion coefficient, the depth of penetration of the pollutant concentration profile can be from four to five times greater (or lesser).
15.4.4 Ductility Clay liners must be sufficiently ductile to prevent cracking from dessication, from differential settlement below the facility, or from interaction with landfill leachate or impounded liquid. This issue is germane to concerns about the influence of leachate on liner hydraulic conductivity. Matrix hydraulic conductivity refers to flow through intra- and inter-particle pores whereas volume defect hydraulic conductivity refers to flow through gross void volume defects such as cracks, fissures, root holes, pinholes, tears, faulty seams, etc. (Gray, 1988). The former involves advective- or diffusioncontrolled flow through microscopic pores, and is described by Darcy's and Fick's Laws, respectively. The latter involves flow through gross void volume defects or macro-pores, and is not
376
LINING SYSTEMS
well described by either of these flow laws.
100
L%'~..._._.
-$
:~
8~
I 60
201 0
0
~ -
"--'---... ", ~*
~
-
-
4.74E-11
.......
*"
~.
1.6E-11
".,. ....
0.5
"..
1
- .....
,, , , .
1.5
2
2.5
Depth into bottom clay liner (m)
Figure 15.4. Effect of variation of diffusion coefficient on calculated chloride concentration profiles after 25 years.
The importance of hydraulic volume defects on hydraulic conductivity and the magnitude of the problem of predicting field leakage rates from compacted clay liners can be evaluated by recognizing the various mechanisms by which cracks could develop. These mechanisms are:
(1)
Mechanically induced cracks: (a) (b) (c)
(2)
Construction cracking due to tensile separation caused by compaction on weak subgrade or in laterally tmconfined areas; Settlement cracking due to excessive settlement or differential movement within a clay mass after placement; and Compaction cracking due to poor compaction at boundaries between successive vertical lifts of compacted clay.
Physico-chemically induced cracking." (a)
(b)
Gel cracking induced by a change in the inter-particle force balance resulting from: (i) Replacement of interstitial water with a lower dielectric organic solvent; and (ii) Exposure, or leaching with a highly saline aqueous solution. Dessication cracking induced by thermal removal (evaporation) of adsorbed water.
The persistence of cracks once they are initiated is a matter of serious concern in the case of clay liners and covers. Anderson and Kneale (1984) studied this phenomenon and found that crack initiation as a result of dessication corresponded to an average soil suction of 6 kPa. Subsequent closure depended on the rain fall pattern. All cracks appeared to close with the advent of winter rains. Following surface closure of cracks, field hydraulic conductivity tests revealed a value of
REQUIRED PROPERTIES OF CLAY LINERS
377
5x 10-5 m/sec for previously cracked locations, and 1x 10-8 rn/sec for intact sites. The implication of these findings is that cracks, once formed, continue to persist in spite of their apparent closure at the surface.
15.4.5 Slope stability Mechanical stability of the compacted clay layer on slopes has received little attention. The compacted clay strength must be adequate to maintain stability of the liner system on the side slopes of the facility.
15.4.6 Adequate Interface Strength The sliding resistance along the interface between the clay and other components of a lining system (e.g., geomembrane) can be very low (Mitchell et al., 1990). Low interface shear resistance may have contributed to the stability failure of a large hazardous waste landfill in California (Seed et al., 1990).
15.4.7 Long Term Stability The long term stability of a clay liner is a major concern in a facility that is to contain waste for several decades. Long term mechanical instability of a clay liner due to creep in the clay or along the interface between the clay and other components of the liner system is a possibility that should be considered. However, the major long term concern for clay liners is their chemical compatibility with leachate or impounded liquids rich in organic and inorganic compounds. It is important that these compounds not interact with the clay in a way that adversely affects its performance. These concerns are discussed in section 15.7.
15.4.8 Constructability A clay liner should be constructed using a soil that can be processed and compacted to the specified uniformity, dry density, and moisture content without excessive difficulty. Adequate trafficability on top of both the uncompacted and compacted material should be considered.
15.5
REQUIRED PROPERTIES OF CLAY LINERS
The US EPA requires that clay liners have a hydraulic conductivity less than or equal 1 x 10"9 m/sec. The minimum thickness of the compacted clay layer of the double liner system should be at least 1 m thick. To meet the low hydraulic conductivity requirement, certain characteristics of soil materials are required. First, the soil should be at least 29% fines (fine silt and clay sized particles). Some soils with less than 20% fines will have hydraulic conductivities below 10.9 m/sec but, at such low fines content, the required hydraulic conductivity value is much harder to meet. Second, plasticity index, which is defined as the numerical difference between liquid limit and plastic limit, should be greater than 10%. Soils with very high plasticity index, say between 30 to 40%, are sticky and, therefore, difficult to work with in the field. When high plasticity index soils are dry, they form
378
LINING SYSTEMS
hard clumps that are difficult to break down during compaction. Third, coarse fragments should be screened to no more than 10% gravel size particles. Soils with a greater percentage of coarse fragments can contain zones of gravel that have high hydraulic conductivities. Finally, the materials should contain no soil particles or chunks of rock larger than 25 to 50 mm in diameter. If rock diameter becomes a significant percentage of the thickness of a layer of soil, rocks may form a permeable window through a layer. As long as rock size is small compared to the thickness of the soil layer, the rock will be surrounded by the other materials in the soil. Further requirements to ensure that the hydraulic conductivity and system stability are achieved and maintained are listed in Table 15.3. The requirements are specified in terms of: (1) operational specifications, (2) testing requirements, (3) design considerations, (4) excavation and embankment constructions, and (5) quality control/quality assurance. It is clear that these requirements may lead to a satisfactory clay liner performance. It is, however, difficult to address all of the relevant factors through regulation. Ideally, the task of ensuring that a clay liner for a specific landfill or impoundment will function properly falls on the design engineer, the regulator, and the contractor, working as a team.
15.6
FACTORS CONTROLLING PROPERTIES OF CLAY LINERS
The performance of clay liners is usually measured in terms of hydraulic conductivity, flexibility and strength. These factors are controlled by the soil structure and can be altered during the life of the clay liner. The initial structure of a compacted soil is based on composition and environmental factors. The compositional factors include mineralogy, particle size and shape, composition and moulding water content. The environmental factors include blending, and method of compaction. The initial soil structure can be altered by physical, chemical and biological processes. The physical processes include consolidation, wet-dry cycles, freeze-thaw cycles, and erosion. The chemical processes include dissolution, mineralogical transformation, sorption and precipitation. Biologically, soil structure can be altered by various bacterial actions occurring in the soil. The effect of these processes may contribute to a final soil structure which leads to the failure of the clay liner. Since clay liners are usually intended to perform for extended periods of time, a primary requirement of a clay liner design is to minimize the negative impact of these processes on long term performance.
15.6.1 Compositional Factors
Mineralogy The soil mineral types determine potential ranges of soil properties, including plasticity, shrink-swell, adsorption, ion exchange, flocculation-de-flocculation, etc. Soil minerals of all types and their physical and chemical properties are discussed in Chapter 4. The most important soil components with respect to the attainment and maintenance of low hydraulic conductivity, durability, and long term stability are clay minerals. Sodium montmorillonite might at first seem to be best barrier material. However, the very high plasticity and water holding capacity of expanded sodium montmorillonite makes it susceptible to collapse when exposed to chemicals of various types. This collapse could lead to an increase in hydraulic conductivity. Thus, non-expansive illitic clays may be preferable in many cases. Changes
FACTORS CONTROLLING PROPERTIES OF CLAY LINERS
379
in clay hydraulic conductivity as a function of clay mineral type are discussed in Chapter 5. Soil minerals that may be soluble in the chemicals in a waste contaminant facility should not be used in liner construction.
Table 15.3: Design and construction requirements of clay liners. DESIGN AND C O N S T R U C T I O N R E Q U I R E M E N T S OF C L A Y L I N E R S (1) O P E R A T I O N A L S P E C I F I C A T I O N S 9 Depth of soil liner. 9 Water content wet of optimum should be specified. 9 Dry density greater than 95 percent of standard Proctor maximum dry density. 9 Lift thickness not more than 150 mm. 9 Number of passes per layer. 9 Surface scratches between lifts. 9 Compaction equipment (e.g., sheep-foot roller, weight of roller). (2) T E S T I N G 9 Adsorption and diffusion characteristics. 9 Grain size, liquid limit, plastic limit, compaction, water content. 9 Hydraulic conductivity. 9 Stress-strain behaviour. (3) DESIGN
9 Seepage control factors: [] Waste disposal characteristics; [] Waste-effluent characteristics; [] Sink-source distance.
[] Attenuation capacity; [] Groundwater characteristics;
(4) E X C A V A T I O N AND E M B A N K M E N T C O N S T R U C T I O N 9 Side wall slope (typical side wall slopes are not steeper than 3 horizontal to 1 vertical). 9 Bottom slope and configuration (bottom slope greater than 2 percent to ensure drainage towards sump). 9 Drainage system (minimum thickness of 0.3 m and hydraulic conductivity of 1 x 10 -2 m/see). 9 Monitoring wells. 9 Field compaction. (5) Q U A L I T Y C O N T R O L / Q U A L I T Y ASSURANCE 9 Soil characteristics. 9 Roller characteristics. 9 Compaction operation characteristics. 9 Testing of clay liner, and granular drainage blanket during construction.
380
LINING SYSTEMS
Mitchell and Jaber (1990) indicate that in the early 1980's, a 0.15 m thick liner, for a liquid impoundment, was constructed using native carbonate soil over a limestone subgrade. Chromic acid was disposed of in the impoundment. The carbonate soil was dissolved by the acid, and large quantities of industrial waste were released to the environment.
Particle Size and Shape The geometry of a flow channel is a function of the size, shape and or orientation of the soil particles. In the case of clay soils, because of the plate-like nature of the soil particles (especially the clay minerals), clay soil hydraulic conductivity in the horizontal direction is different than that in the vertical direction. This is very significant for soil particles with a large diameter to thickness ratio. According to the Kozeny-Carman relationship, the hydraulic conductivity, Kw, of a granular soil, can be obtained in terms of the tortuosity of the flow path, porosity, surface area and flow factor, etc., as follows: CsYn 3 Kw :
rla:2S~2(1 -n 2)
[15.1]
where Cs is a shape factor, with 0.4 generally taken as a standard value, z is tortuosity, which is the ratio of the effective flow path to the actual thickness of the sample being evaluated, n is porosity, Sa is the surface area per soil volume of particles, and y and r I are the density and viscosity of the permeant, respectively. To apply the Kozeny-Carman equation, the following conditions need to be satisfied: (1) a relatively uniform particle size, (2) laminar flow of liquid throughout the pores, (3) validity of Darcy's law, and (4) absence of long- and short-range forces of interaction.
Composition The composition of soil determines the possible range of its properties. In order to obtain a clay liner that will perform satisfactorily, i.e., be stable and have a hydraulic conductivity smaller than 1• 10.9 rn/sec, the material used to construct the liner must be chosen carefully. Soils that are appropriate for use in liner construction should have at least 20% fines, no more than 10% gravel size, and no soil particles or chunks of diameter larger than 25 mm (Goldman et al., 1988). Well graded soils are the easiest to compact to high densities with a uniform distribution of small pores. The plasticity index should be greater than 10% and less than 30%. However, if the plasticity index is greater than 30%, the material becomes sticky when wet and may form large clods. It then becomes difficult to construct, as experienced both in the laboratory and in the field by Elsbury et al., (1988).
Chemistry and Pore Water The composition of both soil and water has a significant effect on the plastic behaviour of a soil. Warkentin (1961) noted that Na-montmorillonite has a liquid limit almost three times larger when water is used as the soil mixing liquid rather than a molar solution of NaC1. It has been also shown that theflow index (the negative slope of water content versus logarithm of number of blows in Casagrande's device) depends upon the nature of the clay minerals and electrolyte concentration (Yong and Warkentin, 1975). The effect of relative concentration of monovalent to divalent cations on the liquid limit of
FACTORS CONTROLLING PROPERTIES OF CLAY LINERS
381
soils was noted by Landreth (1983). In his study, soils containing 25% and 45% clay size particles (less than 2 lam) and solutions of NaC1 and CaC12 in water were used. A combination of two levels of salt concentration, 2 and 50 g/l, and two levels of sodium adsorption ratio (SAR), 2 and 20 (mmoles/1) 1/2,yielded the following set of conclusions: The higher the proportion of fines in a soil, the larger is the effect of pore fluid chemistry; (1) (2) The liquid limit is more affected by liquid chemistry than is the flow index; In a soil containing montmorillonite, both the SAR and the total salt concentration (TSC) (3) significantly affect plastic characteristics. The higher the SAR and the lower the TSC, the larger is the liquid limit; and In a predominantly kaolinite soil, the lower the TSC, the higher is the liquid limit. (4) The above discussion relies only on two liquid phase characteristics. It is, however, to be expected that other characteristics like pH, abundance of non-native ionic species and chelating agents, density, viscosity, dipole moment, dielectric constant, etc., may have a significant impact on soil structure, and thus on the limits and on relevant geotechnical characteristics. The compatibility between clay liners and chemical waste is discussed in Chapter 5.
Moulding Water Content The water content of soil at the time of compaction is an important variable controlling the engineering properties of soil liner materials. Soils compacted at water contents less than optimum (dry of optimum) tend to have a relatively high hydraulic conductivity whereas soils compacted at water contents greater than optimum (wet of optimum) tend to have a low hydraulic conductivity, as shown in Figure 14.9 in Chapter 14. It is usually preferable to compact the soil wet of optimum to achieve minimum hydraulic conductivity. 15.6.2 Environmental Factors
Compaction Method and Compaction Different methods of compaction can lead to large differences in hydraulic conductivities for clay compacted wet of optimum at the same dry density (Mitchell et al., 1965). For example, samples prepared by kneeding compaction, a method that results in a more dispersed structure due to the introduction of large shear strains during compaction, have lower hydraulic conductivity than samples prepared by static compaction. For this reason, a heavy sheep-foot roller that introduces large shear strains during compaction in the field should be used. Furthermore, the liner should be compacted in several layers so that the feet of the compactor can penetrate and import shear strains to all the layers. The influence of compacted effort on soil materials and the resultant effect on hydraulic conductivity are discussed in Chapter 14.
Size of Clods The size of pores in a compacted clay layer depends on the size of particle aggregation, or clods. The larger the clods, the larger the inter-aggregate pores, and the larger the hydraulic conductivity of the compacted clay (Olsen, 1952). The effect of clod size on hydraulic conductivity of a compacted clay is demonstrated by Daniel (1984). The hydraulic conductivity of a compacted clay was found to increase by approximately 2 orders of magnitude for an increase of clod size from 1.4 mm to 9.4 mm. Clay clods should be destroyed in the field either before compaction, by
382
LINING SYSTEMS
mechanically pulverizing them, or during compaction as suggested by Benson and Daniel (1990). Destruction of clay clods during compaction may be achieved by: (1) compacting the soil at a high moulding water content so that the clods are made soft and weak, and can be easily crushed during compaction, or (2) compacting the soil at a lower water content but with an extremely heavy roller that crushes the clods. In either case, minimizing the hydraulic conductivity should be the basis for selecting the technique to be used for destruction of clay clods.
Bonding Between Lifts Lift interfaces have important ramifications with respect to the overall hydraulic performance of a soil liner. In full-scale tests (Brown et al., 1983), a fluid was shown to penetrate the liner through a vertical crack, then travel along the interface between the two clay lifts until it encountered another vertical crack. Such defects in the liner provided the fluid with a series of preferential paths that resulted in a much shorter breakthrough time and larger hydraulic conductivity than expected. To bond lifts together, the surface of the previously compacted lift should be rough so that the newly placed lift can effectively blend into the surface. If necessary, the surface of the previously compacted lift can be roughened by discing the soil to a depth of approximately 25 mm. Discing the soil involves plowing up the soil surface to a shallow depth so that the surface is rough, with no abrupt interface between lifts. Compactors with long "feet" on the drums are useful in blending one lift into another. During the first few passes of the compactor, the feet sink through a loose lift of soil and compact the newly placed lift onto the surface of the previously compacted lift. Using a roller with feet that fully penetrate a loose lift of soil is recommended to bond lifts and to minimize high horizontal hydraulic conductivity at lift interfaces.
15.7
POST CONSTRUCTION CHANGES
The compositional and environmental factors discussed in the previous sections are all controllable by appropriate soil type and use of proper construction procedures. A carefully designed and implemented construction quality assurance plan can result in a liner that initially satisfies or exceeds the required hydraulic conductivity of 1x 10-9 m/sec (Daniel, 1987; Lahti et al., 1987; and Gordon et al., 1989). However, changes may occur after construction that may adversely affect the properties of the compacted clay liner. These changes can be classified into: (1) Physical processes: settlement, wet-dry cycles, freeze-thaw effects, and erosion-dispersion; (2) Chemical processes: dissolution, precipitation, complexation, sorption, and mineralogical transformations, and (3) Biologicalprocesses: biodegradation, and biotransformation.
15.7.1 Physical Processes
Volume Changes Volume changes that may occur following compaction of the soil liner should be estimated because a change in bulk density will result in a change in transport properties. If a soil is marginal in terms of its potential for generating a particular "as compacted" hydraulic conductivity value, and if it has a high swell-shrink tendency, it will likely be rejected as a liner candidate. Generally, volume changes have a detrimental effect upon soil hydraulic conductivity.
POST CONSTRUCTION CHANGES
383
If the moisture content of the soil element under consideration changes, relative to the ascompacted strata, then swelling or shrinkage may occur provided the necessary conditions are present. According to Nayak and Christensen (1971), the following factors determine the intensity of volume change: (1) type and amount of clay mineral, (2) nature of pore fluid, (3) initial placement condition, e.g., structure and density, (4) stress history, (5) temperature, (6) volume change permitted during swelling pressure measurements, (7) shape, size, and thickness of the sample, and (8) time. Basic information on fluid chemistry-clay interaction is discussed in Chapter 5 and in Grim (1962, 1968), Sposito (1989), Yong, Mohamed, and Warkentin (1992), and Mitchell (1993). The tendency of compacted soil to swell can be stated in general terms as: (1) The lower the moisture content and the higher the density, the higher is the swelling potential; (2) An increase in moulding water content, at a given density, causes a decrease in swelling pressure and swell; and (3) An increase in density at low moisture content causes an increase in swell. An increase in density at high moisture content does not alter drastically the swell characteristics.
Figure 15.5. Effect of pore fluid chemistry replacement on soil creep.
Differential Settlement Excessive differential settlement under a clay liner due to landfill loading can result in cracking of the clay layer. This effect is minimized if the clay layer is thick and ductile. The effect of leachate on the resultant axial strain during creep testing was studied by Yong et al., (1985). They examined the effect of five different types ofleachate, namely: 0.025 N sodium silicate (pH = 11.40), 0.025 N sodium chloride (pH - 7.33), 1 N sodium acetate (pH - 5.0), saturated calcium hydroxide solution plus 10% carbon dioxide air (pH = 12.26), and distilled water (pH = 6.94) on the resultant axial strain in an incremental stress controlled drained creep tests. Experiments were conducted with
384
LINING SYSTEMS
distilled water, for the first 6.94 days, after which leachate solutions were introduced. The results obtained are shown in Figure 15.5. The introduction of sodium silicate, which acts as a dispersing agent, caused a significant increase in observed axial strain in comparison to the reference natural untreated sample. Similar results were obtained with 1 N sodium acetate acid leaching solution. The large increases in axial strain, with sodium silicate and sodium acetate solutions, from 3 to 5 times higher than those observed in the reference sample, may be attributed to the degradation of the basic fabric, which in turn resulted in higher compressibility and, hence, higher strain values of the clay sample. It would appear that the development of a higher net repulsion, in the two cases cited, reduced the effectiveness of the initial bonding between particles. Thus, a degradation of the basic fabric unit stability occurred. The likely fundamental items controlling axial strain performance are shown in Figure 15.6. The axes shown in the figure provide an indication of the greater or lesser influence of the four major issues (i.e., attraction, repulsion, fabric reinforcement, fabric degradation) on the response of a clay barrier to chemical attack. With regard to the stability of fabric units and the overall fabric the results, shown in Figures 15.5 and 15.6, indicate that the bonding mechanisms and inter-particle forces are the two major factors which control the structural stability of the clay barrier. Reinforcement, i.e., strengthening of the structural stability of the fabric unit, can occur as a result of: (1) precipitation of more bonding material due to the leaching action of intruding solutions, (2) net decease in repulsive forces because of replacement of ions in the pore water, and (3) a combination of both (1) and (2). Conversely, degradation or weakening of the structural stability of the fabric unit will occur if: (a) the bonds are degraded or destroyed, (b) the net repulsive forces are increased, and (c) a combination of the two preceding factors.
Reinforcement of the fabric unit
225
Repulsive forces dominant
C
EL (/) (/)
i /11
200
~ m m
L-
9H20 ~num~
mmn,,~ ~umm
~','n ~
~mmmmanm~
mmnnnommmmmmmmmmm ~
DW
O3
(~ 175 E
.i.-i
_
150 10
mmmmm m - = mmmum~
i I
Attractive forces dominant
'
i
CH3COONa Degradation of the fabric unit
I
I
I
12
14
16
18
Ultimate strain (%) Figure 15.6. Variation of ultimate stress-strain relationships in drained creep test as a function of different chemical treatments.
POST CONSTRUCTION CHANGES
385
Slope Instability A slope failure that involves liner breakage can result in the localized release of pollutants to the environment. Slope instability or excessive down-slope movements can result in cracking of the clay liner. Such instability can develop when water is added to the landfill, or can be triggered by seismic activity. Slope stability must be ensured by use of appropriate side slope angles, liner system materials, and waste placement techniques that prevent interface slip within the liner system. High interface friction geomembranes are available. Textile reinforcements, such as geogrids anchored at the top of the slope, have also been used to increase the stability of landfill liners (Giroud and Beech, 1989).
Wet-Dry Cycles Drying and wetting cycles have been recognized for some time as one of the major causes of soil liner failures. In the late 1940's and early 1950's, the Prairie Farm Rehabilitation Administration (PFRA) (1951) carried out an extensive testing program on canal liners. It was reported that considerable deterioration was caused by allowing canals to dry out at the end of each growing season. Rollins and Dylla (1970) conducted a major field test program involving the construction of twelve experimental reservoirs located at Yerrington and Fallon, Nevada. It was found that wet-dry cycles resulted in a decrease in dry density and corresponding increase in hydraulic conductivity, with typical seepage rates of 200-300 mm per day. Recently, Boynton and Daniel (1985) desiccated slabs of compacted clay and measured the hydraulic conductivity at different effective confining stress. It was concluded that at low confining stress, the desiccated soils were much more permeable than the control, as shown in Figure 14.11 in Chapter 14. At high confining stress, however, the desiccated soils were no more permeable than the control. It appeared that the application of large compressive stress (> 35 kPa) closed the desiccation cracks that had formed and, in combination with hydration of the soil, essentially fully healed the damage done by dessication. More recently, Johnston and Hang (1992) studied the effect of wet-dry cycles on hydraulic conductivity. Their results have shown that: (1) samples compacted near optimum or dry of optimum underwent very little change in hydraulic conductivity, (2) the lowest hydraulic conductivity obtained for specimens subjected to two wet-dry cycles occurred for specimens compacted near optimum moulding water content, and (3) moulding water content was found to have more impact on changes in hydraulic conductivity than compaction effort for specimens subjected to wet-dry cycles.
Freeze-Thaw Cycles Repeated freeze-thaw cycles of a clay liner can result in a significant increase in hydraulic conductivity of the clay and a loss of strength. In tests on four fine-grained soils, Chamberlain and Gow (1979) found that repeated freeze-thaw cycles caused significant structural changes in the clay such as severe cracking, that resulted in up to two orders of magnitude increase in their hydraulic conductivity. Similar results were reported by Yong et al. (1985) and Mohamed et al., (1991) in relation to the effect of cyclic freeze-thaw on soil structure and the resultant geotechnical properties. More recently, Mohamed et al., (1993) have showed that the increase in hydraulic conductivity was found to be higher for soils with large plasticity index and high moisture content. This can be attributed to the formation of vertical cracks during freezing as well as rearrangement and aggregation of soil particles. The changes in hydraulic conductivity as a function of number of freeze-thaw cycles are discussed in Chapter 14.
386
LINING SYSTEMS
15.7.2 Chemical Processes Dissolution
Clay liner dissolution can be achieved by an infiltrating chemical that dissolves the exposed surfaces of the clay layer. Portions of the clay structure may be dissolved if exposed to organic or inorganic acids and bases, as discussed in Chapter 5. Acids have been reported to dissolve aluminum, iron, alkali metals and alkaline earths while bases will dissolve silica (Grim, 1953). Since clay minerals contain both silica and aluminum in large qualities, they are susceptible to partial dissolution by either acids or bases. The solubility of clays in acid varies with the nature of the acid, the acid concentration, the acid to clay ratio, and the temperature and duration of treatment (Grim, 1968). This in turn could permit even weak acids to dissolve clays under some conditions. The study conducted by Pask et al., (1954), where several clay minerals were boiled in acid, showed that the percentage of solubilization of alumina was 3% for kaolinite, 11% for illite, and greater than 33% for montmorillonite. Dissolution of clay minerals in solution may proceed via one of the following mechanisms: (1) The replacement of adsorbed cations by H § ions from solution. This process can be regarded as a step in the hydrolysis of the clay minerals. For example, Na-clay + HOH-.
(2) (3)
H - c l a y + Na + O H
[15.2]
Diffusion of atoms from the framework to the solid liquid interface and from there to the bulk solution; and De-polymerization of the tetrahedral and octahedral layers. In this process, protons are attached to nucleophillic sites on the surface or in the framework of the mineral, followed by the breaking of Si-O, Mg-O, Fe-O, or A1-O bonds. The process begins at either the broken bond surfaces or the flat oxygen planes.
These processes result in the formation of neutral or anionic species of mono-silica acid and hydrated species of the exchangeable and lattice metallic cations, such as Na, K, Ca, Mg, A1, and Fe. The dissolution of heavy metals that are retained in soils is demonstrated in Figures 15.7 to 15.9 for illite, natural clay and kaolinite, respectively (Phadungchewit, 1990). The effluent pH decreased from 8 to 5 when the cumulative acid applied reached 60 cmol/kg dry soil, and from 5 to 3 when the cumulative acid applied exceeded 150 cmol/kg dry soil, as shown in Figure 15.7. These changes in pH resulted in a change in the movement pattern of the heavy metals. The relative concentrations of the metals, especially lead and copper, were still low at pore volumes 1 to 6. After pore volume 6 (when the cumulative acid applied exceeded 60 cmol/kg dry soil), the effluent pH dropped to less than 5, and the relative concentrations of the metals, particularly zinc and cadmium, increased sharply. The relative concentrations were lower for lead and copper than for zinc and cadmium. These results indicate that the mobility order of the heavy metals in the illite soil column is lead < copper < zinc < cadmium. It can be seen from Figure 15.7 that the breakthrough points of lead and copper were around pore volumes 12 and 11 when the cumulative acid applied reached 120 and 110 cmol/kg dry soil, respectively, and the effluent pH became lower than 4. The breakthrough points of zinc and copper were around the same pore volume of 8 (80 cmol/kg dry soil). The effluent pH was less than 5 at the
POST CONSTRUCTION CHANGES
387
breakthrough points. The relative concentration of the heavy metals increased beyond the breakthrough points (i.e., Ce/Co > 1), an indication that the effluent concentration of heavy metals became higher than the influent concentrations.
Figure 15.8. Breakthrough curves of lead, copper, zinc, and cadmium for natural clay soil from Quebec. Y 1 and Y2 refer to the left and the right vertical axes, respectively.
388
LINING SYSTEMS
Figure 15.9. Breakthrough curves of lead, copper, zinc, and cadmium for kaolinite soil. Y 1 and Y2 refer to the left and the right vertical axes, respectively.
A plausible explanation of these results is that heavy metals which had already been retained in the soil became remobile. The remobilization of heavy metals could be due to the dissolution of heavy metals that were retained in the soil. Another cause of remobilization is the exchange of the heavy metals, which were retained on the soil exchangeable sites, with H ~ ions. The effect of acid dissolution on clay hydraulic conductivity of bentonite clay was reported by Pavilonsky (1985). The study has shown that permeation with acids leads to an approximate 5 to 12 fold increase in the hydraulic conductivity of a montmorillonite specimen in comparison to that determined using water as permeant. The changes in soil hydraulic conductivity could be attributed to: (1) extraction of lattice aluminum ions from the octahedral sheets of the clay minerals, (2) ion exchange on the surface of the clay minerals due to replacement of naturally adsorbed cations of lower valence by the extracted aluminum ions, hence a reduction in the thickness of the diffuse double layer, and (3) increase in effective pore space and, hence, in the tortuosity factor, thereby resulting in a higher hydraulic conductivity. The effect ofpH on the behaviour of clay barriers can be illustrated by examining how a clay mineral hydraulic conductivity changes as pH changes. This is illustrated in Figure 15.10. In acidic conditions, the hydraulic conductivity is high due to flocculation while in basic conditions, it is low due to dispersion.
Precipitation Precipitation occurs when the transfer of solutes from the aqueous phase to the interface results in accumulation of a new solid phase. Precipitation can occur on the surfaces of the soil solids or in the pore water. Surface precipitation is a multilayer adsorption process which occurs on metal oxides, clay solid surfaces and organic matters, at relatively high concentrations of the pollutant
POST CONSTRUCTION CHANGES
389
species involved. Pore water precipitation occurs when the ionic activity product exceeds the solubility product, and at higher concentration of the solutes. The pH of both the soil and soil pore water, and the concentration of the solute are important factors that control precipitation. By contrast with hydroxides, the extent of precipitation of metallic carbonates and sulfides in soils is controlled not only by the solubility products of the individual carbonate and sulfides, but also by the partial pressure of gaseous carbon dioxide and hydrogen sulfide, respectively. This occurs because the content of the carbonate, CO32, and of sulfide anions, S2, is controlled by the amount of CO2 and H2S dissolved in the soil water (Evans, 1989). The amount of CO2 or H2S dissolved in water is controlled by the partial pressure, pCO2 and pH2S, of the respective gases in the soil atmosphere. The content of dissolved gas is related to the partial pressure by Henry's constant.
.~ .........
Kaolinite m
m
m
m
m
m m
m
m
mmm
m
m
mlm
mmm m
mmmm
m
m
m
m
m
m
m
m
m m
m
m
(j
Illite
0 c~ rj co
Montmorillonite
Acid
Base pH
"r
Figure 15.10. Influence of pH on hydraulic conductivity of different clays.
The effect of sodium hydroxide precipitation on the resultant clay hydraulic conductivity was investigated by Lentz et al., (1985) at pH values of 9, 11 and 13. Little effect on the hydraulic conductivity was measured at pH 9 or 11. However, hydraulic conductivity decreased by approximately one order of magnitude for permeant solution of pH 13. The decrease was interpreted to have resulted from the use of tap water rather than distilled water for preparation of the permeant. The tap water contained 300 mg/1 calcium carbonate, and the resulting formation of calcium hydroxide may have blocked some of the processes. Similar results were obtained by Pavilonsky (1985), and Dunn and Mitchell (1984).
Complexation A chemical species in which a cation is covalently bonded to one or more coordinating groups is referred to as a complex ion. The coordinating group is termed the ligand. Water is the most common ligand encountered in aqueous solutions. For simplicity, cations in aqueous solution
390
LINING SYSTEMS
are often written without the coordinated water molecules. However, a more accurate formulation would include such groups. For example, AP § should be written as AI(H20)63§ As an example of the formation of a complex ion (Benfield et al., 1982), consider the interaction between the hydrated zinc ion and ammonia: Zn(H20)42. + M e 3 ~ Zn(NH3)(H20)32. + H 2 0
[15.3]
Zn(NH3)(H20)32. + N H 3 ~ Zn(NH3)x(H20)22+ + H 2 0
[15.4]
Zn(NH3)2(H20)22+ + N H 3 ~ Zn(NH3)3(H20) 2+ + m 2 0
[15.5]
Zn(NH3)3(H20) 2+ + N H 3 ~ Zn(NH3)42. + H 2 0
[15.6]
Eqs. [ 15.3] to [ 15.6] represent the stepwise formation of the various ammonia complexes of zinc. An aqueous solution which contains zinc ions and ammonia will usually contain all these complexes in equilibrium with each other. If complexation is a significant factor in soil interaction with metals, the displacement or transport of the metals will not be totally govemed by adsorptive interactions and precipitationdissolution considerations (de Haan and Swermann, 1978). The change in mobility of the metal is not only due to the fact that the metallic cation is complexed, but also because this complexation may result in a transformation of the cation into an anionic complex. The characteristics of the complexing agent are obviously important. From a thermodynamic point of view, it is necessary to consider the availability of the metallic cations in competition with other cations, and the stability of the complexes. Complexes formed in conjunction with one ligand would be less stable than those formed with multi-ligands. For example, the mononuclear complexes formed between a central metal ion and a number of anions or ligands may be positive, negative or neutral (Butler, 1964). Cadmium combines with chloride ions to form CdCl, CdClf, CdCl3, and CdCl42 complexes. Variably-charged clay particles such as oxides, hydroxides and soil organic matter can potentially form inner-sphere complexes with cations, as discussed in Chapters 4 and 5. The formation of inner-sphere surface complexes results in the formation of covalent bonds. Therefore, the adsorbed species cannot easily exchanged (Evans, 1989). Borderline and soft acids, such as Hga+ and Pb 2§ tend to form inner-sphere complexes. Cations that are attracted to the edge and surfaces of particles without losing their water of hydration form outer-sphere surface complexes. Outersphere complexes are weakly associated cations and participate in cation exchange reactions. Hard acids, such as Na § Ca2§ and Mg 2§ tend to form outer-sphere complexes. Interlayer complex formation results from the entry of chemical species, especially small and polymeric organic compounds, between the layers of clay minerals, resulting in an increase and a
POST CONSTRUCTION CHANGES
391
decrease in the basal spacing of non-expansive and expansive clay minerals, respectively. Studies of interlayer complex formation in kaolinite by Weiss (1962) reported that ammonium acetate could change the basal spacing of kaolinite from about 0.7 to 1.4 nm. Formamide and urea could have similar effects. Interlayer complex formation of acids or bases from aqueous solution is associated with cation exchange. For example, in the case of amines, the protonated species exchange interlayer cations (Vansant and Uytterhoeven, 1973). The influence of chemicals on the hydraulic conductivity, then, depends on the valence of cations and pH which in turn influence the edge and surface charges of the clay particles. The effect of permeating organic molecules in the clay-pollutant system on the hydraulic conductivity can be explained on the basis of log octanol-water partition coefficient and molecular weight of the organic compounds, as discussed in Chapter 5. Generally: (1) The hydraulic conductivity to an organic liquid decreases as the log of its octanol-water partition coefficient increases; (2) The repulsive energy increases with an increase in the octanol-water partition coefficient. Correspondingly, the hydraulic conductivity decreases; (3) The greater the dielectric constant, the greater the hydraulic conductivity, in as much as the dielectric constant serves as an approximate measure of a liquid's hydrophobic or hydrophillic character; and (4) As the molecular weight of the organic chemical increases, the clay soil hydraulic conductivity decreases.
Adsorption With regard to pollutant-soil interaction, the adsorption reactions which occur are processes by which pollutants in solution become attached to the surface of soil (solids) particles through mechanisms which balance the forces of attraction from the soil solid surfaces. These processes are govemed by the surface properties of the soil solids, and the physico-chemical properties of the pollutants. The net energy of interaction due to adsorption of a solute ion or molecule onto soil constituent surfaces is the result of both short-range forces, such as covalent bonding, and long-range forces, such as electrostatic forces. Considering then net negative charge existent at or near the surfaces of clay minerals as the source of the surface-active nature of the particles, the presence of counter-ions (cations) in the soil solution will produce diffuse ion layer, as discussed in Chapter 5. A knowledge of the interactions provides one not only with a better understanding of water-holding capability of the material, i.e., soil-water energy characteristics, but also with an appreciation of the buffeting nature of the material. The effect of exchangeable cations on water-holding capability of bentonite clays can be best illustrated by examining liquid limit changes as the type of cation changes. This is illustrated in Figure 15.11 (Sridharan et al., 1986). As the cation valence increases, the liquid limit decreases, which is consistent with the diffused ion layer theory, as discussed in Chapter 5. The effect of pH on the distribution of Ca 2+, Mg 2+, Na +, K +, H +, and A13+ions is shown in Figure 15.12 (Brady and Weil, 1996). The hydrogen and aluminum exist in two forms: (1) that tightly held by the pH-dependent sites (boundform), and (2) that associated with permanent negative charges on the colloids (exchangeableform). Under very acidic conditions, exchangeable aluminum and hydrogen ions, and bound hydrogen and aluminum tend to dominate. At higher pH values, the exchangeable bases tend to dominate. At intermediate pH values, aluminum hydroxy ions, such as AI(OH) 2+ and AI(OH)2+, are prominent.
392
LINING SYSTEMS
Figure 15.12. General relationship between soil pH and the cations held by soil colloids.
The effect of exchangeable cations on the behaviour of clay barriers can be investigated in view of experimental data from Sridharan et al., (1986). The calculated hydraulic conductivities from
POST CONSTRUCTION CHANGES
393
consolidation experiments are shown in Figure 15.13, and discussed below. Monovalent Effect: At a void ratio of 2, sodium bentonite is nearly twice as permeable as lithium bentonite. At the same void ratio, potassium and ammonium bentonite are nearly six times more permeable than sodium and lithium bentonite. The behaviour of potassium and ammonium bentonite is consistent with their suppressed diffuse double layer thicknesses, as discussed in Chapter 5~ Divalent Effect: Saturation of bentonite by divalent cations makes it ten times more permeable than sodium and lithium bentonite. The adsorption of divalent cations by a clay surface leads to a marked reduction in the thickness of the diffused ion layer. In the mean time, it is expected that an outer-sphere surface complex between exchangeable calcium ions and a pair of opposing siloxane hexagonal cavities is formed, as discussed in Chapter 4. The dual effects of a reduction in the diffused ion layer thickness and the formation of quasi-crystals, due to complexation, substantially increase the effective pore spaces for water flow, yielding high hydraulic conductivities. Comparing the hydraulic conductivities of the divalent bentonite at a void ratio of 2, barium bentonite is nearly twice as permeable as calcium and magnesium bentonites. Trivalent Effect: Trivalent bentonite is in turn more permeable than divalent bentonite. Aluminium bentonite exhibits a higher range of hydraulic conductivity than iron bentonite. This can be explained in view of the higher liquid limit values of iron bentonite (120%) than aluminum bentonite (108%), as shown in Figure 15.11.
Figure 15.13. Influence of exchangeable cations and anion on hydraulic conductivity of bentonite.
394
LINING SYSTEMS
Anion Effect: Treatment of bentonite clay with fluoride degrades the mineral structure and decreases the cation exchange capacity (CEC) of bentonite clay by 60%. The decrease in CEC is attributed to: (1) aluminum, as AI(OH)3, extraction from clay lattice, (2) aluminum exchange with naturally adsorbed cations on clay surfaces, and (3) decrease the number of exchange sites on clay surface. The decrease in CEC and surface area (due to particle aggregation), on fluoride treatment, leads to a decrease in the diffuse double layer thickness and in the liquid limit values. At void ratio of 2, the hydraulic conductivity of fluoride-treated bentonite is approximately twelve times higher than that of its untreated counterpart, as shown in Figure 15.11. This is attributed to the decrease in the thickness of the diffused ion layer due to reduction in both the CEC and the specific surface area. Furthermore, the precipitated AI(OH)3 binds adjacent silicate surfaces together, resulting in particle aggregation and migration of hydroxide ions into solution. The fixed AI(OH)3 does not enter the interlayer space but is probably located on the clay domains or in the pore between domains, affecting their aggregation by electrostatic bonding. Also, due to the interaction of fluoride with bentonite clay, a cryolite mineral, as indicated in X-ray diffraction patterns, could be formed. 15.7.3 Biological Processes Biological processes and their significance in clay barriers geochemistry has been recognized for many years. There are at least three ways in which bacteria in clay barriers may influence the form and concentration of metals in the aqueous phase (Ehrlich, 1978). Bacteria may reduce the oxidized form of certain metals (e.g., Fe 3~ to Fe2§ and may in fact use the oxidized form as a terminal electron acceptor in oxygen-deficient conditions. This represents a type of respiration. Other bacteria may catalyse the oxidation of reduced metals (e.g., Fe 2~ to Fe3§ and may even be able to conserve the free energy released for growth and metabolism. Such bacteria normally require oxygen but not organic carbon, since they can obtain their carbon for biosynthesis from inorganic sources such as carbon dioxide. Depending on the metals involved, such redox reaction may greatly affect the solubility of the metal and, hence, its mobility. The various biological processes are discussed in Chapter 21. In biodegradation processes, the organic pollutants are used as a source by the bacteria, with water and carbon dioxide as the principal products of respiration under aerobic conditions, and water, carbon dioxide, and methane or hydrogen sulfide under anaerobic conditions. Thus, there is the potential for a highly toxic pollutant to be degraded to harmless products. As a broad generality, aromatic hydrocarbons such as those found in petroleum production (benzene, toluene and xylene, for example) are considered to biodegrade under aerobic conditions. On the other hand, halogenated hydrocarbons tend to biodegrade rather slowly under anaerobic conditions and more slowly or not at all under aerobic conditions. It is also recognized that the biodegradation process is not always complete and that, in some cases, the products of degradation may be more toxic than the original pollutants. For example, as a consequence of the biodegradation of DDT, the more toxic DDE has been observed to accumulate as an intermediate in the degradation process. Microorganisms can be either positively or negatively charged. However, most microorganisms, such as microbes, viruses and bacteria, are net negatively charged at the pH of most soils. Hence, microorganisms are considered as pH-dependent. Cells with an exclusive negatively charged surface (carboxyl type) sorb approximately twice as much clay as those with a complex carboxylamino surface. Metal adsorption by bacterial cells may exceed the extracellular concentrations by 3 to 6
OTHER TYPES OF LINING SYSTEMS
395
orders of magnitude (Beveridge, 1985). Such high concentrations have been reported for Co 2+, Cu 2+, Fe 3§ Pb 2§ Ti 2§ Zn 2+, and Ni 1+ (Ehrlich, 1978). Metal uptake is an active process by means of metabolic energy, but metal binding on bacterial cell walls is a physico-chemical process (Beveridge, 1985). In environments which are anaerobic, and contain organic carbon as well as sulfate and sulfate-reducing bacteria, the production of HzS by these bacteria may result in the precipitation of metals (such as Hg and Cd) which have insoluble sulfides. The metals would thus be rendered immobile. Adsorption of pollutants by bacteria will be influenced by the nature of the exchangeable cation in the soil. In terms of microorganism adsorption selectivity, we could arrange the exchangeable cations in the following order: Na +< Li+< NH4< K+< Mg2+< Ca2+< Ba2+< Mn2+< A13+< Fe 3+
[15.7]
In relation to microorganism-pollutant interaction, it appears that heavy metals can be retained strongly by microbial cells (Christie and Costa, 1983). For example, metals may be biologically transformed from a water-insoluble form to a volatile alkyl metallic compound. An active microbial population may also enhance the microbial mobility through simple alteration of the soil physical environment. This could result from a localized pH or redox potential variation in soil micro-sites. Redox potential modification may result from depletion of electron acceptor, such as oxygen, nitrate, nitrite, and sulfate. Mechanisms for pH alteration include carbon dioxide or organic acid synthesis. Indeed, any microbial process that yields acids (organic or mineral) has a major effect on localized mineral solubility. The simple production of carbon dioxide through respiration causes a localized pH drop (Tate, 1987). The reduction of Fe3+to Fe2+by microorganism results in an increase in the negative surface charges of the clay soil and a decrease in swelling potential (Stucki et al., 1984). In summary, microbes can enhance or reduce metal mobility through: (1) modification of site pH or Eh, (2) inorganic and organic acid synthesis, (3) excretion of complexes, (4) catabolism of metal-organic matter complex, and (5) accumulation of extra-celluar slimes.
15.8
O T H E R TYPES OF LINING SYSTEMS
15.8.1 Soil Cement
Soil cement is a compacted mixture of Portland cement, water, and selected in-situ soils. The resulting mixture has greater strength than the natural soil. The hydraulic conductivity varies with particle size distribution of the soil. Fine textured soils (clays and clay loams) result in the lowest hydraulic conductivity with a range averaging near 10.8 to 10.9 m/see (Stewart, 1987). The addition of cement increases soil pH and forms less soluble compounds. The mechanism by which cement changes the properties of soil is explained by the formation of strong nuclei, distributed throughout the mass, or the formation of a skeleton of hydrated cement throughout the voids. For instance, montmorillonite clay supports, at low cement contents, a nucleated structure which changes to a skeleton structure as cement contents increase. Detailed discussion, regarding soil- cement interaction, is given in Chapter 20. The ageing and weathering characteristics of soil cements are
396
LINING SYSTEMS
good, especially those associated with wet-dry, and freeze-thaw cycles. Some degradation has been noted when soil cement is exposed to highly acidic environments (Stewart, 1978). However, soil cements can resist moderate amounts of alkali, organic matter and inorganic salts. One of the main deficiencies of soil cement as a liner material is its tendency to crack and shrink on drying. 15.8.2 Soil Carbonates
Carbonates are some of the most common minerals in most rock types. The three major mineralized groups of carbonates are the calcite, aragonite and malachite groups, all of which include a range of carbonate minerals (Kraus et al., 1959). The most widely known minerals are calcite and dolomite, both of which are from the calcite group. Carbonates, particularly calcium carbonate, tend to be sparingly soluble in water. Typical solubility products in pure water are shown in Table 15.4.
Table 15.4: Typical solubility products of various carbonate minerals Mineral
Mineralo~;ical Group
Formula
Solubility
Dolomite Calcite Strontianite Siderite Rhodochrosite
Calcite Calcite Aragonite Calcite Calcite
MgCO3-3H20 Ca CO 3 Sr CO3 Fe CO3 Mn CO3
2.0• 4.5• 2.5• 3.0• 6.0•
.5 10 -9
-l~ -11 -18
Crushed carbonates vary in their degrees of effectiveness in minimizing migration of potential pollutants in aqueous leachate, depending on the way they are associated with the soil as well as the kind of soil involved. Carbonate layering in clay barrier system is preferred to mixing with soil. Mixing carbonates with soil may result in soil stabilization or cementation that temporarily will inhibit the movement of pollutants. With time, fixed channels form holes through which leachates can move into the natural soil and geotechnical material without first having had much of an opportunity to react with the barrier materials. In leachate pollutant control, the purpose of carbonates is not necessarily to neutralize the acidity of soil, but to react directly with potentially toxic constituents of the leachate before they enter the soil as well as raise the pH level of the leachate passing through which, in turn, lowers the solubility of most heavy metals. Fuller and Warrick (1985) have indicated that the use of carbonate liners over soil is preferred over other methods. The acid-consuming reactions of solid phase carbonates under anaerobic conditions may be expressed as: (a)
For calcite H ++ C a C O 3 --. C a 2§ + H C O 3
[15.8]
OTHER TYPES OF LINING SYSTEMS (b)
397
For Dolomite 2 H ++ C a M g ( C 0 3 ) 2 ~ C a
2 § +
Mg2§ + 2HCO 3
[15.9]
Leachates high in acid, total organic carbon (TOC), and soluble salts possess a greater carbonate requirement than more dilute leachates (Alesii et al., 1980). Hence, fresh leachates consume carbonate more rapidly than older leachates. Consumption of carbonate is directly proportional to the leachate acidity. The greater the acidity, the greater the thickness of lime stone required. This is due to differences in the degree of dissolution. The most important properties that influence the way carbonates contribute to the effectiveness of liner systems are: (1) particle size distribution, (2) quality of the carbonate, (3) thickness of the carbonate layer, and (4) compaction. 15.8.3 Soil Amorphous The oxide/hydrous oxide minerals are generally found as principal constituents of highly weathered tropical soils, such as laterite and bauxite. Common crystalline forms of those minerals include haematite, goethite, gibbsite, boehmite, anatase and quartz. They differ from layer silicate minerals in that their surfaces essentially consist of broken bonds, and in an aqueous environment, these bonds are satisfied by hydroxyl groups of dissociated water molecules. The nature of the surface of these oxides and hydrous oxides, in an aqueous environment, exhibit variable charge properties. Allophanes are an example of these types of minerals which may contain 2-3% oxides of iron, titanium, phosphorous, calcium, magnesium, and potassium, as well as significant quantities of sodium. Specific surface area for allophanes varies between 300 and 700 mZ/g, and has a high (about 150 cmol/kg) cation exchange capacity which increases sharply with an increase in pH.
Figure 15.14. Retention of lead, copper, zinc, and cadmium by iron hydrous oxide gel.
398
LINING SYSTEMS
Over the past 15 years, numerous laboratory studies of adsorption of inorganic pollutants on hydrous oxides, specially iron and aluminum oxides, have been conducted. Figures 15.14 and 15.15 show, for various heavy metals, the amount retained by hydrous oxide gels of iron and aluminum as a function of solution pH. The selectivity sequence for iron is lead > copper > zinc > cadmium while for aluminum, the selectivity sequence is copper > lead > zinc > cadmium. The results indicate that iron and aluminum hydrous oxides can be used as barriers to adsorb inorganic pollutants. Fuller and Warrick (1985) indicate that a thin layer of less than 1 mm thickness of hydrous oxide of iron reduces the migration rate of certain heavy and trace metals.
Figure 15.15. Retention of lead, copper, zinc, and cadmium by aluminum hydrous oxide gel.
15.8.4 Soil Sealants
The hydraulic conductivity of some soils can be significantly reduced by the application of various chemicals or leachates. They may be water borne, mixed in place, spray applied, or injected below the soil surface (Gooding et al., 1967; Jones, 1971). Water-borne or spray-on polymer soil sealants can reduce hydraulic conductivity of earth-lined impoundments. However, the sealing effect is confined to the upper few centimetres and can be significantly diminished by the effects of wet-dry and/or freeze-thaw cycles. Types of sealants include resinous polymer, diesel fuel mixtures, petroleum-based emulsions, powdered polymers which form gels, and monovalent cationic salts. Soil sealants utilizing monovalent cations, such as sodium carbonate, sodium pyrophosphate and sodium silicate, chemically reduce the effective porosity of the soil due to the high tendency to form dispersed structure. Soil treated with sodium silicate and sulfuric acid prior to compaction shows a significant seepage reduction and is compatible with sulfuric acid-bearing wastes (Clark and Moyer, 1974). Soil sealants based on polymers are generally mixtures of linear and cross linked polymers of approximately the same molecular weight. The linear portion sorbs to the soil, forming
OTHER TYPES OF LINING SYSTEMS
399
a flexible network. The cross linked polymer particles can flow, and thus conform to and permeate the soil pores. The polymer is usually mixed in a low pH water/acid solution and sprayed on an unfilled site as a low viscosity slurry. The low pH allows the slurry to penetrate the surface and form a deeper seal. The main disadvantages of polymer seals are: (1) low strength, and (2) exposure to salts, acids, and multivalent cations causes the polymers to shrink, increasing the hydraulic conductivity and decreasing the effectiveness of the seal (Parks and Rosene, 1971). For example, soil hydraulic conductivity to water was reduced when latex was used as a soil sealant. However, the latex was subjected to damage by microbiological attack, frost action, and vegetation (Uniroyal, 1972).
15.8.5 Flexible Membrane Liners Prefabricated liners based upon sheeting of polymeric materials are of particular interest for the lining of waste storage and disposal impoundments. As a group, these materials exhibit extremely low hydraulic conductivity. They have found substantial use in water impoundments in reservoirs and are being used in the lining of sanitary landfills and various waste disposal facilities. Polymers used in the manufacture of lining materials include a wide range of rubber and plastics, differing in polarity, chemical resistance, basic composition, etc. They are generally classified into: (1) Rubbers (elastomers) which are generally cross linked (vulcanized); (2) Plastics which are generally unvulcanized, such as polyvinyl chloride (PVC); (3) Plastics which have a relatively high crystalline content, such as the polyolefins; and (4) Thermoplastic elastomers, which do not need to be vulcanized. The polymeric materials most frequently used in liners are polyvinyl chloride (PVC), chlorosulfonated polyethylene (CSPE), chlorinated polyethylene (CPE), butyl rubber, ethylene propylene rubber (EPDM), neoprene, and high density polyethylene (HDPE). The thickness of polymeric membrane for liners ranges from 20 to 120 mils ( 1 mil = 0.0254 mm). Most polymeric lining materials are based on single polymers. However, blends of two or more polymers, e.g., plastic-rubber alloys, are being developed and used in liners. Therefore, it is difficult to make generic classifications based on individual polymers of the liners. Most of the membrane liners currently manufactured are thermoplastic. Thermoplastic polymers can be heat-sealed or seamed with solvent containing dissolved polymer to increase the viscosity and reduce the rate of evaporation. Crystalline sheetings, which are also thermoplastic, can only be seamed by thermal or fusion methods.
Types of Flexible Membrane Liners (FMLs) Polyvinyl Chloride (PVC): PVC polymer is produced from vinyl chloride monomer by suspension and emulsion polymerization processes. Suspension polymerization occurs between 5-12 bars at 60 ~ C and emulsion polymerization occurs between 5-8 bars at 50 ~ C. The final product of the polymerization processes contains: (1) 55% chloride, (2) 25% to 35% plasticizers, which make the sheet flexible and rubber-like, (3) 1% to 5% of a chemical stabilizer, and (4) various amounts of other additives. The presence of plasticizers increases the PVC resistance against water, alkaline solutions, non-oxidizable acids and hydrocarbons. Plasticizer loss during service is a source of PVC liner deterioration. The basic mechanisms for plasticizer loss are volatilization, extraction, and
400
LINING SYSTEMS
microbiological attack. Ultraviolet exposure has a direct effect on PVC polymer. Carbon black prevents ultraviolet attack but causes the absorption of solar energy, raising the temperature to a high level and causing vaporization of plasticizer. Soil cover material used to bury the liner protects the PVC liner from ultraviolet exposure and reduces the rate of plasticizer loss. PVC membranes are the most widely used of all polymeric membranes for waste impoundments. They show good chemical resistance to many inorganic chemicals (Chan et al., 1978). However, they are attacked by many organic chemicals, particularly hydrocarbons, and solvents.
Chloro-sulfonated Polyethylene (CSPE): It is a family of polymers prepared by reacting polyethylene in solution with chlorine and sulfur oxide. Presently available CSPE polymers contain from 24% to 43 % chloride and from 1.0% to 1.5% sulfur, i.e., 1 SO2C1 per 100 ethylene groups. They are used in both thermoplastic (uncross linked) and cross linked (with metal oxides) composition. Thermoplastic CSPE is more sensitive to temperature than the commonly-used elastomers. CSPE is generally tougher at room temperature, but softens more rapidly as temperatures are increased (Morton, 1973). CSPE is characterized by ozone resistance, light stability, heat resistance, good weatherability, and resistance to deterioration by corrosive chemicals and bacteria. Also, CSPE tends to harden on aging, due to cross linking by moisture, ultraviolet radiation, and heat. Chlorinated Polyethylene (CPE): It is produced by a chemical reaction between chlorine and high-density polyethylene. Presently available polymers contain 25% to 45% chlorine and 0% to 25% crystallinity. CPE is compounded and used in both thermoplastic and cross linked compositions. Since CPE is a completely saturated polymer (no double bonds) it weathers well, is not susceptible to ozone attack, and has good tensile strength and elongation. CPE is resistant to deterioration by many corrosive and toxic chemicals. It also has good resistance to bacteria. Continuous exposure of CPE to aromatics will shorten the service life of the liner, hence CPE is not recommended for containment of aromatic hydrocarbons. Polyethylene: It is ethylene-based thermoplastic crystalline polymer. It is made in three major forms: (1) low density polyethylene (LDPE), (2) linear low-density polyethylene (LLDPE), and (3) high-density polyethylene (HDPE). The monomer ethylene (CH 2 = CH2) is polymerized with a peroxidic initiator at a pressure of 1000 bars or higher and at a temperature of about 200~ to produce LDPE. The use of a high pressure produces a highly branched polymer (low crystallinity). Examples of the final products are: (1) In gas form: methane, ethane, butane; (2) In liquid form: pentane, decane; and (3) In solid form: tetradecane and polyethylene. The final product of LDPE exhibits low reactivity, low density and excellent flexibility and is, therefore, not recommended for waste containment purposes. The properties of a polyethylene are largely dependent upon crystallinity and density. The high density polyethylene polymers exhibit the greatest resistance to oils, solvents, and permeation by water vapour and gases.
OTHER TYPES OF LINING SYSTEMS
401
Performance Requirements of FMLs The performance requirements of a FML used in a hazardous waste management unit include low hydraulic conductivity, chemical compatibility, mechanical compatibility, and durability. Based upon the designed use of the unit, the designer must make decision on the composition, thickness, and construction (fabric-reinforced or unreinforced) of a FML. Composition of the liner is based primarily on chemical compatibility. Mechanical compatibility and hydraulic conductivity determine the thickness of the FML sheeting. It should be noted that liner performance does not correlate directly with any one property (e.g., tensile strength). The construction of a continuous water-tight FML is critical to the containment of hazardous waste and is heavily dependent on the seams that bond the sheeting together. The seams are the most likely source of failure in a FML. Seams can be manufactured both in the factory and in the field. The quality of field seams is difficult to maintain since the installer must deal with changing whether conditions, such as temperature, wind, and precipitation, as well as construction site conditions. Several bonding systems are available for the construction of factory and field seams in FMLs. Bonding systems include solvent methods, heat seals, heat guns, dielectric seaming, extrusion welding, and hot wedge technique. The selection of a bonding system is dependent primarily on the polymer making up the sheeting. To ensure the integrity of seams, a given FML should be seamed using the bonding system recommended by the FML manufacture.
Hydraulic Conductivity: The primary function of a linear system in a hazardous waste management unit is to minimize and control the flow of hazardous waste from the unit to the environment, particularly the groundwater. The hydraulic conductivity of a FML made of a particular polymer may change upon exposure to waste or leachate, depending on the composition of the waste contained by the FML. Since different plastics and rubbers exhibit various degrees of compatibility with different chemicals, a number of materials are used to manufacture FML sheeting. The material is selected based on expected chemical exposure during its service life. The calculated hydraulic conductivities of FMLs, from water vapour transmission (ASTM E-96), range from 10 -12m/sec to 1015 m/sec (Cadwalladar, 1988). Mechanical Compatibility: A FML must be mechanically compatible with the designed use of the lined facility in order to maintain its integrity during and after exposure to short term and long term mechanical stresses. Short term mechanical stresses may include equipment traffic during the installation of the liner system, as well as thermal expansion and shrinkage of the FML during operation of the unit. Long term mechanical stresses usually result from the placement of waste on top of the liner system or from differential settlement of the subgrade. The tensile properties of FMLs are assessed via ASTM D412, D638, and D882. Mechanical compatibility requires adequate friction between the components of a liner system, particularly the soil subgrade and the FML, to ensure that slippage does not occur at the slopes of the unit (Richardson and Koerner, 1987). Specifically, the foundation slopes and the subgrade materials must be considered in design equations in order to evaluate: (1) the ability of the FML to support its own weight on the side slopes, (2) the ability of the FML to withstand downdragging during and after filling, and (3) the best anchorage configuration for the FML.
402
LINING SYSTEMS
Chemical Compatibility: Chemical compatibility of FMLs and waste liquids is a critical factor in the service life of a liner system. Chemical compatibility is evaluated indirectly using swelling resistance via ASTM D570, ozone resistance via ASTM D1149, and ultraviolet light resistance via ASTM D3334. Chemical compatibility requires that the mechanical properties of the FML remain essentially unchanged after the FML is exposed to waste. If the seams between the sheets are made with materials other than the sheet parent material, they also must be compatible with the waste liquids. Incompatibility is due primarily to the absorption of waste constituent by the FML, the extraction of components of the FML compound by wastes or leachate, or reaction between FML constituents and waste or leachates. Before, during and after operation of a landfill, the polymers used in construction of the FML can breakdown and change crystallinity. The polymer can break down when oxygen reacts with it to induce cross linking or scission of the polymer chain. Ultraviolet light causes scission of the polymer. Also, applied heat from thermal-based seaming causes the polymer to degrade. The polymer can be altered markedly (swelling and softening) when it absorbs organic chemicals. Koemer (1990) indicates that 20% to 40% changes in chlorosulfonated polyethylene (CSPE) properties can be obtained when the CSPE is immersed in acetone at room temperature for a period of approximately 100 days. Short term exposure probably does not affect polymer integrity (Mohamed et al., 1993). However, there is no knowledge about long term combined effects of softening and mechanical stresses (i.e., creep). Koemer (1990) has indicated that when testing FML compatibility with leachate, a marked change in FML performance is an indication of the unsuitability of the FML. However, the lack of a change does not necessarily mean that the correct FML has been found since it is quite possible that the FML was under incubation for a short period of time. Since incubation periods of longer than 6 months are generally unrealistic, accelerated aging testing is attractive. Such aging is usually accomplished by incubation at an elevated temperature, e.g., up to 50~ C in the US EPA 9090 test procedure (US EPA, 1986). No universal rules have, however, been established. Schneider (1987) has indicated that hydrolysis is one of the most important chemical processes of polyethylene terephtalat (PET) degradation. Hydrolysis can take place only after water has been adsorbed and diffused into the polymer. The rate of hydrolysis is indirectly related to concentration. The reaction takes place much faster in an acid or alkaline medium. It was indicated that the pH of NaOH solution has very little effect on the hydrolytic decomposition of PET. However, if NaOH is replaced by Ca(OH)2, degradation is greatly accelerated. This acceleration might be explained by a change in electrolytic structure. In any case, it is a phenomenon that must be taken into account when assessing the long term properties of FMLs. Evaluation of the lifetime performance of FMLs is covered in Koemer et al. (1991).
15.9
SUMMARY AND CONCLUDING REMARKS
The main objective of a lining system is to contain the waste in a manner that is protective of human health and the environment. To meet this objective, lining systems must maintain their low hydraulic conductivity over the service life. Clay-based lining systems must be able to attenuate the movement of leachate and prolong the release of the pollutants in the leachate. The performance of clay liners is usually measured in terms of hydraulic conductivity, flexibility and strength. These factors are controlled by the soil structure and can be altered during
SUMMARY AND CONCLUDING REMARKS
403
the life of the clay liner. The initial structure of a compacted soil is based on composition and environmental factors. The compositional factors include mineralogy, particle size and shape, composition and moulding water content. The environmental factors include blending, and method and effort of compaction. These factors are generally well understood. The initial soil structure may, however, change with time due to occurring physical, chemical and biological processes. The physical processes include consolidation, wet-dry cycles, freeze-thaw cycles, and erosion. The chemical processes include dissolution, mineralogical transformation, sorption and precipitation. Biologically, soil structure can be altered by various bacterial actions occurring in the soil. These processes may contribute to a final soil structure which results in the failure of the clay liner. Since clay liners are usually intended to perform for extended periods of time, a primary requirement of a clay liner design is to minimize the negative impact of these processes on long term performance. Because no endeavour of mankind can be undertaken without some degree of risk, there is always a risk that the lining system will fail to perform up to expectation. The best that can be expected from a designer is to ensure that the risk posed is extremely low. Monitoring systems are generally installed around the lining systems to determine if the lining system is performing in accordance with expectation. The minimum technology requirements established by regulatory agencies are designed to minimize the risk associated with failure of waste containment systems. It is generally assumed that as long as the minimum technology requirements are met, the environmental risk will be minimum. Since soil is a living material that continuously changes due to various environmental conditions, such as physical, chemical, electrical, and biological, the designer should account for the timedependent changes in the lining system so as to ensure that the risks associated with the selected design remain low.
This Page Intentionally Left Blank
CHAPTER
SIXTEEN
SOIL V A P O U R E X T R A C T I O N
16.1
INTRODUCTION
Soil vapour extraction (SVE), also known as in-situ soil venting (ISV), is an in-situ remediation technique for sites polluted with organic chemicals. It is designed to remove, from the vadose (unsaturated) zone, pollutants with vapour pressure greater than 0.5 mm Hg and Henry's constant greater than 0.01. This includes volatile organic compounds (VOCs) and some semi-volatile compounds such as diesel fuel, kerosene, and heavy naphtha. Typical volatile and semi-volatile compounds are: (1) Volatilecompounds: Toluene, benzene, xylene, chloroform, hexane, tetrachloroethylene, methylene chloride, dichloroethylene, trichloroethylene, cyclohexane, ethylacetate, acetone, and methanol. (2) Semi-volatilecompounds: Dichlorobenzene, trichloropropane, and chlorobenzene. (3) Hydrocarbons:Jet fuel, gasoline, diesel, kerosene, and heavy naphtha. Soil vapour extraction technologies have been used to remove vapour from landfills since the 1970's. During the 1980's, SVE was applied extensively to recover pollutants from the vapour phase of unsaturated soil located beneath leaking underground storage tanks. In recent years, the technology has been applied to mitigate the adverse impacts from uncontrolled hazardous waste sites. The technology can be useful for sites where there are interferences or obstructions, such as buildings and highways that generally have sub-bases of porous material. Since it transfers pollutants from soil and groundwater to air and to condensed wastewater streams, SVE is generally used in conjunction with other technologies. There are practical limitations on the final soil pollution levels that can be achieved with soil vapour extraction systems. Knowledge of these limits is necessary to realistically set cleanup criteria and design effective systems. Maximum efficiency of a venting operation is limited by the equilibrium partitioning of pollutants between soil matrix and vapour phases. Therefore, understanding the basic interaction mechanisms between pollutants and soils, and the impact of various environmental parameters on the interaction process is very important for a successful application of the soil vapour extraction technique. These various elements are the focus of discussion in this chapter.
16.2
T E C H N O L O G Y DESCRIPTION
SVE uses the principle that pollutants with high vapour pressures will exist in part as an equilibrium gas phase in the pore spaces of the soil. By applying a vacuum to a well installed in the 405
406
SOIL VAPOUR EXTRACTION
polluted soil mass, pressure gradients are formed throughout the soil mass, inducing gas flow towards the well. The gas flow into the well contains pollutants in vapour form which are either vented to the atmosphere or treated. At a steady state pressure distribution throughout the soil, the in-flux of clean air into the soil will equal the flow into the well. The in-flux of clean air induces a non-equilibrium phase in the pore spaces which is countered by volatilization of pollutants that exist as other phases within the soil mass. Thus, with continued application of the vacuum, all volatile pollutants will eventually enter the gas phase and be removed from the soil.
Figure 16.1. In-situ soil vapour extraction system.
A basic in-situ SVE system, shown in Figure 16.1, couples vapour extraction wells with blowers or vacuum pumps to remove pollutant vapours from zones permeable to vapour flow. The vacuum developed through a screened casing in the extraction well boring results in air being drawn from the atmosphere through the soil to the well. Above ground equipment often consists of blowers or vacuum pumps, control valves to adjust air flow, pressure gauges and flow metres at wellheads, an air-liquid separator, a pressure gauge and flow metre at the pump, and a vapour treatment unit. Vapour treatment systems such as catalytic and thermal destruction systems, activated carbon adsorption, and biological gas treatment systems, are installed if air treatment is required. To access residual pollutants below groundwater level, a groundwater pump and skim well are used. The groundwater recovery well is used to depress the water table to expose saturated soil to vapour flow. More complex SVE systems incorporate trenches, horizontal wells, forced air injection wells, passive air inlet wells, low hydraulic conductivity or impermeable surface seals, multiple level vapour extraction wells in single boreholes and moisture, pollutant, and temperature monitoring systems. The success of SVE, as measured using response variables, depends on site conditions, soil properties, chemical properties, and control variables (Johnson et al., 1990), which are discussed
SITE CONDITIONS
407
in the following sections.
16.3
SITE CONDITIONS
Site conditions that effect the success of SVE systems include volatile organic pollutant distribution, depth to groundwater, infiltration rate, location of heterogeneities, temperature and atmospheric pressure. 16.3.1 Volatile Organic Pollutant Distribution
The VOC distribution in the soil, both vertically and horizontally, is important in deciding whether or not SVE is preferred over some other remediation technique. Soil vapour extraction becomes cost effective over soil excavation when (Charbeneau et al., 1992): (1) polluted soil volume is greater than 500 cubic yards, (2) pollution has penetrated more than 6 or 9 m, and (3) pollution has spread over an area greater than hundreds of square metres at some particular depth. The vertical depth of pollution zone influences the well depth. The well passes through the polluted zone and typically extends below the water table. The lateral extent of pollution will influence the required number of wells, the need for injection wells, the magnitude of the vacuum applied to the wells, and the need for any surface cover. 16.3.2 Groundwater Table
When the groundwater table is located at depths below 10 to 15 m and the pollutants extend down to the groundwater table, SVE system may be one of the only ways of remediating the soil (Charbeneau et al., 1992). In other cases, however, the groundwater table may have to be lowered to increase the volume of the unsaturated zone. In cases of high groundwater table or when the pollutants are limited to shallow depths, horizontal trends or perforated pipes may be preferable over wells for extracting pollutants (Ghassemi, 1990). The lateral distribution of immiscible pollutants may be affected by the water table (Dolan, 1990). When pollutants are discharged into the ground, they migrate downwards under the force of gravity and capillary action with some lateral dispersion. Upon reaching the water table, the miscible pollutants will dissolve in the groundwater while immiscible pollutants will float on the groundwater surface. If the weight of pollutants exceeds that of the pore water contained in the capillary zone, the pore water and possibly the water table may be suppressed. As buoyant forces act to restore the pore water or water table, the immiscible pollutants are forced to migrate laterally. This effect of the groundwater table on the lateral distribution of pollutants will be apparent with seasonable water table fluctuations, with groundwater pumping, and after a fresh spill of pollutants. Floating immiscible pollutants and dissolved pollutants will also migrate laterally in the direction of groundwater flow. Vertical groundwater fluctuations during lateral migration will act to extend the vertical distribution of pollutants.
408
SOIL VAPOUR EXTRACTION
16.3.3 Infiltration Water infiltration into the polluted zone will contribute to the downward leaching of pollutants to either the water table or some impervious stratum. Infiltration of water into the soil affects groundwater table fluctuations and the moisture content of the unsaturated zone. Pollutants will, to some degree, partition into the soil moisture depending on the amount of soil water present. A high soil moisture content will decrease the rate of vapour migration because vapours will compete with water for pore space. Infiltration of air into the polluted zone limits the lateral influence of SVE extraction wells. Natural surface barriers (e.g., a well packed surface cover or a frost layer) and artificial surface seals (e.g., polymer-based liners or asphalt, concrete and clay caps) may contribute to the increase of the extraction well's radius of influence by reducing air infiltration into the soil from the atmosphere. For shallow treatment zones (less than 5 m), the surface seal will significantly affect the vapour flow path and may be added or removed as required to achieve a desired result. However, for extraction wells screened at depths below 8 m, the influence of surface seals becomes less significant (Johnson, et al., 1990(a)). Surface seals also prevent water infiltration into the polluted soil mass, thereby reducing pollutant leaching and preventing water saturation of soil pores.
16.3.4 Location of Heterogeneity Locations of heterogeneity are of particular importance since they influence gas movement within the soil, determine the location of pollutants, and influence the location of wells (Dolan, 1990). Various soil strata have differing properties, notably air permeability, organic content, clay content, adsorption site density, and moisture content. These properties influence the mobility of pollutants and their location in the soils. An increase in organic content or adsorption sites will result in a decrease in volatilization. A high clay content will result in a lower soil hydraulic conductivity, hence decreasing mobility. The development of a vacuum in clayey soils is more difficult to establish than in a sandy soil. Low hydraulic conductivity layers (less than 10.9 m/s) tend to act as barriers or confining layers which may hinder or aid SVE. A horizontally stratified soil may promote SVE by limiting the rate of vertical inflow from the ground surface, thereby extending the influence of the applied vacuum horizontally from the point of extraction. Natural cracks or fissure, disturbances from excavations, and coarse-grained media (larger than 2 mm) promote migration of gases and may allow for short circuiting of gas flows around polluted regions within the soil. As mentioned, a high soil moisture content will decrease the rate of vapour migration because vapours will compete with water pore space. Other types of heterogeneity include underground structures like walls and backfilled utility trenches. Such structures divert or short circuit gas flow, thereby affecting the performance of SVE system.
16.3.5 Temperature Soil temperature influences the volatilization of the pollutants within the soil. The more volatile a pollutant, the more susceptible it is to removal by SVE. Compounds that have a higher vapour pressure and higher solubility tend to be more volatile. Vapour pressure tends to increase with temperature. For example, the vapour pressure for a volatile compound like benzene increases
SITE CONDITIONS
409
from 0.037 atm at 0 ~ to 0.137 atm at 27 ~ (Charbeneau et al., 1992). Over the same temperature range, the vapour pressure of n-dodecane increases from 2.8x 10.5 to 2.3x 10 4 atm. Assuming the gas is treated as ideal with a relatively constant enthalpy of vaporization, the Claussius-Capeyron equation can be used to predict the temperature dependence of vapour pressure (Johnson et al., 1990(b)): -
l,~
In
,
'
[16.11
Pb,i
where Pf is vapour pressure, T, is temperature of the soil, Tb,, is the boiling temperature at a pressure of the i th component of the mixture of pollutants present in the soil, and Tr,~is the reference temperature at which the vapour pressure P,~(T~,~) is known. Temperature will also affect the flow of vapour throughout the soil (Johnson et al., 1990(b)). For a given vacuum pressure, the vapour flow rate through the soil will be inversely proportional to the vapour viscosity, p. The vapour viscosity at low pressures and temperatures not near the critical point depends on temperature as: Pb, i
P(T1) : P(T2)
d T 1 + 273 T2 + 273
[16.2]
For a well operating in a soil at the same vacuum pressure, Q, but differing temperatures:
~ T 2 + 273 Q(TI) : Q(T2)
16.3.6 Atmospheric
T~ + 273
[16.3]
Pressure
Atmospheric pressure varies as a function of elevation and weather condition. Standard sea level pressure is designated as 1 atm and can vary by as much as 0.027 atm due to weather changes (Strahler and Strahler, 1992). Atmospheric pressure decreases exponentially with increased elevation to about half its sea level value at an altitude of 5 km. The pressure distribution obtained throughout a soil during SVE will depend to a degree on the atmospheric pressure. The steady state pressure distribution P(r) at a radius r from an extraction well may be estimated using: 2
2
2
ln(r/Rw)
P2(r) - Pw = (Patm - Pw) ln(R1/Rw )
[16.4]
where Pw is the vacuum applied to a well of radius Rw, and R 1 is the extraction well's radius of influence, defined as the radius where the pressure equals the ambient atmospheric pressure. During SVE application, extraction wells may operate with vacuum pressure up to 3.33 atm (Ghassemi, 1990).
410 16.4
SOIL VAPOUR EXTRACTION SOIL PROPERTIES
Soil properties that affect the success of SVE systems include air and water permeability, porosity, organic carbon content, soil structure, soil moisture characteristics, particle size distribution, and soil mineralogy. 16.4.1 Soil Hydraulic Conductivity The vapour flow through a confined porous stratum of thickness, m, is governed by the following relationships: O(~.pm)
_
- -X7(pmv)
Ot
k bt
v = - -- (VP
+ Pg)
[16.5]
[16.6]
where p is the vapour density, g is the gravity constant, e is the vapour-filled void fraction of the soil, v is the Darcian vapour velocity, ~t is the vapour viscosity, P is the vapour phase pressure, m is the stratum thickness, and k is the hydraulic conductivity of the soil. Thus, the gas velocity through a body of uniform soil will be directly proportional to the soil hydraulic conductivity. Hydraulic conductivity will depend on a number of factors such as particle size distribution, soil structure, porosity, soil moisture content, mineralogy of the soil, and the type of permeant, as discussed in Chapters 5, 14 and 15. Soils with smaller particle size usually have smaller-sized flow channels and, therefore, lower hydraulic conductivity. The porosity of the soil reflects the soil's mineralogy, chemistry, stress history, and overburden pressure. Compacted clayey soils tend to have a lower porosity and hydraulic conductivity compared to sandy soils. Glacial till formed below continental glaciers can have a very low porosity and hydraulic conductivity. Porosity will decrease with depth due to overburden pressures, thus one can expect hydraulic conductivity to decrease with depth. Soil structure involves the spatial distribution of particles, pores, cracks, and channels within a soil. A sandy soil may have a random structure with isotropic hydraulic conductivity. A clay soil, depending on its depositional history (flocculated versus dispersed deposition, compaction history), may have an anisotropic hydraulic conductivity, thus stratification of the soil mass at both the microand macro-scale will affect the hydraulic conductivity of the soil mass, as discussed in Chapter 15. Soil moisture will compete with other liquids and gases for voids within the soil, thereby limiting their hydraulic conductivity. Soil mineralogy includes primary minerals (e.g., quartz and feldspar), clay minerals, carbonates and sulfates, metal oxides and hydroxides, and soil organic matter, as discussed in Chapter 4. Carbonates, soil organic matter, as well as amorphous hydroxides, oxides and silicates may cement soil particles together, thereby affecting the hydraulic conductivity of the soil. Finally, hydraulic conductivity, k, of soils depends on the liquid or gas permeating the soil and may be expressed as: k
kipgg
bt
[16.7]
where k,p is the intrinsic permeability of the soil, and p and ~t refer to the particular liquid or gas
SOIL PROPERTIES
411
flowing through the soil. At 20 ~ C, water has p = 998.2 kg/m 3 and ~t = 1.002x 10.3 N.s/m2 and dry air has p = 1.205 kg/m 3 and ~t = 1.81 • 10.5 N.s/m 2. The air flow rates induced by SVE are higher than groundwater flow rates induced by pumping (Travis and Macinnis, 1992). Organic pollutants in soil may reside in five different locations: (1) as a free liquid in between soil particles, (2) in the macro-pores as a vapour phase, (3) dissolved in the soil moisture, (4) adsorbed onto the surface of soil particles and soil organic matter, and (5) sequestered in the interior of the soil matrix or the micro-pores. It is noteworthy that while the free liquid, vapour, and adsorbed phases are the most amenable to SVE, the sequestered phase is the most difficult to remediate. Also, it should be noted that when organic pollutants have been in contact with soils for several years, a substantial portion of the pollutant can be found in the soil matrix. To remove organic pollutants from the soil matrix during SVE, these pollutants must diffuse from the soil micro-pores to the macro-pores where a majority of the SVE induced vapour flow occurs. Therefore, it appears that the structure of the soil being treated and the location of pollutants within the structure have a significant bearing on the success of SVE. The vapour pressure, p v (pore), of a pollutant held in a soil by capillary forces may be expressed as: P~V(pore) = P~v(7) exp
rRT
)
[16.8]
where PiV(T) is the pollutant vapour pressure at temperature, T, o is the surface tension, V m is the molar volume of the pollutant, r is the radius of curvature of the liquid in the soil pore due to surface tension, and R is the gas constant. When pollutants are held in soil micro-pores with diameters on the order of 10.8 m, pollutant vapour pressure is significantly reduced, thereby trapping the pollutants in the soil matrix. The effect, however, is much less pronounced in the macro-pores of the soil where SVE is most applicable.
16.4.2 Soil Organic Matter Content The soil organic matter content has a significant bearing on the adsorption of VOCs in soils. In particular, highly reactive insoluble poly-functional humic substances with their high specific surface areas and charge densities are known to be the soil component most responsible for the adsorption of organic chemicals in soils (Morrill et al., 1982). Humic substances rarely exist alone in soils and are typically found in contact with other soil components. In many cases, the interaction of humic substances with soil components is the major determinant of the adsorptive capacity of soils for pollutants. The ability of soil organic matter to promote the formation of aggregates increases soil hydraulic conductivity and water holding capacity and may be significant with as little as 1-3% soil organic matter. A list of the functional groups found with humic substances and the mechanisms associated with them is shown in Table 16.1 (Sposito, 1989). Aliphatic and aromatic organic pollutants containing positively charged nitrogen bonded in tetrahedral coordination can adsorb onto negatively charged humic substance via cation exchange. COOH and NH groups on humic macro-molecules can form hydrogen bonds with electronegative atoms on pollutants like O, N, and F. The bonding between organic pollutants and uncharged portions of humic macro-molecules is attributed to van
412
SOIL VAPOUR EXTRACTION
der Waals attraction forces. Overall, adsorption of organic chemicals generally: (1) increases with length of molecule, (2) increases with adsorbent charge density and specific surface area, and (3) varies with the number and type of functional groups involved.
Table 16.1: Reaction mechanisms associated with humic functional ~roups Mechanism
Functional groups
9 Cation exchange 9 Protonation [] Anion exchange [] Water bridging 9 Cation exchange 9 Ligand exchange 9 Hydrogen bonding 9van der Waals attraction
Amino, ring NH, heterocyclic N (Aromatic ring) Amino, heterocyclic N, carbonyl, carboxylate Carboxylate Amino, carbonyl, carboxylate, alcoholic OH Carboxylate, alcoholic OH, amines, carbonyl Carboxylate Amino, carbonyl, carboxyl, phenolic OH Uncharged organic units
16.5
CHEMICAL PROPERTIES
Chemical properties that affect the success of SVE systems include Henry's constant, solubility, adsorption equilibrium, air and water diffusivity, density, viscosity, and octanol-water partition coefficient. 16.5.1 Henry's Constant and Solubility As mentioned earlier, VOCs in a soil can exist in several phases. This may be expressed in terms of vapour, free liquid, soil moisture, and sorbed phases using the following expression (Johnson et al., 1990b):
zp V M
-
RT
+ x,MHc + Y,Mw + k,vi
Ms mw
[16.9]
where M, is the total mass of pollutant i in the soil, Mn(. is the total mass of volatile organic pollutants in the soil, Mw is the mass of soil moisture in the polluted soil, Ms is the mass of polluted soil, x, is the mole fraction of pollutant i in the free liquid phase, z, is the mole fraction of i in the vapour phase, y, is the mole fraction of i in the soil moisture phase, P is the total pressure in the pore vapour, e is the vapour-filled void fraction of the soil, V is the volume of polluted soil, R is the gas constant, T is the absolute temperature, mw is the molecular weight of water, and k; is the soil adsorption coefficient for pollutant i. The prescribed model of pollutant partitioning in soil is based on several assumptions: (1) Pollutants are distributed uniformly through what is assumed to be a homogeneous soil;
CHEMICAL PROPERTIES (2) (3)
413
All phases are assumed to be in equilibrium with each other; and The vapour may be treated as an ideal gas, the free liquid phase as an ideal mixture, and the soil moisture phase as non-ideal. With the phase equilibrium assumption, one may relate x,, y,, and z~ using: z p = x, p vi :
O~y,pVi = C i R T
[16.10]
where P[ is the vapour pressure in the vapour phase of the ith pollutant, ~ is the activity of the ith pollutant, and C~ is the molar concentration of the ith pollutant in the vapour. Note that the sum of Xl equals one when the free liquid phase is present, and is less than one when the free liquid phase is absent. Partitioning of a volatile pollutant between gas and water phases may be described using Henry's Law as: p V = Hy,
[16.11]
where Hi is Henry's constant for the ith pollutant. Henry's constant can be used as an indicator of the s t r i p a b i l i t y of a volatile pollutant from the soil moisture phase (Dolan, 1990). Compounds with higher solubility and high vapour pressure volatilize most easily. During operation, SVE system will preferentially remove the more volatile low molecular weight organic compotmds. Pollutants with vapour pressures greater than 0.5 mm of mercury or Henry's constant greater than 0.01 are considered easily removable by SVE (Charbeneau et al., 1992). 16.5.2 Adsorption Equilibria and the Octanol-Water Partition Coefficient
Adsorption of volatile organic pollutants onto soil can occur from free liquid, aqueous, or gas phases and depends on several soil factors including adsorption site availability, organic content, moisture content, grain size, porosity, hydraulic conductivity, surface tension and temperature. An adsorption coefficient kl of a particular soil may be expressed as: k~ -
co.
Ceq
[16.12]
where C~d is the adsorbed concentration in soil, and Ceqis the equilibrium concentration in the pore fluid. Volatile organic pollutants with k~ greater than 1000 are considered to stay adsorbed onto soils. As mentioned, the organic content of soil can be responsible for a significant percentage of the VOC adsorption in soil. Soil adsorption coefficients may be estimated using the Karickhoff equation: k~ = 0.63 kow,~foc
[16.13]
where kow., is the octanol-water partition coefficient for the ith pollutant, andfoc is the organic content of the soil. Pollutants withfow.~ greater than 4 are considered more likely to be adsorbed onto the soil
414
SOIL VAPOUR EXTRACTION
and less likely to volatilize. A pollutant mobility index, M1, may be defined as: MI=
log (S, p~") k
ow,i
[16.14]
where S, is the pollutant's aqueous solubility. Pollutants with MI greater than 5 are considered very mobile while pollutants with MI less than -10 are considered immobile.
16.5.3 Diffusivity The diffusivity of a pollutant is an important consideration in SVE when either a free-liquid pollutant phase is floating on the groundwater table or pollutants are trapped within an impermeable soil lens surrounded by permeable material. In these cases, SVE induces gas flow around, but not through, these polluted zones and volatile pollutant removal is dictated by diffusion processes. Diffusion of pollutants occurs in a direction perpendicular to the gas flow, thereby creating a concentration boundary layer. The total pollutant removal rate from the polluted zone will depend on the vapour velocity, the free product thickness, and the vapour flow path and will be limited by either vapour phase diffusion, liquid phase diffusion or some combination thereof. Overall, the efficiency of applying SVE to a heterogeneous polluted soil mass will never be as high as an application to a homogeneous soil mass. In the case of a free-floating phase on the groundwater table, it may be more efficient to remove the free phase by pumping and removing the residual by SVE. In the case of low hydraulic conductivity soil regions, removal by SVE may be the only feasible remedial alternative regardless of the poor efficiency due to diffusion.
16.5.4 Density and Viscosity As discussed previously in section 16.4, the vapour flow of volatile organic pollutants through a confined porous stratum of thickness m is related to the density, p, and viscosity, ~t, of the volatile organic pollutants as follows: 0(cpm) = -V.(pmv) Ot
[16 15]
k v = - --"-'.__tvp + pg)
[16.16]
and
Treating the volatile organic pollutants as an ideal gas allows us to assume:
P-Patm(~atml
[16.17]
where Patmis the density of the pollutant at a reference vapour phase pressure Ofpatm. Eq. [ 16.15] states that the vapour phase flow velocity will increase as the viscosity of the vapour phase
CONTROL VARIABLES
415
decreases. As discussed, viscosity is temperature dependent in a manner described by Eq. [ 16.2]. The penetration of pollutants into the saturated zone will be greatly enhanced in the case of an organic compound with a density greater than water. As a dense non-aqueous liquid reaches the water table, a slight depression of the water table surface is likely to occur. Since water table fluctuation causes a rise in the phreatic surface elevation, saturated solute conditions will exist at the top of the horizontal flow system. This will cause greatly increased pollution of the saturated zone (Cole, 1994; Baehr et al., 1989). In the case of a recent spill of an immiscible pollutant with density less than water, when relatively steady groundwater flow prevails, a zone may exist in the capillary fringe of floating product. The infiltrating water, under draining conditions, will reach an equilibrium with the immiscible liquid, resulting in a saturated solute condition. The potential for solute transfer from the unsaturated zone to the saturated zone is greatly increased as a result of groundwater table fluctuation. The result of this scenario is that saturated solute concentration exists at the top of the horizontal flow area of the saturated zone (Fetter, 1993; Lyman et al., 1992; Hoag et al., 1984). The density of the pollutant greatly affects its mobility by restricting its availability in one phase over the other. As a result, SVE performance will be influenced.
16.6
CONTROL VARIABLES
Control variables that affect the success of SVE system include air withdrawal rate, well configuration, extraction well spacing, vent well spacing, ground surface covering, pumping duration, and inlet air volatile organic pollutant concentration and moisture content. 16.6.1 Air Withdrawal Rate
The air withdrawal rate, Q, during SVE, is defined as the volumetric flow rate of gases being drawn from the extraction well due to an applied vacuum. Assuming gravity flow is negligible, the soil structure is unaffected by pressure changes through the soil, and the changes in soil hydraulic conductivity due to drying out the soil are negligible, one may derive expressions for the air withdrawal rate for a SVE system operating on a confined porous stratum of thickness m. Combining Eqs. [16.15], [16.16] and [16.17] yields: [16.18]
which, in radial coordinates, may be approximated by:
kpatm
ot
r Or
--&-r ]
[16.19]
where p ' = p - Patm" Subject to the following boundary conditions"
p/ = 0 and
asr~
[16.20]
416
SOIL VAPOUR EXTRACTION _
Limr-.O ( rop /
Q~
[16.21]
2~mk
Eq. [ 16.19] has a solution:
p/ -
QP 2 rcmk
f7
e -X dx 2~ x 4ktPatm
[16.22]
which may be solved numerically to give p' (r, t). Of particular interest is the time required to reach a steady state pressure distribution throughout the soil. For a typical sandy soil, the time to steady state will depend on factors such as porosity, viscosity and hydraulic conductivity. At steady state, given the following boundary conditions:
P :Pw
at r : R w
[16.23]
at r = R,
[16.24]
and P - Pat,,, a solution to Eq. [ 16.22] may be expressed as:
P2(r)
-
2 Pw
-
( 2 Po,m-
21 ln(r/Rw) ln(RI/R w)
&,
[16.25]
where Pw is the vacuum applied to a well of radius Rw, and R, is the radius of influence of the well which is defined as the radius where the pressure equals the ambient atmospheric pressure. The corresponding flow velocities, v (r), in the soil may be written as:
k
Pw ) pw )
[16.26]
ln(RJR)
which in turn may be used to define the air withdrawal rate, Q, as:
pw[1//2/
Q =2~RwV(Rw) H=H(~_~kg )
where H is the vacuum well screen length.
ln(R )_~
[16.27]
CONTROL VARIABLES
417
For multi-layered soil systems, Q equals the sum of the flow rates through each layer:
Q : ~
Q, : 2~Rw ~
v,(Rw)H
[16.281
where the subscript i refers to layer i. With this development of equations describing SVE, it is seen that the radius of influence, R,, is dependent on the properties and characteristics of the surrounding soils~
16.6.2 Well Configuration Extraction wells used in SVE are very similar in construction to a groundwater monitoring well (Dolan, 1990). The well screen size is chosen to maximize flow into the well without entraining fine soil particles. Gravel filter packs with or without geotextiles are installed around the screen to reduce entrainment of fine soil particles. Extraction wells usually penetrated to a depth below the polluted soil and down to the water table. Throughout the pollution zone, wells are screened. The annular space at the top of the well is grouted to seal the well space from the atmosphere. A vacuum is applied to the extraction well using an injection or regenerative type explosion-proof blower. The extracted vapours may be passed into a treatment system.
16.6.3 Well Spacing and Surface Covering The location of extraction wells should be such that flow through the polluted soil is maximized and flow through unpolluted soil is minimized. If one extraction well is sufficient for an SVE application in a homogeneous soil mass, the well is usually located in the geometric centre of the polluted soil zone. In the case of soil mass characterized by inhomogeneities, the extraction well is located to maximize the air flow through the polluted zone. In some SVE applications, multiple extraction wells are required. The required number of extraction wells can be defined from either one of the following methods.
First method: The number of required wells is given by: Nwell s -
gspill
tCQ
[16.29]
where Nwell s is the number of wells, M,pitt is the mass of pollutants spilled, t is the estimated time required for the remediation, and C is the estimated pollutant removal rate expected for a single extraction well operating at an air withdrawal rate, Q. In this method, vapour concentration changes with time, and residual pollutant concentration after time t are ignored in estimating the value of C.
Second method: The number of wells required is given by: A Nwells
-
~RP2
[16.30]
418
SOIL VAPOUR EXTRACTION
where Apis the total area of pollution, and R~ is the radius of influence. When multiple wells are used, it is important to consider the combining impact of the wells on the vapour flow behaviour. For example, if three wells were located in a homogenous soil in an equilateral triangular configuration, a stagnant region of little vapour flow would exist at the centre of the configuration. Such a problem could be alleviated by locating a passive vent well or a forced injection well at the stagnant location. A passive well is simply a well open to the atmosphere while a forced injection well is a well into which air is pumped. It is important to locate the vent and injection wells within the extraction well's radius of influence. This is especially true in the case of injection wells since one does not wish to force vapours off-site. Passive wells or trenches may also be used around the perimeter of a polluted site in order to prevent the off-site migration of pollutants. Trenches are, however, limited to shallow depth applications, as discussed in Chapter 13. Surface seals such as polymer-based liners or asphalt, concrete and clay caps are sometimes used to increase the extraction well's radius of influence by reducing air infiltration into the soil from the atmosphere. For shallow treatment zones (less than 5 m), the surface seal will significantly influence the vapour flow path and may be added or removed as required to achieve a desired result. However, for extraction wells screened at depths below 8 m, the influence of surface seals becomes less significant. Surface seals also prevent water infiltration into the polluted soil mass, thereby reducing pollutant leaching and preventing water saturation of soil pores. In some sites, a well packed surface cover or a frost layer may act as a surface seal. Ideally, installation of extraction wells, air inlet wells, and surface seals during SVE application should be done incrementally in order to optimize the configuration. 16.6.4 Pumping Duration Vapour extraction systems can be remained on line anywhere from 6 weeks to 3 years depending on the properties and characteristics of the polluted soil and the cleanup objective (Dolan, 1990). Cleanup is quicker in more permeable soils and may be cost prohibitive in low hydraulic conductivity soils. Intermittent blower operation may be more efficient in cases where pollutant transport is limited by diffusion through water or air. However, while such operation results in reduced energy consumption and increased vapour concentrations at the extraction well, the total cleanup times may be considerably increased. The pressure from the outlet side of the extraction pump may be used to force air into injection wells thereby reducing costs. 16.6.5 Inlet Air Concentration and Moisture Content
Air inlet moisture content will have a bearing on the mass balance of soil moisture during SVE. If the inlet air is saturated with water and the soil moisture level is high enough that soil pores are saturated with water vapour, then the soil moisture content will remain unchanged. If the inlet air is less than saturated but the soil contains enough water to saturate the soil pores, then SVE will lower the moisture content of the soil by removing water as vapour. At low soil moisture content, the removal of soil moisture will result in increased air permeability of the soil as less soil pores are occupied by water in liquid form. Inlet air volatile organic pollutant concentration will have a similar effect on the SVE process. If the inlet air volatile organic pollutant concentration is high, then less volatile organic pollutants from the soil will be removed by the SVE process. A high inlet volatile organic pollutant
RESPONSE VARIABLES
419
concentration will limit pollutant mass transfer from the adsorbed, free liquid, and aqueous phases to the vapour phase and, if high enough, may result in an increase in the total amount of volatile organic pollutant in the soil. Essentially, a high inlet air volatile organic concentration will degrade the efficiency of SVE system.
16.7
RESPONSE VARIABLES
Response variables that can be monitored to gauge the success of SVE application are related to soil, extracted air and the system. These variables are discussed below. 16.7.1 Soil-related Variables
Pressure Gradient As discussed, application of a vacuum to an extraction well will result in gas flow from the surrounding soil into the well. Pressure gradient arises in the soil due to resistance by the soil to gas flow. The pressure gradient distribution around an extraction well is govemed by Eqs. [ 16.14] and [16.15]. Distribution of Volatile Organic Pollutants The final volatile organic pollutant distribution in the soil upon SVE completion depends on the nature of the pollutants, the nature of the soil, and the duration and intensity of the SVE application. Volatile organic polluted soils may contain a single pollutant, as in the case of a solvent spill, or contain a mixture of pollutants, as in the case of a gasoline spill. In the case of a mixture of pollutants, the final volatile organic pollutant distribution will be dependent on the components of the mixture. Partitioning of the pollutants and their removal by SVE will depend on the chemical properties of the pollutants and how they interact with the soil system. Partitioning of pollutants to the vapour phase favours more volatile pollutants with high Henry's constant. Thus, these pollutants will be preferentially removed from the soil during SVE application. As an example, benzene with a vapour pressure of 0.10 atm is removed preferentially over toluene with a vapour pressure of 0.029 atm which in turn is preferentially removed over xylene with a vapour pressure of 0.0066 atm. The partitioning of the pollutants between sorbed, vapour, and free liquid phases will vary over time. The free liquid phase will exist only when sufficient pollutants are present. The removal of pollutants will increase with temperature and flow rates through the polluted zone. The spatial inhomogeneities in soil properties and the partitioning of pollutants will also result in spatially preferred removal of pollutants. Mass loss rates from a free floating liquid phase or from an impermeable clay lens in a sandy soil mass will depend on diffusion processes, since vapour flow will not pass through these zones. As well, removal of pollutants sequestered in the soil matrix will also be limited by diffusion processes. Finally, the location of extraction wells, air inlet wells, and surface barriers will have a bearing on the vapour flow rates through the polluted soil, and thus the final volatile organic pollutant distribution in the soil.
420
SOIL VAPOUR EXTRACTION
16.7.2 Extracted Air-related Variables
Concentration The concentration of the extracted air is dependent on the same factors that influence the final volatile organic pollutant distribution in the soil, i.e., nature of the pollutants and of the soil, and the duration and intensity of the SVE application. The final distribution of volatile organic pollutants in the soil may be derived by integrating over time the extracted air concentration in the extraction well. In the case of a site polluted with many types of volatile organic pollutants, the extracted air will initially contain high concentrations of low molecular weight pollutants with high vapour pressures and Henry's constants. As these pollutants are preferentially removed from the polluted soil, the extracted air will contain less volatile components and the mass extraction rate will decrease. Extracted air concentration will also be affected by the vapour flow paths through the soil, and by the spatial inhomogeneities in the soil. Air and water diffusion will limit mass transfer thereby affecting the concentration of pollutants in the vapour flow entering the extraction well. After a long period of SVE operation, the pollutants removed may originate, by diffusion, from the soil matrix, impermeable soil masses, and free floating liquids. Moisture With soils containing relatively large quantities of soil moisture, one may expect the extracted gases to be 100% saturated. Loss of moisture from the soil, however, will occur with SVE, possibly causing an increase in the air permeability of the soil. In cases where the initial soil moisture is greater than 20%, changes in soil properties with moisture removal are not expected to be significant. With lower initial soil moisture content, however, soil moisture removal may be significant enough to alter the biological activity in the vadose zone and lower the soil's air permeability. Temperature The extracted air temperature is usually a function of the climate characteristic of the site, and will vary daily and seasonally. In some cases, the temperature of the soil is artificially raised to promote increased volatilization of pollutants and speed up the SVE process. Steam stripping is similar to SVE with the exception that 150 to 200 ~ C steam is injected into the soil using injection wells and subsurface injection equipment. Few months are typically required to raise the temperature of the soil to a desired level, and several months of SVE are required to remove the pollutants. 16.7.3 System-related Variables
Power Usage The power usage for a typical SVE operation is attributed to pumping and treatment equipment. Power used for pumping will be directly proportional to the gas flow rate and the duration of pumping. Pumping efficiency, defined as the power required to remove a quantity of pollutants, will depend on the location of the extraction wells relative to the pollution in the soil, the number of wells used per volume of polluted soil, the use of surface seals to extend the radius of influence, the use of air inlet wells, the partitioning of the pollutants in the soil, and the properties
MODELLING OF VAPOUR REMOVAL RATE
421
and characteristics of both the pollutants and soil. Higher efficiencies will be obtained with increased soil temperature, reduced air infiltration from the surface, more volatile pollutants, and higher soil hydraulic conductivity. In cases where diffusion processes limit pollutant mass removal rates from the soil, intermittent pumping may be used to improve efficiencies and reduce power consumption. Using the outlet pressure of the pump and treat system to force clean air into injection wells may improve overall system efficiency.
16.8
MODELLING OF VAPOUR REMOVAL RATE
Models for predicting the maximum removal rate have been presented by Martley and Hoag (1984) and Johnson et al., (1988). The former considered only compositions in a residual free phase, while the latter considered the effects of sorption and dissolution processes. The following elaboration is based on the latter approach. In the model development, it is assumed that: (1) local equilibrium exists between vapour, free liquid, sorbed, and dissolved phases, and (2) the pollutant is uniformly distributed throughout a given amount of soil at all times. The total mole balance of component i during the venting operation is:
aM, -
-
QC~
-
[16.31]
B~
dt
where M~ is the total number of moles of i in the soil. This includes free liquid in the soil pores, moles sorbed to the soil or dissolved in the soil moisture, and moles found in the pore vapour. Q is the volumetric flow rate of air and hydrocarbon vapours into the vacuum well, C, is the molar concentration of i in the vapour entering the vacuum well, and t denotes time. The rate of degradation of i, whether due to biological or chemical processes, is lumped into the term B~. The total mole balance of component i is given by:
zp V M
-
RT
Ms + X,MHc + y , M w + k y ~
mw
[16.32]
where x, is mole fraction of i in free liquid phase, z~ is mole fraction of i in the vapour phase, P is total pressure in the pore vapour (atm), c is void fraction occupied by vapour, V is volume of polluted soil (m3), R is gas constant (82.1 • 10.6 m3/mole ~ T is absolute temperature in soil (~ MHc is the total moles in free liquid phase, M w is the total moles in soil moisture phase, yl is mole fraction of i dissolved in soil moisture, k~is distribution coefficient for i species, Ms is total mass of contaminated soil (g), and mw is the molecular weight of water (18 g/mole). The first term on the fight hand side of Eq. [16.32] represents the number ofmoles o f / i n the vapour phase. The second term is the number of moles of i in the free liquid phase. The third term is the number of moles dissolved in the soil moisture. The last term is the number of moles sorbed to the soil particles. In writing this term, it was assumed that the total number of moles in the soil moisture is approximately equal to the number of moles of water. To calculate the equilibrium distribution between phases, it is assumed that the vapour phase behaves as an ideal gas, the free liquid hydrocarbon phase behaves as an ideal mixture, and the soil moisture phase is non ideal. Then x,, y~, and z~ are related by:
422
SOIL VAPOUR EXTRACTION ziP = x, p Vi = a y i p V
= CflT
[16.33]
where p v is the pure component vapour pressure of component i, and a, is the activity coefficient of i in water. Both p v and tr, are functions of temperature. Substituting Eq. [ 16.25] into Eq. [ 16.32], we obtain the following expression: P ~ve-v
Mw
x~
+ MHc + RT
k ,M s +
t~t
8(Mw) = 0
8(Mw) = 1
]
6(M w) = M~
t~m w
]
[16.34a]
if M w = 0
[16.34b]
if M w > 0
[16.34c]
The mole fraction, x,, of i in the hydrocarbon free liquid phase is equal to:
xi -
Mi,HC
[16.35]
The equilibrium distribution between all phases can be determined for any set of M, by iteratively solving Eqs. [ 16.33 ] and [16.34a, b, c] subject to the constraint: xi = 1
[16.36]
The solution can be simplified by assuming that Mw is equivalent to the number of moles of water in the soil moisture. When the computed set of M~ is high enough, a free liquid phase is present, hence the pollutant is distributed in four phases. If all the M, are small, the pollutant is distributed in three phases, namely soil moisture, sorbed, and vapour phases. In that case, Eq. [ 16.33] reduces to: tZi Yi = P R 6 V + M w + k, M s 8(Mw) T a~ a~ m w
[16.37]
For a given set of M,, if all the products a, y,, calculated from Eq. [ 16.36], are less than unity, the equilibrium distribution does not include a free liquid phase.
CASE STUDY 16.9
423
CASE STUDY
The following hypothetical spill condition is used to illustrate the mathematical model predictions. The polluted porous material is characterized as a sandy soil with a bulk density of 1.5 Mg/m 3 and a porosity of 0.4. The soil has an organic carbon fraction of 0.01, and soil moisture content of 10% by dry weight. The volume of gasoline spilled is assumed to be 1500 litre (400 gal) resulting in 4.0107x 10.6 m 3 of polluted soil. The estimated residual gasoline in the soil is 2% by dry soil weight or 20000 ppm of total petroleum hydrocarbon content. The bulk density of the liquid gasoline is assumed to be 0.8 Mg/m 3. The vacuum well flow rate is assumed to be 34 m3/hour (20 cfm). It is also assumed that only 25% of the air actually flows through the polluted soil, yielding an effective flow rate of 8.5 m3/hour (5 cfm). The relative humidity of the incoming air is 100%. With these initial conditions, the model was calibrated and tested. Then, it was run by varying the variables of interest, such as temperature, porosity, moisture content, organic carbon fraction and air flow rate. The calculated results are presented in the following sections.
Figure 16.2. Predicted gasoline removal as a function of time and temperature.
16.9.1 Temperature The calculated gasoline removal as a function of time and temperature is shown in Figure 16.2. Over a period of few days, gasoline removal is very high. As time increases, the removal continues to increase with a low rate. It is also noted that the removal rate increases with temperature. The predicted relative mass removal, (Qm(t)/Qm(t-0)), for gasoline, is shown in Figure 16.3. It is shown that as the temperature increases the time taken for a phase change to occur decreases. For example, at 20 ~ C, prior to day 295, the distribution of gasoline in the soil was in 4-phase system
424
SOIL VAPOUR EXTRACTION
namely free liquid, soil moisture, vapour and adsorbed phases. At 295th day, a phase change occurs from a 4-phase to a 3-phase system. Prior to this transition, a sufficient quantity of volatile gasoline components are present to support a free liquid gasoline phase. However, after 295 days there is not sufficient gasoline components in the soil mass to support a free liquid phase recovery and gasoline is partitioned in the adsorbed, soluble, and vapour phases. Similarly, for 10~ C a phase change occurs at 550 days while for 5~ C it takes 1615 days. This means that as the temperature increases, the removal rate increases.
Figure 16.3. Predicted relative mass removal of gasoline as a function of time and temperature.
16.9.2 Air Flow Rate
Air flow rate plays a major role in the removal of volatile organic compounds. For the following analysis, three flow rates were used (i.e., 17, 34, and 51 m3/h or 10, 20 and 30 ft3/min (cfm)). The calculated results are shown in Figure 16.4. It can be seen that the higher the flow rate, the faster is the removal of gasoline by SVE. As the air flow rate is increased, more air is introduced into the pores of soil particles. Consequently, more hydrocarbon compounds, which were adsorbed, are displaced by the incoming air, resulting in faster volatilization. The higher the air flow rate, the shorter is the time at which a phase change occurs. As shown in Figure 16.4, for a flow rate of 17 m3/h, a phase change occurs at approximately 480 days while for 34 and 51 m3/h, it takes 295 and 180 days, respectively. Increasing the air flow rate alone does not always enhance the system performance since it is primarily dependent on the soil property as well as the blower capacity. In general, increasing air flow rate implies additional energy requirement, hence an added remedial cost.
CASE STUDY
425
Figure 16.4. Predicted relative mass removal for gasoline as a function of time and flow rate.
16.9.3 Organic Carbon Fraction
In order to study the effect of the organic carbon fraction (foc) on the removal of volatile organic compounds, four different values were used (i.e., foc = 0.01, 0.1, 1, and 10). The calculated results, depicting the relationship between the relative mass removal of gasoline, time, and organic carbon fraction content, are shown in Figure 16.5. The results indicate that as the organic carbon fraction increases, the removal rate decreases. This is due to the increased adsorption capacity through hydrophobic bonding between organic chemicals and soil organic carbon, as discussed in Chapter 5. As a result, more pollutants are adsorbed by the soil system, thus restricting partitioning in the vapour phase. Consequently, pollutant availability for extraction is reduced. The time required for a phase change to occur is accelerated as the organic carbon fraction content is increased. For organic carbon fraction contents of 0.01, 0.1, 1, and 10, it is predicted that phase changes will occur after 295, 115, 75, and 55 days, respectively. In general, as organic fraction content increases, volatilization decreases, hence reducing the amount of pollutants to be removed by the SVE system. 16.9.4 Soil Moisture Content
Soil moisture content plays a major role in the removal of volatile organic pollutants by SVE. High soil moisture content decreases the diffusion rate of vapour due to competition with liquids for pore space. However, insufficient moisture causes an adsorption increase because organic pollutants become the wetting fluid. It is known that soil water wets the soil particles and displaces organic pollutants. But when the water is removed, the organic pollutants adhere to the soil, enter some of the soil pores, and physically bound to soil surfaces. As a result, the organic pollutants are not available for volatilization, and the diffusion rate of organic pollutants into soil pores is reduced.
426
SOIL VAPOUR EXTRACTION
When soil air concentration is reduced, the removal rate of organic pollutants is decreased. It should be noted that vapour extraction dehydrates the soil, causing a decrease in the removal rate. To improve performance, the SVE system may be shut down periodically or moist air may be injected into the soil.
Figure 16.5. Predicted relative mass removal for gasoline as a function of time and organic carbon fraction content.
16.10 ISSUES RELATED TO SVE APPLICATION Once SVE has been identified as a viable candidate for remediating a polluted site, the principal activities that must be considered are discussed below.
16.10.1
Feasibility Study
This requires preliminary information about geology of the site, the chemical and physical characteristics of the pollutants, and regulations related to the level of cleanup. There are five major activities/issues to be resolved: (1) Definition of expected range of air flow rate for a single well: This parameter may be estimated with the previously discussed flow equations; Estimation of removal rates for each well: In order to satisfy this requirement, an estimation (2) of the total vapour connected to the air flow is required. Then, the product of flow and concentration will determine the estimate of a single well removal rate; Will this removal rate be acceptable? The rate obtained from (2) is compared with the (3) required rate, which is obtained by dividing the total mass to be recovered by the product of number of wells and mass removal rate per well;
INTEGRATED SYSTEMS (4)
(5)
16.10.2
427
What residual pollution will remain at the site following cleanup, and will it meet the regulatory requirements? Given the air flow rates and a model describing removal rates as a function of time, the soil concentrations of pollutants may be determined at the end of the remedial period. These levels must be compared with the regulatory criteria; and What are the negative effects that may develop as a result o f vapour extraction use? Evaluation of the effects of water table variations must be determined from the use of SVE technology. Furthermore, the effects of other soil remediation techniques should be evaluated and compared with SVE.
Physical Testing
Physical testing must be completed to establish accuracy of model predictions and give an overview of the effectiveness of SVE. Evaluation of pumping tests gives site specific data necessary for finalizing the design of the vapour extraction and water control systems. Monitoring the vapour concentrations during well pump test helps in validating earlier calculations and modelling.
16.10.3
Final System Design
Site specific data is used in conjunction with derived equations to estimate removal rates. Then, using the single well extraction rates, the number and proximity of wells are configured. Vacuum pressure is calculated and the corresponding pumps are chosen. Also, depending on the nature of the pollutants, the appropriate vapour treatment unit is used.
16.10.4
Monitoring
It is recommended that the following parameters be monitored at the site: (1) air flow rates, (2) pressure at each extraction and injection well, (3) ambient soil temperature and air temperature, (4) water table elevation, (5) vapour concentrations and composition at each extraction well, and (6) soil gas vapour concentrations at various distances from the extraction wells. This information will be useful in determining the degree of cleanup.
16.11 INTEGRATED SYSTEMS Based on the previous discussion on SVE site and system parameters, it is apparent that employing other physical, chemical, and thermal processes in parallel with the technique is beneficial. For example, soil heating will increase rates of volatilization, thereby improving mass removal rates of SVE. Several integrated systems are discussed below. 16.11.1
Air Sparging
Air sparging could be used to extend the application of SVE technology to water saturated soils. Air is injected under pressure below the water table, as shown in Figure 16.6. The injected air travels through the saturated zone and strips VOCs from the soil and groundwater into the vapour phase. In addition, the injected air increases the oxygen supply rate to the groundwater, resulting in
428
SOIL VAPOUR EXTRACTION
increased aerobic biodegradation rates. Once air is injected into the saturated zone, its advective flow is governed largely by the applied pressure, buoyant forces, vertical and horizontal hydraulic conductivity distributions in the saturated zone, immiscible fluid displacement, and capillary properties of the soil. Detailed discussion of this technology can be found in Norris et al., (1994). Although the use of air sparging has increased, considerable research and demonstration must be conducted before a consistent and reliable design approach can be realized and before a full understanding of its effectiveness and feasibility can be achieved.
Figure 16.6. Integrated air sparging and vapour extraction systems.
16.11.2
Bioventing
Bioventing is the process of aerating soils to stimulate in-situ biological activity and promoting bioremediation. Since the rate of oxygen supply is increased during the application of SVE, an increase in aerobic biodegradation of VOCs will occur (Wilson and Ward, 1986; Ostendorf and Kampbell, 1989). To convert a hydrocarbon such as benzene into CO2, H20, and biomass, a mass ratio of 3:1 g-oxygen/g-benzene is required (Ely and Heffner, 1988; Staps ,1989; Hinchee et al., 1989). Bioventing clearly represents a positive advance in vapour extraction-based applications at sites having aerobically biodegradable pollutants in the unsaturated zone (vadose zone). This technology extends the potential application of SVE to sites polluted with semi-volatile fuel hydrocarbons. Detailed discussion of this technology can be found in Leeson and Hinchee (1997) and Norris et al., (1994).
INTEGRATED SYSTEMS
429
Case Study
According to Dupont et al., (1991), a spill of approximately 100,000 litre of JP-4 jet fuel occurred when an automatic overflow device failed at Hill AFB in Ogden, Utah. Pollution was limited to the upper 0.2 m of a delta out-wash of the Weber River. The soil was composed of mixed sand and gravel with occasional clay clusters to a depth of 20 m. The groundwater level was located at a depth of 200 m. The average soil moisture was less than 6% by weight. The maximum soil JP-4 concentration was 20,000 mg/kg soil and the average concentration was 400 mg/kg. The pollutant had migrated to a depth of 20 m below the surface.
60 .~
~rj 4~
5O
c~ 40
30 0
E ~ 2O ~
Volatilized
10
Biodegraded 0
0
5
10
15
2O
25
Time (months)
Figure 16.7. Cumulative hydrocarbon removal from Hill AFB site (Dupont et al., 1991).
Extraction wells were drilled to 20 m depth below the ground surface and were screened from 3 to 18 m below the surface. The SVE system was activated in December, 1988 at vacuum extraction rate of 40 m3/hour. The off-gas was treated by catalytic incineration. The venting rate was gradually increased to 2500 m3/hour and maintained for a period of 12 months (until November 1989). Then the ventilation rate was reduced to 500-1000 m3/hour to provide aeration for bioremediation. During the 24 month period between December 1988 and November 1990, more than 1.0• 106 m 3 of soil gas was extracted from the site. The cumulative hydrocarbon removal from the site is shown in Figure 16.7. The Figure shows that about 53,000 kg of JP-4 as carbon had volatilized and 43,000 kg had biodegraded. 16.11.3
Thermal Enhancements
The efficiency of SVE can be enhanced by heating the subsurface. The improved performance is due to the increase in (1) pollutant vapour pressure with temperature, hence an increase in volatilization, and (2) biodegradation rate because of increased microbiological activities.
430
SOIL VAPOUR EXTRACTION
A number of thermal enhancement processes being used or under development are discussed below.
Steam Stripping Steam stripping is the process of injecting steam at 150 to 200 ~ C into the soil to increase the vaporization rate and solubility of pollutants, and reduce interfacial tension between pollutants and soil. The injected steam travels some distance from the injected point and then condenses. The energy lost because of cooling and condensation is transferred to the soil. Pollutants in the soil may be (1) vaporized due to increased vapour pressure, (2) become dissolved in the condensate front due to increased solubility, and (3) displaced due to a reduction in viscosity and capillary forces. These processes must be controlled to minimize possible detrimental effects, such as pollutant smearing or enhanced vapour transport away from the source area. Because the petroleum industry uses steam injection for oil recovery, there is extensive experience with the technique. Steam injection can be applied to semi-volatile as well as to volatile compounds. The organic pollutants can be collected as a separate phase for reprocessing and reuse. Depending on the extent of pollution, the amount of pollutant present, and soil characteristics, treatment times can be significantly reduced compared to those of ambient temperature treatment systems. Treatment times ranging from hours to days have been reported (La Mori, 1989; Udell and Stewart, 1989). In soils with low hydraulic conductivity, the steam flow may be too small to allow practicable treatment. Success of in-situ steam stripping operations can also be limited due to soil heterogeneity. In addition, steam injection systems are unable to heat formations to temperatures significantly greater than 100~ C. Hence, it will not be efficient for removal of higher boiling point compounds, such as some aliphatic and aromatic fractions of jet fuels and gasoline, chlorobenzene, trichloroethylene, dichloroethane, and tetrachloroethane.
Radio Frequency Heating Radio frequency heating is a process of increasing soil temperature to promote vaporization of pollutants. It has the potential of increasing subsurface temperature well above 100~ C, allowing more rapid removal of higher boiling point compounds. In this process, energy is delivered to the subsurface via radio frequency waves, which excite molecular motion and induce heating. In addition to standard soil vapour extraction system, the system utilizes electrodes or antennae connected to a radio frequency generator for transmitting waves into the soil formation. Some of the energy is absorbed for heating. The exciter array electrodes can be inserted into holes drilled into the formation or positioned on the soil surface. Radio frequency heating occurs through ohmic and dielectric mechanisms. Ohmic heating results from a voltage drop pushing electrons up into the conduction band and moving them through the soil mass, producing resistance heating. Dielectric heating results from distortion of the atomic or molecular structure in response to an applied electric field. Typically, dipoles in a polar substance are randomly oriented. The application of an external electric field will cause dipoles to align.
16.12 SUMMARY AND CONCLUDING REMARKS Soil vapour extraction (SVE) can be effectively used for removing a wide range of volatile chemicals under a wide range of conditions. The process can be part of a remediation effort and used
SUMMARY AND CONCLUDING REMARKS
431
along with other treatment processes. As with any in-situ process, it is usually impossible to do a complete materials balance on a given site because most sites have an unknown amount of volatile organic pollutants in the soil and in the groundwater. As a result, it is difficult to establish when the cleanup has been completed. To ascertain if pollutants are still vaporizing, SVE processes could be operated on an intermittent basis for a period of time after the concentrations in the extracted vapours have approached zero. This technique can help to limit the soil boring and groundwater data on residual pollution required to demonstrate that cleanup objectives have been attained. Although many aspects of the SVE technology are understood, additional information which would be of assistance to engineers and operations personnel needs to be developed. Documentation of case studies where difficulties resulting from the mis-application of the technology or operating problems would be useful. All too often, only success stories are documented. However, failure or marginal successes can provide more insight into factors which need to be considered in other applications. These can be particularly useful when a thorough investigation/assessment of the failure identifies the causes of that failure. It is known that temperature affects the volatility of organic compounds and therefore the efficiency of the SVE process. Under what conditions and for what pollutants would it be costeffective to consider technologies such as radio frequency heating to increase the soil temperature in order to extract organic pollutants or speed remediation? Would heated air injection serve the same purpose, and what effect would it have on extraction rates? How would implementation of such techniques change the overall efficiency of the process for different soil types? Little attention has been paid to the effect of naturally occurring soil organic matter on extraction rates or the ultimate removal efficiencies. Since such soil components can adsorb organic pollutants, to what extent do they reduce the effectiveness of the process? Since buildings and other interferences can be found at many sites where volatile organic pollution occurs, it is likely that it will be necessary to remediate polluted soils near these structures. What special techniques (i.e., horizontal drilling) could be used under these conditions? Would using SVE result in transporting vapours into the building or could they be vented away? Other technical issues that need to be resolved include (1) the effectiveness of forced or passive vapour injection wells, (2) other means of controlling vapour flow paths, (3) well documented demonstrations for removing pollutants from low hydraulic conductivity soils, (4) the possibility of decreasing residual pollutant levels in water saturated zones by air sparging/vapour extraction, and (5) optimal operation schemes for multiple vapour extraction well systems.
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CHAPTER
SEVENTEEN
SOLVENT EXTRACTION PROCESSES
17.1
INTRODUCTION
Chapters 12 to 16 dealt with remedial measures for: (a) extracting pollutants from grotmdwater via pump and treat system (Chapter 12), (b) restricting off-site pollutant migration via subsurface containment techniques (Chapter 13), coveting systems (Chapter 14), and lining systems (Chapter 15), and (c) extracting pollutants from the vapour phase via vapour extraction systems (Chapter 16). Non of these remedial measures involved treatment of the source of pollution. This chapter is the first of five chapters (Chapters 17 to 21) that deal with in- and ex-situ treatment techniques for polluted soils. Treatment of polluted soils encompasses processes which alter the chemical structure of the soil constituents, and produce a residue less hazardous than the original or render the constituents easier to remove from the soil. These remedial methods may be divided into three broad categories: (1) those that are aimed at removing the pollutants (extraction processes), (2) those which destroy the pollutants (thermal processes), and (3) those which transform the pollutants (biological processes). Polluted soils may be excavated and treated on-site or elsewhere, a process known as ex-situ treatment. Alternatively, they may be treated in-situ, i.e., without excavation. The feasibility of insitu treatment is dependent on the site geology and hydrology, and on the characteristics of the soil itself and of its pollutants. Due to their greater hydraulic conductivity, sandy soils are more amenable to in-situ treatment than clayey soils. Care must be taken to ensure that treatment reagents do not migrate away from the treatment zone. Various pollution extraction technologies, based on soil washing or flushing, are available for the treatment of polluted soils. Solvent- or surfactant-based chemical reagents may be used to extract the pollutants. The principles of solvent-based extraction processes are discussed in this chapter.
17.2
PROCESS DESCRIPTION
Solvent extraction serves to separate pollutants, water and particulate solids into various fractions. Pollutants are transferred to and carried off by the liquid phase. Two principal mechanisms may be utilized: (1) pollutants are dissolved in the extracting agent, with or without chemical reaction, and (2) pollutants are dispersed in the extracting phase as suspended or colloidal particles, with or without prior mechanical treatment. The polluted particles are then separated from the clean soil particles by making use of the difference in their particle size, settling velocity, surface properties or combinations of these. 433
434
SOLVENT EXTRACTION PROCESSES
Polluted soils may be excavated for the purpose of washing, or the site may be flooded with the appropriate flushing solvent solution and the leachate collected in a series of shallow well points or subsurface drains. If the soil is to be excavated, pre-treatment is usually necessary to remove large objects and to reduce the size of the hard clods of soil, as illustrated in Figure 17.1. Water or solvent may be added to the soil to make it pumpable. The soil is then mixed intensively with the extracting agent, and pH adjustment may be required at this stage. Once the pollutants have been transferred to the fluid phase, the soil and the extracting agent are separated. The soil is washed again with a clean extracting agent or water to remove the original extracting fluid. The larger particles in the fluid phase are separated as much as possible and then re-washed. Finally, the polluted extraction fluid is cleaned and, where feasible and economical, is reused. In some cases, the purification step may be applied directly to the unseparated suspension of soil particles, and the separation of the soil particles carried out following this step.
M akeup water
Recycled water
ii
Chemicals
Extracting agents
i
Polluted soil
Soil prepparation
Soil washing Prepared process soil ,.._
Blowdown
water I
,._ W a s t e w a t e r treatment ......
9W a s h i n g 9Rinsing
~
9Size separation Oversized reacts
I
Polluted soil
Treated
water
Clean soil
Figure 17.1. Aqueous soil washing process.
Extraction methods in which the pollutants are generally not dissolved in the extracting liquid and are separated from the soil based on different settling velocities can be quite effective. Since pollutants are often adsorbed to clay particles and humus, reasonable cleaning can result when these clay fractions and humus are separated from the soil. Selection of extracting agents, other than water, must take into account a number of factors. Of primary concern is the safety of the agent to human and the environment. Clearly, it is undesirable to pollute further the environment with excess or uncontrolled amounts of either extraction fluid, or undesirable side products. The desired degree of pollutant removal must also be considered. The ease of handling of the agent, and its potential to be purified and recycled are additional major factors. Economic considerations, of course, will also influence the final decision.
EXTRACTION TECHNOLOGIES 17.3
435
EXTRACTION TECHNOLOGIES
17.3.1 Metal Mining Solvent extraction of metals from polluted soils is a technique which has received considerable attention lately, and it appears to have great potential. There are three solvent extraction technologies in the mining industry which have potential use for cleanup of metal-polluted soils: (1) vat leaching, (2) solution mining, and (3) heap leaching. Vat leaching and solution mining have received considerable attention in the waste treatment industry while heap leaching has been largely ignored. This is of particular interest since, in the mining industry, the techniques of choice are solution mining and heap leaching. Vat leaching, the more complex and expensive technique, is only considered if heap leaching is not a viable alternative. The vat leaching technologies used in soil remediation are based on soil washing. Vat leaching is a solvent extraction technique. It is a fines separation procedure, and would be more appropriately calledfines separation since the clay and silt materials are washed out of the polluted soil, separated from the liquid washing stream, and collected (Lauch, 1989). The pollutant is frequently adsorbed by the fine fraction in the soil. This process removes the pollutant from the bulk of the soil. The major disadvantage of this technology is that, although soil washing reduces the volume of soil to be treated, it does not completely cleanup the soil. The remaining polluted fines must still be treated. In solution mining, the polluted soil is excavated, and placed in an agitated vessel with an extraction solution. When the soil is sufficiently clean, the solids are separated from the liquid, and the liquid is processed or reused. Unlike the fines separation method discussed above, the polluted material is actually removed from the soil phase and introduced into the liquid phase. The pollutant must then be removed from the liquid phase. This is a common characteristic of a solvent extraction treatment processes. Like the fines separation technique, this reactor-based technology has good process control and QC/QA assessment. However, it is energy intensive and requires a high level of operational skills. This technology operates with a solid:liquid ratio between 1:10 and 1:70, and appears to be restricted to soils with fines concentrations of less than 30% (defined as fines with particles smaller than 60 ~tm). Soil flushing is a technology for in-situ extraction of toxic materials, and is very similar to the solution mining technique used in the mining industry. An extraction fluid is applied in-situ to the un excavated polluted soil. At the base of the polluted soil zone, the flushing fluid is recovered using subsurface drainage pipes, trenches, wells or well points (Kunz and Gee, 1989). The advantage of this system is the ability to treat large amounts of material without handling the material itself. The disadvantages are: (1) since the soil is anisotropic and nonhomogeneous, soil flushing posses potential preferential flow problems, (2) there is a potential for groundwater pollution, (3) the extraction fluid may cause soil dispersion and plug the aquifer, hence reducing the hydraulic conductivity, and (4) QC/QA assessment is difficult. In the uranium mining industry, where soil flushing has been used extensively, groundwater pollution has been a serious problem. Field testing has shown that complete restoration, as defined by the state regulatory agencies, of solution mining sites has not been achieved with any reasonable degree of aquifer flushing (Tweeton, 1981; Deutsh et al., 1985; Charbeneau, 1984; Yelderman et al., 1983).
436
SOLVENT EXTRACTION PROCESSES
The third solvent extraction technique, heap leaching, has been largely ignored as a hazardous waste treatment technique. In the heap leaching process, metal bearing ore of interest is excavated and mounted on a pad. The metals are removed by passing extraction fluid through the ore using some type of a liquid distribution system. Traditionally, a simple sprinkler system was used. Recently, drip irrigation systems have been employed (Trexler et al., 1990). The extraction fluid is collected in a solution pit and processed to remove the metal of interest. The advantages of heap leaching are: (1) it is operationally simple, (2) QC/QA assessment is practical, (3) it is not energy intensive, (4) liquid to solid ratio used is low, (5) it can be performed on-site on a large scale, (6) it is applicable to soils with high percent of clay content, and (7) it eliminates the need for long term monitoring. The disadvantages are: (1) multiple handling of the polluted material, and (2) process control is difficult if one needs to adjust temperature, and impossible if one would like to modify the pressure in the heap. The precious metals industry has recently introduced the use of geothermal resources to heat the leach pile and accelerate the leaching process (Trexler et al., 1990).
17.3.2 Metals Mining Methods Versus Hazardous Waste Treatment Techniques Considering the wealth of information available on heap leaching in the mining industry, it might appear that the outcome of heap leaching of toxic metals can be pre-determined (Hanson et al., 1992). However, there are a number of major differences between metal mining methods and hazardous waste treatment techniques. The primary differences are physical and geochemical in nature. In the mining industry, the metals of interest are usually associated with solid matrix of particles larger than 10 to 25 mm in diameter. In contrast, in polluted soil, metals are adsorbed onto the surface of the fines portion, consisting of particle sizes less than 1 ~tm. As discussed in Chapter 5, heavy metals are retained in the fines portion of soil particles in five different associations, namely: (1) exchangeable sites of clay minerals, organic materials, and amorphous materials, (2) associated with carbonates, (3) associated with amorphous or poorly crystallized Fe, A1, and Mn oxides, (4) associated with soil organic matter, and (5) residual fraction within the lattice of silicate minerals. These associations are discussed in section 17.4. The nature of metal association with a solid matrix significantly affect the dissolution time of the metal. The size of the soil particles will affect the application rate and the distribution of the extraction fluid in the leaching pile. In addition to the system's physical and hydrodynamic differences, the metals of interest are generally different. In the mining industry, metals are extracted from the rock matrix until an economic constraint is reached. In the hazardous waste treatment industry, on the other hand, pollutants are extracted until a regulatory treatment is achieved. In spite of the fact that the two systems (i.e., metal mining and hazardous waste) are different, there is much that the hazardous waste treatment industry can learn from the mining industry experience with metal removal from rock matrix using extraction fluids. Table 17.1 (Hanson, et al., 1992) contains a list of the leached metals, and the extraction technique and fluids used. Similarly, Table 17.2 is a listing of heavy metals, soil type, and reagents which may be used to treat soils polluted with the corresponding heavy metal.
SELECTIVE SEQUENTIAL EXTRACTION
437
Table 17.1: Metals which have been successfully mined using solvent extraction. Metal
Process
Extraction fluid
Lithium (Li) Beryllium (Be) Sodium (Na), Magnesium (Mg) Aluminum (AI) Vanadium (V) Chromium (Cr) Manganese (Mn) Cobalt (Co) Nickel (Ni)
Vat Vat
H20 after roasting at 900~ H2SO4 then H20
Vat Vat Vat Heap Heap Heap Vat
SO 2 and temperature Boiling 25% HC1
Copper (Cu) Zinc (Zn)
Heap Vat
Arsenic (As) Selenium (Se) Zirconium (Zr) Niobium (Nb) Molybdenum (Mo) Silver (Ag) Cadmium (Cd) Tin (Sn) Hafnium (Hf) Tantalum (Ta) Tungsten (W) Gold (Au) Mercury (Hg) Lead (Pb) Uranium (U)
Heap Heap Vat Vat Heap Heap Vat Vat Vat Vat Heap Heap Heap Vat Solution/ Heap
17.4
Comments
Boiling H2SO 4
Acid SO2 5% by wt. H2SO4,pH = 1.5-3.0 Heat, pressure, mixing NH4, H2SO4, HC1 & NO3 H2SO4, pH = 1.5-3.0 Chlorine & 02 heat and pressure Alkaline NaCN Dilute HC1 Acid boil, alkaline 3.03% NaOCH NaCN H2SO4 and mixing Roasting, H2SO4 & NaCI Dilute HC1 Acid boil, alkaline HC1 NaCN NaCN FeC1, heat, pH = 3.0 Alkaline
Laboratory Laboratory Cu by-product
Commercial
U tailings Au by-product
Laboratory Commercial
Laboratory Commercial Au by-product Commercial Tailings
SELECTIVE SEQUENTIAL EXTRACTION
17.4.1 Description of the Technique
The selective sequential extraction (SSE) method is based on the fact that different association of heavy metals that are retained in soil (e.g., as oxides, hydroxides, carbonates, bound with soil organic matter, etc.) can be extracted selectively using appropriate extraction reagents (Tessier et al., 1979). SSE uses appropriate chemical reagents in such a manner that different heavy metal fractions can be extracted from the soil solids once the binding phase is destroyed. The metal
438
SOLVENT EXTRACTION PROCESSES
associations present are transformed selectively onto species, and then an analytical procedure which can detect the species is applied. The SSE method of analysis is a potentially useful tool for extracting the solid forms of metals. In most cases, the extraction reagents are taken from soil analysis (Jackson, 1958). These reagents are classified as (Tessier et al., 1979): (1) concentrated inert electrolytes, (2) weak acids, (3) reducing agents, (4) complexing agents, (5) oxidizing agents, and (6) strong acids.
Table 17.2: Metals which have been successfully removed from polluted soils using solvent extraction. Element
Soil Type
Reagent
Reference
Pb
Artificial Soil
DTPA, EDTA.
Pb, Zn, Cu, Cd
Kaolinite, smectite, illite
Ag, Cr, Cu
Kaolinite, smectite
Pb
Sand and silty clay
EDTA (0.001 M, pH 7), oxalic acid (0.1 M pH 3.3); ammonium oxalate (0.1 M), ammonium nitrate (0.1 M), nitric acid (0.1 M), sodium citrate (0.01 M). HNO3, Ca(NO3)2, EDTA, Fulvic acid. EDTA (0.01-0.1 M)
Karamanos et al., (1976). Farrah and Pickering (1978).
Cu, Zn, Cd, Ni, Pb, Cr Pb, Zn, Cu, Cd Pb, Cd, Cu
Artificial soil
EDTA, citric acid
Contaminated soil Kaolinite
Pb, Cu, Zn, Cd
Illitic soil
CaC12 NTA, EDTA, EGTA, DCyTA EDTA (0.001-0.1 M, pH 4-8)
Pb, Cu, Zn, Cd, Ni
Naturally polluted soil Kaolinite
Hg
EDTA (0.1 M) Citric acid
Flemming et al., (1990). Peters and Shem (1992). Leidmann et al., (1994). Asami (1995). Hong and Pintauro (1996). Mohamed and Castellan (1997a&b); Castellan (1996). Mohamed and Trasente (1996). Singh et al., (1996).
The sequence of application of the reagents differs between different researchers since there is no accepted procedure which dictates which reagents should be used ahead of the other. A typical sequential extraction procedure, for fractionation of heavy metals in soils, is shown in Table 17.3 (Phadungchewit, 1990). The five different associations which appear to be common to most research groups include:
SELECTIVE SEQUENTIAL EXTRACTION
439
Table 17.3: Sequential extraction procedure used in the fractionation of heavy metals retained in soils Reagent and extraction method
Heavy metal phase released from soil
I. Extraction of soil: At room temperature with 8 ml of potassium nitrate (KNO3) for 1 hour with continuous agitation.
Exchangeable cations
II. Extraction of residue from step I: At room temperature with 8 ml of 1 M sodium acetate (NaOAc) adjusted to pH 5 by acetic acid (HOAc) for 5 hours with continuous agitation.
Carbonates
III. Extraction of the residue from step 11: With 20 ml of 0.04 M hydroxylamine hydrochloride (NH2OH.HC1) in 25% (v/v) HOAc at 96+ 3 ~ C for 6 hours with occasional agitation.
Oxides and hydroxides
IV. Extraction of residue from step III: lSt: With 3 ml of 0.02 M H N O 3 and 5 ml of 30% H202 adjusted to pH 2 with HNO3 at 85+ 2 ~ C for 2 hours with continuous agitation; 2"d: With 3 ml of 30% H202 adjusted to pH 2 with HNO3 at 85+ 2~ for 3 hours with intermittent agitation; and 3ra: With 5 ml of 3.2 M ammonium acetate (NH4OAc) in 20% (v/v) HNO3, diluted to 20 ml at room temperature, with continuous agitation for 30 minutes.
Bound to organic matter
V. Digestion of residue from step IV: lSt: With a 5:1 mixture of hydrofluoric acid (HF) and perchloric acid (HC104); and 2 "a : Dissolve the residue left from 1st step with 12 N hydrochloric acid (HC1) and dilute to 25 ml.
Residual fraction
Exchangeable Metals in this group are considered to be non-specifically adsorbed and ion exchangeable, i.e., they can be replaced by competing cations. The sorbent soil solids are mostly clay minerals, organic materials, and amorphous material. Natural salts, such as MgC12, CaC12, BaC12 and NaNO3, are commonly used as ion-displacing reagents to promote the release of ions physically bound by electrostatic attraction to the negatively charged sites on the soil particle surfaces. Because of the affinity of groups ii and III cations (valence of 2 or 3) for most surface sites, cations in the reagent solution have to be present in larger concentrations than the metal being extracted. In practice, concentrations higher than 1 M are widely used although lower concentrations are sometimes
440
SOLVENT EXTRACTION PROCESSES
favoured if natural leaching conditions are to be simulated. With regard to the dissolution of the soil solids from contact with neutral electrolytes, little evidence exists to suggest that this occurs. Pickering (1986) reported that MgC12 sediment leachate contained only low levels of A1, Si, and organic carbon, confirming the weakness of the neutral salts in interaction with the clay surfaces, sulfides and soil organic matter. Since neutral salt solutions are applied at neutral pH, dissolution of Fe or Mn oxides are not likely, and only minimal dissolution of carbonates is expected. Other types of salts, such as NH4C1 and NHaOAc, may dissolve considerable amounts of compounds such as CaCO3, MgCO3, BaCO3, MgSO4, and CaSO4. NH4OAc can, in addition, cause some dissolution ofMn oxyhydrates and metal oxide coatings. Mn oxides and carbonates can be removed by using hydroxylamine hydrochloride solution buffered with NaOAc (Chester and Hughes, 1967). Carbonate Association Metals precipitated or co-precipitated as natural carbonates can be released by application of an acid. The generally used extraction reagent is acidified acetate. A solution of 1 M HOAcNaOAc (pH 5) appears to be sufficiently efficient in dissolving calcite and dolomite, releasing the metals bound to them without dissolving the soil organic matter, oxides, or clay mineral particle surfaces. Metal Oxide Association The metals considered are those which are attached to amorphous or poorly crystallized Fe, A1, and Mn oxides. Ferromaganese, ranging from completely crystalline to completely amorphous, occurs as coatings on soil particles. The varying degree of crystallization results in several types of association with the heavy metals: (1) exchangeable forms via surface complexation with functional groups (e.g., hydroxyls, carbonyls, carboxyls, amines, etc.), and (2) precipitation and coprecipitation. A more complete recounting of the various interaction/retention mechanisms can be found in Chapter 5. Because of the considerable difficulties in separating the contributions from various functional groups, or even in attempting to quantify each of the groups present within the soil system, it is not feasible to provide a detailed quantitative functional group association (complexation) with the various pollutants. It is useful to restate once again that in choosing the specific reagent, one tries to select the reagent that would perform a known specific function, thus permitting one to deduce, from the results obtained, a direct association between heavy metal removed and mechanism responsible for initial retention. As noted previously, we have no assurance of 100% selectivity. Thus, whereas amorphous ferromaganese oxyhydrates, for example, can be dissolved under the effect of redox gradients in sediments, leading to a marked increase in metals in the interstitial water, it is necessary to choose an extraction method capable of differentiating between amorphous and crystalline oxides. The reagent selected for oxyhydrates should not attack either the silicate minerals or the soil organic matter. Such a reagent has been used by Chester and Hughes (1967). It consists of a combination of an acid reducing agent (1 M hydroxylamine hydrochloride) with 25% (v/v) acetic acid. Confirmation of the inability of this technique to reduce the natural organic carbon in sediment samples has been reported by Tessier et al., (1979). However, it was pointed out that this acidreducing combination may release A1 and Si if hot digestion is used as part of the extraction technique. The primary objective, however, is to ensure that the Fe or Mn will not precipitate when
SELECTIVE SEQUENTIAL EXTRACTION
441
released into solution. The modifications introduced by Belzile et al., (1989) to the Tessier et al., (1979) procedure should lessen the risk. Organic Matter Association The binding mechanisms for metals in association with soil organic matter include adsorption and complexation. Approximately one-third of the total binding capacity of marine humic acids, for example, is due to cation exchange, with the remainder due to complexing sites. Because of the different types of binding mechanisms, some overlapping effects will be obtained with those methods designed to release exchangeable cations. However, the general technique used with respect to metal binding to organic matter is to obtain release of the metals as a result of oxidation of the organic matter. Residual Fraction This metal fraction is generally retained within the lattice of silicate minerals, and can become available only after digestion with strong acids at elevated temperature. The residual materials consist of silicates and other resistant materials. 17.4.2 Retention Phases
The amounts of heavy metals extracted as exchangeable metals from the different constituents in the natural clay soil from Quebec are shown in Figures 17.2(a) to 17.2(d) (Phadungchewit, 1990). From these Figures, it is evident that the amounts of heavy metal fractions retained by each soil constituent changes with soil pH. At high pH level, retention of Pb appears to be higher than Cu, Zn, and Cd. Except for Cd, all the other tested heavy metals (e.g., Pb, Cu, and Zn) appear to be almost or fully retained at the highest pH level. When the pH of the soil solution is greater than 4, heavy metals are retained in soil as exchangeable fractions, forming new hydroxides (precipitation) or bound to the existent oxides and carbonates, resulting thereby in high amounts of total heavy metals retained. But as the amount of acid input increases or when pH becomes less than 4, only the exchangeable metal fraction predominates, resulting in lower heavy metal retention. The results of the calculations using the geochemical model (M1NTEQ), for distribution of the probable percentage of precipitated and dissolved species of heavy metals in the leachate solution at different pH values can be seen in Table 17.4. The calculations show that, as the pH increases, the form of heavy metal species in the solution changes from a simple to a more complex form. For example, about 80% of Pb remains as a free cation (Pb 2+) at pH 1-5. At pH 6, 70% of Pb precipitates and more species of Pb are formed in solution. As the pH increases, a higher percentage of Pb precipitates, and the percentages of Pb hydroxide species increase. The same kinds of results are obtained for Cu, Zn and Cd. However, Zn precipitates at higher pH values than Pb and Cu whereas Cd does not appear to precipitate even at pH 8. It can be expected from these calculated percentage distributions of heavy metals that the retention mechanisms of heavy metals in soils would change according to the difference in the speciation of heavy metals as the soil solution pH changes. It is acknowledged that equilibrium models such as MINTEQ or other geochemical moods (e.g., GEOCHEM and PHREEQE) may not give exactly the same percentages because the model predictions depend on the values of thermodynamic constants used in database. However, geochemical models can be used as a guide for selecting different types of chemical extracts for each phase.
442
SOLVENT EXTRACTION PROCESSES
Figure 17.2(a). Lead retention in Quebec natural clay soil in different phases by sequential extraction analysis when heavy metals were applied compositely with leachate.
Figure 17.2(b). Copper retention in Quebec natural clay soil in different phases by sequential extraction analysis when heavy metals were applied compositely with leachate.
SELECTIVE SEQUENTIAL EXTRACTION
443
Figure 17.2(c). Zinc retention in Quebec natural clay soil in different phases by sequential extraction analysis when heavy metals were applied compositely with leachate.
Figure 17.2(d). Cadmium retention in Quebec natural clay soil in differem phases by sequential extraction analysis when heavy metals were applied compositely with leachate.
444
SOLVENT EXTRACTION PROCESSES
Table 17.4: MINTEQ results for Metal Pb precipitated (%) Pb dissolved (%) Dissolved Pb species (%): pb 2+ Pb CI§ PbNO 3 PbOH § Pb(OH)2 ~ Cu precipitated (%) Cu dissolved (%) Dissolved Cu species (%): C u 2+
p i l l ]2
ties distribution for leachate
.
]3[4
,7 ' 8
,5
,6
0
0
100
100
0 100
0 100
0 100
70.4 29.6
99.7 0.3
100 0
88.3 9.6 2.1
88.1 9.8 2.0
88.1 9.8 2.0
88.1 9.8 2.0
88.0 9.8 2.0
88.6 10.1 2.1
78.8 9.3 1.9
1.1
9.9
40.5 4.9 1.0 1.7 51.8
0
0
0
0
0
100
100
100
100
100
92.5 7.5
99.8 0.2
0
99.2
99.2
99.2
99.2
99.2
97.0
43.8 2.9 52.7
98.6
CuOH § Cu(OH)2 ~ Zn precipitated (%) Zn dissolved (%) Dissolved Zn species (%): Zn 2+ ZnOH § Zn(OH)2 ~ Cd precipitated (%) Cd dissolved (%) Dissolved Cd species (%): Cd 2+ CdC1§ CdOH +
1.2
100
0
0
0
0
0
0
0
97.3
100
100
100
100
100
100
~ 100
2.7
99.2
99.2
99.2
99.2
99.2
99.2
98.2
0 100
0 100
0 100
0 100
0 100
0 100
0 100
78.8 20.5
78.5 20.9
78.5 20.9
78.5 20.9
78.5 20.9
77.6 21.7
77.3 21.9
85.7 6.4 6.4 0
100 76.0 22.1 1.0
It can be seen from Table 17.4 that as the pH value of the solution increases, a higher percentage of metals precipitate, and hydroxy species are formed. The formation ofhydroxy species enhances the retention of heavy metals by the hydroxide phase. The experimental results show that the pH at which heavy metal retention by the hydroxide phase begins is lower than the pH at which the metal starts to precipitate and form hydroxy species. For example, in Figure 17.2(a), it is seen that retention of Pb in soil by the hydroxide phase begins when the pH is about 4 and, as shown in Table 17.4, Pb begins to precipitate and form hydroxy species (PbOH § when the pH is about 6. This suggests that it might be easy for heavy metals to form hydroxy species in the presence of soil particles, probably due to the fact that the soil particle surfaces provide good nucleation sites where precipitates can grow at a fast rate. The other possibilities are: (1) since the operationally defined hydroxide phase also involved the dissolution of Fe-Mn hydroxides, the presence of these hydroxides may explain the increased retention of Pb, and (2) as suggested by James and Healy (1972), the
SELECTIVE SEQUENTIAL EXTRACTION
445
addition of an OH group to the metal ion reduces the free energy required for adsorption. A further observation of the results shown in Figure 17.2(a) reveals that the retention of Pb as exchangeable ions increases as the pH increases, indicating the influence of pH on variably charged soil constituents. However, when the soil solution pH exceeds 4, exchangeable Pb drops, and Pb retention by hydroxide and carbonate phases increases, as shown in Table 17.4. At pH values of 1 to 4, Pb is present in the solution as a free cation (pb2§ The dominant mechanism of Pb retention in the soil is by cation exchange, and the amounts retained increase as pH increases. However, when the soil solution pH increases to a certain level, Pb begins to form hydroxy species, resulting in the beginning of Pb retention in the soil by the hydroxide phase. From this point onward, the amounts of Pb 2§ decrease, as evidenced by the drop in the curve of Pb exchangeable form in Figure 17.2(a). The results obtained with respect to Cd retention (Figure 17.2(d)) show that the amounts of Cd retained in the hydroxide phase are smaller in comparison with Pb, Cu, and Zn, for the same forms of retention. The reason for this might be found from the information in Table 17.4 which shows that Cd begins to form hydroxy species at a very high pH, and that no precipitation occurs even at pH 8. In addition, it would appear that the amount of C1- present in the leachate affects the formation of Cd hydroxy species by forming Cd-C1 complexes (Yong and Sheramata, 1991). This results in smaller amounts of Cd being retained in the soil by the hydroxide phase. Therefore, Cd exhibits the highest mobility among the four tested heavy metals. The presence of carbonates in Quebec natural clay soil contributes measurably to the retention of the heavy metals, through precipitation with the carbonates, as seen in Figures 17.2(a) to 17.2(d). These results correspond to observations reported by Udo et al., (1970) and Yanful et al., (1988). The higher the carbonate content of the soil, the greater the amount of heavy metals which can be retained by the carbonate phase (Yanful et al., 1988). The amount of heavy metal retained in the carbonate phase becomes negligible when the pH decreases to less than 3. At low pH levels, the dissolution of carbonates originally present in the soil appears to be responsible for the decreased amount of heavy metals retained by the carbonate phase (Buckman and Brady, 1969).
17.4.3 Mobility and Bioavailability Selective sequential extraction is usually conducted to determine the mobility/retention and bioavailability of heavy metals in polluted soils. Bioavailability means "easily available to plants or microorganisms." It can be determined by exchanging heavy metals with neutral salts such as MgCI2, CaC12 or BaC12. From the previous discussion, it is clear that Cd and Zn are exchangeable with these neutral salts because they are loosely bound to soil surface. Lead Lead exists principally in the +2 oxidation state in soils. Under oxidizing conditions, Pb 2+ ion becomes less soluble as soil pH is raised. At high pH, Pb retention mechanisms are: (1) complexation with organic matter, (2) chemisorption on oxides and silicate clays, and (3) precipitation as carbonate and hydroxide species. In alkaline soils, solubility may increase by formation of soluble Pb-organic and Pb-hydroxy complexes. Lead is the least mobile heavy metal in soils, especially under reducing or non-acidic conditions. Most of the lead in polluted soils appears to be unavailable to plant tops. Plant Pb 2+ concentrates in the roots, translocating very little from the roots to tops as long as the plant is actively
446
SOLVENT EXTRACTION PROCESSES
growing. Toxic effects of lead on plants have not often been observed, but a hazard to animals exist because of the inherently higher toxicity of this metal to animals. The health concern with leadpolluted soils arises mostly from the contamination of plants by soil particles and ingestion of soil by humans and grazing animals. The acute toxicity of lead in human is non-existent. However, chronic toxicity of lead is numerous, which can damage brain and neurological centres.
Copper Most of the colloidal material of soils (oxides of Mn, A1, and Fe, silicate clays and humus) adsorb Cu 2§ strongly, and increasingly so, as the pH is raised. Organically complexed CuR+is bound more tightly than any other divalent transition metal. Therefore, the mobility of these complexes is rather low, and the bioavailability is limited. For this reason, farmers have been able to apply large amounts of Cu salts to organic soils over time without toxicity effects to crops. Because of the high affinity of Cu 2§ for soil colloids, copper is rated a low mobility metal in near neutral soils, undergoing virtually no downward migration from the polluted soil. In most alkaline soils, where free Cu 2§ solubility is exceedingly low, soluble complexes of Cu 2§ (most importantly hydroxy, carbonate and organic matter complexes) form and increase the total copper solubility. Consequently, mobility may be significant under high pH conditions. Most of the total dissolved copper in surface soils, over a fairly wide range of pH and particularly at higher pH, is in the form of Cu-organic complexes. Copper toxicity is generally found in mining areas. Short term exposure to copper gives symptoms of coughing and more prone to infection of the lungs. Long term damage causes liver failure and affects the blood system. Zinc
Zinc has only one possible oxidation state is soils, i.e., Zn 2§ In acidic and aerobic soils, Zn is retained on clays and organic matter via exchangeable forms. Therefore, Zn has medium mobility. At higher pH, however, chemisorption on oxides and alumino-silicates and complexation with humus lower the solubility of Zn 2§ Consequently, Zn mobility in neutral soils is very low. If soils are slightly alkaline, Zn-organic complexes can become soluble and raise mobility, even though the activity of the free Zn 2+ion is extremely low. In strongly alkaline soils, Zn-hydroxy anions may form and increase solubility. In soils polluted with high levels of Zn, precipitation of Zn oxide, hydroxide, or hydroxy carbonate may limit Zn solubility at pH 6 or higher. In acidic soils, toxicity of Zn to plants is most likely to occur.
Cadmium Cadmium, which is more soluble than zinc in acidic soils, is rated as having medium to high mobility. The high mobility is attributable to the fact that Cd 2+ adsorbs rather weakly on organic matter, silicate clays and oxides unless the pH is higher than 6. Above pH 7, Cd 2§ can co-precipitate with C a C O 3 o r precipitate as C d C O 3. Therefore, mobility and bioavailability of Cd in neutral to alkaline soils is low. High bioavailability in soils and very high toxicity to animals and human has made Cd the element of greatest concern in considering the value of spreading sewage sludge on land as a soil improvement technique. Results of selective sequential extraction, shown in Figures 17.2(a) to 17.2(b), reveal that Cd and Zn can be very easily leached out from the soil while Pb and Cu are difficult. This is because Cd and Zn are mainly bound to carbonate fraction while Pb and Cu are bound to organic matter and/or
EXTRACTION VIA HYDRAULIC ACID
447
Fe/Mn oxide fractions, which are difficult to leach out than carbonate fraction. The residual fraction of heavy metals is not very high. The experimental results indicate that selective sequential extraction is an important tool for understanding the interaction, retention, and mobility of heavy metals in soils. Therefore, based on mineralogical and chemical analysis of soils, specific chemical reagents can be identified for extracting heavy metals from different phases.
17.5
EXTRACTION VIA HYDROCHLORIC ACID
This method of treatment utilizes hydrochloric acid (HCI) to extract heavy metals from polluted soils. In an experimental program (Mohamed, 1992), illitic silty clay soils were polluted by adding a mixture of Pb, Zn, Cu, and Cd in nitrate forms. The initial concentration of each metal ion was 1000 ppm. Soil suspensions of 1:10 ratio of the polluted soil to distilled water, adjusted to specific pH values by adding HC1, were prepared. The suspensions were centrifuged and filtered to separate the solid from the liquid phase. After each washing, the liquid phase was removed for chemical analysis and a new fresh solution was added to the solid phase. The procedure was repeated for a total of nine washings. The removal efficiency of various heavy metals as a function of number of washings and pH is shown in Figures 17.3 (a) and 17.3 (b).
Figure HC1.
17.3(a). Removal of heavy metals from polluted illitic silty clay soil with pH 3 solution of
The removal efficiency was greatest for Cd followed by Zn, Cu and then Pb. Comparing the pH 3 and pH 4 mixtures, the removal efficiencies of Zn and Cu were considerably high at pH 3 while only a minor difference was observed in the case of Cd and Pb. With each successive washing using
448
SOLVENT EXTRACTION PROCESSES
the pH 3 solution, a relatively steady decrease in the pH of the supernatant was observed. By the ninth washing, the supernatant of the pH 3 mixture had attained the desired value of pH 3. At this point the acid buffering capacity of the soil had broken down and extraction of the heavy metals would be expected to occur with further washings. The same trend was obtained on using the pH 4 solution to wash the soil. The main difference observed, when comparing the results for the two sets of washings, was that the successive reduction in the pH of the supernatant was not as accelerated as when the pH 3 solution was used. The reason for this is that the weaker acid has a less harsh effect on the soil and, therefore, requires a longer period of time to overcome the effects of the buffering capacity of the soil.
Figure 17.3(b). Removal of heavy metals from polluted illitic silty clay soil with pH 4 solution of HC1.
As discussed earlier, Cd was the most effectively removed by both acid solutions, followed by Zn, Cu, and finally Pb. This is precisely the expected order for the success of the extraction process, based on the selective sequential extraction results discussed previously. It can also be seen from Figure 17.2(a) that between pH 3 and pH 4, about 40% of Pb is retained by natural clay soil while, as shown in Figure 17.2(d), less than 10% of Cd is retained. Above pH 4, most of the Pb ions are found in the exchangeable, carbonate, and hydroxide fractions. Below pH 4, Pb ions are retained mostly as exchangeable cations. Therefore, the pH 4 solution would be capable of removing lead from the carbonate and hydroxide phases. The extraction of Cd from illitic silty clay soil is similar except that it occurs at about pH 6. Therefore, for Cd, pH 6 solution would be capable of extracting most of the Cd associated with carbonate and hydroxide forms. In the experiments reported by Tuin and Tels (1988), a Pb removal efficiency of 96% with pH 1 soil washing solution was achieved. This is attributed to the dissolved alumina sheet in the clay structure, as discussed in Chapter 5. Soil washing using 2 M HC1 for 24 hours was successful in
EXTRACTION VIA CHELATING AGENTS
449
removing 91% of Pb, 31% of Zn and 45% of Cu (Benschoten et al., 1994). Also, heavy metal release was found to be greater with HC1 than with HNO3 extraction, possibly due to the formation of soluble metal-chloride complexes (Tuin and Tels, 1990). The main disadvantage of using strong acids as extraction reagents is the adverse effects on the chemical structure of soils, such as mineral dissolution, as discussed in Chapters 5 and 15.
17.6
EXTRACTION VIA CHELATING AGENTS
17.6.1 Concept and Definitions Metals in aqueous solutions interact physically and chemically with water and other components in the solution. A coordinate bond is established when an anion or a polar molecule donates an electron pair to a metal cation. The electron donor is often referred to as a metal complex. The formation of a coordinate bond is known as complexation or coordination. Ligands can vary from the simple to the complex. Ligands such as chloride and ammonia can only donate one pair of electrons to the metal cation and are, therefore, referred to as uni-dentate. More complex molecules or ions, however, contain several atoms capable of donating electron pairs to the metal and are, therefore, referred to as multi-dentate. Ethylene-Diamine-Tetra acetic-Acid (H 4 EDTA) is an example of a multi-dentate containing six donor atoms. When a multi-dentate ligand forms a hetero-atomic ring structure with a metallic cation, the ligand is referred to as a chelating agent, and the ring structure is referred to as a chelate. Likewise, the formation of the ring structure is referred to as chelation. Typically, the more donor atoms involved in the chelation of a metal cation by a single ligand, the higher the stability of the resulting complex (Bell, 1977). Figure 17.4 shows examples of uni-dentate and multi-dentate coordination.
H3N ~~, j
H3N
Pt
H3N
H3N
Uni-dentate tetrammine platinum (11)
H2 i H2
H20~%
Cl
H2 CH 2 I
Co N H2
I
CI
N / H2
CH
2
Multi-dentate dichloro bis (ethylenediamine) cobalt (111)
Figure 17.4. Examples of uni-dentate and multi-dentate coordination.
Since metal cations in aqueous solution are hydrated, coordination with a ligand involves the replacement of water molecules in the metal's coordination shell. With uni-dentate ligands, water
450
SOLVENT EXTRACTION PROCESSES
molecules are replaced in a step by step fashion as follows: m ( H2 0 ) xn ++ L ~ M ( n 2 0 ) x _ l L n ++
n20
M(H20)x_,L "+ + L ~ M(H20)x 2 L2 "+ + H 2 0
[17.1]
[17.2]
With bi-demate ligands, two water molecules are replaced in each coordination reaction step. If the coordination number equals the number of donor atoms in the ligand, all water molecules are removed in a single reaction step (Bell, 1977). Chelation of metals by a particular chelating agent will be hindered by competing reactions such as metal cation hydrolysis and ligand protonation (Bell, 1977). Hydrolysis depicted as: M(n20)x n+~
M(n20)x_ 1
(on)(n-l)++ /--/aq+
M(HzO)x-i (OH) ("-1)+ ~ M(H20)x - 2 (OH)2("-2)++ H aq *
[17.3]
[17.4]
is favoured under alkaline conditions for metal cations with low polarizing power. Protonation of ligands may be depicted as: L + H3 O+ ~ L H *
+ H20
[17.5]
It should be noted that a bi-dentate may be protonated twice, a tri-dentate trice, etc. Other competing reactions include the reaction of the metal cation with other ligands. Equilibrium constants may be defined for chelation and competing reactions. The equilibrium constant K/for the reaction depicted in Eq. [ 17.1 ] may be expressed as: O~MLO~H20
K~ =
[17.6] aMa L
where a M, a L, all20 and aML, are the activities of the hydrated metal cation, ligand, water and metalligand complex, respectively. Since all2o may be considered a constant in an aqueous solution, a thermodynamic stability constant/s may be defined as:
EXTRACTION VIA CHELATING AGENTS
KIT_
Kl
_
all2 o
r
=
aM eL
[ML] YML
451
[17.7]
[M] y M [L] yL
where Yx is the activity coefficient for component X. It should be noted that K 1r is a thermodynamic constant and does not vary with ionic strength. It is related to the standard free energy of change, A G ~, of the reaction: AG o:
_ RT
ln(K~ 7") : AH o_ T A S o
[17.8]
where T is temperature, R is the gas constant, and AH~ and AS~ are, respectively, the enthalpy and entropy changes associated with the reaction. The stability constant K1C, as a function of concentration, may also defined as:
Klr
[ML]
[M] [L]
[17.9]
17.6.2 Classification and Properties of Chelating Agents Chelating agents, by definition, must have at least two donor atoms capable of binding to the same metal cation. A chelating agent with Z donor atoms is therefore capable of forming Z-1 chelating rings. Donor atoms are electronegative and usually come from Group V (e.g., N, P, As, Sb) and Group VI (e.g., O, S, Se, Te) elements. The ability to act as a donor atom tends to fall off as the atomic number increases, and complex stabilities of Group VI elements follow the order P > As > Sb (Dwyer and Mellor, 1964). It should be noted that while Group VII elements (e.g., C1, F1, Br, I ) are quite electronegative, they typically behave either as uni-dentate ligands in mono-nuclear complexes or bridging atoms in poly-nuclear complexes. In chelating agents, Group VII elements do not act as donor atoms (Dwyer and Mellor, 1964). Donor atoms in chelating agents are often part of functional groups that may be classified as either acidic or basic. Acidic groups include carboxylic (- CO2OH ), phenolic ( - OH), sulphonic (SO3H), and phosphoric (- PO(OH)2). Basic groups include amino (- NH2), imino (- NH), carbonyl (-O), and alcohol (-OH) group. It should be noted that the presence of donor atoms is not sufficient to make a molecule a functional chelating agent (Bell, 1977). Donor atoms must be properly located on the structure of the chelating agent, and molecular bonds must be sufficiently flexible to allow adaptation of the chelating agent to the metal cation. In addition, functional groups on the chelating agent must not hinder the metal atom from reaching the chelating site. As noted above, chelating agents may be classified as bi-dentate, tri-dentate, tetra-dentate and so forth. Bi-dentate chelating agents can be classified as those with two basic groups, those with two acidic groups, and those with one basic and one acidic group, as shown in Figure 17.5. Structural characteristics of various bi-dentate chelating agents provide insight into the nature of chelating
452
SOLVENT EXTRACTION PROCESSES
agents and, hence, formation of guidelines for the synthesis of new and more exotic ligands.
H
N
I
2
H2C
I
H2C
~-
N H2 1,10-phenanthroline
Ethylendiamine
with two basic groups
(a) Bi-dentates
Oxalate
OOC-----
COO
Succinate
._ OOC
~
Glutarate
... OOC
H2 --. C ~
Catechol
~
Malonate
OOC.--C
H2
---COO
._
H2 H2 C .-- C ... C O O
H2 C ~
H2 _. C -- COO ~,o
OH Salicyclic
OH
acid
OH
OH
(b) B i - d e n t a t e s
inlllnlnnnnlllnlnnlllllllllllllnnm
Salicylaldehyde
m
~ (c) B i - d e n t a t e s
with two acidic groups mu~mmnnBmmmE~
c
m ~ m ~ l
/H ~o
OH
with one acidic and one basic group
Figure 17.5. Structure of a selected group of bi-dentate chelating agents.
Ethylenediamine and propylenediamine are aliphatic chelating agents with two basic nitrogen groups. It should be noted that propylenediamine is associated with stability constants less than those of ethylenediamine mainly because of its longer chain structure. 1,10-phenanthroline, on the other hand, is associated with stability constants greater than those of ethylenediamine mainly because of its hetero-aromatic structure. It should be noted that methyl group substitution onto 1,10-
EXTRACTION VIA CHELATING AGENTS
453
phenanthroline has been observed to hinder chelation (Bell, 1977). Figure 17.5(a) depicts these chelating agents and Table 17.5 shows stability constants associated with the formation of various chelates.
Table 17.5: log stability constants for various chelates formed by bi-dentate chelating agents in an aqueous medium 1 M with respect to KC1 or KNO~ (Bell, 1977) Cu 2+ Ni 2+ Zn 2+ Mn 2+ Co2+ Ligand Ethylenediamine
2.75
5.94
Propylenediamine 4.13
1,10-phenanthroline
7.25
7.51
10.72
6.39
9.98
8.8
9.25
5.79
6.55
Bi-dentate with two acidic groups, shown in Figure 17.5(b), may be organic or inorganic. Oxalic, malonic, succinic, and glutaric acids are all capable of acting as chelating agents. The stability constants associated with these acids decrease as the number of methylene groups in the acid increase (Bell, 1977). Catechol and salicyclic acid are examples of aromatic bi-dentate with two acidic groups. Inorganic anions which have three or four oxygen atoms also have a chelating ability (e.g., CO32, SO42, PO43",and CRO42).Bi-dentates with one acidic group and one basic group are mostly organic.
H H2C| --..-C| ----'CH21 N
H2
N
H2
N
H2
H . N ---ill 2
H2i - ' - ' H2c
~
N
N/
CH 2
H2 H2 1,2,3-tria m inopropane
N-(2-am inoethyl) ethylenediamine
Figure 17.6. Multi-dentate chelating agents; (a) 1,2,3-triaminopropane; and (b) N- (2-aminoethyl) ethylenediamine.
Multi-dentate chelating agents usually contain N and O as donor atoms, as shown in Figure 17.6 (Bell, 1977). Aliphatic structures include 1,2,3-triaminopropane and N-(2-aminoethyl) ethylenediamine. The stability constants associated with N- (2-aminoethyl) ethylenediamine are greater than those associated with 1,2,3-triaminopropane. The increase in stability is attributed to the donor atom linkage in the aliphatic structure of N- (2-aminoethyl) ethylenediamine which gives it
454
SOLVENT EXTRACTION PROCESSES
increased structural flexibility (Dwyer and Mellor, 1964). While aliphatic structure with more than three nitrogen donors can be synthesized, the resulting basic structure would promote competitive protonation, increase the pH, and promote the formation of metal hydroxides (Bell, 1977). This problem, however, is alleviated by incorporating acidic and basic groups onto the same molecular structure as in the case of sexa-dentate EDTA and other aminopolycarboxylic acids -- two highly effective agents. These agents contain several carboxylalkyl groups bound to one or more nitrogen atoms. Ethylenediaminetetraacetic acid (EDTA) and 1,2-diaminocyclohexanetetraacetic acid (DCTA) belong to the aminopolycarboxylic acids group. The basic structure of EDTA and DCTA are shown in Figure 17.7.
EDTA
H2
O,CH,C
H
HO2CH2C/NH+
C H2
CH2CO2H CH2COa"
DCTA H2C ~
I
H2C~c
H2 C~ ~C H2
H C ~N"
I ~a
H~ ~
CH2CO2" CH2CO2H
+H ~
CH2CO2"
~
CH2CO2 H
H
Figure 17.7. Structure of a select group of aminopolycarboxylic acids: (a) Ethylenediaminetetraacetic acid (EDTA); and (b) 1,2-diaminocyclohexanetetraacetic acid (DCTA).
17.6.3 EDTA Stability Complexes in Solution EDTA is available either as H4Y with a water solubility of 2 g/1 at 22 ~ C or as Na2H2Y.H20 with a water solubility of 108 g/l at 22 ~ C (Bell, 1977). The notation Y refers to the fully deprotonated form of EDTA. In aqueous solution, EDTA undergoes the following reactions:
H4Y ~ H3Y-+ H+;
K4 =
[17.10] [H3Y- ] [H +]
EXTRACTION VIA CHELATING AGENTS
HaY- ~ HzY 2-+ H+;
K3 =
H2 Y2--~ Hy3-+ H+;
K2 =
Hy3- ~ y4-+ H+;
Kl =
455
[H3r-]
[17.11]
[HEr2-] [H+]
[H~r 2-]
[17.12]
[HY 3-] [H+]
[HY3-]
[y4-] [ n +]
[17.13]
Equilibrium constants in a 0.1 ionic strength o f K N O 3 solution at 20 ~ C are" loglo (K4) = 2.07, loglo (K3) = 2.75, log,o (K2) = 6.24, and log~o (K0 = 10.34.
100 o~ tO
= 0
80 -
40
I--
20
a uJ
/ :, .~ H : , Y
|
\/ ~./
Hy3- X ~
,, 9
y4-
60
0.
Or) <
~
0 0
2
4
6
8
10
12
pH Figure 17.8. Speciation of EDTA in aqueous solution as a function of pH (Bell, 1977).
EDTA speciation is pH dependent, as shown in Figure 17.8. As pH increases, the first two protons to be removed from the structure come from the carboxyl groups. The last two protons, however, are removed from the two nitrogens. At pH 3 to 5, EDTA exists mainly as H2Y2- and reacts with metals to form complexes according to:
456
SOLVENT EXTRACTION PROCESSES M "+ + H2 Y2-
~
MY (n-4)+ + 2H +
[17.14]
However, at pH 7 to 10, EDTA exists mainly as H Y 3 and reacts with metals according to" M " * + H Y 3-
~
MY (n-4)+ + H *
[17.15]
The stability constant K ~ for complex MY is defined as:
XM Y -
[MY]
[M][Y]
[17.16]
The stability constants for various EDTA-metal complexes are shown in Table 17.6. These logarithmic equilibrium values reveal a numerical description of the reaction between the metal and the chelating agent. Usually, the larger the value, the greater is the tendency for the chelate to form and, hence, the efficient of the extraction (Wilson et al., 1994). It is evident from this table that EDTA is an effective chelating agent for a wide range of metal cations. However, in comparing these values, the greater ones may not always prove to be the more effective metal-chelate complexes.
Table 17.6: Stability constants for EDTA complexes at an ionic strength of 0.1 KNO 3 at a temperature of 20 ~ C (Bell, 1977 Cation
log,0 (K~,)
Cation
lOglo (Kmr)
Cation
logl0 ( K ~ )
Mg 2+ Ca 2+ Sr 2+ Ba 2+ V 2+ V 3+
8.69 10.96 8.63 7.76 12.70 25.90 18.77 14.04 14.33 25.10 16.50
C o 2+
16.31 18.62 18.80 16.50 18.00 21.80 15.50 15.98 16.40 16.61 16.10
Eu 3+ Gd 3+ Tb 3+ Dy 3§ Ho 3+ Er 3+ Tm 3+ y b 3+ Lu 3+ Sm 3+
17.35 17.37 17.93 18.30 18.74 18.85 19.32 19.51 19.83 17.14
V O 2+
Mn 2+ Fe 2+ Fe 3+ Cd 2+
(a)
Ni 2+ C u 2+ Z n 2+
pb 2+ Hg 2+ La 3+ Ce 3+ pr 3+ Nd 3+ AI 3+
The relative ability of the transition metal ions to form complex ions follows the order: For the divalent cations: Mn 2§ < Fe 2§ < Co 2§ < Ni 2§ < Cu 2§ > Zn 2§
EXTRACTION VIA CHELATING AGENTS (b)
457
For the trivalent cations: Cr 3+ = Mn 3+> Fe 3+ < Co 3+
Copper (II) is the strongest complexing divalent cation. Iron (III) is the weakest complexing trivalent transition metal ion. Iron (III) is, also, stronger than other trivalent cations such as A13+ (Bohn et al., 1979). Chemists generally consider that stability constants less than 109 indicate weak complex ions. This is the range of values for most ion pairs. Competition and high concentrations of more weakly complexing cations can overcome tendencies dictated by high stability constants. This situation, illustrated in Figure 17.9, shows the distribution of EDTA complex ions under representative soil solution conditions. The Fe-EDTA complex predominates in acid solutions because of the great stability of Fe-EDTA complexes. The EDTA ligand prefers Fe (II and III) despite the higher Ca 2+and Mg 2+ concentration in the soil solution. In alkaline soils, however, the high Ca 2+ and Mg 2§ concentrations and the low solubility of iron (III) hydroxide favours the formation of Ca-EDTA and Mg-EDTA complexes. This general picture holds true for many complexing ligand but shifts according to the values of the specific stability constants (Bohn et al., 1979).
Figure 17.9. Mole fraction of EDTA in various complexes versus pH (Norvell, 1974).
This concept of stability constants, represents the affinity of metal ions for EDTA in aqueous solution in the absence of their interfering metal ions. If the solution contains other species reacting with the metal or EDTA, the stability constants, including the free energy of the system can vary considerably from the standard values. This can be modified by the concept of conditional constants, developed by Schwarzenbach (1957). Incorporation of competing reactions into an equilibrium constant may be accomplished using a conditional constant KCCMVdefined as:
458
SOLVENT EXTRACTION PROCESSES
MY
[Mr
Ycc]
[17.17]
where [M co] and [YCr represent all forms of the metal cation and EDTA respectively that are not complexed as MY. The concentrations [M r and [Yet] may be related to concentrations [M] and [Y] respectively using the appropriate Schwarzenbach a s coefficient defined as:
as(Y)- [Y~q [r]
[17.18]
as(M) - [M c]
[17.19]
[M]
For an aqueous system, [W] may be written as: [yq
= [y4-] + [Hy3-] + [H2Y2-] + [H3Y- ] + [H4Y ]
[17.20]
and a,. (Y) as'
as(Y ) : 1 + K,[H+] + K, K2[H+] 2 + KIK2K3[H+] 3 + KIK2K3K4[H+] 4
[17.21]
From the values of K1 to K4 given earlier, a, (Y) can be calculated as a function of solution pH. For example, at pH = 7, a~ (Y) = 1 + 101034 x l 0 "7 + log a,. (Y) = 3.3 7
101658 x l 0 "14 +
101933 xl0 21 + 10 214 xl0 -28 = 2.37 x l 0 3
In a similar fashion, as (M) may be defined for reactions involving the complexation of the metal cation with various ligands in solution. Knowing the appropriate values for the Schwarzenbach a~. coefficients, the conditional constants, U c ~ , can be calculated using: KCCMy _
xM, ~(M) ~(r3
[17.22]
The conditional constants of various competing reactions are shown in Table 17.7 (Ringbom, 1963).
EXTRACTION VIA CHELATING AGENTS Table 17.7: Conditional constants of metal-EDTA complexes.
PH Metal Ag A1 Ba Bi Ca Cd co cu Fe(I1) Fe(II1) WII) La Mg Mn Ni Pb
Sr Th Zn log cl,v(Y)
0
2
3
3.0
5.4
5.3
8.6
10.6
1.0 1.0 3.4
3.8 3.7 6.1 1.5 11.5 9.2 1.7
6.0 5.9 8.3 3.7 13.9 11.1 4.6
11.8 2.2 7.9 7.8 10.2 5.7 14.7 11.3 6.8
3.4 2.4
1.4 6.1 5.2
3.6 8.2 7.4
5.5 10.1 9.4
1.8
5.8 1.1
9.5 3.8
12.4 6.0
21.4
17.4
13.7
10.8
1.4
5.1 3.5
1
8.2 6.5
4 0.7 7.5
6 2.8 10.4 3.0 13.6 5.9 11.7 11.5 14.0 9.5 14.6 11.1 10.6 3.9 9.2 13.8 13.2 3.8 16.7 11.7
7 3.9 8.5 4.4 14.0 7.3 13.1 12.9 15.4 10.9 14.1 10.5 12.0 5.3 10.6 15.2 14.5 5.2 17.4 13.1
8 5.0 6.6 5.5 14.1 8.4 14.2 13.9 16.3 12.0 13.7 9.6 13.1 6.4 11.7 16.3 15.2 6.3 18.2 14.2
9 5.9 4.5 6.4 14.0 9.3 15.0 14.5 16.6 12.8 13.6 8.8 14.0 7.3 12.6 17.1 15.2 7.2 19.1 14.9
10 6.8 2.4 7.3 13.9 10.2 15.5 14.7 16.6 13.2 14.0 8.4 14.6 8.2 13.4 17.4 14.8 8.1 20.0 13.6
11 7.1
12 6.8
13 5.0
14 2.2
7.7 13.3 10.6 14.4 14.0 16.1 12.7 14.3 7.7 14.3 8.5 13.4 16.9 13.9 8.5 20.4 11.0
7.8 12.4 10.7 12.0 12.1 15.7 11.8 14.4 6.8 13.5 8.2 12.6
7.7 11.4 10.4 8.4
7.3 10.4 9.7 4.5
15.6 10.8 14.4 5.8 12.5 7.4 11.6
15.6 9.8 14.4 4.8 11.5
14.5 7.9
5 1.7 9.6 1.3 12.8 4.1 9.9 9.7 12.2 7.7 14.8 11.3 8.8 2.1 7.4 12.0 11.4 2.0 15.8 9.9
10.6 8.6 20.5 8.0
7.6 8.5 20.5 4.7
4.6 8.0 20.5 1.0
8.6
6.6
4.8
3.4
2.3
1.4
0.5
0.1
0
0
0
10.6
460
SOLVENT EXTRACTION PROCESSES
17.6.4 Remediation Via Chelating Agents The importance of chelates to site remediation lies in their ability to remove metal cations from soils and enable the extraction of the ligand-metal complex during solvent extraction treatment process. Large amounts of Pb (80%) from artificially polluted soil were extracted using chelating agents, such as diethylenetriaminepentaacetic acid (DTPA) and EDTA, at specified pH values (Karamanos et al., 1976). EDTA and DTPA form extremely stable complexes with heavy metals. The ability of seventeen different chemical solutions to extract heavy metals (Pb, Zn, Cu, Cd) from artificially polluted clays such as kaolinite, illite and montmorillonite, at either pH 5 or 7, was examined by Farrah and Pickering (1978). Of the reagents used, only EDTA (0.001 M, pH 7) quantitatively released all four metals from the three clays. Oxalic acid (0.1 M, pH 3.3) totally displaced at least three metals from each clay. Other reagents, such as ammonium oxalate (0.1 M), ammonium nitrate (0.01 M), nitric acid (0.1 M) and sodium citrate (0.01 M) effectively displaced one or more heavy metals from individual clays. Also, it was reported (Connick, et al., 1985) that the use of 0.14 M EDTA resulted in a 63, 93, 94, 100, and 86 percent removals of Pb 2§ Zn 2§ Ni 2§ Cd 2§ and Cu 2+, respectively. Evangelista and Zownir (1989) conducted a bench-scale soil washing to explore the feasibility of using EDTA as an extracting agent for soil decontamination. It was found that the binding energies of heavy metals to soil changes with time until equilibrium is reached, thus the results may differ those obtained using aged soils. Also, it was concluded that EDTA is an effective extraction reagent for the removal of lead from coarse fraction (greater than 0.1 mm), removing up to 97%. This is sharp in contrast with control experiments with distilled water which removed no more than 29.8%. In another study conducted by Esposito et al., (1989), it was found that the optimum concentration is 3:1 molar ratio of tetrasodium EDTA to total pollutant metals present in the soil. It was found that the efficiency of removal varied according the soil particle size, as shown in Table 17.7.
Table 17.7: EDTA removal effectiveness as a function of particle size. Particle size diameter
Removal effectiveness (%)
>2mm 250 lam- 2 mm < 250 lam
95.9 - 98.1 85.1 - 98.4 21.9 - 82.2
Sands and silty clays artificially polluted by Pb were investigated for the mobility of Pb in the presence of EDTA (Peters and Shem, 1992). For independent pHs between 4 and 12, Pb removal ranged from 58 to 64% for EDTA concentrations of 0.01 to 0.1 M. Also, Mohamed and Trasente (1996) used Na-EDTA at 0.1 M concentration and variable pH solutions from 2 to 12 to evaluate the removal efficiency of Pb, Cu, Zn, Cd and Ni from naturally polluted soils. The results indicated that after 3 washings, at pH 4, the maximum percent removals were Pb (100%), Cu (89%), Cd (60%) and Ni (54%). The low removal of heavy metals was attributed to the competition of Ca and Mg for EDTA complexation (Brown and Elliott, 1992; Ringbom, 1963). EDTA and citric acid were used
EXTRACTION VIA CHELATING AGENTS
461
to extract heavy metals such as Cu, Zn, Cd, Ni, Pb and Cr from polluted soils at neutral pH (Leidmann et al., 1994). EDTA was able to mobilize up to 90% of Cd, Cu, Pb and Zn, but only 2545% of Ni. Citric acid (0.26 M) was found to extract 50 - 80% of Cu, Cd, Zn and Ni whereas only 30 - 40% of Cr and Pb were extracted at neutral pH. The effect of pH and EDTA concentrations on Pb removal from polluted illitic soil is shown in Figure 17.10 (Castellan, 1996). It should be noted that the initial Pb concentration in the polluted samples was 5000 mg Pb/kg soil. The 10.5 M EDTA solution extracted approximately 34%, at a pH value of 4, of the total concentration of Pb in the soil. As the pH of the soil solution increased above 4, the concentration of Pb removed from the soil decreased sharply, and attained 0% removal at pH values above 5.5. The 10-4 M EDTA solution was slightly more effective in removing Pb from the polluted soil. A maximum Pb removal of 36% of the total Pb in the soil was attained at pH 4. As the pH of the solution became more alkaline, the removals were less pronounced. For pH values above 7.5, the 10-4 M EDTA solution was not effective at removing Pb from the polluted soil. Similar Pb removals were also achieved for the 10.3 M EDTA solution. Extraction tests with 10.2 M EDTA solution showed significant Pb removals for a wide pH range. As shown in Figure 17.10, at pH 4.7, 44% of the total Pb in the polluted soil was removed. The Pb removals decreased as the soil pH increased and reached a minimum of 36% at pH 8. The 10-1 M EDTA solution removed 96% of the total Pb in the polluted soil. The maximum Pb removal occurred at pH 4, although the pH did not seem to have a significant effect on the removal efficiencies. Previous studies by Allen and Chen (1993) have also shown that the solution pH did not greatly influence the extraction efficiencies at EDTA high concentration.
Figure 17.10. Effect of pH and EDTA concentrations on Pb removal from polluted illitic soil.
The maximum removals, at pH 4, of Pb, Cu, Zn and Cd versus the concentration of EDTA for illitic soil polluted with single-specie heavy metals are shown in Figure 17.11 (Mohamed and
462
SOLVENT EXTRACTION PROCESSES
Castellan, 1997a). The concentration of each single-specie was 5000 ppm. The 101 M EDTA solution was capable of removing Pb (96%), Cu (96%), Zn (100%) and Cd (93%), in the indicated amounts, from the polluted soil. With 10.5 M EDTA solution, the results were Pb (35%), Cu (36%), Zn (37.5%), and Cd (35%). It seems that, at low concentration, EDTA was capable of removing bound soil pollutants, which suggests that a low concentration chelating agent has the potential to decontaminate polluted soils. However, with more dilute chelate solutions, the soil pH becomes increasingly important if significant removals are to be obtained.
Figure 17.11. Maximum removals, at pH 4, of Pb, Cu, Zn and Cd versus the concentration of EDTA for illitic soil polluted with single-specie heavy metals.
The results of the EDTA extraction tests performed on multi-species heavy metal polluted illitic soil are shown in Figure 17.12 (Mohamed and Castellan, 1997b). The Figure depicts the maximum removals of heavy metals at pH 4 as a function of EDTA concentration. Each singlespecie in the mixture has a concentration of 5000 ppm. The obtained removals with different EDTA concentrations follow the order: (1) 10 .5 M: Cd > Cu = Zn > Pb (2) 10 .4 M: Cu > Cd > Zn > Pb (3) 10 .3 M: Cd > Cu = Zn > Pb (4) 10 2 M: Cu > Cd > Pb > Zn (5) 10 l M: Pb > Cu = Cd > Zn Based on the stability constants shown in Tables 17.6 and 17.7, the order would be Cu > Pb > Zn - Cd, which is not consistent with the experimentally obtained order at any concentration. This is not surprising since the data shown in Tables 17.6 and 17.7 did not account for the role of the reactive clay surfaces and concentration of competing cations.
EXTRACTION VIA ORGANIC ACIDS
463
Figure 17.12. The maximum removals of Pb, Cu, Zn and Cd at pH 4 as a function of EDTA concentrations for illitic soil polluted with multi-species heavy metals.
17.7
EXTRACTION VIA ORGANIC ACIDS
Low molecular weight (< 500) organic acids (bi-dentate chelating agents) such as oxalic, tartaric, citric, acetic, fumaric and succinic acids are present in natural soils in top layer near the ground surface and can form complexes with metals and, hence, affect the metal mobility and speciation (Fox and Comerford, 1990). High concentrations of oxalic acid and decreasing amounts of formic, citric and acetic acids were found in the soils. Organic acids, such as citric and oxalic acids, form more stable complexes with metals than do formic and lactic acids. Biological compounds in organic molecules such as glycine, citric acid, tartaric acid and gluconic acid have chelating properties and can react with heavy metals (Lo et al., 1992). When heavy metals form complexes with the soil organic matter (humic acid), they form stable chelates and lower the mobility of metals. On the contrary, metal mobility and bioavailability increase when they combine with organic acids (e.g., acetic acid). Several kinds of organic acids, namely formic, acetic, malonic, lactic acid and citric acids were used to extract rare earth metals from soils (Matsuyama et al., 1996). Acetic acid is often used to replace the more strongly adsorbed cations of heavy metals as well as to extract the available Pb fraction from soils. This reagent can remove slightly higher amounts of Pb than exchange reagents such as ammonium acetate (Karamanos et al., 1976). The main advantages of using organic acids to extract heavy metals from polluted soils are: (1) biodegradability, (2) less destructive to soil structure than mineral acids, (3) more specific to heavy metals and less likely to leach soil nutrients, and (4) low cost.
464 17.8
SOLVENT EXTRACTION PROCESSES S U M M A R Y AND C O N C L U D I N G R E M A R K S
As discussed, metal ions are held to soil in different forms (i.e., exchangeable, carbonates, metal oxides and organic matter associations, and residual fraction) with variable bonding strengths. Thus, although they are not by any means all irreversibly bonded, they are very difficult to extract from the soil completely without a selective sequential extraction procedure. Release of these ions is usually better at lower pH's in the presence of organic acids or chelating agents capable of converting metal ions into anionic complexes. If a metal is strongly adsorbed to manganese and iron oxides in soils, chelating agents or acids such as hydroxylamine, or sodium dithionate/citrate, can be used. When precipitated hydrous oxides are present, a more concentrated EDTA solution will be appropriate. It is noteworthy that not all sorbed metal ions are ever extracted.
CHAPTER EIGHTEEN
SURFACTANT EXTRACTION PROCESSES
18.1
INTRODUCTION
Surfactants are surface active agents which are used to reduce interfacial tension and increase solubility of non-aqueous phase liquids. Surfactants have been used in the petroleum industry for a number of years to enhance the secondary extraction of oil. The oil industry is usually able to recover, in the first instance, between 20 and 60% of the available oil by pumping. The remaining portion is tightly held by capillary forces and requires a slightly more complex procedure, such as surfactant injection, to mobilize it. A logical extension of this technique may be employed for the treatment of polluted soils. One significant difference between surfactant use in secondary oil recovery and in treatment of polluted soils is that, in the former, there is rarely any requirement to prevent pollution by the surfactant. However, in treatment of polluted soils where the purpose is to prevent any degradation of the environment, the selection of a suitable surfactant must take into account its toxicity. Surfactant extraction technology, as applied to site remediation, relied on the use of surfactants to increase the solubility of hydrophobic organic pollutants and, hence, increase the removal efficiency on application of the washing fluid. Studies have shown that soils polluted with fuel oil, jet fuel, and waste oil from underground storage leak can be effectively treated by the surfactant extraction process. Removal efficiencies of 90 to 98% for hydrocarbons have been achieved when heat and surfactants were added to the wash water. The process can be used to remove volatile organic compounds and other organic pollutants having relatively high vapour pressure and low solubility. Removals of 90 to 99% or more of volatile organic compounds can be achieved by simple washing. Removal rates for semi-volatile organic pollutants tend to be lower, on the order of 40 to 90%, and the addition of surfactants to the wash water is often required to accelerate the treatment process. The surfactant extraction process is appropriate for treating soils that contain at least 50% sand and gravel but is relatively ineffective in treating soils that are rich in clay and silt-sized particles. Soils with high cation exchange capacity and soil organic matter content tend to interact with organic chemicals, which can limit the ability of the surfactant extraction process to effectively remove pollutants from soils. In surfactant-based in-situ soil washing, the surfactant, organic pollutant and soil system (solid and gas phases) are the three basic interacting components. Each component interacts with the others in the manner illustrated in Figure 18.1. The interaction mechanism between various components are listed along the line connecting the specific components. These interaction mechanisms can be physical, chemical or biological in nature. A successful application of the surfactant extraction technique is highly dependent on a proper understanding of the various interaction mechanisms between the soil-organic pollutant-surfactant systems. This chapter is 465
466
SURFACTANT EXTRACTION PROCESSES
concerned with the principles of surfactant-based extraction processes.
Figure 18.1. Processes involved in partitioning of non-aqueous phase liquid (NAPL) and surfactants into solid, liquid and gaseous phases.
18.2
SURFACTANT PROPERTIES
18.2.1 Definition and Types The term surfactants comes from the descriptive phrase surface active agents. Surfactant molecules have two distinct regions (moieties): hydrophobic (water disliking) and hydrophilic (water liking), as shown in Figure 18.2. Thus, surfactant molecules migrate to interfaces where both portions of the molecule can be in their preferred phase. For example, a surfactant will accumulate at an oil-water interface with its hydrophobic moiety (lipophilic tail) in the oil phase and its hydrophilic moiety (polar or ionic head) in the water phase. Both moieties of the molecule are then in their preferred phase, and the free energy of the system is minimum. Accumulation of surfactants at interfaces alters the nature of the interface, resulting in the designation of these molecules as
SURFACTANT PROPERTIES
467
surface active agents.
A
Cationic
4 { ....
Anionic
4{---~
L_
(1) r L_
Amphoteric -~---/
r r 0
Nonionic
Hydrophobic moiety
(~ r r
- ~ ....
// Hydrophilic ....... moiety
Figure 18.2. Schematic representation of surfactant molecules.
As discussed previously, one end of a surfactant molecule is polar (or ionic) and, hence, water soluble (hydrophilic). The other end is non polar (organic soluble) or oleophilic, thus promoting aqueous solubilization of compounds of low water solubility. At room temperature and atmospheric pressure, many organic compounds are insoluble in water, and form a phase system with a large interfacial tension. If a surfactant is added, the interfacial tension is reduced because of the ability of a surfactant to interact with both the organic and water phases: the hydrophobe (tail portion) will interact with the organic phase while the hydrophile (the head part) will interact with water. Surfactants are typically classified, according to the nature of their head group, as cationic, anionic, nonionic and amphoteric (both cationic and anionic groups), as illustrated in Figure 18.2. (1) Anionic surfactants normally have a sulphonate-sulphate or phosphate functional group as their solubilizing group. Aqueous anionic surfactant has a negative charge which enhances the dispersion process. They often form unstable emulsion. (2) Cationic surfactants are small groups of softening and coating agents with a positively charged polar solubilizing group -- usually an amino or a quaternary nitrogen functional group. Commercially available cationic surfactants are either sulphonate-based (biodegradable) or sulphate-based (non-biodegradable). (3) Nonionic surfactants do not have a charge on their solubilizing group. Nonionic surfactants normally contain a polyoxyethylene group which acts as the solubilizing group. Two commercially-available nonionic surfactants are ethoxylated alcohol (biodegradable) and ethoxylated nonylphenol (non-biodegradable). Nonionic are the most commonly used
468
SURFACTANT EXTRACTION PROCESSES surfactant in environmental applications.
Surfactants are also characterized by their hydrophile-lipophile balance (HLB) number, which is a measure of dispersion and emulsification performance. Surfactants with high HLB value are hydrophilic while surfactants with low HLB values are lipophilic. The physico-chemical properties that are of importance to the interaction between surfactants, NAPL, and soils are listed in Table 18.1, and are discussed below.
Table 18.1: Physico-chemical properties of importance to surfactant-NAPL-solid interactions PHYSICO-CHEMICAL PROPERTIES SOLID 9 Temperature 9 pH
9 Mineralogy 9 Organic content 9 Amorphous content 9 Specific surface area 9 Cation exchange capacity 9 Particle size 9 Porosity 9 Hydraulic conductivity 9 Pressure distribution 9 Moisture content
NAPL 9 Temperature "pH 9 Aqueous solubility 9 Octanol partitioning 9 Polarity 9 Dielectric constant 9 Molecular weight 9 Viscosity 9 Density 9 Biodegradability 9 Volatilization
SURFACTANT 9 Cationic, anionic, nonionic 9 Solubilization 9 Detergency 9 Interfacial tension 9 Surface tension 9 Biodegradability
18.2.2 S u r f a c e T e n s i o n
Surface tension reduction is a measure of surface wettability, which is defined as the tendency of one fluid to spread on or adhere to a solid surface in the presence of other immiscible fluids. It is a major factor controlling the location, flow and distribution of fluids in soils. In a porous medium that has definite preference for either of the fluid phases, the wetting phase has a tendency to occupy the smallest pores and to contact the substrate while the non wetting phase is consigned to the centre of the larger pores. Organic pollutants adsorbed to soil surfaces are present in thin molecular layers rather than in thick discrete phases as found in oil-bearing strata. Very low surface tension, in the order of 1 dyne/cm or lower, is sufficient to cause spontaneous oil flow. The higher the ability of a surfactant to minimize surface tension the better its efficiency. The surface tension of a surfactant solution is determined by its concentration in water which, in turn, determines the interfacial tension between the aqueous solution (water and surfactant) and the non-aqueous phase liquid. Interfacial tension, which arises from an imbalance of molecular forces, characterizes the contact between two phases. Figure 18.3 shows the variation of surface tension with concentration for various commercial surfactants. The surface tension decreases as the
SURFACTANT PROPERTIES
469
concentration of the surfactant increases until a critical concentration is reached, after which the surface tension does not decrease. The concentration at which that occurs is referred as the critical micelle concentration of the surfactant.
Figure 18.3. Variations of surface tension with concentration for different commercial surfactants.
Figure 18.4. Schematic showing the arrangement of surfactant molecules at critical micelle concentration.
470
SURFACTANT EXTRACTION PROCESSES
18.2.3 Critical Micelle Concentration
The critical micelle concentration (CMC) delineates the point at which the surfactant molecules become arranged into clusters or aggregates (also known as micelles). Micelles have hydrophobic interiors and hydrophilic exteriors, as shown in Figure 18.4. The hydrophobic interiors of the micelle act as a sink for non-aqueous phase liquids. Micelles have organized, dynamic chemical structures with such geometrical shapes as spheres, oblate spheroids, and prolate spheroids (Edwards et al., 1991). CMC depends on surfactant type, temperature, and electrolyte concentrations (Rosen, 1989). Nonionic surfactants, for example, have lower CMCs than ionic surfactants. Micelle formation distinguishes surfactants from amphophilic molecules (e.g., alcohols) that exhibit a much lower degree of surface activity and do not form micelles. The average number of surfactant molecules within a micelle is known as the aggregation number. The shaded area in Figure 18.5 (Rosen, 1978) is the CMC band, which is the optimum concentration at which the surfactant would displace non-aqueous phase liquids.
Figure 18.5. Changes in some physico-chemical properties of a surfactant aqueous solution in the neighbourhood of the critical micelle concentration.
18.2.4 Solubilization and Detergency
The increased aqueous solubility of organic compounds beyond CMC surfactant concentrations is referred to as solubilization. As the surfactant concentration increases, additional micelles are formed and the pollutant solubility continues to increase as shown in Figure 18.5. The solubilization effect is the spontaneous dissolving of the pollutants within the surfactant micelles and results in a stable colloidal solution with the lowest possible sum of free energies of its component parts. Solubilization can increase the effective solubility of the solute in water by two to three orders of magnitude (Wayt and Wilson, 1989). The increase in solubility will be essentially linear once the
SURFACTANT PROPERTIES
471
CMC has been reached and will substantially enhance the removal of organic pollutants by the passage of water. Highly hydrophobic chemicals such as DDT, however, undergo some increase in apparent aqueous solubility as a result of solubilization into individual surfactant molecules (Edwards et al., 1991). Various parameters that affect the micellization process will be discussed latter. Micelles are also significant in detergent processes. The detergency of a surfactant can be thought of as the cleaning power since it is a measure of the chemicals ability to remove pollutants from solid surfaces and subsequently keep them suspended in the aqueous media. This could also be thought of as the ability of the surfactant to disperse the pollutant. As the surfactant concentration increases the detergency effect increases until CMC surfactant concentration is reached, as shown in Figure 18.5.
18.2.5 Dispersion In addition to its ability to cause the displacement of organic pollutants, a surfactant can also undergo hydrolysis and form flocs, may create viscous emulsions with petroleum products, and can cause dispersion of soil colloids. All of these processes are likely to result in a reduction in the hydraulic conductivity of the soil. This is obviously undesirable for groundwater extraction since the rate of pumping will be closely controlled by the hydraulic conductivity of the soil. The amount of dispersion of soil colloids that a surfactant causes has been found to be closely linked to surfactant concentration, with minimal dispersion occurring below the CMC. The ability of surfactants to mobilise hydrophobic organic compounds, such as oils and PCB's, is due to their tendency to displace these materials from the soil by detergent action and by enhancing the solubility of the pollutants in aqueous solutions. The likelihood that oil will be displaced by the flow of groundwater is controlled by the ratio of the viscous pressure drop across the micelles to the capillary pressure across the oil-water interface, as given by Eq. [ 18.1 ] (Sitar et al., 1987):
ktwVLhDt > 4
[18.1]
where ktw is dynamic viscosity of the fluid, v is Darcy's velocity, L h is the micelle length in the direction of flow, D, is the throat radius of pore space between soil aggregates, o is the interfacial tension, and kw is the soil hydraulic conductivity.
18.2.6 Micelle Formation As discussed previously, micelle formation is controlled by such parameters as temperature, counter ions, pH, and type and concentration of the organic chemicals involved. These parameters are discussed below.
Temperature In case of ionic surfactants, it is often observed that the solubility of non-aqueous organic chemicals will undergo a sharp, discontinuous increase at some characteristic temperature,
472
SURFACTANT EXTRACTION PROCESSES
commonly referred to the K r a f f i t e m p e r a t u r e , Tk. Below that temperature, the solubility of the surfactant is determined by the crystal lattice energy and heat of hydration of the system. Above T~ the solubility of the surfactant monomer increases to the point at which micelle formation begins and aggregated species become the thermodynamically favoured form. A schematic representation of the temperature-solubility relationship for ionic surfactants is shown in Figure 18.6 (Myers, 1988). It should be noted that nonionic surfactants do not exhibit a Krafft temperature. As temperature increases, the solubility of nonionic surfactants decreases. In some cases, phase separation will occur, producing a cloudy suspension of surfactant. The temperature at which that occurs is referred as the c l o u d p o i n t of the surfactant.
Total surfactant Micelles r
CMC
0
. u
.
L_
\~lk
/ / / -~
Monomers
-I--I
t(!) o r
Monomer solubility curve
o
o " Krafft temperature
Temperature
Figure 18.6. Solubility-temperature relationship for typical ionic surfactants.
Counter Ions
The extent of ion pairing in a system increases as the polarizability and valence of the counter ion increase. However, greater ion separation results when the hydration radius is large. For a given hydrophobic tail and anionic head group, the CMC decreases in the following order: +
+
Li + > N a + > K + > Cs + > N ( C H 3 ) 4 > N ( C H 2 C H 3 ) 4 > Ca 2+ = m g 2+
[18.2]
In the case of cationic surfactants such as dodecyl trimethyl ammonium halides, the CMC's are found to decrease in the following order (Mukerjee, 1967)" F- > Cl- > Br->
I-
[18.3]
SURFACTANT PROPERTIES
473
As the counter ion is changed from monovalent to di- and trivalent, the CMC decreases rapidly. The CMC's of various salts of dodecyl-sulfate are shown in Table 18.2 (Lange and Schwuger, 1980).
Table 18.2: CMC's of various metal ions of dodecyl-sulfate Counter Ion
Temperature (~
Li + Li § Na + Na + K§ C~+ 1/2 Ca 2+ 1/2 Mg 2+ 1/2 Zn 2+
CMC (mm)
25 4O 25 40 40 40 54 25 60
8.8 10.5 8.1 8.9 7.8 6.9 2.6 1.6 2.1
Electrolytes
In aqueous solution, the presence of an electrolyte causes a decrease in the CMC of most surfactants. The greatest effect is obtained for ionic surfactants. Nonionic surfactants exhibit a much smaller effect. The effectiveness of a given ion at altering the micellization process can be qualitatively related to the radius of hydration of the added ions, with the contribution of the cations and anions being approximately additive. In general, the smaller the radius of hydration of the ion, the greater is its effect on the CMC. The order of effectiveness of anions at decreasing the CMC has been given by Ray (1971) as follows: 1 S O 2- > F - > B r O 3 > C l - > B r - > N O 3 > I 2
[18.4]
For cations, the order is N H 4 > K + > N a ~ > Li + > _ 1 Ca z+ 2
[18.5]
pH
Since the carboxylate group is not fully ionized near or below the acid dissociation constant (PKb), the electrostatic interactions between head groups will vary with the solution pH. This, in turn, will influence the formation of micelles. A similar result will be observed for the cationic alkylammonium salts near and above the base dissociation constant (pKb), resulting in a decrease in the CMC. When the surfactant is in ionized form, excess acid or base will act as a neutral electrolyte. It is expected that pH will have no effect on the CMC of nonionic surfactants. At low pH, a material containing carboxyl and amine groups would act as a cationic surfactant while at high pH it would
474
SURFACTANT EXTRACTION PROCESSES
act as anionic surfactant. If the cation is a quaternary ammonium salt, no pH sensitivity would be expected, as would be the case for a strong acid anionic group.
Organic Chemicals Organic chemicals that have low water solubility can be solubilized in micelles to produce systems with substantial organic content where no solubility would occur in the absence of a surfactant. In the process of solubilization of water immiscible organic in micelles, the size of the aggregate and the curvature of its surface can change significantly. This, in turn, will affect the energy of interaction between head groups at the micelle surface and the hydrophobic interactions among the hydrocarbon tails. The combined effects of the presence of the solubilized material is usually to produce a slight decrease in the measured CMC of the system (Shinoda et al., 1961). Organic additives with substantial water miscibility such as the lower alcohols, dioxane, acetone, glycol, and tetrahydrofuran would not be expected to partition into the interior of the micelle when present in small amounts. The effect of these additives on the CMC, therefore, would be expected to be relatively minor (Harkins et al., 1949). As the alkyl groups go beyond carbon two (C2), the inherent surface activity of the alcohol may become significant. The energy requirements for bringing the hydrophobic tail into solution may decrease, leading to an increase in the CMC. However, the added organic material will result in a reduction in the dielectric constant of the solvent mixture. Such an effect would tend to decrease the CMC of ionic surfactants due to their lower solubility and the reduced repulsion between adjacent head groups at the micellar surface. The net effect on the CMC will, therefore, depend on the relative magnitude of the two opposing trends. In any case, the effects are usually relatively minor until a substantial additive concentration is reached.
18.3
ORGANIC POLLUTANT PROPERTIES
18.3.1 Solubility The solubility of a solvent is the amount of solute that must be added to a given quantity of the solvent, at a specific temperature, to obtain a saturated solution. The solubility product, K~z of a saturated solution represents the maximum possible value of the product of the ion concentrations, for a given set of conditions. Solubility equilibria are useful in predicting whether a precipitate will form under specified conditions, and in choosing conditions under which two chemical substances in solution can be separated by selective precipitation. Substances which are more soluble are more likely to adsorb by soils and less likely to vaporize from water. On the one hand, substances with no hydrogen bonding groups or little polar character, such as hydrocarbons or halogenated hydrocarbons, usually have very low solubilities compared to compounds such as alcohols which are capable of interacting with water. Small molecules are apparently more readily accepted into the water structure, and display higher solubility than large molecules (Benefield et al., 1982).
18.3.2 Hydrolysis Hydrolysis is the breakdown of minerals under the influence of H § and OH ions in water. The hydrolysis of soil solution is controlled by pH, type and concentration of cations, oxidation state of the cations and redox environment, and temperature. Hydrolysis proceeds effectively at high
ORGANIC POLLUTANT PROPERTIES
475
temperature, low organic content, low pH values, and low redox potentials. Increasing hydrogen ion concentration in the presence of free acids speeds chemical disintegration. Also, most highly charged cations are strongly hydrolysed in aqueous solution, and have low pK values (pK value is the logarithm of the reciprocal of the dissociation constant (k) of an electrolyte). The pK value is considered to be a significant characteristic in so far as it affects the equilibrium between most chemical species and controls the degree of ionization and solubility of compounds. The smaller the pK value, the higher the degree of ionic dissociation and, hence, the more soluble is the substance. It also influences the transport and adsorption of chemical species in soils. The pK value can also be used to indicate the strength of acids and bases. Weak acids do not favour the formation of H § ions because they have high pH values.
Table 18.3: Listed half-life of various organic compounds Group/Compound
Half-Life (days)
9Pesticides DDT 81 DDT 4400 Dieldrin 3800 Endosulfan/Endosulfan sulfate 21 Heptachlor 1 9Halogenated Aliphatic Hydrocarbons Chloroethane (ehtyl chloride) 38 1,2-Dichloropropane 180-700 1,3-Dichloropropene 60 Hexachlorocyclopentadiene 14 Bromomethane (methyl bromide) 20 Bromodichloromethane 5000 9Halogenated Ethers bis (chloromethyl) ether <1 2-Chloroethyl vinyl ether 1800 9Monocyclic Aromatics Pentachlorophenol 200 9Phthalate Esters Dimethyl phthalate 1200 Diethyl phthalate 3700 Dinbutyl phthalate 7600 Sources: Callahan et al., (1979); Mills et al., (1985); Ellington et al. (1988).
@ pH 9 @ pH 3-5 @ pH 7 @ pH 7 @ pH 7 @ pH @ pH @ pH @ pH @ pH @ pH
7 7 7 7 7 7
@ pH 7 @ pH 7 @ pH 7 @ pH 7 @ pH 7 @ pH 7
Hydrolysis rates are commonly reported in terms of half-life (i.e., the number of days or years for half of the original concentration of the substance to be hydrolysed). Predicted half-lives of various hazardous organic compotmds range from days to years, as shown in Table 18.3. Such factors as pH, temperature, and the presence of other ions affect the rate of hydrolysis of organic
476
SURFACTANT EXTRACTION PROCESSES
compounds. Many classes of organic compounds hydrolyse in aqueous solutions. Some classes are, however, resistant. Table 18.4 summarizes the organic functional groups that are potentially susceptible to hydrolysis and those which are generally resistant.
Table 18.4: Amenability of organic functional ~roups to hydrolysis Potentially susceptible
Generally resistant
9 Alkyl halides 9 Amides 9 Amines 9 Carbarnates 9 Carboxylic acid esters 9 Epoxides 9 Nitriles 9 Phosphonic acid esters 9 Phosphoric acid esters 9 Sulfonic acid esters 9 Sulfuric acid esters
9 Alkanes 9 Alkenes 9 Alkynes 9 Benzenes/Biphenyls 9 Polycyclic aromatic hydrocarbons 9 Heterocyclic polycyclic aromatic hydrocarbons 9 Halogenated aromatic/PCB 9 Dieldrin/Aldrin and related halogenated hydrocarbons 9 Pesticides 9 Aromatic nitro compounds 9 Aromatic amines 9 Alcohols 9 Phenols 9 Glycols 9 Ethers 9 Aldehydes 9 Ketones 9 Carboxylic acids 9 Sulfonic acids
Source: Guswa et al., (1984); and Harris (1982).
18.3.3 Vapour Pressure The vapour pressure of a liquid or solid is the pressure of the gas phase at which equilibrium is achieved with the liquid or solid phase at a given temperature. A liquid with a high vapour pressure, such as gasoline, is said to be volatile, and will evaporate quickly. The transport of a compound from the liquid to the vapour phase is called volatilization. Evaporation depends on the equilibrium vapour pressures, dispersion of emulsions, solubility and temperature. Chemicals with relatively low vapour pressures and high water solubility are less likely to vaporize (Scott and Smith, 1980). Volatilization potential has been considered a significant factor in the disposal of compounds with vapour pressure greater than 10 .3 mm Hg at room temperature (Fuller and Warrick, 1985).
SOIL SYSTEM PROPERTIES
477
18.3.4 Chemical Alternation
Chemical alteration of an organic compound in the environment could arise from one or more of the following reactions: (1) redox behaviour, (2) hydrolysis, and (3) photochemical breakdown. The extent to which an organic compound breaks down to simple molecules will determine its persistence and toxicity. Studies have shown that some of these transformation processes can convert a compound into a derivative that may be substantially more hazardous and persistent. Examples are the photochemical degradation of hydrocarbons and nitrogen oxides to produce a smog that has a more direct and active effect on the environment and its inhabitants. Thus, chemical alterations may magnify the environmental effects of organic compounds.
18.4
SOIL SYSTEM PROPERTIES
18.4.1 Activity
Adsorbed organic molecules may be oxidized by iron which exist in the solid part of the soil. Organic molecules may, also, interact with the absorbed species by clay surfaces (e.g., water of hydration or enzymes). The intrinsic properties of the solid surfaces, such as its surface charge density or cation exchange capacity, CEC, will affect both adsorption and surface transformations, as discussed in Chapters 4, 5, and 6. For the many transformations which require an interaction between an exchangeable cation and its hydration water, the CEC will dictate the catalytic capacity of the solid surfaces. The surface charge density may affect surface interactions in a number of ways. For example, a higher surface charge density means a higher surface concentration of exchangeable counter ions which, in turn, implies a lower probability of direct interaction of nonionic species with the charged surface. A surface property which is of great importance for many adsorption-enhanced trans~ormatmns is surjace aciaiiy. 1 D e suriace aClolty of clay minerals was implicated in the surface catalysed degradation of several pesticides (e.g., Brown and White, 1969). The capacity of clays to catalyse ethylacetate and sucrose was enhanced by the presence of exchangeable protons (McAuliffe and Coleman, 1955). Surface acidity is influenced by the presence of exposed structural acidic groups at the surface, strength of the acidic group, nature of the exchangeable cation, moisture content, and temperature. Fripiat (1968) discussed the increase in the acidity of water brought about by the high electric field at clay mineral surfaces and the related enhancement of surface reactions which involve protons. For the catalysed polymerization of styrene on a clay surface, the surface reactions which are initiated by the formation of carbonium ion may involve a proton transfer from sites associated with tetrahedrally-coordinated aluminum (Bittles et a1.,1964). For a given type of flow through soil, the fluid/solid interactions play an important role in the flow process. The types of interactions are, generally, functions of properties of permeant and its composition, and amount and ratios of active clay minerals. It is well established that these parameters, among others, greatly influence the flow process. As a result of surfactant fluid injection, the non-aqueous phase liquid, NAPL, and clay particles are dispersed into the surfactant phase. The initialization of the NAPL detaching process, as well as the migration of the dispersed colloidal particles, are due to several forces, namely: (1) shear forces acting on the interface between the carrier fluid and the NAPL, and (2) forces arising
478
SURFACTANT EXTRACTION PROCESSES
from the different types of interactions between the NAPL and the substrate materials. These forces are greatly influenced by the type and the nature of the substrate surface activity (Mohamed and Yong, 1992).
Figure 18.7. Percentage of non-aqueous phase liquid (NAPL) removal as a function of montmorillonite content and hydraulic gradient after an injection period of 80 minutes.
The percentage of NAPL uptake depends strongly on the percentage of active clay minerals in the sample. This can be explained from the results reported by E1 Monayeri (1983); Yong et al. (1991a and b); and Mohamed and Yong (1992), shown in Figures 18.7 to 18.9. In these experiments, the montmorillonite to sand ratio was changed from 5 to 25% by weight. Specimens were initially saturated with bunker oil (NAPL) and packed into lucite cylinders; then surfactant solution consisting of emulgin 05 and emulgin 010 at a ratio of 9:1 was allowed to flow through the compacted specimens under hydraulic gradients varying from 0.2 to 2. It can be seen from these figures that the percentage of NAPL removed, or the tendency of NAPL to move, increased as the initial montmorillonite content decreased. This was observed for all tested samples under different hydraulic gradients. This may be attributed, on one hand, to the fact that increasing the montmorillonite content decreased the injected fluid velocity, lowered the momentum transfer to the stagnant NAPL and, hence, lowered the removed NAPL portion. On the other hand, increasing the montmorillonite content increased the adsorption of NAPL by the solid phase, in turn resulting in lower escaping tendency of NAPL. The escaping tendency of hydrocarbons is related to their solubility in the liquid phase, and their tendency to be adsorbed by the solid phase (Antonov, 1953). The movement of NAPL, in the vicinity of clay particles, occurs without any movement of the latter (Hunt, 1979). It is evident from Figures 18.7 to 18.9 that the flow mechanism is influenced by substrate composition. As clay content increases, the NAPL removal is decreased. The increase of the
SOIL SYSTEM PROPERTIES
479
clay/NAPL ratio resulted in a high mass-to-surface area ratio, increased diffusion path length and, hence, a lower rate of NAPL removal.
Figure 18.8. Percentage of NAPL removal as a function of montmorillonite content and hydraulic gradient after an injection period of 160 minutes.
Figure 18.9. Percentage of NAPL removal as a function of montmorillonite content and hydraulic gradient after an injection period of 250 minutes.
480
SURFACTANT EXTRACTION PROCESSES
18.4.2 Porosity Soil porosity, grain size and pore distributions are among the most important factors governing the flow process through soil. Marczal (1968) pointed out that if a mass contains more than 5% of fine particles (e.g., silt and clay), the behaviour of the grain assembly is determined by the fine particles due to their high specific surface area. The flow of immiscible fluids in a porous medium may be conveniently classified into two categories: (1) steady state, i.e., all macroscopic properties of the system are time-invariant at all points, and (2) unsteady state, i.e., properties change with time. Unsteady state behaviour is attributed to the migration of both NAPL and clay particles, under the influence of the applied hydraulic gradient, through the NAPL/soil mixture samples. Once emulsification occurs, the migration of both NAPL and clay particles takes place through the carrier fluid, i.e., the injected surfactant. The injected fluid carries these materials from one part of the sample to another. The migration rate is highly influenced by the flow rate of the carrier fluid, and by the coalescence and retention of the moving particles which may occur during the flow (Mason and MacKay, 1963). It would be expected that the leaching out of NAPL and fine clay particles from the sample would continue under the influence of surfactant injection. Accordingly, a continuous increase in porosity should be observed in the specimen.
Figure 18.10. Volumetric fluid content variations with time for soil with 5% montmorillonite and 95% sand.
The change in porosity as a function of time for various clay/sand ratios and hydraulic gradients is shown in Figures 18.10 and 18.11. It is seen that the pore volume changes with time, indicating an unsteady state condition. For a hydraulic gradient of 2, the porosity did not change (-~0.43) after an injection period of 80 minutes, as shown in Figure 18.10. However, after an injection
SOIL SYSTEM PROPERTIES
481
period of 250 minutes, the porosity changed to 0.44. The change in porosity along the flow direction was observed for all tested samples. The changes, however, are not identical. The differences in the rate of change of porosity in terms of time and flow direction may be attributed to: (a) the difference in total input energy, due to injection periods. The energy is used for emulsifying the NAPL and/or clay, and thus less clay is leached out at the shorter time interval, and (b) the possible retention of the suspended clay particles. Porosity distribution along the flow path varies with time, due to the change in clay content. For example, with 5% montmorillonite clay content, as the hydraulic gradient increases, the rate of porosity change increases. This was true only up to 160 minutes of surfactant injection. As the injection time increased beyond 160 minutes, the rate of change of porosity decreased. This can be attributed to the fact that the clay migration levelled off. Clay particles which had been removed from some of the pores, created properly connected path lines. Similar behaviour was observed at higher clay contents (i.e., 15% and 25%). The percent removal of clay particles was lower at higher clay content (25%) than at lower clay contents (i.e., 5 and 15%). Thus, the larger the number of loose clay particles, i.e., particles not rigidly attached to any position, the greater the number that will move under the application of a hydraulic gradient.
Figure 18.11. Volumetric fluid content variations with time for soil with 25% montmorillonite and 75% sand.
18.4.3 Moisture Content
Moisture content of a soil system is a critical variable. For remedial application and design purposes, it is necessary to know if the pollution problem exists in the unsaturated or saturated zones. The unsaturated zone is defined as the portion of the soil between the ground surface and the water table and includes the capillary fringe. More generally, the zone is defined as that region in which
482
SURFACTANT EXTRACTION PROCESSES
the pressure head is less than zero. Many of the chemical transformations that organic pollutants may undergo in the unsaturated zone are acid-base catalysed reactions. The extent of the dissociation of the acid-base groups at the surfaces of the soil phase may have a strong effect on the persistence of many organic pollutants. Surface acidity is the parameter which is most frequently used to define the efficiency of a surface as an acid catalyst. When a liquid phase is present, its pH may become the most important parameter in controlling the acid-base catalysis. The liquid phase may either supply acid-base groups (e.g., protons) to the surface or serve as a sink for such groups. Thus, the pH of the liquid phase influences not only the acid-base catalysed reactions in soil solution but also the properties of the catalytic sites at the surface itself (Konrad and Chesters, 1969). Moisture content has a drastic effect on the rate of many surface catalysed transformations. Under sufficiently dry conditions, the moisture content will determine which sites on the clay surface will be anhydrous and which will be hydrated. The hydration status of a site will strongly influence the interaction of an adsorbate molecule with that site. For example, Theng, 1974 found that the formation of the protonated benzidine-montmorillonite complex when benzidine came into contact with the dry clay was inhibited by the addition of water. Also, Saltzman et al., (1976) found that oven dried Na-kaolinite, which is almost anhydrous, catalysed parathion hydrolysis much less than the same clay at their hydration level. The unsaturated zone often contains greater amounts of organic matter and metal oxides than the saturated zone. Pollutants can adsorb onto these materials, making their rate of movement substantially less than in the saturated zone. Furthermore, materials adsorbed by soil constituents can act as a source of pollution to the saturated zone even after remediation. Moisture content may affect the distribution of organic pollutants between different phases in the porous medium. The lower the moisture content, the higher is the fraction of the solute that is adsorbed. The drier the porous medium the more of its volume is occupied by the gas phase. The fraction of the gas volume may strongly affect the distribution of volatile substances between the phases and, hence, the interaction of such substances with the surface of the solid phase. Transport of gases and pollutants through the unsaturated zone can be an important consideration in certain field situations. The vapours are transported through the unsaturated zone and eventually may diffuse into the atmosphere. The key physical processes that affect the transport of gases in the unsaturated zone are diffusion and advection, with diffusion playing the larger role. Transport of gases is the result of their larger diffusion coefficient (10 .5 m2/sec) compared to that of solutes (10.9 m2/sec). Many volatile organic chemicals have equilibrium concentrations that are high enough to increase the density of the vapour phase to 1.5 Mg/m 3. This high density, in principle, should cause the pollutants to sink to the capillary fringe. 18.4.4 Soil Organic Matter Content Humic substances (HS) are the most widespread natural non-living organic materials in the terrestrial and aquatic environment. HS occur not only as the major fraction of soil organic matter (OM), but also in marine, river and lake waters and sediments, sewage effluents of various nature, peat, coal and lignite. The amount of carbon on earth occurring as humic substances (60• 1011 tones) has been estimated to exceed about 10-fold of that occurring in living organisms (7• tons) (Senesi and Chen, 1989). Humic substances, due to their poly-electrolytic nature, interact with man-made organic
SOIL SYSTEM PROPERTIES
483
chemicals. The presence of a great variety of chemical reactive functional groups (shown in Table 18.5), in the molecular structure of humic substances renders them able to chemically bind various organic compounds. Humic substances may interact with organic chemicals by various mechanisms such as physical and chemical adsorption, partitioning, solubilization, hydrolysis, and photosensitization. Chemical bonds of various strength are suggested to form between humic substances and man-made organic chemicals, ranging from weak, partially reversible, physical associations to strong, irreversible covalent bonds (Stevenson, 1972; Hayes and Swift, 1978). All these processes will have evident implications on the persistence of man-made organic chemicals in the environment, as discussed in Chapter 6.
Table 18.5- Content of major functional groups in humic substances (Senesi and Chen, 1979) Medium .
.
.
9 Soil 9 Lake Water 9 Ground Water 9 Sea Water 9 River Water 9 Wetland water
Humic substance . . . Fulvic acid Humic acid Fulvic acid Fulvic acid Fulvic acid Fulvic acid Humic acid Fulvic acid Humic acid
Carboxyl (meq/g)
Phenolic (meq/g)
5.2 - 11.2 1.5 - 5.7 5.5 - 6.2 5.1 - 5.5 5.5 5.5 - 6.0 4.0 - 4.5 5.0 - 5.5 4.0 - 4.5
0.3 - 5.7 2.1 - 5.7 0.3 - 0.5 1.4- 1.6 -1.5 2.0 2.5 2.5
Chen and Schnitzer (1978) showed that the surface tension of soil humic substance solutions is concentration- and pH-dependent. Increasing solution pH and concentration of humic substances lower the surface tension of water, thus increasing soil wettability. At high pH, the surface activity ofhumic substances will increase due to the formation ofhydrophilic sites (COOH and phenolic OH groups). This, in turn, affects the interaction phenomena of humic substances with both hydrophobic and hydrophilic organic chemicals.
18.4.5 Buoyancy Forces The dimensionless capillary number (Nc~) and Bond number (NB) are generally used to assess the impact of viscous and buoyancy forces, respectively, on the mobilization of NAPLs in porous media (Morrow and Songkran, 1981). These numbers are expressed as: Nc a
-
lawv Oow
;
Ne :
gke(Ow- Po) Oow
[18.6]
where law is the viscosity of water, v is Darcy's velocity, Oow is the interracial tension between the
484
SURFACTANT EXTRACTION PROCESSES
organic chemical and water phases, ke is the effective permeability, 9w and Po are the density of water and organic chemical, respectively, and g is the acceleration due to gravity. The utility of these numbers can be evaluated with the use of the experimental results reported by Abriola et al., (1995). A series of soil column experiments was performed to quantify the onset of tetrachloroethylene (PCE) mobilization in Ottawa sand. Residual saturations of PCE were established in soil columns. After the entrapment of PCE and initial water flushing, aqueous surfactant solutions (4% wt.) of Witcono12722, Aerosol MA/OT 100 and Aerosol AY/OT 100 were pumped through a column in a down flow mode. This allowed for a sequential reduction in the interfacial tension (IFT) between PCE and the aqueous phase from 47.8 to 0.09 dyne/cm. To minimize mass removal due to micellar solubilization, only 1.5 pore volumes of each solution was injected through the column. The PCE de-saturation curve for 20-30 mesh Ottawa sand, expressed in terms of capillary and Bond numbers, is shown in Figure 18.12. Mobilization effects were estimated by measurement of the mass of displaced free product, which was large relative to the mass solubilized in the aqueous phase. The PCE saturation remains essentially constant until the IFT was lowered to 0.09 dynes/cm with Aerosol AY/OT solution. During this phase of the experiment, the Bond number was slightly greater than the capillary number.
20
15 oo
~
0
10
~
5
~
04 ~
0
Q
Nca
V
NB
O0
o~-
u.l
9 LL
a.. 0
L__
1E-06
1E-05
1E-04
1E-03
1E-02
Capillary and Bond numbers
Figure 18.12. Tetrachloroethylene (PCE) desaturation for 20-30 mesh Ottawa sand (Abriola et al., 1995).
SOIL SYSTEM PROPERTIES 18.4.6
Hydraulic
485
Conductivity
The results of hydraulic conductivity and dispersion tests for Dougherty sand with 0.05 mole/kg surfactant solutions are shown in Table 18.6 (Allred and Brown, 1995). The largest decrease in hydraulic conductivity, k, was 58% for the A4 anionic surfactant while the smallest was 14% for N 1 nonionic surfactant. On average, the decrease was greater with anionic surfactants than nonionic surfactants. The nonionic surfactants caused minimal dispersion while anionic solutions produced effluent containing as much as 7680 mg/kg nonvolatile solids. Clearly, anionic surfactant solutions have the greatest potential for dispersion and mobilization of colloids within the tested sand.
Table 18.6" Changes in hydraulic conductivity and dispersion due to surfactant permeation for Dougherty sand (Allred and Brown, 1995). Surfactant
Trade name
Symbol
MW
viscosity gm/(cm .s)
Changes in k (%)
Dispersion (mg/kg)
Anionic
9Na-Lauryl Sulfate
Witcolate A PWD
A1
288
0.0114
-47
6750
9Na-Alpha Olefin Sulfonate
Witconate AOS
A2
324
0.0114
-35
7680
9Na-Dodecyl Benzene Sulfonate
Aldrich Cat. #28995-7
A3
348
0.0111
- 54
7439
9Na-Laureth Sulfate (3EO)
Witcolate ES-3
A4
437
0.0117
- 58
5471
9Alkyl Polyoxyethylene Glycol Ether
Witconol SN-90
N1
490
0.0119
-14
194
9Alkylphenol Ethoxylate
Witconol NP- 100
N2
640
0.0152
- 44
125
9Alky Polyoxyethylene Glycol Ether
Witconol 1206
N3
835
0.0131
- 22
139
Nonionie
Note that the sign "-" indicates a decrease in hydraulic conductivity, k, and MW is molecular weight of the surfactant.
486 18.5
SURFACTANT EXTRACTION PROCESSES INTERACTION MECHANISMS OF VARIOUS COMPONENTS
18.5.1 Soil-Organic Pollutant Interaction Soil-organic pollutant interaction mechanisms are discussed in Chapters 5 and 6 and will not be repeated. 18.5.2 Pollutant-Surfactant Interaction The nature of surfactant molecules, having both hydrophilic and hydrophobic groups, is responsible for their tendency to concentrate at interfaces and thereby reduce the free energy of the system with which they interact (Myers, 1988). The primary mechanism of energy reduction in most cases will be adsorption at various interfaces. However, when all available interfaces are saturated, the overall energy reduction may continue through other mechanisms, as shown in Figure 18.13. These mechanisms are aggregation or micelle formation, crystallization, layer formation, and dispersed species. Since most common surfactants have a substantial solubility in water, surfactant characteristics can change significantly with changes in the length of the hydrophobic tail, the nature of the head group, the valency of the counter ion, and the solution environment.
Figure 18.13. Processes involved in the reduction of surface and interfacial energy via surfactant action.
INTERACTION MECHANISMS OF VARIOUS COMPONENTS
487
The overall stability of the newly formed materials in solution increases as the temperature increases. This effect is the result of the physical characteristics of the solid phase (i.e., the crystal lattice energy and heat of hydration of the material being dissolved). The crystal lattice energy and heat of hydration are temperature-dependent. When surfactants are crystallized, it is common for the crystalline form to retain a small amount of solvent in the crystal phase. In the case of water, the material would be a hydrate. The presence of solvent molecules associated with the head group allows for the existence of several unique compositions and morphological structures that are different from the structure of the anhydrous crystal. As water or other solvent is added to a crystalline surfactant, the structure of the system will undergo a transition from the highly ordered crystalline state to one of greater disorder, usually referred to as a liquid crystalline. It should be noted that the packing of hydrocarbon chains into a crystalline alignment is difficult due to the many possible variations in configuration of the units of the chain. This is, in ttma, attributed to the relatively low melting point and low crystallinity of most hydrocarbons. The dilute solution aggregation of surfactants into clusters or micelles is a direct consequence of the thermodynamic requirements of the particular surfactant and solvent system under consideration. The transitional phases between the basic micelle and true crystals are natural consequences of the removal of water from the micellar system. The addition of NAPL, electrolyte, or polar non-surfactant solute to an aqueous surfactant solution can lead to the formation of new phases with distinct properties. These new components will alter the thermodynamic balance of the system and will, as a result, alter the nature of the aggregated species present. The extent of the effect will vary with concentration, structural features, and specific interactions. The forces and changes involved in surfactant aggregation in non-aqueous solvents differ considerably from those in water based systems. The orientation of the surfactant relative to the bulk solvent will be the opposite to that in water. In addition, the micelle, regardless of the nature of the surfactant, will be de-ionized in solvents of low dielectric constant and, thus, will have no significant electrical properties relative to the bulk solvent. Electrostatic interactions will play an important role in the aggregation process. The primary driving force for the formation of micelles in aqueous solution is the hydrophobic effect-- the drive to minimize the unfavourable interactions between water and the hydrophobic tail of the surfactant molecule. In non-aqueous solvents, it is unlikely that there will be any significant change in the interfacial energy between the surfactant tail and the solvent, even if one is a hydrocarbon and the other is aromatic. A more significant energetic consequence of nonaqueous micelle formation is the reduction of unfavourable interactions between the ionic head group of the surfactant and the non-polar solvent molecules. Such an effect might be called a hydrophilic effect. Unlike aqueous micelles in which interactions between the hydrophobic tails contribute little to the overall free energy of micelle formation, ionic, dipolar, or hydrogen bonding interactions between head groups in reversed micelles are the primary driving forces favouring aggregation. The forces controlling surfactant interactions with organic chemicals are van der Waals or dispersion forces, the hydrophobic effect, dipolar and acid-base interactions, and electrostatic interactions. The relative importance of each type of interaction will vary with the nature of the organic chemical and surfactant.
488
SURFACTANT EXTRACTION PROCESSES
18.5.3 Soil-SurfactantInteraction Adsorption is the primary soil-surfactant interaction mechanism. As previously discussed, surfactants will adsorb on clay minerals and dissolved humic materials in soils. The sorption is influenced by the solution pH, cationic strength, and temperature. Also, surfactants interact through the mechanism of biodegradation. Native microorganisms present in a soil system may be capable of biodegrading a surfactant. Interaction mechanisms are discussed in Chapters 5 and 6.
18.6
MICELLE FORMATION IN SURFACTANT SOLUTIONS
In the following discussion of the formation of micelles in a surfactant solution, the solution is considered to contain only a single surfactant (no solubilizate) (Wayt and Wilson, 1988). The free energy, AG,, ~ , of the following process: n A ~- A
[18.7]
where n is number of surfactant ions, and A is micelle concentration, consists of a surface free energy, G~ a translational entropy distribution, G ~ tro,.,and an electrical portion, G ~ et~cThe surface free energy contribution is determined from" [ 18.8]
= '~A [h ,,- h 4n ]
AG ~
where YAis the surface free energy of the hydrophobic surface of the surfactant that is in contact with water, and h, is the hydrophobic surface area of a micelle and is given by:
h
= 4n
4~ ]
3
n
-
n
[18.9]
SA
where vAis the molecular volume of the surfactant molecule = (4/3)~ (r ,)3/n, r is the radius of the spherical micelle, S A is the total area of the micelle = 4~ {3 v,/4~ }2/3 nZ/3, hA is the hydrophobic area of a monomer. Substituting, Eq. [ 18.9] into Eq. [ 18.8] yields:
A Gsurface = y A
4 ( 3va ~
22
-3 n ~ - n(SA+hA )
l
[18.10]
The maximum value of n is obtained when the hydrophobic surface area, h,, is zero, thus nmax is given by: 2 VA nma x = 3 6 n
[18.11]
MICELLE FORMATION IN SURFACTANT SOLUTIONS
489
This gives an upper bound on the size of the micelles. The electrical free energy change, calculated within the framework of the Debye-Huckel theory, discussed in Chapter 5. The electrical potential around a spherical charge distribution in an electrolyte solution is given by: AGde C, is
,(r) = IX exp (-K:r)
[18.12]
r
where r is given by: K
8"gCsZ e K T2e 2} 89
=
[18.13]
and c, is electrolyte concentration, e is electronic charge, 4.803• 101~ esu, z is valence, K is Boltzmann constant, 1.381• 10-16 erg/~ e is dielectric constant of solution, and T is absolute temperature. The electrical neutrality requirement of the system (micelle plus ionic spheres) can be shown to yield: !
Ir=r,, = -
dr
4xo e
[18.141
where o is micelle surface charge density, which is given by: He o
=
-
4xr ff
[18.151
where r, is radius of a micelle containing n surfactant ions. From Eq. [18.12], we have: =
-
tX
[--~12 + "~nn]exp (-Krn)
dr* I r =rn
[18.16]
r n
which, together with Eq. [ 18.14], yields" ne exp (Kr) IX
=
-
e (1 + K r )
[18.17]
Hence, ne exp [-r, (r - rn) ] ,(r)
=
-
e (1 + K r ) r
[18.18]
and lib(r) :
He _
e (1+ Kr) r,,
[18.19]
490
SURFACTANT EXTRACTION PROCESSES The electrical free energy is given by: o
AG
A
etec. =
G n elec. - A G
[18.20]
1
etec.
where A Gnetec. is the electrical free energy in the micelle, which is given by: n2e 2 A Gnelec
=
[18.21]
2e (1+ Krn) r n
and A G 1elec. is the electrical free energy of a single surfactant ion, which is given by: AG l
_ elec
2
e 2e (1 + Krl) r I
[18.22]
As the surfactant ions become micellized, they undergo a loss of translational degree of freedom which leads to a decrease in translational entropy. According to Wayt and Wilson (1989), the translational free energy is given by: AG~
=
-
( n - l ) T AS~ ~
[18.23]
where AS1~ is the standard translational entropy change per surfactant ion. By combining the three terms contributing to the standard free energy change of micelle formation, we obtain: A G ~n
= A G ~surface (n) + AG~ 2
(n) + AG~
(n)
2
= 3tA [4~rl 3 n 3_ (SA+ hA) n] e
+
[18.24]
?72
2
1
2e
1
(1 + Krl) r I
(1 + r, r l n -5) rl n -5
n - 2,3 ..... n max
- (n - 1 ) T A S I ~
The equilibrium constant for the reaction given by Eq. [ 18.7] is given by: -AG
K
= exp
on
[,4.1
KT
[A]"
;
where [A] is micelle concentration, and [A,] = n [A].
n - 2, 3 . . . . .
nma x
[18.25]
MICELLE FORMATION IN SURFACTANT SOLUTIONS
491
A mass balance gives"
Atotal
nmax
= [A] + ~
n=2
n[An] [18.26]
nmax
= [A] + ~
,,X[A]"
n=2
Considering the mean aggregation number as the average number of surfactant ions in the clusters in which a surfactant ion is associated, the aggregation number, nA~, can be expressed as: "max n [A] FlAG = ~ n n=l A total nmax
[A]+ ~ nZk [A]"
[18.27]
n=2
Atotal
Figure 18.14 shows the relationship between the aggregation number, nAc, relative to the maximum number, nmax, and the nominal surfactant concentration, [A],o,at. The surfactant concentration at which nAc starts to increase is the critical micelle concentration (CMC). Below CMC, micelle formation is negligible while above CMC, micelle concentration, [A], is essentially unchanged as [A],o~at increases.
E
CMC i iiii
Nominal surfactant concentration,
[A]total
Figure 18.14. Aggregation number versus nominal surfactant concentration.
492 18.7
SURFACTANT EXTRACTION PROCESSES M I C E L L E SOLUBILIZATION
Solubilization is crucial to the surfactant flushing technique. A simple theory which would permit calculation of the solubilizing power of a given surfactant for a particular hydrophobic compound would, therefore, be extremely useful in making a preliminary assessment of the feasibility of surfactant flushing in any particular situation. Let us assume that the concentrations of the pollutant in the interiors of the micelles and in the aqueous phase outside the micelles are related by the simple partition law (Gannon et al., 1989): Cm
k~a -
[18.28]
Ca
where Cm is the concentration of pollutant in the hydrophobic phase inside the micelles (mole/l), c, is the concentration of pollutant in the aqueous phase outside the micelle (mole/l). Eq. [18.28] is a good approximation for the distribution of solutes between water and solvents immiscible in water. Since the interiors of the micelles are rather like micro-droplets of hydrocarbon solvent, this would appear to be a reasonable approximation to use for calculating solubilizate concentration in the micellar interiors. Since we wish to determine the maximum pollutant concentration in the surfactant solution, we assume that the aqueous phase is in equilibrium with solid (liquid) pollutant, and that the pollutant concentration in the aqueous phase is, therefore, the saturation concentration, c,. The pollutant concentration in the micelle is then given by: Cm
[18.29]
= km a c s
The total effective concentration of contaminant in the solution is then given by: Ca + Cm
Cr =
[18.30]
where Vr is the total volume of the solution. If the surfactant solution is relatively dilute (<5%), then we can approximate the number of moles of solute in one litre of solution by (CsX1 litre). The volume of micelle interior phase in one litre of solution is given by: V
= 1 • (c,,,.f- C M C )
• Vta a
[18.31]
where Vm is volume of micelle interior phase, Cur./is the total molar surfactant concentration, CMC is the critical micelle concentration of the surfactant, and V,,a is the the molar volume of the surfactant hydrocarbon tail. The number of moles of pollutants dissolved, N, in this volume of micellar phase is given by: N : 1 • (cs,,,.f - C M C )
x Vtaitx c m
[18.32]
MICELLE SOLUBILIZATION
493
Hence, the total effective concentration of a pollutant in the solution takes the following relations: for C,,ri> CMC: cr : c
[1 +
kmaVta,t (c u,.f - CMC)]
[18.33]
for C,,ri <- CMC:
CT
=
cs
[18.34]
A plot of cv versus cs,rjis therefore predicted to be flat at surfactant concentrations below the CMC, and to be linearly increasing with surfactant concentration above the CMC, with a slope of cs kma V,,t. Values for kmashould be rather similar to octanol-water partition coefficients. The slope of the plot, which is a measure of the surfactant's solubilizing power for the solute under consideration, can then be calculated independently of any measurements of solubilization. Gannon et al., (1989) reported that the value of km,, could be determined by experimental means using the following equation to measure the partitioning between water and a hydrophobic organic compound (octanol) in a two phase system:
k
-
ma
v(co ) w
Vh
~
_ 1
[18.35]
Cw
where Vw is volume of water, V h is volume of hydrophobic phase, C~ is the initial solute concentration in water, and Cwis the final solute concentration in water.
(D 0
Slope = MSR
E
..13 "1 0
O3
CMC
A
Surfactant concentration
(mole/I)
Figure 18.15. Typical relationship between solubility and surfactant concentration for calculating molar solubilization ratio (MSR).
494
SURFACTANT EXTRACTION PROCESSES
The molar solubilization ratio (MSR), illustrated in Figure 18.15, is defined as the number of moles of organic compound per unit increase in the micellar surfactant concentration. In terms of MSR, Eq. [ 18.28] can be rewritten as (Edwards et al., 1991):
kmo
MSR (1 + MSR) (SpAH,CMc Vw)
[1 8.36]
where
MSR : Senn'M1c- SeAI4'CMC CSurf- CMC
[ 18.37 ]
and SeAn.~c is the initial solubility of the PAH (mol/1) in micellar solution when csurfis greater than CMC, SeAH. cuc is the solubility of the PAH at the CMC, Csurjis surfactant concentration at which SeAi~.ulc is determined, and Vw is the molar volume of water = 0.01805 (l/mole) at 25 ~ C.
18.8
SURFACTANT FLOW IN POLLUTED SOIL
During surfactant injection through porous media, the mobile materials (NAPL and suspended solids) are brought into contact with the fluid stream. The rate of flow of the mobile materials is dependent on the surfactant flow rate. In the following developments, the fluids are assumed to be diffused through the porous medium. The concentration profiles can be determined by solving the classical diffusion equation:
Ot
Ox
~
[18.38]
where 0 is the fluid volumetric content, t is time, x is coordinate axis, and D is moisture diffusivity. In addition to moisture movement, NAPL and suspended solids are dispersed. Therefore, during surfactant injection, multi-phase flow will occur. It is also believed that during the flow, there will be a time lag between the surfactant front and the suspension solution front (El Monayeri, 1983; Yong et al., 1990, and 1991). By considering the mass balance equation for the injected surfactant, the continuity equation can be written as follows:
dO e + div (qe) : 0 Ot
[18.39]
where qe is velocity of the injected surfactant, and O~ is the surfactant volumetric content. In order to evaluate the soil-surfactant flux, qe, relative to a stationary coordinate system, it is assumed that qe is equal to the sum of the two velocities: the first, qL,,is related to the movement
SURFACTANT FLOW IN POLLUTED SOIL
495
of slug (NAPL and/or suspended solids) and, implicitly, its surfactant c o n t e n t 0 e while the second, qes, is related to the movement of surfactant relative to the moving soil particles. Thus, qe - qe,+ 0~qs
[18.40]
Combining Eqs. [18.39] and [18.40] gives: 80 e
[18.41]
- div qes + Oediv q~+ q~ g r a d 0 e
Ot
In order to evaluate qs, it is assumed that it consists of two vectors: the first, q~, is related to the suspended solid flux, and the second, qo~, is related to the movement of NAPL relative to the suspended solid matrix. Thus, [18.42]
qs = qoc + Ooqc
where 0o is NAPL volumetric content. Eq. [ 18.42] can be written as: div qs = div qoc + 0 o div qr
qc grad 0 o
[18.43]
Substituting Eq. [ 18.43] into Eq. [ 18.41 ] yields: dOe Ot
- div qes+ 0 e [ d i v qo~+ 0 o div q + qc grad 0o]+ qs grad 0 e
[18.44]
Simplifying Eq. [ 18.44], by considering one-dimensional flow in the horizontal direction, yields: C30e
8t
Oqes +
- -
8x
Oqoc
0
e
OX
.. Oqc
O0 o]
By ignoring the second-order terms (qoc grad [18.45] becomes: 00 Ot
Oqes -
Ox
+ 0e
O0 e
qc-ffxJ + [q~
+ lO~
8qoc
ax
Oe) , ( 0 o
+ 0 0
qc grad
0%
o e c~X
O~
Ox
O e) ,
[18.45]
and (qc grad 0o), Eq.
[18.46]
In order to evaluate the NAPL particle flux, qoc, and the clay particle flux, qc, it is assumed
496
SURFACTANT EXTRACTION PROCESSES
that these fluxes are in direct proportion to the pore water pressure gradients.
(1)
Suspended Clay Particle Flux, qc, Following the above assumptions, we may write the continuity equation as follows:
OPc + div (Pcqc) = 0 Ot
[18.47]
where qc is the velocity of suspended clay particles, Pc is bulk density of suspended clay particles = wt of clay particles/gross soil volume = G, VcT~/VT, G,. and V~are specific gravity and volume of suspended clay particles, respectively, and Vr is the volume of voids plus volume of solids. Eq. [ 18.47] may then be written as: 0
GsVcYe) Vr + Pc div qc + qc grad Pc = 0
10V~
V
Ot
qc + divqc + --;7-,grad V = 0 v
[18.48a]
[18.48b]
Simplifying Eq. [10.48b], by considering only one-dimensional flow, yields:
OVj Oqc OV~ + + qc - 0 Ot Ox Ox where OV~"= OV~/Vc = change of clay concentration
Oqc Ox
-
[18.49]
(cm3/cm3).Therefore,
OVj OVj + qc Ot Ox
[18.50]
By neglecting the second-order term qc (OVc*/ Ox), Eq. [18.50] reduces to:
Oqr
OVj
Ox
Ot
[18.51]
Assuming that the suspended solids move only under the action of pore fluid pressure, u, we may write the soil particles flux, qc, as:
SURFACTANT FLOW IN POLLUTED SOIL Ou
497 [18.52]
qc = - kc -~x
where k~ is the hydraulic conductivity of the soil. For a one-dimensional problem, we have: Oqc
oZv~ -
-
OVc OD
[18.53]
D
OX
C Ox2
OX
OX
where Dc is clay particle diffusivity = kc (Ou / OVc* ). For simplicity, D~ is taken to be constant over a small time interval, At. By neglecting the second term in the R.H.S. in Eq. [ 18.53], we obtain: Oqc
O2 V c -
[18.54]
- Dc
OX
OX 2
Combining Eqs. [ 18.51 ] and [ 18.54]:
DC
o v; Ox 2
or;
-
[18.55]
Ot
(2) Motion o f NAPL Globules, qoc, Following the same argument used in the case of suspended clay particles movement, the following equation can be obtained:
DO
02Vo* OX 2
-
0Vo*
[18.56]
at
where Vo* is the change of NAPL volumetric content, and Do is the NAPL globule diffusivity = kc ( Ou / OVo* ). Thus, substituting Eqs. [ 18.55] and [ 18.56] into Eq. [ 18.46], we obtain:
at
at
0e]
-~x)
- 0e at
- 0~
o-t
[18.57]
Solving Eq. [ 18.57] yields the surfactant volumetric content, 0e, at different time intervals and along the flow path lines, x.
(1)
A knowledge of the distribution profile of 0 e c a n be used for: Calculating the flow rate of the surfactant at any time, t, and position, x, from the surfactant injection point; and
498
SURFACTANT EXTRACTION PROCESSES
(2)
From an experimentally determined relationship between the surfactant flow rate and the NAPL detaching efficiency, the optimum distance between the injection and NAPL collection points can be determined.
18.9
SUMMARY AND CONCLUDING REMARKS
In surfactant-soil washing process, it is desirable to use a biodegradable surfactant so that any surfactant left unaccounted for will simply degrade into harmless components. On the other hand, if the surfactant degrades too quickly, it will be ineffective for cleaning the soil. Thus, biodegradation tests should be performed using the selected surfactant and native microorganisms from the polluted site. These tests will show how long a surfactant will survive before the onset of biodegradation. Several factors impede the application and flow of surfactant in soil systems. Heterogeneities in soil type, porosity, and hydraulic conductivity will alter the fluid flow in soils and add uncertainty to the flow path of surfactant solutions. These factors make it difficult to obtain a uniform, predictable, areal sweep of injected surfactant solution through the desired soil area. Soil pore blockage will also affect the flow of surfactant solutions. Such blockages can occur if emulsions or floes are formed. It should be possible to avoid such undesirable phase behaviour by performing proper laboratory phase studies using the surfactants and pollutants. Surfactant phase behaviour is known to be a strong function of pH, temperature, cationic strength, and organic concentration. Unfortunately, these are variables that are not completely controllable at most in-situ remediation sites. Therefore, it may not always be possible to avoid undesirable phase behaviour and subsequent pore blockage. Pore blockage can also occur by mobilization of fines (e.g., clay-size particles). Fines can be mobilized and redistributed by the natural hydraulic gradient of the soil. Structurally, the fines consist of organic carbon and clay minerals and have a large capacity for sorption when compared to other soil particles. Therefore, a large portion of a pollutant can adsorb on the fines. Once the fines have redistributed in the soil system, new flow paths are created. Surfactant solutions that have been injected into the soil system will follow the newly created flow paths because they offer less resistance than the tortuous paths that go through the original soil structure. When the fines portion of the soil are by-passed due to the newly created flow paths, the surfactant becomes ineffective in cleaning the polluted soil. Injection of surfactants into the subsurface will be allowed only when the resulting risk is deemed acceptable. Unfortunately, little is known about the migration and fate of surfactants in subsurface environment. Receptor exposure to surfactants may be deemed acceptable if these surfactants have direct food additive status. Also, exposure to surfactant metabolites may be deemed acceptable by virtue of the nature and composition of these surfactants. Thus, by using direct food additive surfactants, any risk associated with surfactant injection may be deemed acceptable, even under worst case scenarios for subsurface transport and fate processes. However, limiting surfactant selection to those used as food additives may be questionable, especially in light of the improved performance of other surfactants and the cost and nature of testing necessary for obtaining direct food additive status (Sabatini et al., 1995). Thus, other risk-based approaches for surfactant screening may be warranted as surfactant-based subsurface remediation technologies experience widespread implementation.
CHAPTER
NINETEEN
ELECTROCHEMICAL REMEDIATION
19.1
INTRODUCTION
In 1939, Leo Casagrand used electrokinetic process to stabilize railroad in Salzgitter, Germany. Since then, the process has been investigated and used by geotechnical engineers to improve the stability of excavated earth materials, increase the strength of pile foundations, stabilize soft sediments, and dewater sludge, dredged sediments, and mine tailings. Recently, geoenvironmental engineers have used the electrokinetic process to divert polluted plume, modify groundwater flow, contain polluted soil and groundwater, repair failing containment barrier systems, and treat in-situ polluted soils. In-situ electrochemical soil remediation is a recent technology that may be applied to fine grained, low hydraulic conductivity soils, for which most other remediation techniques are inefficient. Electrochemical remediation, variably named electrokinetic soil processing, electroreclamation, electroremediation or electrorestoration, uses direct electrical currents to extract heavy metals, radionuclides, certain organic compounds, or mixed inorganic species and organic wastes from soils. In addition, the technique may be used to mobilize and transport pollutants to a treatment or collection zone, or to degrade, i.e., treat, pollutants by delivering nutrients, microorganisms or chemicals to the pollutants in-situ. The application of a low direct electric current to soil for an extended period of time results in several changes, such as pH, redox potential and electrolyte concentration, in the soil medium. These changes may impact the nature of the clay surface chemistry and the success of the electrochemical remediation. Understanding the basic electrochemistry and the various changes within the system are the fundamental requirements to the success of electrochemical remediation.
19.2
CONDUCTION OF ELECTRICITY
1 9 . 2 . 1 0 h m ' s Law The fundamental relationship governing the conduction of electricity is Ohm's law. This law states that when electricity flows through an electrical conductor, the current density, ie, is proportional to the field strength:
ie
d4e = - ke d x
499
[19"11
500
ELECTROCHEMICAL REMEDIATION
where i e is the current density (ampere/m2), which is defined as the ratio of the current, I (ampere), and the cross sectional area, A (m2), of the electrical conductor, ke is the electrical conductivity (siemens/m), ~e is the electrical potential (volt), and x is distance (m). 19.2.2 Mobilities
Electrical conduction may be viewed from the standpoint of the charge carriers themselves. First we treat the case in which there is only one kind of charge carrier, with the carriers travelling from left to right. The electric current equals the rate at which charge crosses any plane perpendicular to its flow and is given by: I-
dQ dt
[19.2]
_ NAct4Qiv i
where I is the electrical current (ampere), Q is the electric charge (coulomb), t is time, N A is Avogardo's constant (6.022• 1023 mole1), c, is the concentration of the charge carrier (mole), A is the cross sectional area, Q, is the charge of the carrier, and v, is the average velocity of the carrier in the direction of the current. The term NAG designates the number of carriers per unit volume. The velocity of a charge carrier is proportional to the electric field strength as follows: dr e
vi = + ui
ax
2i
- -
d~ e
u
Iz, I ' ax
[19.3]
where u, is the mobility of the carrier (m 2 V 1 s-l), z, is the charge number (-2 for sulphate ion and 2 for magnesium ion), and Iz, I is the absolute value (2 for sulphate ion and magnesium ion). The group of terms - z , / I z , I simply adjusts the sign in the equation to take account of the sign of the carrier charge. Ion mobilities at extreme dilution in aqueous solution at 298 ~ K are shown in Table 19.1 (Oidham and Myland, 1994). Eq. [19.2] can be represented in another form as: I = -
Iz, lu,c,
AF dr
[19.4]
where F is Faraday's constant = N A Q , / z , = N A Qe = (6.0220x 10 23 mole ~) (1.6022x 1019 C) = 96485 C mole ~, and Qe is the elementary charge. Faraday's constant is numerically equal to the charge carried by one mole of univalent cations. Eq. [19.4] is valid only for a conductor containing only one kind of charge carrier. If there are several, then Eq. [ 19.4] is replaced by: d~ e I :
- AF
2., dx
i=1
Iz, lu,c,
[19.5]
ELECTROCHEMICAL REACTIONS
501
in which the summation is over all charge carriers. Note that for both anions and cations, each of the three terms ]Zl], ui, and c, is positive. Thus, the motions of anions and cations provide contributions of similar sign to the total current.
Table 19.1: Ion mobilities at extreme dilution in aqueous solution at 298 ~ K Cation
u,~ (m 2 S"1 V -l)
Anion
ur (m 2 s -1 V 1)
H§ K§ NH4 + Ag + Cu 2+ Mg 2+ Zn 2+ Na + Li +
362.5x10 -9 76.2x 10 -9 76.1 x 10 -9 64.2 x 10 "9 58.6x10 -9 55.0x10 -9 54.7x 10 .9 51.9x10 -9 40.1 x 10 -9
OH SO42BrCI NO 3" C032CIO 4-
204.8x10 -9 82.7x10 -9 81.3 x 10 .9 79.1 x 10 .9 74.0x10 -9 71.8x10 -9 69.8• .9 46.1 • 10 -9 33.5 • 10 .9
HCO 3 C6H~COO-
The fraction of the total current carried by one particular charge carrier is known as the
transport number, t~, and is given by:
Iz, Lu,e, ti
=
Iz,)u c,
[19.6]
i=l
Ion motion of the kind we have been discussing is known as migration. Other transport mechanisms will be discussed latter. Comparing Eqs. [ 19.1 ] and [19.5], the electric conductivity can be written as: k
: F i=1
19.3
Iz, lu,c,
[19.7]
ELECTROCHEMICAL REACTIONS
19.3.1 Electrolysis A schematic representation of the experimental assembly, typically used in syntheses, is shown in Figure 19.1. The switch (S) completes the circuit when current, which can be detected and measured by an ammeter, then flows through important to realize that neither the solvent (i.e., water or an organic solvent) nor
electrochemical closed. Electric the system. It is the solute alone
502
ELECTROCHEMICAL REMEDIATION
is able to conduct electricity to any significant degree. Using electrochemical terminology, the conducting solution is known as the electrolyte and the positive and negative metal conductors are the anode and cathode, respectively. The process which occurs at the electrodes may be referred to as anodic and cathodic reactions and involves electron transfer to or from electroactive species. The overall system, i.e., a direct current (dc) source, electrodes, electrolyte, and electroactive species, constitutes an electrolytic cell.
Figure 19.1. Experimental assembly typically used in electrochemical syntheses.
An important consequence of the passage of an electric current during an electrolysis was demonstrated by the early studies of Faraday, the results of which may be considered in the form of two laws of electrolysis: (1) The amount of chemical change produced by an electric current, i.e., the weight of any substance that undergoes a reaction at an electrode, is proportional to the quantity of electricity passed; and (2) The weights of different substances reacting during electrolysis in which the same quantity of electricity is passed is proportional to their chemical equivalent weights. The quantity of electricity that is passed in an electrolysis is equal to the product of the current strength and the time for which that current passes: w -
/tW F
[19.8]
where w is the weight (g) of the electroactive compound that has reacted, I is the current strength
ELECTROCHEMICAL REACTIONS
503
(Amperes), t is the time of passage of the current (sec), and W is the equivalent weight of the reactant. Deviations from Faraday's law can be attributed to (Ross et al., 1975): (1) chemical or physical reactions of the products of the electron transfer process with the electrodes or the electrolyte, (2) simultaneous electrode reactions involving impurities in the electrolyte, and (3) alternative reaction pathways for the primary product, e.g., in organic electrode processes a free radical may be formed. During electrolysis, the current flow through a system involves: (1) electronic conduction through the external connectors and the metal electrodes, (2) ionic conduction across the solution between the electrodes, and (3) transfer of electrons across the metal-solution interfaces. These processes are discussed below. Dissolution of a solute in the solvent results in the dissociation of the salt into ions. Electric conduction in the solution is the movement of these ions, the cations (positively charged species) moving toward the cathode and the anions (negatively charged species) moving toward the anode. Electric conduction, therefore, involves a net transfer of matter. At the electrode-solution interface, there is a transfer of electrons from the cathode to the electroactive species, which are reduced, and from the electroactive species, which are oxidized, to the anode. The overall electrochemical reaction is a heterogeneous redox reaction, i.e., reduction at the cathode and oxidation at the anode. The inorganic redox process may be conveniently represented as an electron transfer between the reactants, but in organic chemistry the situation is seldom so clearly defined. Organic reactions involve electron movements accompanied by the making and breaking of covalent bonds. Electrochemistry introduces a "third body" into the reaction, namely, the electrode, which acts as an electron source for a reduction process and an electron sink for an oxidation process. Electron transfer to or from the organic molecule occurs at the electrode surface. 19.3.2 Electrode Potential
The importance of electrode potential in electrochemical reactions was realized through the electro-reduction of nitrobenzene to aniline, as shown in Figure 19.2, (Haber, 1898).When a metal is immersed in an electrolyte, there is a tendency for the metal atoms to ionize, and for the resulting cations to move into a layer of the solution adjacent to the metal. The loss of positively charged ions leaves an excess of electrons at the metal surface. These electrons oppose the layer of cations on the solution side of the interface, as shown in Figure 19.3, and represents the formation of an electrical double layer.
Nitrobenzene
Figure 19.2. Electro-reduction of nitrobenzene to aniline.
.._
Aniline
504
ELECTROCHEMICAL REMEDIATION
Figure 19.3. The formation of an electrical double layer: (a) immersion of electrode, (b) initial formation of a compact layer, and (c) compact and diffuse ion layers of the double layer.
An important property of an electrical double layer is the steady difference of electric potential across the interface between the metal and the solution. The redox reaction is given by: M
~
M 2§
+ ze
[19.9]
where M represents the metal, z is valence, and e is electronic charge. The establishment of an equilibrium is an inherent characteristic of all chemical reactions. The driving forces leading to equilibrium conditions depend on the type of reaction, but they are always governed in the thermodynamic sense by the tendency of the free energy of the system to approach zero. The tendency toward electron loss or gain in the redox system, expressed in Eq. [ 19.9], can be viewed as an electrical driving force or as the electrode potential. It is known that alkali and alkaline earth metals cannot attain equilibrium with their ions in aqueous solutions. In contrast, noble metals such as gold or platinum group metals remain visibly unchanged when immersed in most aqueous solutions. Electrode potentials are expressed relative to an arbitrarily selected standard. The accepted zero of electrode potential, at any temperature, is defined as the potential corresponding to reversible equilibrium between hydrogen gas at one standard atmospheric pressure and hydrogen ion at unit activity. This potential is often known as the normal reversible hydrogen electrode potential, and is given by: +
2 H2 ~ H
+ e
[19.10]
The physical properties of the metal and the solution which contribute to the magnitude of the potential can be illustrated by a thermodynamic free energy cycle, as shown in Figure 19.4, in which the following processes are considered (Ross et al., 1975): (1) metal atoms are sublimed into the gas phase, (2) the gaseous atoms are ionized, and (3) the gaseous ions are dissolved in the solution and the electrons "adsorbed" into the metal surfaces.
ELECTROCHEMICAL REACTIONS
505
The change in the chemical potential Ag ~, and the electrode potential is given by: Ag ~ = A G ~
Mgas
[19.11]
+ I + AG(MZ+sol,.)- zedp e
I
M z§
/1G s,bl
Z! G ( g /1p~
M
,.,J-/
ze
soln.)
MZ+soln. +
-
zeqbe
ZeM
/I G ~
= s t a n d a r d free e n e r g y of s u b l i m a t i o n of the metal;
I
= total ionization potential;
/1G (MZ+soln.)
= s t a n d a r d free e n e r g y of solvation o f the m e t a l ion" = electron w o r k function of the metal; a n d
/1p~
= c h a n g e in the c h e m i c a l potential.
Figure 19.4. Schematic representation of the thermodynamic free energy cycle.
19.3.3 Thermodynamics of Electrode Potential For a reversible electrolytic cell, the change in the free energy, AG, at constant temperature and pressure is related to the electromotive force of the cell AG
= - ZFEmf
[19.12]
where z is the number of electrons involved in the overall cell reaction, and Era1is the electromotive force. In electrolysis of water, the overall cell reaction is written as" 2H20-*
2H2 + O2
[19.131
However, the individual electrode processes involve four electrons since at the cathode and the anode
506
ELECTROCHEMICAL REMEDIATION 4H + + 4e-~ 2H 2
4OH--. 2H20
+
02 + 4e
[19.14]
, hence z = 4.
Figure 19.5. Daniell cell.
The free energy can also be expressed in terms of activities of the reactants and products. Consider Daniell cell, shown in Figure 19.5, which the middle barrier prevents the mixing of the two solutions. The barrier may consist of finely porous materials such as filter paper or compacted porous media. Their function is to allow the passage of ions but to inhibit the mixing of the solutions. The reaction in this case is given by:
Zn(s) + CuZ+(aq)~ Zn2*(aq) + Cu(s)
AG : -
RT In (k) + RT ln | az"2*'aq) ac~(~'I
[19.15]
[19.16]
azn(s ) acu 2.(aq) ]
Combining Eqs. [19.12] and [19.16]:
RT l n ( k ) _ R T lnlaznz_*(aq_2)acu(s____~) I Emf- zF zF azn(s) acu2.(aq)
[19.17]
ELECTROCHEMICAL REACTIONS
507
where the subscripts (s) and (aq) refer to solid and aqueous phases, respectively, k is the equilibrium constant for the reaction, and a is the activity coefficient. Applying the condition that the activity fraction is unity, Eq. [ 19.17] simplifies to:
RT Emf = E Omy - zF ln(k)
[19.18]
and E~ is the standard electromotive force for that cell. The expression for the cell becomes"
Emf
=E o RT ln(az,,~*(aq) acu(s)) mf- z f azn(s) acu2+(aq)
emf then
[19.19]
The cell reaction may be considered as the sum of the individual electrode reactions so that the anodic process is the dissolution of zinc metal:
Zn(s) ~ ZnZ+(aq) + 2e-
[19.20]
and the cathodic process is the deposition of copper:
Cu2+(aq) + 2 e - ~ Cu(s) It follows that the reversible electrode potentials, written in a form similar to Eq. [ 19.19]"
[19.21]
l'I/[zn(s)lZn2+(aq)]and lkI/[cu(s)lCu2+(aq)], may be
=~o [Zn(s)lZnZ+(aq)]-RT ln(aznZ+(aq)) lYl~[Zn(s)lZnZ+(aq)] zF azn(s)
[19.22a]
RT lt~[Cu(s)lCu2+(aq)]= ~o [Cu(s)]CuZ+(aq)]- zF In
[19.22b]
acu(s) ) acu 2+(aq)
where ~O/znr162 and ~otc,ur are the standard reversible electrode potentials for the electrode reactions. Eqs. [ 19.22a] and [ 19.22b] may be generalized in the form:
E = E ~ - RT In II activities of the products z---F II activities of the reactants
[ 19.23]
in which E and E ~ are now used for both reversible cell emf and single electrode potential, respectively, and the symbol I-[ indicates the product of activities.
508
ELECTROCHEMICAL REMEDIATION
19.3.4 Electrochemical Reaction Process
Before electrolysis, the aqueous solution is generally homogeneous, although the potential difference at the metal-solution interface causes a preferential population of either cations or anions in the double layer. If voltage is applied to the cell, current flows and alters the electrode equilibrium. Consider the situation at the electrode. The rate of transport of the electroactive species to the electrode surface and the rate of electrochemical reaction at the surface are finite. Therefore, there is a period of time during which the concentration of the oxidized and/or reducible species at the surface adjust to a new equilibrium condition and the rate of transport of the species reaches a constant value. After this initial period, a steady state is reached. It is convenient to assume that in this steady state a diffusion layer exists, i.e., a layer of solution near to the electrode across which a concentration gradient exists (Figure 19.6) with respect to the electroactive species. Outside this layer, the species are transported by convection.
Figure 19.6. A schematic representation of the formation of the diffusion layer.
A consequence of the reaction of electroactive species is that its concentration at the electrode surface is significantly reduced and a gradient is established between the reaction layer close to the electrode and the bulk solution. Diffusion of the reactant therefore occurs from the bulk electrolyte to the electrode. On the other hand, the concentration of the product in the reaction layer is higher than that in the bulk electrolyte so that the product tends to diffuse away from the electrode surface. The establishment of these gradients of concentration is shown in Figure 19.6. Following the transport of electroactive species to the electrode, an adsorption equilibrium is established. Electron transfer occurs to or from activated species in the electrode double layer. Hence, the adsorption or orientation of the molecule, and particulary the orientation of the functional group, is of prime importance. The next step in electrochemical reaction is an electron transfer to or
TRANSPORT OF ELECTROACTIVE SPECIES
509
from the organic molecule oriented (or adsorbed) at the surface. Reactions subsequent to the electron-transfer step may occur at the electrode surface. However, desorption is also possible. In summary, an electrochemical reaction can be regarded as consisting of four steps: (1) transport of electroactive species to the electrode, (2) adsorption or orientation at the surface, (3) charge transfer, which usually occurs to or from the adsorbed or oriented species, and (4) subsequent reaction, which may be chemical in nature with desorption from the interface.
19.4
T R A N S P O R T OF E L E C T R O A C T I V E SPECIES
Three modes of transport of an electroactive species may be identified. These are migration, diffusion and convection. Though mixed transport is common, the three mechanisms are quite different from each other and arise from distinct causes. Each mechanism may be associated with the gradient of a particular property. Migration occurs in response to an electrical potential while diffusion occurs in response to a gradient of activity or concentration. Convection occurs in response to a gradient of pressure. These transport mechanisms are discussed below.
19.4.1 Migration Migration is a transport mechanism that occurs because charged particles experience a force when placed in an electric field. This force moves cations down the field (i.e., to regions of less positive potential) while anions are moved in the opposite direction. Migration applies only to ionic solutes. An ion of charge number, zi, and mobility, u,, moves at an average velocity, v,, in a field of strength - ddOe/dx, according to:
v~ =
zi d~ e lz~l u~ dx
[19.24]
Transport occurs in the direction of the electric field, which we have chosen to be along the positive x coordinate. The migration flux density is given by: Zi jmig
i
= Vici = _
Iz~[
d~ e uici
dx
ziF -
RT
DFi
d~ e
dx
[19.25]
where u, = Iz, IFD~/RT, according to Nernst-Einstein law, and D i is the diffusion coefficient. It should be noted that the cations travel down a potential gradient and anions climb up it. Any ionic solution must contain at least two species of ions and the migratory flux densities of these different ions are seldom independent. The two or more flux densities are coupled because the same field drives each and by the need to preserve electroneutrality (z+ c ~= - z_ c_). The sum (u~ + u) of two ionic mobilities can be determined from the measurements of the electrical conductivity of a solution of a binary electrolyte, i.e., one that dissolves to give two species of ions only, a cation and an anion as:
510
ELECTROCHEMICAL REMEDIATION k u++ u
-
=
ke : -
F (z+c+)
F (z_c_)
[19.26]
For a sufficiently dilute solution, the mobilities of the ions decline linearly with the square root of the ionic strength. 19.4.2 Diffusion
Diffusion occurs in response to a gradient of activity. Experimental results indicate that the magnitude of the diffusive flux density is proportional to the activity gradient. Since the activity coefficient varies between a point in solution and a neighboring point, the flux can be related to concentration gradient as described by Fick's first law: 4 atii :
-
D, dc
[19.27]
where D, is the diffusion coefficient. The minus sign in Eq. [ 19.27] arises because diffusion occurs down a concentration gradient. 19.4.3 Convection
Convection is a very efficient transport process compared with diffusion or migration. Convection is the movement of solute under the influence of a pressure differential. We may differentiate two types of convection. Forced convection results from a deliberate actions such as stirring, bubbling gases and rotating the electrodes. Natural convection is usually caused from vibration or density gradient due to temperature or concentration gradients. Eq. [19.25] relates the migration flux density to the potential gradient while Eq. [19.27] relates the diffusive flux density to the concentration gradient. The convective flux density is given by Poiseuille's law as: jco,,v _ Qwci _ _ Ac, d P A 8 ~ r I dx
[19.28]
where Qw is the volume flow rate (m 3 s~), and r I is the viscosity of the solution (kg m -1 s-l).
19.5
ELECTROKINETIC PHENOMENA
The electrokinetic phenomena described in this section are dependent on the surface potential of clay particles, discussed in Chapters 4 and 5. When electrical charges are generated at an interface, specifically a solid-liquid interface, a diffuse layer of electric charges is formed on the liquid-phase side of the interface. If an external electric field is applied to this system, a relative movement of the
ELECTROKINETIC PHENOMENA
511
two phases takes place with respect to each other. Since the sign of the charge is opposite on both sides of the interface, the electric force acts on the two phases in opposite directions. Conversely, if a relative motion is caused by an external force tangential to the interface, an electric field will be generated. These phenomena, illustrated in Figure 19.7, are called electrokinetic phenomena. Electroosmosis and electrophoresis are examples of the first kind while streaming potential and sedimentation potential are examples of the second kind.
Figure 19.7. Occurrence of electrokinetic phenomena. Solid phase and liquid phase are subjected to forces acting on them in opposite directions.
All electrokinetic phenomena are caused by the flow of charge and liquid in an electrical double layer relative to the clay particle surface. In treating these phenomena theoretically, it is assumed in many cases that the double layer takes the form of a plane condenser and the electric charges on the liquid side are in a plane parallel to the interface. This is called the Helmholtz layer. The electrokinetic phenomenon in which the solid phase is at rest can be described by Poisson's equation as: d2tp _ 4~co dx 2
[19.29]
where ~ is surface potential, o is surface charge density, e is the dielectric constant of the medium, and x is distance. Although the dielectric constant is not necessarily constant throughout the system, it is usually assumed to be so.
512
ELECTROCHEMICAL REMEDIATION
Figure 19.8. Schematic diagram of the potential distribution from the clay particle surface showing different potentials.
Let us assume that the electrical double layer has the structure depicted in Figure 19.8. That is, the Stem layer is in contact with the interface, whose potential is ~,, and the diffuse layer extends outward from the end of the Stem layer. There is an abrupt change in potential in the Stem layer, but this should be distinguished from that in ~ potential. It should also be remembered that the dielectric constant of the liquid in this layer is much lower than that of the bulk water. The slipping plane is assumed to be located on the outside of the Stem layer. It can be assumed that the movement of the liquid relative to the solid surface never begins immediately next to the clay particle surface, but that a stationary layer of the liquid with a thickness of one to several molecules is always present. The outside boundary of this layer is called the slipping plane. The electrical potential in the slipping plane plays a decisive role in theories of electrokinetic phenomena and is usually called the zeta potential, ~. In view of the presence of the Stem layer, there still remains some uncertainty about whether the location of the slipping plane is inside or outside the layer or on the outer boundary of the layer. However, the plane is generally thought to be located on the outside. 19.5.1 Electroosmosis
Electroosmosis is the movement of an electrolyte solution relative to a stationary charged surface due to an applied electric field. The pressure necessary to counter-balance electroosmotic flow is termed the electroosmotic pressure. A typical electroosmotic fluid flow in a capillary is shown in Figure 19.9. When the capillary tube is negatively charged, the applied electric field exerts a force, along the cathode direction, on the excess positively charged ions near the surface. The positively charged ions drag the electrolyte solution and flow occurs towards the cathode.
ELECTROKINETIC PHENOMENA
513
Figure 19.9. Electroosmotic flow in a capillary.
When an electric field parallel to the interface is applied, a stationary state is soon reached. In this stationary state, a layer dx at a distance x from the solid surface moves with a uniform velocity parallel to the wall, as shown in Figure 19.10. There are two different forces acting on the layer. These two forces must equilibrate.
Figure 19.10. Schematic diagram showing the forces acting on the electric charges in the diffuse ion layer.
The force acting due to the external electric field, ~e, is given by ~eo dx per unit area. This force acts separately on each of the ions in the layer and is transferred to the layer as a whole by the internal friction of the liquid. The friction force exerted on the layer by its neighboring layers moving with different relative velocities is given by:
514
ELECTROCHEMICAL REMEDIATION
x+dx
x
per unit area. As a consequence, in the stationary state:
d2v
(l)e(J dx
1]
1] (
x
dx
dx 2
[19.31]
where v is the velocity of the layer, dx. By using Eq. [19.29] we obtain: f'dOe d 2 1 l l _
4~
dx 2
d2v - rl~ dx 2
[19.32]
Then, Eq. [ 19.32] is integrated over the whole region, in which there is movement of the liquid, in two steps. The first integration gives: dv E~e dl[I + CI = _ n - 4~: dx dx
[19.33]
where the integration constant C~ can be determined to be zero using the boundary conditions that dt~/dx = 0, and d v/dx = 0 when x ~ . Recalling that v = 0 and ~ = ~ on the slipping plane, a second integration results in: -
e~e 4re
~ + C 2 = - rlv
[19.34]
with integration constant C2 = ~ee~/4. Hence, we get:
I~_e(l) (~- Ill) _ _ lqv 4n
[19.35]
which reduces to: rlv= _ e_ed~ ~ 4r~
when x approaches infinity. Eq. [ 19.36] gives the osmotic velocity of the liquid.
[19.36]
ELECTROKINETIC PHENOMENA
515
Let us calculate the volume of liquid transported by electroosmosis. If a capillary in which there is an osmotic flow has a constant cross-sectional area Ao.,, the volume of liquid transported per unit time is given by: V
: v Aos -
~4~e~ Ao~ 4TcrI
[19.37]
In Eq. [19.37] it is assumed that the whole mass of the liquid in the capillary moves with a constant velocity of voo.Since the velocity of the liquid in the diffuse double layer is lower than v~ , as seen from Eq. [ 19.35], the assumption above is correct only when the double layer thickness is very small compared to the diameter of the capillary. This condition is usually satisfied because the double layer thickness is generally smaller than 10 .7 m and the diameter of the capillary is in most cases larger than 10 .5 m. Consider Ohm's law in the following form: ~)e
=
~I
-
I Ao3 "
[19.38]
where f2 is the resistance (Ohm), and ~ is the specific conductance of the liquid (Siemens). Substituting Eq. [ 19.38] into Eq. [19.37], the volume of liquid transported per unit time is given by:
~s-
~: 4~rl~
[19.39]
which is called the H e l m h o l t z - S m o l u c h o w s k i equation.
19.5.2 Streaming Potential When a pressure difference is applied between the ends of a porous material (Figure 19.11), a flow of liquid is induced through the porous material. As this flow carries the charge of the double layer with it, a potential difference arises between the ends of the porous material. The potential difference produces a conduction current directed opposite to the liquid flow and the two are soon balanced. In the stationary state, the convection and conduction currents are equal. The convection current is proportional to the pressure difference, P, while the conduction current is proportional to the potential difference. Therefore, the streaming potential, ~.,.,r, is proportional to the pressure, ~p. The proportionality constant is given by the Helmholtz-Smoluchowski equation as: ~p = 4TcrlZ
[19.40]
This relation (Eq. [ 19.40]) is independent of the dimensions of the porous material under the following conditions: (1) the flow of liquid through the porous material is laminar, which is satisfied
516
ELECTROCHEMICAL REMEDIATION
in most cases, (2) the radius of the porous material section is much larger than the thickness of the double layer, and (3) the specific conductance determines the magnitude of the conduction current and is equal to that of the bulk solution.
Figure 19.11. Streaming potential within a capillary tube.
It is interesting to note that electroosmosis and streaming potential are related to each other through a constant, as revealed by comparing Eqs. [19.39] and [19.40], in the following way: I
~p
4~rlX
[19.41]
19.5.3 Sedimentation Potential
An electric field is created when charged particles move relative to a stationary liquid. The movement of the particles could be under gravitational or centrifugal fields. This phenomenon is sometimes called the Dorn effect or the migration potential. A sedimentation potential in a cylinder is shown in Figure 19.12. The sedimentation potential, llJsed, is given by:
~sea =
ANsa 3ge~ 3Xrl
[19.42]
where X is the specific conductance of the liquid, g is the gravitational constant, A is the difference in density of the solid and liquid, N, is the number of solid particles in a unit volume, and a is the radius of the particle when it is assumed to be spherical.
ELECTROKINETIC PHENOMENA
517
Figure 19.12. A schematic illustrating the sedimentation potential.
19.5.4 Electrophoresis The phenomenon in which colloid particles migrate under the influence of an electric field is called electrophoresis. This motion of colloid particles is very similar to that of ions in an electric field. A typical particle electrophoresis is shown in Figure 19.13. Due to the presence of the anode and the cathode terminals, an electric field, E, is established from left to right. The negatively charged colloidal particle migrates towards the anode.
Figure 19.13. Electrophoresis of a colloidal particle.
According to Smoluchowski, electrophoresis is regarded as the inverse of electroosmosis. In electroosmosis, the liquid moves while the solid is at rest. In contrast, the solid is assumed to move in the stationary liquid in electrophoresis. In both cases the movement of liquid relative to the
518
ELECTROCHEMICAL REMEDIATION
solid surface is governed by the forces acting on the double layer. Consequently, electrophoretic velocity is given by Eq. [ 19.36].
19.6
ELECTROKINETIC REMEDIATION
In standard electrokinetic treatment process, a common form of which is shown in Figure 19.14, compacted polluted soil specimen is placed between the anode and the cathode. As direct current passes through, water flows from the anode to the cathode. The collected water at the cathode is then analysed for specific chemical species.
Figure 19.14. Standard electrokinetic treatment process.
(1) (2)
(3) (4) (5)
The use of electrokinetic processes to transport chemical species relies on several processes: Ion Migration: the movement of the charged ions to the respective electrode as a result of the applied electric field; Ion Diffusion: The movement of ions as a result of being concentrated in the region of the electrode by the applied electric field; Convection of Pore Fluid: The movement of fluids containing the chemical species towards the electrode where the fluid is being collected; Electrodic Reaction: The reaction between the electrolyte and the electrodes; and Electrochemical Reactions of Chemical Compounds: The interaction between chemical compounds and constituents which are the result of electron transfer or the separation of cations and anions in the electrolyte.
Of these five processes, convection of the pore fluid and ion migration will benefit the decontamination process because of the movement of water and ions. However, ion diffusion will prevent ion migration due to the building up of like charges in the regions of the electrodes. Cation
ELECTROKINETIC TREATMENT
519
extraction at the cathode may minimize the ion diffusion effects. The electrodic and electrochemical reactions are very site specific and mainly influenced by the reaction of the chemical compounds in the electrolyte solution. The various parameters which influence the electrokinetic treatment process are discussed below.
19.6.1 pH Variation As discussed earlier, when a low level direct current creates a voltage difference between electrodes, different reactions occur in the soil-water system. At the cathode, hydrogen gas is produced and the surrounding pore fluid solution becomes basic with pH in excess of 12 while at the anode, oxygen is yielded. These reactions are given by Eq. [ 19.14]. The production of H + ions at the anode creates an acid front, as shown in Figure 19.15 (Hamed et al., 1991), which moves from the anode to the cathode by: (a) migration due to electrical gradient, (b) diffusion due to concentration gradient, and (c) convection of the pore fluid due to the prevailing electroosmotic flow and hydraulic gradient. As the acid front moves from the anode to the cathode, the H + ions exchange with the adsorbed cations in the diffuse ion layer around clay particles. This cation exchange process releases the heavy metals into the pore fluid, which are then advanced toward the cathode by advection and diffusion. In addition, low solution pH dissolves heavy metal precipitates (e.g., carbonates and oxides), and increases heavy metal concentration in the pore fluid.
Figure 19.15. Variation of pH as a function of distance and time for kaolinite during the application of electrokinetic remediation.
520
ELECTROCHEMICAL REMEDIATION
However, at the cathode, a base front is developed. The hydroxide ions formed at the cathode migrate toward the anode. These hydroxide ions tend to form metal hydroxide precipitates when they encounter the metal ions which are being transported to the cathode by the electroosmotic flow. It should be noted that at much higher pH values, the solubility of metals increases once again due to the formation of soluble complexes which can have neutral, positive or negative charges, as shown in Figure 19.16. The ionic mobilities ofH + and OH are 362.5x 10-9 and 204.8x 10-9 m2 s1 V1 , respectively, as shown in Table 19.1. Therefore, the acid front is expected to be faster than the base front and precipitation may occur in the vicinity of the cathode.
Figure 19.16. Lead chloride speciation as a function of pH. pHp is referred to lead precipitation pH.
19.6.2 Redox Potential It is known that the oxidation state of a metal can be affected by the supply of oxygen. Oxygen is generated at the anode as water dissociates and may, therefore, diffuse through the soil and change the redox potential during the electrokinetic treatment process. At the cathode, the redox potential is low and pH is high. This condition favours the formation of PbO22 and PbO2H complexes at the cathode. These negative lead complexes will, then, migrate toward the anode. However, at the anode the redox potential is high and pH is low, hence the formation of Pb 2§ Pb(OH)2+, and PbC12+ complexes is favoured. These positive lead complexes will, then, migrate toward the cathode and their mobility in soil will be highly dependent on soil pH.
19.6.3 Surface Charge of Clay Minerals As discussed in Chapter 4, there are two types of surface charged minerals, namely the constant surface charge (fixed charge) and the constant surface potential clay minerals (variable
ELECTROKINETIC TREATMENT
521
charge). Each type has different adsorption properties. Imperfection of the lattice structure, such as substitution of aluminum for silicon in a silicate structure or substitution of trivalent aluminum for divalent magnesium in the brucite layer, may lead to an excess of negative charge or positive charge, respectively, at the clay particle surface. Since this imbalance of charge is due to the defect of the crystal and is not influenced by an external factor, the mineral has a fixed charge. Fixed charge minerals include illite, montmorillonite, and chlorite. When the surface charge is generated by the adsorption of ions onto the surface, the net charge is dependent on the ambient pH. Such a mineral is often called a pH-dependent charge or variable charge mineral. Variable charge minerals include kaolinite and halloysite. Since kaolinite shows little isomorphous substitution, more attention should be paid to the broken bonds at the edges of the clay minerals. The broken bonds attract hydrogen or hydroxyl ions from the pore water depending on the pH of the pore water. At a pH below the zero point of charge (pH 4.2), most of the sites on the kaolinite become positively charged by attracting extra hydrogen ions. Consequently, the repulsion forces between the kaolinite and the positively charged metal ions in solution are increased. This, in turn, will contribute to an increase in ion migration, and reverse osmosis, i.e., the direction of water flow is reversed due to positive charges on kaolinite surface.
19.6.4 Soil Buffering Capacity The principal features which establish the usefulness of the buffer capacity assessment centre around the acidity or alkalinity of the initial soil water system. As noted in Figure 19.17, the pH-acid titration (Phadungchewit, 1990) curves for four soils and a "blank", representing a pure solution (without soil), show the ability of the soils to maintain their natural pH value over different ranges of acid input. The characteristic shapes shown for the kaolinite soil solution follow the same pattern of the "blank", indicating the inability of kaolinite to buffer against acid input.
Figure 19.17. pH-acid titration curves for some clay soils and a blank solution.
522
ELECTROCHEMICAL REMEDIATION
When the soil system pH falls rapidly upon addition of acid, we can interpret this to mean that the soil system shows good removal efficiency of metal ions during the application of electrokinetic treatment. Hence, kaolinite performs better than montmorillonite, natural Quebec soil, and illitic soil, as shown in Figure 19.17. The distribution of lead remaining in the kaolinite-based soil and pore fluid pH along the tested specimen after 27 days of electrokinetic remediation are shown in Figure 19.18 (Li, 1997). More than 57% of the length of the kaolinite specimen has 64% to 76% removal. However, within the adjacent part to the cathode, which constitutes about 22% of the specimen length, the lead concentration increases up to 309% of its initial value. There is a similar trend with the pH curve. At low pH, the lead concentration is low while at high pH, the lead concentration is high. This is attributed to the acid front generated at the anode and propagated through the specimen, providing a low pH environment at the anode and high pH environment at the cathode and, hence, favouring lead removal at low pH and lead precipitation at high pH.
Figure 19.18. Distribution of pH and lead in 80% sand- 20% kaolinite mixture after 27 days of electrokinetic remediation.
The distribution of lead remaining in illitic soil after 43 days of electrokinetic remediation is shown in Figure 19.19 (Li, 1997). The results indicate that lead is removed, only, from a small section next to the anode. As discussed earlier, illitic soil starts to release lead in soil pore fluid when the soil pH is below 5, hence we cannot expect any removal as long as soil pH is greater than 5, as shown in Figure 19.19. The results indicate that it is difficult to remove lead from illitic soil by electrokinetic treatment process, even if it is applied for an extended period of time due to the higher buffering capacity, as discussed earlier. In this situation, chemical reagents should be used to solubilize the precipitated heavy metal and reduce the buffering capacity of the soil. Such chemical reagents are discussed by Mohamed (1997a), Mohamed et al., (1995c), and in Chapter 17 (solvent
ELECTRODIALYSIS REMEDIATION
523
extraction processes).
Figure 19.19. Distribution of pH and lead in 80% sand- 20% illite mixture after 43 days of electrokinetic remediation; Y(1) and (Y(2) refer to the left and the right vertical axes, respectively.
19.6.5 Control of pH and Redox potential Effective control of pH at the cathode can be achieved by rinsing the cathode with water or dilute acid. This is done to remove the hydroxyl ions as they are being generated, hence preventing the high pH wave from entering the soil. In this manner, the cathode is effectively moved into the low pH region on the Eh-pH diagram. Similarly, the anode can be rinsed to promote neutral conditions. The redox potential in the soil can be modified by saturating the anodic solution with oxygen that will be transported into the soil by electroosmosis. Therefore, adjustment of the pH and oxygen concentration might enable the oxidation-reduction process to proceed at a rate that is fast compared to the expected duration of the treatment.
19.7
ELECTRODIALYSIS REMEDIATION
19.7.1 Electrodialysis Electrodialysis is a membrane process in which dissolved ions are removed from soil pore fluid through membranes under the driving force of a direct current. Electrodialysis membranes are ion exchange membranes and are of two types: cation exchange membranes that essentially allow only cations to pass through, and anion exchange membranes that allow only anions to pass through
524
ELECTROCHEMICAL REMEDIATION
(Shaffer and Mintz, 1980). In standard electrodialysis, a common form of which is shown in Figure 19.20, flat membrane sheets are arranged to form parallel channels. The membranes are arranged so that cation exchange membranes and anion exchange membranes alternate, with electrodes at each end. The ioncontaining solution, for example sodium chloride, flows through the channels. When an electric field is applied transverse to the membrane, cations such as Na § pass through the cation exchange membranes and anions such as CI pass through the anion exchange membranes. With reference to Figure 19.20, this reduces the salt content in the channel formed by the left pair of the membranes, termed the dialysate channel, and increases in the channel formed by the right pair of membranes, termed the concentrate channel.
Figure 19.20. Standard electrodialysis demineralization process.
The salt solution is pumped through the dialysate and concentrate channels, with salt being removed continuously along the length from the dialysate channel and transferred to the concentrate channel. A dialysate and concentrate channel with the associated membranes are termed a cellpair. A typical electrodialysis stack may have 50 to 300 cell pairs between a single pair of electrodes, and a number of stacks may be used in series to achieve the desired level of salt removal. Using the same principle, the polluted soil could be placed in the dialysate, as shown in Figure 19.21. Since the generated hydrogen ions at the anode, due to water hydrolysis, are not allowed to cross the anion exchange membranes, acid solution is formed in the anodic cell. In the same manner, the generated hydroxyl ions at the cathode are prevented from crossing the cation exchange membranes, hence base solution is formed in the cathodic cell. The acid and base solutions could be pumped through the anodic and cathodic cells with continuous metal ion removals from the polluted soil in the dialysate.
ELECTRODIALYSIS REMEDIATION
525
Figure 19.21. Electrodialysis soil remediation.
(1) (2) (3) (4) (5)
By using the above described arrangement, we could achieve: Complete isolation of the electrodes; Control soil pH and redox potential in the polluted soil; Eliminate buildup of pH gradient within the soil; With the same external applied voltage, multiple dialysate could be placed between the anode and the cathode, hence increasing the removal efficiency and decreasing the treatment cost; and Production of useful chemical products in the anodic and cathodic compartments through different membrane arrangements.
Therefore, with the use of electrodialysis we could overcome the shortcoming of the electrokinetic treatment process.
19.7.2 Treatment Efficiency Variations of relative lead concentrations (Ce/Co) in the cathodic cell as function of time and chemical reagents used are shown in Figure 19.22, where Ce is lead concentration variation with time, and Co is initial lead concentration in the compacted soil specimen in the dialysate. For tap water reagent (pH 4), the shape of the curve indicates continuous increase of lead concentration in the cathodic cell with time. However, for sodium acetate reagent (pH 5), the curve has three distinct zones. At the end of each zone, lead removal starts to decrease and system efficiency is reduced. To increase lead removal, the concentrated electrolyte solution in the cathodic cell was removed and new fresh brine solution was added. For both chemical reagents, as time increases, relative lead concentrations are increased. In
526
ELECTROCHEMICAL REMEDIATION
the early stages, the use of sodium acetate reagent resulted in up to 10-30% greater lead removal than tap water reagent. At the end of the test, lead removals of 92% and 80% were obtained for sodium acetate and tap water reagents, respectively.
Figure 19.22. Relative lead concentrations in the cathodic cell as a function of time during the application of electrodialysis remediation (Mohamed, 1997b).
Figure 19.23. pH variations within illitic soil after the application of the electrodialysis remediation (Mohamed, 1997b).
SUMMARY AND CONCLUDING REMARKS
527
Soil pH variations with distance from the cathode, for both tap water and sodium acetate reagents, are shown in Figure 19.23. Average soil pH of 7.5 and 6.5 were determined for tap water and sodium acetate reagents, respectively. Unlike the standard electrokinetic remediation discussed earlier, the electrodialysis remediation maintains a uniform pH variation within the compacted soil specimen, as shown in Figure 19.23. Therefore, with the use of electrodialysis, the following two problems associated with the standard electrokinetic remediation are eliminated: (1) soil acidification and clay surface charge reversal in compacted soil specimen near the anodic side, and (2) production of alkaline soil and metal ion precipitation in the compacted soil specimen near the cathodic side. Total lead concentrations with distance from the anode, for both tap water and sodium acetate reagents, are shown in Figure 19.24. For tap water reagent, after a testing period of 90 days, the remaining total lead concentration in the compacted specimen varies between 20 mg/kg of soil near the anodic compartment to 30 mg/kg of soil near the cathodic compartment. This concentration level is below the background lead concentration required by various regulatory agencies. For sodium acetate reagent, the remaining total lead concentration in the compacted soil specimen was less than 10 mg/kg soil, with higher concentrations near the cathodic cell.
Figure 19.24. Lead concentration variations within illitic soil after the application of the electrodialysis remediation (Mohamed, 1997b).
19.8
SUMMARY AND CONCLUDING REMARKS
The implementation or development of electrochemical treatment techniques for soil remediation and site rehabilitation requires a good working knowledge of how pollutants are retained within the soil-water system. Evaluation of the effectiveness of a decontamination procedure can usefully benefit from a closer consideration of how the pollutants are held (retained and bonded) to the soil constituents. As discussed in Chapters 4 and 5, pollutant retention mechanisms for the soil
528
ELECTROCHEMICAL REMEDIATION
constituents differ, depending on mineral type, amorphous material, soil organic matter, and the type and nature of pollutants. An evaluation of the relative ease of removal of a pollutant can be made if one obtains a better appreciation of how the pollutants are held to or within the soil constituents. Therefore, quantification of the internal energy of the system and how it can be affected by the application of various external variables needs to be made. The use of electrokinetic processes for treatment of polluted soils has attracted considerable attention, partly because experiences with electroosmotic procedures in soil dewatering and partly because of the relative simplicity of the field application method. The procedure relies heavily on the fundamental aspects of electrochemistry that govern the movement of ionic species, i.e. anions and cations, to the selective electrodes. For clay soils, diffuse double layer mechanisms developed in the soils indicate that extra energy requirements are needed to maintain movement. The use of chemical reagents and ion selective membranes to enhance electrchemical remediation has potential, but remains to be fully assessed.
CHAPTER
TWENTY
SOLIDIFICATION/STABILIZATION PROCESSES
20.1
INTRODUCTION
Solidification/stabilization (S/S) as a method of treatment for liquid or semi-liquid wastes can be traced back to the disposal of low-level radioactive waste in the 1950s. It has become more important since the US Resource Conservation and Recovery Act (RCRA) of 1976 made disposal of hazardous liquids or sludge in landfills unlawful (land-ban). S/S technologies are also used for on-site or in-situ remediation of polluted soils. The term solidification is given to those operations which improve the physical and handling characteristics of the waste while stabilization refers to those operations which result in rendering a waste less toxic through chemical fixation of pollutants. The S/S technology involves mixing a waste with additives/binders to chemically stabilize it and physically contain it, producing a waste form suitable for shallow land disposal. The S/S treatment is generally appropriate for materials containing inorganic, semi-volatile and non-volatile organic chemicals. Laboratory experiments and field experience have demonstrated the ability of the S/S matrix to decrease the mobility of pollutants by a combination of physical and chemical mechanisms. The exact nature of these mechanisms is not well understood. Long term testing is difficult because of the undefined environmental factors that affect the S/S matrix. Without a fundamental understanding of these mechanisms, it is difficult to predict the long term performance of binder-waste systems.
20.2
S/S PROCESSES
S/S processes are often classified based on the principal additives used to obtain a solid matrix. A number of processes based on inorganic and organic additives are listed in Table 20.1.
20.2.1 Inorganic-based Processes The two principal types of inorganic-based processes are cement processes and pozzolanic processes (lime, kiln dust, fly ash). A pozzolan is a material that contains silica or silica and alumina. It has little or no cementation value itself, but under certain conditions can react with lime to produce cementitious material. Cement and pozzolanic-based processes or a combination of cement and pozzolans are the methods of choice in the S/S industry today. This is attributed to the low cost of the materials, their applicability to a wide variety of waste types, and the ease of field operation. A cement-based waste form is a porous matrix whose hydraulic conductivity is a function of the pore structure and the amount of water originally present in the waste. The leaching of a pollutant depends on either it remains in solution in the pore system or is immobilized through 529
530
SOLIDIFICATION/STABILIZATION PROCESSES
chemical reactions. Cement-based processes create an alkaline environment suitable to the containment of several toxic metals.
Table 20.1: Classification of solidification/stabilization processes Inoqganic-based processes
Organic-based processes
9 Portland cement 9 Soluble silicate-cement 9 Pozzolan-lime 9 Pozzolan-cement 9 Clay-cement 9 Gypsum
9 Bitumen
9 Urea formaldehyde 9 Poly-butadiene 9 Polyester 9 Epoxy 9 Polyethylene
20.2.2 Organic-based Processes
The basic types of organic-based processes are thermoplastic and thermosetting. Thermoplastic processes involve blending waste with melted asphalt, polyethylene, or other thermoplastic additives. Liquid and volatile phases associated with the waste are driven off, and the waste is contained in a mass of cooled, hardened thermoplastic (US EPA, 1986). Thermosetting processes involve mixing the waste with reactive monomers, which join to form a solid incorporating the waste. Urea formaldehyde is one thermosetting resin that has been used. The major problem associated with organic-based processes is the use of hydrophobic additives which are not compatible with water-based waste. Therefore, a special treatment is required to form an emulsion prior to treatment by an organic additive. In addition, organic-based processes are subject to deterioration via biodegradation or exposure to ultraviolet light. This, in turn, impacts on the long term stability of the organic-based processes.
20.3
CEMENT
TYPES AND COMPOSITION
Different types of cement are available for a variety of purposes. The bulk of cement produced is, however, the standard Portland cement. Portland cement is made by heating together lime stone and clay or some other sources of silica at about 2700 ~ forming a mass called clinker. A small amount of gypsum is added and the clinker is ground to a fine powder. Other members of the Portland cement family that are in general use are Portland blast-furnace slag and Portlandpozzolan cement.
C E M E N T TYPES A N D C O M P O S I T I O N
531
Table 20.2: Cement types and compositions (Conner, 1990) Type 9ASTM Type I 9ASTM Type IA 9ASTM II 9ASTM IIA 9ASTM Type III 9ASTM IIIA 9ASTM IV
9ASTM V 9ASTM Type IS, IS-A
9ASTM Type IP, P
9Masonry cements
9Natural cements
9Expanding cement
9High alumina
9 Waterproofed 9 Sorel cement
Description 9 General purpose Portland cement, and usually the least expensive. Used most commonly in S/S processes (CSA "Normal"). 9 Same as Type I, but contains air entraining agents for improved resistance to freeze-thaw and scaling. 9 General use where moderate sulfate attack is expected, or where moderate heat of hydration is required (CSA "Moderate"). 9 Same as Type II, but with air entraining agents. 9 Used where high early strength is required, and in cold weather (CSA "High early strength"). 9 Same as Type III, but with air entraining agents. 9 Low heat of hydration, used in massive structures where temperature rise must be controlled. Develops strength more slowly than type I (CSA "Low heat of hydration"). 9 Used where soils and groundwater have high sulfate contents. Develops strength slowly (CSA "Sulfate resisting"). 9 These are Portland/blast furnace slag cements, the latter designation containing air entraining agents. The slag reacts in the presence of Ca(OH)2 and gypsum in the cement paste. Low early strength, but hardens slowly to the same values as Type I. It is more common in Europe than in the United States. 9 Pozzolan-containing cements made by inter-grinding of suitable pozzolan with the cement clinker. Similar uses and properties as Type IS; also, air entraining grades. 9 Used as mortar in bonding brick and masonry. Usually contains one or more of hydrated lime, lime-stone, chalk, shell, talc, slag, or clay. Good workability, plasticity, and water retention. 9Made from same materials as Portland cement, but at lower temperatures below the sintering point. More like a hydraulic lime. Not restricted in magnesia content. 9 Contains slag and calcium sulfo-aluminate cement (50% gypsum, 25% bauxite and 25% chalk). Expands slightly on hydration due to the formation of ettringite. 9Not a Portland cement. Made by fusing limestone and bauxite with small amounts of silica and titania. Develops very high strength quickly, but may have long term stability problems. Sets slowly. 9 Made by inter-grinding small amounts of calcium or aluminum with cement clinker. 9 Also known as magnesium oxy-chloride cement. Not a Portland cement. Made by adding MgO to a solution of magnesium chloride. The reaction product is hard, but not water resistant.
532
SOLIDIFICATION/STABILIZATIONPROCESSES
Table 20.3: Chemical composition of Portland cements (Comer, 1990) Ignition Insoluble Loss material Characteristics Type 0.6 0.1 I Normal Portland I1 Moderate heat evolution ----111 High early strength --___ Low heat evolution ___ --IV V Sulfate resistance -__ --Rapid hardening 0.9 0.2 Super rapid hardening 0.9 0.1 Jet cement 0.6 0.1
SiO, 22
___ --___ ___
21 19.7 13.8
A1,0, Fe,O, CaO 5.1 3.2 65 --- ___ ---
---
--___ ___
4.9 5.1 11.4
2.8 2.7 1.5
---
---
___ ---
--66 65 59
MgO SO, 1.4 1.6 2.5 --2.0 --1.8 --1.9 --1.1 2.5 2.0 3.0 0.9 10.0
Table 20.4: Mineralogical composition of Portland cements (Comer, 1990) Type I I1 I11 IV V
Characteristics Normal Portland Moderate heat evolution High early strength Low heat evolution Sulfate resistance Rapid hardening Super rapid hardening Jet cement S, = specific surface area. C,S = 2CaOSi0, C,S = 3 CaOSiO, C,A = 3 CaOA1,0, C,AF = 4 CaO A1,0, Fe,O, CS = CaSO, H = H,O
C,S 27 31 19 49 43 11 5 0
=
dicalcium silicate tricalcium silicate tricalcium aluminate tetracalcium aluminoferrite
= = =
C,A 11 5 11 4 4 8 9 22
Free C,AF CaO 8 0.5 0.4 13 0.7 9 0.2 12 0.5 9 9 --8 ___ 5 ---
C,S 45 44 53 28 38 66 68 52
CS 3.1 2.8 4.0 3.2 2.7
___
___
--_
Specific Gravity 31.7
---
---
----3.13 3.14 3.04
s
A
(m2/g) 32.20~10-*
___ ----_ -__
43.40 x 1O-* 59.50 x 1O'* 53.OO x 1O-*
CEMENT HYDRATION
533
Only those cements classified as Portland cements have seen substantial use in waste solidification/stabilization technology (Conner, 1990). Other cement types, such as alumina cement and Sorel cement have seen relatively little use. The American Society of Testing and Materials (ASTM) provides eight types of Portland cement while the Canadian Standards Association (CSA) provides five. These, along with a number of other regular cements, are listed and briefly described in Table 20.2. The composition of some of the different cement types are shown in Tables 20.3 and 20.4. Composition may be expressed in the more familiar form, as elements or stoichiometric compounds, or in terms of cement nomenclature, which is more useful in discussing the hydration reactions of cement.
20.4
CEMENT H Y D R A T I O N
When Portland cement reacts with water, heat evolves as shown in Figure 20.1 (Bye, 1983). The complexity of the hydration process is immediately apparent from the three peaks in heat liberation rate, which suggest that there are three maxima in the rates of the hydration reactions. The first peak is by far the highest but of short duration and followed by the dormantperiod in which the heat liberation rate is relatively low. The physical changes in the paste during this period are readily detected by an increasing stiffness which can be quantified by means of a penetrometer. The chemical and physical processes occurring are, therefore, of practical importance since they lead to a decrease in workability. At normal water to cement ratios, the time limit on placing the solidified matrix is reached about half-way through the chemical dormant period.
3.5 3
~
25.
"-"
2
~"
15-
-r"
1
49- - "
-
-
9
0.5-
0 0
L 5
h 10
I 15
l 20 Time
t 25
I 30
i 35
40
(hours)
Figure 20.1. Variations in the rate of heat evolution in the hydration of ordinary Portland cement (water to cement ratio of 0.4); the recorded heat evolution rate after 30 seconds was 200 watt/kg.
534
SOLIDIFICATION/STABILIZATION PROCESSES
To understand the reactions occurring in the hydration of cement, and to link them with the setting and development of strength in the hardened paste, we need to know: (1) How the hydration reactions of the individual compounds in cement contribute to the heat peaks, such as in Figure 20.1; (2) The causes of the considerable changes in heat-evolution rate; (3) How the products of the hydration reactions pack together to fill space; and (4) The nature of the bonds between the hydration products in the hydrated paste. Historically, two mechanisms of hydration have provided the basis for the interpretation of experimental observations. Le Chatelier (1882) suggested that cement hydration occurred by the dissolution of the anhydrous phases followed by the crystallization of the hydrates in an interlocking mass. Michaelis (1893), on the other hand, thought that solidification of a paste occurred by the formation of colloidal material which hardened with time. There are elements in both hypotheses that are still relevant. Modem views on the rate and mechanism of cement hydration are based on results obtained in investigations of the hydration of laboratory synthesized samples of the individual compounds present. Studies have also been made of the interactions occurring when assemblies of the phases are hydrated. Techniques such as scanning electron microscopy (SEM) and high-voltage electron microscopy (HVEM) have been employed in studies of hydration rate. 20.4.1 Kinetics of Cement Hydration
The usual method of measuring the rate of chemical reaction involves the determination of changes in the concentrations of reactants or products with time. It is possible, as shown in Figure 20.2, to follow the consumption of each of the anhydrous phases in cement hydration (Copeland and kantro, 1964). 20.4.2 Hydration of Phases in Portland Cement
In discussing the hydration reactions of cement, it is useful to use the composition of the reaction products in terms of cement nomenclature, as in Table 20.4.
Tricalcium Silicate (C~S) The products of hydration of this phase at ambient temperature are calcium silicate hydrate (tobermorite) and calcium hydroxide. The reaction may be written as:
2(3Ca0.Si02) + 61120-~ 3CaO.2SiO2.3H20 + 3Ca(OH)2 tricalcium silicate water tobermorite gel calcium hydroxide (2C3S)
(6H)
(C3S2H3)
[20.1]
(3 C/-/)
Calcium silicate hydrate, formed by the hydration of C3S, is poorly crystalline, with only a few broad, weak bands in its X-ray diffraction pattern. This and its uncertain composition have resulted in the general use of the notation C-S-H to represent it. The hyphens indicate that the
CEMENT HYDRATION
535
composition is indefinite. Because C-S-H has a high specific surface area, it is considered a gel. The formation and growth of C-S-H can be observed from the increase in the peak intensity of its major (d = 4.15/~) and secondary (d = 2.67 ,&) basal spacings.
Figure 20.2. Degree of hydration of the principal phases in a sample of ordinary Portland cement (Copeland and kantro, 1964).
Figure 20.3(a). Hydration of C3S through solution, with dissolution-diffusion-crystallization.
536
SOLIDIFICATION/STABILIZATION PROCESSES
The classical mechanisms of hydration are shown in Figures 20.3 (a) and (b), with possible rate-determining steps numbered. The through-solution mechanism, shown in Figure 20.3 (a), is usually identified by the formation of a crystalline product at points remote from the dissolving phase. The solution produced is supersaturated with respect to the hydration product; nucleation and growth of the latter will be rate-determining if the rate of dissolution of the hydrating phase is high.
Figure 20.3(b). Hydration C3S at the surface, with diffusion of ions through a product layer.
The second mechanism, shown in Figure 20.3(b), often described as topo-chemical because of the special relationship between product and reactant, is usually identified by the kinetics of growth and the orientation of the crystal in the product layer. The low solubility of silica and its very high supersaturation with respect to C-S-H would ensure that the product was formed close to the surface even if a through-solution mechanism operated. Consequently, distinction between the two mechanisms is virtually impossible.
(1) (2)
(3) (4)
In summary, the C3S hydration stages are as follows: Immediately after immersion of C3S in water, both lime and silica appear in solution. The concentration of the latter in solution rises rapidly and the hydration proceeds as pH increases; It is generally accepted that a hydration product is formed on the surface of C3S, but views differ as to whether it covers the surface completely; The calcium ion concentration and pH in the aqueous phase of a paste of C3S increase to a maximum; After calcium solution saturation, crystallization of tobermorite takes place;
CEMENT HYDRATION
(5) (6)
(7)
537
Due to formation of C-S-H, heat evolution is accelerated; When the surface area of the C3S or the temperature is increased, the length of the dormant period is decreased. Addition of organic and inorganic chemicals will affect the dormant period. For example, addition of saturated lime solution increases the period (Kondo and Daimon, 1968) while addition of oxalic acid decreases it (Odler and Dorr, 1979); and The reactivity of C3S can be enhanced by rapid cooling (Skalny and Maycock, 1974), hence shortening the dormant period. However, surface hydration and carbonation due to exposure to moist air decrease the reactivity of C3S.
Dicalcium Silicate (C2S) The products of the hydration of C2S are C-S-H and calcium hydroxide, as described by the following reaction: 2(2Ca0.Si02) + 4H20-~ 3CaO.2SiO2.3H20 + Ca(OI--I)2 dicalcium silicate water tobermorite gel calcium hydroxide (2C2S) (4H) (C3S2H3) (CH)
[20.21
The proportion of the tobermorite produced is approximately one-third of that produced in a C3S reaction, as seen in Figure 20.2. Therefore, the contribution of C2S to the second peak in the hydration of Portland cement, shown in Figure 20.1, is quite small. C2S contribution to early strength development is also small; however, it makes significant contribution to late strength development.
Tricalcium Aluminate (C3A) Tricalcium aluminate reacts rapidly with water to form a crystalline hydration product: 3CaO.Al203 + 6H20-~ 3CaO.AI?O3.6H20 tricalcium aluminate water tricalcium aluminate hydrate (C3A) (6H) (C3AH6)
[20.3]
Tricalcium aluminate hydrate (CAH) is formed as platelets with hexagonal symmetry, and morphology resembling that of tobermorite. If lime is present in solution, as is the case when C3S is hydrating simultaneously, the formation oftetracalcium aluminate hydrate (C3A.Ca(OH)2.12H20), as described by Eq. [20.4], is favoured. Conversion of (C3A.Ca(OH)2.12H20) to C3AH6 is inhibited and hydration proceeds more slowly.
3CaO.Al203 + 12H20 + Ca(OH)2-~ 3CaO.AI203.Ca(OH)2.12H20 tricalcium aluminate water lime tetracalcium aluminate hydrate (C3A) (12/-/) (CH) (C4AHI3)
[20.4]
538
SOLIDIFICATION/STABILIZATION PROCESSES
In the presence of calcium sulfate, the product of hydration, as described by Eq. [20.5], is an aluminosulfate (C3A.CaSO4.32H20) known as ettringite -- a mineral, with prismatic crystals and hexagonal cross section.
3CaO.Al203 + 26H20 + 3CaSOa.2H20-. 3CaO.AlzO3.3CaSO4.32H20 tricalcium aluminate water gypsum ettringite (C3A) (26H) (C3S3H6) (C6AS3H32)
[20.5]
Figure 20.4(a). Solubility diagram of calcium showing the ettringite formation boundaries; calcium with log [AI(OH)42] = -2M, and log [SO42] = -3M (adapted from Mohamed et al., 1995).
Ettringite formation induces cracks, due to high tensile stresses, in concrete (Mehta, 1969, 1982). It is, therefore, considered as an expansive mineral. As will be discussed latter, ettringite induces swelling in stabilized soils, and decreases the hydraulic conductivity under no-volume change conditions. Substitution of Fe 3§ for A13+ is common as well as adsorption of anions such as chromate. Much of the water in the structure is loosely held and associated with the sulfate ions. Water is held in channels between columnar units which are parallel to the needle axis. This structure is identified by an empirical formula (Ca3AI(OH)612H20) 3+ (Moore and Taylor, 1968). Ettringite dissolves in water in the absence of lime and sulfate to form calcite, gypsum, and alumina gel as shown in Figures 20.4(a) and (b) (Mohamed et al., 1995).
CEMENT HYDRATION
539
Figure 20.4(b). Solubility diagrams of aluminum showing the ettringite formation boundaries; aluminum with log [Ca2+] = -2M, and log [SO42-] - -3M (adapted from Mohamed et al., 1995).
The retardation of C3A hydration in the presence of calcium and sulfate ions in the aqueous phase has long been considered due to formation of ettringite. Retardation is, also, greater if lime is present in solution (Forsen, 1938).
Tetracalcium Aluminoferrite (CuAF) The hydration products formed by the ferrite phase are usually described as being similar to those formed by C3A, with A13+ ions partly substituted by Fe 3+. In the presence of lime, calcium aluminate hydrate is formed according the following reaction:
4CaO.AI203.Fe203 + 10H20 + 2Ca(OH)2-. 6CaO.A1203.Fe203.12H20 tetracalcium aluminoferite water lime calcium aluminoferrite hydrate (C4AF) (1 OH) (C2H2) (C6AFH12)
[20.6]
The hydration of ferrite in the presence of gypsum produces an iron-substituted ettringite. In order to have a general concept of Portland cement hydration, we assume, as a first approximation, that the major phases in Portland cement hydrate independently. The hydration of C3S is accelerated by soluble sulfate, and the lime produced influences the course and rate of hydration of C3A and C4AF. Studies have indicated that C3S makes the major contribution to the second peak shown in Figure 20.1. The hydration of C3A to form ettringite also makes a significant
540
SOLIDIFICATION/STABILIZATION PROCESSES
contribution to the second peak. If the supply of sulfate ions is greatly reduced, before all the C3A has reacted, then the third peak appears. The role of the ferrite phase is less well established. The reactions contributing to the first peak are complex because all the major phases initially react rapidly with water.
20.5
FACTORS INFLUENCING THE SET OF PORTLAND CEMENT
The following discussion will focus primarily on C3S and C3A -- the two phases which play the most important role in the setting of Portland cement. 20.5.1 Accelerating Admixtures
C3S Accelerated Admixtures The role of an accelerator is to promote the dissolution of the calcium cations and anions from cement, thus promoting the formation of CSH and CH nuclei (Taylor, 1990). The mechanisms inducing these accelerating reactions are not well understood. The following are the reported accelerators for C3S hydration: (1) Sulfate, chromate, and thiosulfate. These are stronger accelerators than chlorides, calcium, or potassium (Regourd, 1982); Calcium salts. These usually have a more pronounced accelerating effect than potassium salts (2) (Mattus and Mattus, 1996); Calcium aluminates, magnesium oxide, and sodium bicarbonates (Conner, 1990); (3) Sodium and calcium carbonates (Taylor, 1990); and (4) Sodium fluoride, sodium aluminate, and potassium carbonate (Jones, 1990). (5) C3A Accelerated Admixtures The hydration of C3A is accelerated in the presence of fluoride (Skalny and Young, 1980), and triethanolamine. The effect of triethanolamine is similar to that of sugar. When added in large quantities, it provokes a flash set (high heat evolution with rapid set). It should be noted that fluoride and triethanolamine retard the hydration of C3S. 20.5.2 Retarding Admixtures
Portland cement set can be retarded when the following compounds are incorporated in the cement matrix (Jones, 1990; Ramachandran, 1984; Regourd, 1982): (1) Inorganic compounds, such as: (a) sodium salts of phosphoric, boric and oxalic acids, (b) chloride salts, and (c) heavy metals, such as copper, lead, arsenic, and zinc; and (2) Organic compounds, such as phenols, glycols, alcohols, carbonyls, carboxylate, chlorinated hydrocarbons, oil, and grease.
C3S Retarded Admixtures The hydration of C3S is retarded in the presence of: (1) fluorides and phosphate salts (Skalny and Young, 1980), and (2) lignosulfonates, sodium gluconate, sugar and many other organic acids.
SOIL-PORTLAND CEMENT INTERACTION
541
C3A Retarded Admixtures
The hydration of C3A is retarded in the presence of: (1) sodium and potassium ions, (2) chromate, sulfate, and carbonate salts, and (3) sugar. It is known that the addition of a solution of 1% sugar completely inhibits the setting of cement paste.
20.6
SOIL-PORTLAND CEMENT I N T E R A C T I O N
When soil pore water encounters cement, hydration of the cement occurs rapidly. The major hydration, or primary cementitous, products are hydrated calcium silicates (C3S2H3), hydrated calcium aluminate (C3AH6), and hydrated lime (Ca(OH)z). The first two of these are the main cementitious products formed. The third, hydrated lime, is deposited as a separate crystalline solid phase. These cement particles bind the adjacent cement grains together during hardening and form a hardened skeleton matrix which encloses unaltered soil particles. The silicate and aluminate phases are internally mixed, so it is likely that none is completely crystalline. Part of the Ca(OH)2 may also be mixed with other hydrated phases since it is only partially crystalline. The tricalcium silicate reaction that produces the primary cementing product is: c 3 s + H ~ o ~ c 3 s Y 3 + Ca(01-1) 2
primary cementitious products
[20.71
In addition, the hydration of cement leads to a rise in the pH value of the pore water. The strong bases dissolve the soil silica and alumina from both the clay minerals and amorphous materials on the clay particle surfaces, in a manner similar to the reaction between a weak acid and a strong base. The hydrous silica and alumina will then gradually react with the calcium ions liberated from the hydrolysis of cement to form insoluble compounds (secondary cementitious products). This secondary reaction is known as the pozzolanic reaction. The pozzolanic reactions which take place in soil-cement stabilization, to produce secondary cementitious products, are given as: Ca(OH)2-. Ca 2+ + 2(OH)-
[20.8]
Ca 2+ + 2(OH)- + SiO 2 -~ CSH soil silica
secondary cementitious products
[20.9]
542
SOLIDIFICATION/STABILIZATION PROCESSES Ca 2§ + 2(OH)- + A1203 -~ soil alumina
CAH secondary cementitious products
[20.10]
In order to have additional bonding forces produced in the cement-clay mixture, the silicates and aluminates in the material must be soluble. The solubility of the clay minerals is equally affected by the impurities present, the crystalline degree of the materials involved, the grain size, etc. In the above equations, the cementation strength of the primary cementitous products is much stronger than that of the secondary ones. The cement hydration and pozzolanic reactions can last for months, or even years, after the mixing, and so the strength of cement-treated clay is expected to increase with time.
20.7
SOIL-LIME INTERACTION
The major strength gain of lime-treated soil is mainly derived from three reactions, namely: dehydration of soil, ion exchange, and pozzolanic reaction. Other mechanisms, such as carbonation, cause minor strength increase. Short term reactions include hydration (for quicklime) and flocculation (ion exchange). Longer term reactions are cementation and carbonation. The use of lime as a stabilizing additive is mainly due to its well known effects when mixed with soils. The natural stabilizing agents for cohesive soil are calcium hydroxide, hydrated lime, or slaked lime. Calcium hydroxide is not itself a binder, but will produce a binder (consisting mainly of calcium silicate hydrate) by slow chemical reactions, principally with the silicates in the clay mineral of cohesive soils (Assarson et al., 1974). 20.7.1 Hydration
A large amount of heat is released when quicklime (CaO) is mixed with clay. This is due to the hydration of quicklime with the pore water of the soil. The increase in temperature can be high enough to cause the pore water to boil (Broms, 1984). The reaction is given by: CaO + H20-* Ca(OH) 2 + HEAT (280 Cal/gm o f CaO)
[20.11]
The calcium hydroxide, Ca(OH)2, resulting from the hydration of quicklime or use of calcium hydroxide as the stabilizer, dissociates in the water, as given by Eq. [20.12]. As a result, calcium concentration and pH of the pore water will increase, and SiO2 and A1203 from clay particles will dissolve. Ca(OI-I)2
~
Ca 2+
+
2(OH)-
[20.12]
CONTROLLING FACTORS IN SOIL STABILIZATION
543
These processes, i.e., hydration and dissociation, will result in ion exchange, flocculation, and pozzolanic reactions that are discussed below. 20.7.2 Ion Exchange and Flocculation
When lime is mixed with clay, sodium and other cations adsorbed on the clay mineral surfaces are exchanged with the calcium. This change in cation complex affects the structural component of the clay mineral. The reaction of the carbon dioxide in soil air and the free water produces carbonic acid (H2CO3). Within a period of a couple of minutes up to some hours after mixing, the calcium hydroxide is transformed, due to the presence of H2CO 3 (Kezdi, 1979). The reaction results in the dissociation of lime into Ca 2+ and OH, hence modifying the electric surface forces of the clay structure. The change in the soil structure is a consequence of cation exchange of monovalent alkali ions by dissociated divalent calcium ions in the pore water. This results in the flocculation of clay particles. The clay plasticity is reduced, making it more workable and potentially increasing its strength and stiffness. 20.7.3 Pozzolanic Reaction
The increase in shear strength of stabilized soils with time is mainly due to pozzolanic reactions. Calcium hydroxide, in the soil water, with silicates and aluminates (pozzolans), from clay particles, form cementing materials or binders (Diamond and Kinter, 1965). Calcium ions react with the dissolved SiO2 and A1203 and form hydrated gels of calcium silicate hydrate (CSH) and calcium aluminate hydrate (CAH). These gels bind the soil particles in a manner similar to that produced by the hydration of Portland cement. However, the lime binding process is much slower than that of the cement process. 20.7.4 Carbonation
Lime reacts with carbon dioxide and forms relatively weak cementing agents, such as calcium carbonate and magnesium carbonate (Ingles and Metclf, 1972).
20.8
CONTROLLING FACTORS IN SOIL STABILIZATION
20.8.1 Soil-Portland Cement Stabilization
Type of Cement: Stabilization of clay soils by different types of Portland cement has been investigated by various researchers. Type III Portland cement yields better improvement of soil than Type I. However, Type I Portland cement is more commonly used for soil stabilization because it is more readily available and cheaper. Cement Content: In general, it has been found that the greater the cement content, the greater is the strength of the cement-treated soil (Broms, 1984; Mohamed et al., 1992).
544
SOLIDIFICATION/STABILIZATION PROCESSES
Curing Time: The strength of cement-treated soil increases with time. The rate of increase is generally rapid in the early stages of the curing period. Thereafter, it decreases with time. Soil Type: The effectiveness of cement-treated soil decreases with increasing, clay content, moisture content, and soil organic matter content. The improvement decreases with increasing plasticity index of soils. Curing Temperature: An increase in temperature accelerates the chemical reaction and solubility of silicates and aluminates, thus increasing the rate of strength gain of the treated soil. Clay Minerals: Hilt and Davidson (1960) observed that montmorillonite and kaolinite minerals were effective pozzolanic agents, as compared to illite, chlorite or vermiculite. The amount of secondary cementitious materials produced during pozzolanic reaction of clay particles and hydrated lime (Ca(OH)2) is dependent on the amount of mineral composition of the clay fraction as well as the amorphous silica and the alumina present in the soil. Soil pH: Pozzolanic reactions are favoured by high pH values, to produce secondary cementitious products, as given by Eqs. [20.8] to [20.10]. 20.8.2 Soil-Lime Stabilization
Type of Lime: The efficiency of lime stabilization depends, in part, on the type of lime material used. Quicklime is generally more effective than hydrated lime (Kezdi, 1979). Unslaked lime or quick lime is more effective than hydrated lime since it can adsorb water from the soil, resulting in an increase in temperature which is favorable to strength gain (Broms, 1984). Lime Content: The strength of properly cured soil-lime mixtures increases as the lime content is increased. Methods for determining the optimum lime requirement have been proposed. Eades and Grim (1966) suggested that the percentage of lime required to maintain a soil pH of 12.6 is the optimum lime content. Hilt and Davidson (1960) reported that the amount of lime is proportional to the amount of clay fraction present in soils, and is independent of the adsorbed cations on clay surfaces. This relationship is given as: The optimum lime content (%) = [% of clay/35] + 1.25. Curing Time: The strength of stabilized soil gradually increases with time through pozzolanic reactions. At early times, the rate of increase is rapid. Soil Type: Clay content should not be less than 20%, and the sum of the silt and clay fractions should be more than 35%. The plasticity index of the soil should be greater than 10 (Broms, 1984). Clay Mineral: The quantity of lime needed to stabilize a soil is highly dependent on the type of clay minerals present in the soil (Eades and Grim, 1966). Kaolinitic and monmorillonitic clayey soils are effectively stabilized with lime alone while illitic clays require the addition of fly ash to obtain a significant strength gain (Hilt and Davidson, 1960). Soil pH: To initiate pozzolanic reaction, a minimum pH of 10.5 is necessary (Davidson et
KINETICS OF REACTANTS AND PRODUCTS IN STABILIZED SOIL
545
al., 1965).
20.9
KINETICS OF REACTANTS AND PRODUCTS IN STABILIZED SOIL
20.9.1 Clay Minerals
As discussed previously, the addition of lime or cement to a soil increases its pH and the solubility of the aluminates and silicates in the soil. Two techniques may be employed to investigate this phenomenon. In the first, the concentrations of aluminum and silicon in the soil pore fluid, after the addition of lime, are obtained via chemical analysis. In the other technique, X-ray diffraction (XRD) method is used to identify the major clay minerals in the soil. The second approach was used to quantify the solubility of palygorskite mineral with time in the lime-treated specimens. The results are shown in Figure 20.5. The Figure shows the relationship between palygorskite peak intensity and time for soil treated with 10% lime content. The major reflection indices (10.4, 6.33, 5.38, and 4.46 A) for identifying palygorskite clay mineral are shown in the Figure. It should be noted that the results shown for zero time represent the reference values, i.e., without treatment. As curing time increases, palygorskite peak intensity decreases. This is due to the increased solubility of alumina and silica fractions of the palygorskite clay mineral. More detailed information about the structure of palygorskite clay can be found in Chapter 4. The results clearly indicate the decrease of palygorskite mineral with time.
Figure 20.5. Variation of palygorskite peak intensity with time for soil stabilized with 10% lime content.
546
SOLIDIFICATION/STABILIZATION PROCESSES
20.9.2 Cementing Agents (CSH and CAH) In a similar manner to the palygorskite study, the formation and growth of CSH and CAH can be obtained by observing the peak intensity of their major basal spacings with time, as shown in Figure 20.6. The formation and growth of both CSH and CAH are enhanced with time as a result of the increased solubility of both silica and alumina.
Figure 20.6. Variation of CSH and CAH peak intensities with time for soil stabilized with 10% lime content. Y(1) and Y(2) refer to the left and right vertical axes, respectively.
20.9.3 Ettringite The formation of ettringite mineral requires a high pH environment where sufficient calcium, aluminum, and sulfate are present. In previous studies (Mohamed et al., 1995; Solem and McCarthy, 1992; Mitchell and Dermatas, 1992; Kumarathasan et al., 1990; Hassett et al., 1989), it was shown that the pH condition required to optimize the formation of this mineral is between 11.5 and 12.5. Addition of lime to a soil increases pH, and calcium and aluminum concentrations in soil pore fluid. Since soil pore fluid was rich in sulfate (130 meq/100g of dry soil), all the necessary conditions for formation of ettringite were satisfied in the tested specimens. The results are shown in Figure 20.7. The formation of ettringite increased with time, and contributed to an increase in the swelling of the stabilized soil. In a non-volume change environment, the increased swelling would contribute to a decrease in the hydraulic conductivity, hence less infiltration of leachates through the stabilized matrix. On the other hand, in a volume change environment, high swelling would cause a heave of the stabilized soil matrix. The latter case is responsible for various reported failures of stabilized roads.
INORGANIC WASTE-PORTLAND CEMENT INTERACTION
547
Figure 20.7. Variation of ettringite peak intensities with time for soil stabilized with 10% lime content.
20.10 INORGANIC W A S T E - P O R T L A N D CEMENT I N T E R A C T I O N 20.10.1 Cations
Lead
The amphoteric (pH dependency) nature of lead makes it difficult to immobilize in cementbased waste forms (Mattus and Mattus, 1996). Solubilities of lead hydroxides, carbonates and sulfides are given in Table 20.5. It is seen that the solubility order of a metal precipitate, from most soluble to least soluble, is carbonates > hydroxides > sulfides. As solubility increases, the metal ion mobility increases. Since the pH of the pore fluid in the cement paste is greater than 12, pH control is an important factor in lead retention in the solidified matrix (Wiles, 1992). Lead retards the set of Portland cement through the formation of insoluble compounds around the silicate phases, hence preventing access of water (Brown and Bishop, 1990; Ortego et al., 1989; Thomas et al., 1981). Some authors have indicated that phosphate can be useful in retaining lead in cement-based waste matrix. In the study performed by Ortego (1990), 10% by weight lead nitrate, lead nitrate plus sodium sulfide, and lead nitrate plus sodium phosphate were introduced in the cement paste and leached with acetic acid. The values for the cumulative percent extracted from the matrix were 11.1% from the cement alone, 15.01% from the matrix containing sulfide, and 4.29% from the matrix containing phosphate. Zinc
The hydration of C3S is delayed in the presence of zinc due to the formation of amorphous zinc hydroxide, Zn(OH)2 (Arliguie and Granet, 1990). Solubilities of zinc hydroxides, carbonates and sulfides are given in Table 20.5. The effect of zinc on the hydration of C3Ais highly dependent
548
SOLIDIFICATION/STABILIZATION PROCESSES
of the percent of sulfate in the cement. Up to 2.5% of sulfate, zinc has an accelerating effect on C3A hydration. When the concentration of sulfate exceeds 2.5%, C3A hydration is retarded. At a sulfate concentration of 5.5% or greater, C3A hydration is completely inhibited (Arliguie and Granet, 1990). A study of the micro-structure of zinc-containing waste forms reveals that the porosity is increased as a result of the formation of large crystals of ettringite and calcium monosulfoaluminates (Thashiro, 1979).
Table 20.5: Comparative solubilities of metal hydroxides, carbonates and sulfides
Metal
Hydroxides
Cadmium, Cd 2§
2.3x 10.5 8.4 x 10-4 2.2x 10l 2.2x 102 8.9x 10i 2.1 1.2 3.9x 10.4 6.9x 10.3 13.3 1.1 x l 0 -4 1.1
C h r o m i u m , Cr 3+
Cobalt, Co 2§ Copper, Cu 2+ Iron, Fe 2+ Lead, Pb 2+ Manganese, Mn 2§ Mercury, Hg 2§ Nickel, Ni 2§ Silver, Ag § Tin, Sn 2§ Zinc, Z n 2+
Solubility (mg/1) Carbonates 1.0• 10-4 ---
------7.0x 10.3 --3.9x 102 1.9x 101 2.1 x 10! ---
7.0x 10.4
Sulfides 6.7x 101~ No precipitate 1.0x 10.8 5.8x 1018 3.4x 10.5 3.8x 10-9 2.1 x 10.3 9.0x 10.2o 6.9x 10.8 4.7x 1012 3.8• .8 2.3 x 107
Source: US EPA/600/D-91/088
Cadmium Previous studies clearly indicated that cadmium can be efficiently fixed in cement-based waste forms (Butler et al., 1990; Brown and Bishop, 1990). The cadmium-cement system generates Cd(OH)2 precipitates, which provide sites for the nucleation of CSH and Portlandite, Ca(OH)2. The presence of sulfide enhances the retention of cadmium due to the formation of CdS precipitates (Ortego, 1990). Solubilities of cadmium hydroxides, carbonates and sulfides are given in Table 20.5. Chromium Previous studies have indicated that chromium must be reduced from chromium (VI) to chromium (III) prior to any stabilization in order to form the insoluble chromium (III) (Wiles and Barth, 1991). Chromium (III) is most insoluble between pH 5 and 13. The hydration of C3A is increased with the addition of chromium (Tashiro, 1979). In the study by Mollah et al. (1992), where 10% (w/w) of chromium solution was introduced in Portland cement, Cr 3+ was incorporated in the CSH through a substitution of Si 4+ by Cr 3+.
INORGANIC WASTE-PORTLAND CEMENT INTERACTION
549
20.10.2 Anions
Chloride The accelerating effect of chloride on C3S hydration depends on the nature of the cation which is associated with the concentration of the chloride salt and the mobility of chloride ion (Regourd, 1982). The mobility of chloride ion follows the order: MgC12 > CaC12 > LiC1 > KC1 > NaC1. It is reported (Lea, 1971) that CaC12 has a dual effect on Portland cement set. CaC12 retards the set of Portland cement at concentrations less than 1% (w/w) but accelerates it at high concentrations. The acceleration phenomenon is related to the ability of CaC12 to shorten the time needed to achieve supersaturation with respect to Portlandite, Ca(OH)2 (Vidick et al., 1989). From the viewpoint of long term performance of the solidified matrix, chlorides could change the physical integrity via ionic substitutions. For example, the interaction of MgC12 with Ca(OH)2 produces magnesium hydroxides precipitates, Mg(OH)2(s ), according the following reaction: Ca(OH)2 + MgCl 2 -~ Mg(09)2(s)
+
CaCl 2
[20.13]
The soluble CaC12 can then react with C3A to form monochloroaluminate:
CaCl 2 + 3CaO.Al203 + 10//20--, 3CaO.Al203.CaCl2.1OHzO(s ) tricalciumaluminate monochloroaluminate C3A C3A. CaClz. I OHaO(s)
[20.14]
Monochloroaluminate is unstable in the presence of sulfate and will form expansive, sulfatebased ettringite (C3A.3CaSO4.32H20) (Mattus and Mattus, 1996).
Fluoride It is reported that: (1) fluoride acts as an accelerator to C3Ahydration but as a retarder to C3S hydration (Skalny and Young, 1980), and (2) fluoride precipitates as calcium fluoride (CaF2) (Conner, 1990). Fluoride precipitation contributes to an increase in calcium concentration in the pore solution, thus presenting a problem for the stabilization process. Carbonate Carbonate ions present in waste solution can change portlandite into calcite, CaCO3, or penetrate the crystalline network of hydrated calcium silicate, C-S-H, causing it to become amorphous and, hence, increasing the strength of the solidified matrix (Regourd, 1982). In addition, the presence ofNa2CO 3 or CaCO 3 decreases the formation of ettringite but enhances the formation of a new crystal -- thaumasite, CaCO3.CaSO4.CaSiOs.15H20 (Regourd, 1982). The decrease in ettringite formation accelerates the set of Portland cement.
s~te Calcium sulfate is used in the cement industry, as an additive to clinker, to control the rate of set and the hydration of the aluminate phase. To control the hydration of C3A, calcium sulfate, in
550
SOLIDIFICATION/STABILIZATION PROCESSES
the form of gypsum (CaSO4.2H20), is most commonly used. As discussed previously, the addition of gypsum contributes to the formation of an expansive ettringite mineral, which retards the set of Portland cement. However, the presence of calcium sulfate is considered the most dangerous waste species due to its adverse effect upon the durability of the waste matrix (Clifton, 1989) as indicated by the following examples (Mattus and Mattus, 1996): (1)
Formation of sulfate-based ettringite due to reaction of sulfate with aluminate hydrates: 2(3CaO.A1203.12H20 ) + 3(Na2SO4.10H20 ) -. 3CaO.A1203.3CaSO4.31H20 + 2AI(OH)3 + 6NaOH + 17H20
[20.15]
Ettringite has a negative impact on the solidified matrix due to the formation of cracks and, hence, an increase in the hydraulic conductivity of the solidified matrix.
(2)
Formation of gypsum due to the attack of portlandite by sulfate as: Ca(09)2
Na2S04.10H 20 -. CaSO4.2H20 + 2NaOH + 8H20
+
[20.16]
Since the size of gypsum crystals is larger than that of portlandite, stresses will be developed and cracks will be initiated. The gypsum formed may, also, react with C3A hydrates to form ettringite.
(3)
Destruction of calcium silicate hydrate (C-S-H), which is the main cementing agent in a hydrated cement paste, due to reaction of C-S-H with magnesium sulfate as: 3CaO.2SiOz.3H20 + 3MgSO 4 + 61-120-+ 3(CaSO4.2H20 ) + 3Mg(OH)2 + 2SiO 2
[20.17]
Phosphate Phosphate forms a protective hydroxide coating around cement grains, thus preventing hydration at normal rate. This is why phosphate is considered a cement destroyer (Lea, 1971). The role of associated cations seems to be of great importance. Calcium phosphate, which is very insoluble, does not seem to present as many problems as the more soluble forms (Kertesz et al., 1990).
ORGANIC WASTE-PORTLAND CEMENT INTERACTION 20.11
551
ORGANIC WASTE-PORTLAND CEMENT INTERACTION
Concerns have been raised regarding the applicability of cement-based S/S technology to organic-bearing wastes. Few studies have been reported on the performance (physical and chemical stability) of S/S matrix containing hazardous organic wastes. In the stabilized matrix, it is difficult to identify positive results due to: (1) sorption effects, (2) dilution by reagent addition and leachate, (3) volatilization of volatile and semi-volatile constituents, and (4) chemical reactions between organic chemicals and inorganic cementitious solidifying agents. During the cement-based S/S process, volatile organic compounds (VOCs) will reach their boiling points since VOCs have boiling points below 150 ~ C. Semi-volatile organic compounds may also volatile during cement hydration. Aside from adsorption, volatilization and biodegradation, possible interaction mechanisms between organic chemicals and inorganic cementitious solidifying agents are hydrolysis, oxidation, reduction and salt formation (Conner, 1990). Hydrolysis refers to the reaction of a compound with water. This usually results in the exchange of a hydroxyl group (-OH) for another functional group at the reaction center. Many organic compounds are resistant to hydrolysis. The list includes some esters, many amides, all nitriles, and some carbamates and alkyl halides. Those less resistant to hydrolysis include alkyl and benzyl halides, polymethanes, substituted epoxides, aliphatic acid esters, chlorinated acetamides, and some phosphoric compounds (Conner and Smith, 1996). Oxidation is said to occur when an electrophilic agent, such as a clay surface, attacks an organic molecule and removes an electron, hence forming a free radical. Free radical reactions require much less energy than oxidation of a polar compound or cleavage of covalent bond. Many substituted aromatics (e.g., benzene, naphthalene, and phenol) undergo free radical oxidation, as discussed in Chapter 21. Chlorinated aromatics and polynuclear organics are unlikely to be oxidized (Conner, 1990). Reduction is said to occur when an organic chemical experiences a net gain of electrons. Reduction alteration of organic pollutants in waste is the least understood of all the reactions. If the measured system redox potential is less than that of the original organic constituent, reduction takes place. The majority of organic species encountered in organic S/S systems retard or inhibit the setting and curing of the final product. Oils, grease, and non polar organics inhibit the setting and may decrease long term durability of cement-based S/S. The retarding effect may be due to the nonionized (OH) group which is adsorbed, by bonding, to the surface of cement particles (Lea, 1970). The greater the number of OH groups per molecule, the greater is the possibility of adsorption and retardation. Adsorption of organic complexes forms a protective layer which slows down C3A hydration (Young, 1972). Generally, polar organics have little effect on setting although some alcohols may retard setting (Conner, 1990). Organic compounds may decrease long term durability of the solidified matrix. The long term impact of some organic compounds on solidified matrices by Portland cement Type I is shown in Table 20.6 (Spooner et al., 1984). In the study performed by Mohamed and Lu (1997), 20% (w/w) concentration of motor oil was mixed with Portland cement Type I. Specimens were tested after different periods of hydration ,i.e., t = 0 and 90 minutes. Specimens were then tested using Fourier Transform Infrared Spectrometer (FTIR) to quantify the hydration and interaction mechanisms. The reduced results from various FTIR spectra are shown in Figure 20.8, in terms of the hydration index for different
552
SOLIDIFICATION/STABILIZATION PROCESSES
conditions of the stabilized matrices. The hydration index was calculated based on absorption of the Si-O group since it represents the hydration of C-S-H. For specimens hydrated for a period of 90 minutes at room temperature (0%-90 and 20%-90, shown in Figure 20.8), the addition of oil contributed to a reduction of about 70% of the C-S-H hydration.
Table 20.6: Long term impact of some organic compounds on solidified matrices by Portland cement Type I Organic compound 9 Alcohols & glycols 9 Aliphatic & aromatic hydrocarbons 9 Organic acids & acid chlorides 9 Phenols 9 Chlorinated hydrocarbons
Observed impact Decreased durability; destructive action occurs over a long time period. Increased set time; no significant effect on durability. No significant effect on set time; decreased durability; destructive action occurs over a long time period. No significant effect on set time; decreased durability; destructive action occurs over a long time period. Increased set time; decreased durability; destructive action occurs over a long time period.
Figure 20.8. Variations of Portland cement hydration index with % oil content and hydration time (H) in minutes.
EVALUATION OF POTENTIAL LEACHABILITY
553
These results can be explained by considering the interaction mechanisms between motor oil and hydrated cement. Since the oil consists of large molecules with long chain lengths, it is expected that the primary mode of interaction will be van der Waal's attractive force. Oil adsorption on the cement particle surfaces can be characterized from the appearance of C-H stretching group at the wave-number band 2800 to 3000 cm l. This observation suggests the occurrence of an intercalation phenomenon, i.e., penetration of oil into the inner spacing of the polymerization structure of C-S-H. The C-H bands are shifted to higher frequencies in the mixture, as compared to their frequencies in oil, due to the formation of strong hydrogen bonding between hydrogens of the C-H group and the oxygen or the hydroxyl of the silicate layer. Adsorption of oil on the cement surface leads to an increase in both frequency and energy of the newly formed bond and, hence, an increase in the spacing of the pore structure of the hydrated cement paste. This, in turn, would contribute to a reduction in the unconfined compression strength of hydrated oil-cement mixtures (Shannon, 1993). In addition, Si-H bonding is inhibited by oil addition.
20.12 EVALUATION OF POTENTIAL LEACHABILITY 20.12.1 Leaching Tests Leaching tests are used to predict the mobility of pollutants in the solidified matrix. In view of the variety of possible landfill scenarios, it is not surprising that no single leaching test procedure or protocol can duplicate all possible field conditions. Ideally, the treated waste would be tested with the actual surface and environmental conditions in the area under study. In practice this is rarely achieved because of lack of definite knowledge of the appropriate environmental conditions. Given this reality, leaching test procedures have been developed by various countries. In this discussion we will be concerned with a diffusion modelling test, developed by the American Nuclear Society (ANS 16.1) (ANS, 1986), which can be used to predict the long term performance of the solidified matrix. ANS 16.1 Procedure The ANS 16.1 test consists of a procedure in which the leachate is sampled and replaced by a fresh leachate solution at designated intervals. Specimens are prepared in monolithic form, usually as a cylinder with a length to diameter ratio of 0.2 to 5. The minimum recommended dimension is 10 mm. The leachate solution used is demineralized water with an electrical conductivity of less than 5gmho/cm at 25 ~ C and total organic carbon content of less than 3 ppm. The specimen is suspended in the leachate contained in the leach test vessel such that at least 98% of the surface of the specimen is in contact with the leachate. The leachate volume should be at least 10 times the surface area of the specimen. Periodically, the specimen is removed, rinsed, and immediately re-immersed in another fresh leachate batch. Leachate replacements are at 2, 7, and 24 hours from the initiation of testing. Subsequent leachate sampling and leachate replacement are made at 24-hour intervals for the next 4 days. Three additional intervals of 14, 28, and 43 days may be used to extend the test to 90 days. A sample of the leachate is taken at the end of each leachate interval to determine the amount of species of interest present in the leachate volume.
5 54
SOLIDIFICATION/STABILIZATION PROCESSES
20.12.2 Leaching Mechanism A solidified waste is a porous solid matrix which is partially saturated with water. The pore water in the solid matrix is in chemical equilibrium with the solid phase. When the solid matrix is exposed to leaching conditions, the equilibrium is disturbed. The resulting difference in chemical potential between the solid and the leachate solution causes a mass flux between the solid surface and the leachate. This in turn causes concentration gradients that result in bulk diffusion through the solid matrix (Conner, 1990; Cote et al., 1989). Transport can either be by diffusion of metal ions from the solid matrix surface into the bulk aqueous phase, or by dissolution into the water in matrix pores. Consequently, the porosity and integrity of the waste form is of major concern. Leaching will take place when the chemical constituents in the solid matrix pore fluid exist in ionic form. The amount of dissolution which occurs is dependent on the solubility of the constituent and the chemical makeup of the pore water, particulary its pH. Under neutral pH leaching conditions, the leaching rate is controlled by molecular diffusion of the soluble species. Under acidic conditions, however, the rate will also be governed by the rate of penetration of hydrogen ions into the solidified matrix, since this establishes the speciation and solubility of the pollutants present. Acid attacks pozzolanic-based paste through permeation of pore structure and ion dissolution. Acid consumes most of the calcium hydroxide in the leached layer and leaves a highly porous structure. Experimental results have shown that the thickness of the leached layer is usually less than 10 gm. Diffusion across this layer can be considered a steady state process since the leached layer provides little resistance to diffusion. At the leaching front, diffusion of hydrogen ions proceeds as if the porous medium is infinite and dissolution reactions occur in the pores. Proton transfer reactions are usually very fast, with half-lives of less than milliseconds. Hence, dissolution reactions can be treated as diffusion-controlled fast reactions. The whole process then can be described as steady state diffusion across the leached layer, and unsteady state diffusion-controlled fast reactions in the porous leaching front (Bishop, 1991).
20.12.3 Leachate Transport Modelling Once metal ions become soluble, they are transported from the solid matrix through the leached zone to the leaching solution by molecular diffusion. The flux of the constituent within the solid can be described by Fick's first law: Oc J~ = - D~ -~x
[20.181
where Jc is the flux, c is the concentration of constituent, Do is molecular diffusion, and x is distance. A semi-infinite medium diffusion model with uniform initial concentration and zero surface concentration can be used to interpret the kinetic data generated from serial batch leaching tests (Godbee et al., 1980). The equation takes the form:
-Co
ci
= 2
1 i 5 tn2
[20 19]
EVALUATION OF POTENTIAL LEACHABILITY
555
where c, is pollutant loss during leaching period (mg), Co is initial amount of pollutant in specimen (rag), V is volume of specimen (cm3), S is surface area of specimen (cm2), tn is time (sec), and De is effective diffusion coefficient (cm2/sec). From Eq. [20.19], De can be calculated from the slope of a plot of (~c/Co) versus ~/tn. The linear relation for a given pollutant indicates that the release of the pollutant from the solid matrix to the aqueous solution is controlled by diffusion. As part of the American Nuclear Society leachability test (ANS 16.1), the leachability index is recommended as a standard method for evaluating solidified waste forms (ANS, 1986):
log/e/ where LX is leachability index (dimensionless), [3 is a constant with an assigned value of 1 cm2/sec. The leachability index is used to compare the relative mobility of pollutants on a uniform scale which varies from very mobile for values of 5 ( De = 105 cm2/sec) to immobile for values 15 (D e = 10~5 cm2/sec) or greater. Sample Problem 1: Specimens were prepared by mixing Portland cement and lithium polluted soil. Each specimen measured 5 cm in height and 1 cm in diameter. The initial amount (Co) of lithium in the specimen was 44.4 mg. Leaching procedures, as described by ANS 16.1, were performed and the test results are shown in Table 20.7. The results reported in the Table are the sampling time, concentration in the leachates for each leaching period (c), and cumulative pollutant loss during each leaching period (~c). It is required to calculate the effective diffusion coefficient and the leachability index.
Table 20.7: Results of ANS 16.1 leaching test t (h)
c, (mg/1)
~c, (mg)
~ c / c o (%)
2 7 24 48 72 96 120 192 264 336 504 672
1.150 0.745 1.303 0.779 0.596 0.370 0.305 0.694 0.493 0.613 0.655 0.276
1.725 2.842 4.796 5.964 6.858 7.413 7.870 8.911 9.650 10.570 11.552 11.966
3.88 6.40 10.80 13.43 15.44 16.69 17.72 20.06 21.73 23.81 26.02 26.95
556
SOLIDIFICATION/STABILIZATION PROCESSES Solution: From Eq. [20.19], the effective diffusion is given by: 71
.=
De
:
4t
Co
[20.21] 71;
4•215215 : 3.683 x 10 .8
cm
2/sec
Based on Eq. [20.20], the leachability index, LX = 7.43. Therefore, the mobility of lithium is considered medium.
20.12.4 Factors Affecting Leachability There are number of factors which can affect the leachability of a particular S/S waste form (Bishop, 1991). These factors are characterized as waste form factors and leachate factors. Waste form factors include pollutant binding mechanism, alkalinity, surface to volume ratio, porosity, tortuosity, and durability. Leachate factors include composition, leachate volume to waste surface area ratio, flow rate and temperature. The composition of the waste form determines the physico-chemical properties and the leaching mechanisms. Every effort should be made to minimize the potential for leaching by improving the quality of the waste form. Alkalinity is needed in the final product to maintain metals in their most insoluble form and to buffer against acid dissolution. One of the principal factors governing diffusion of soluble metals from the waste form is the surface to volume ratio. The larger the solidified matrix, the smaller the surface to volume ratio and the smaller the potential for leaching. The role of internal fractures in calculating the correct surface to volume ratio has not as yet been determined. Porosity and tortuosity govern to a large extent the rate of diffusion to be expected. A highly porous matrix will have a higher effective diffusion coefficient for a particular pollutant than a less porous one whereas a waste with a large tortuosity factor will have a reduced effective diffusion coefficient. Waste durability is an important factor because if an initially intact and acceptable monolith weathers poorly over time, the porosity and surface to volume ratios will increase markedly, resulting in increased leachability. The composition ofleachate determines the reactions which will occur within the waste form. Acid dissolution, oxidation-reduction reactions and metal complexation can all occur, depending on the chemical composition of the leachate. Increased leachate volume to surface area ratios and flow rates will increase leaching because diffusing substances will be removed from the monolith surface more rapidly and concentration gradients in the solid will be greater. Temperature increase results in an increase in reaction rate, hence an increase in leachability.
SUMMARY AND CONCLUDING REMARKS 20.13
557
S U M M A R Y AND CONCLUDING R E M A R K S
The long term durability of cement has been well established in conventional construction projects. However, the long term durability of cement-based stabilization/solidification (S/S) is unknown. Stabilized/solidified materials can be disposed in landfills to provide secondary barriers between natural waters and the wastes. Contaminant release may be initiated when the secondary barrier permits natural waters to come into contact with the waste forms. The question is not whether the waste forms eventually release pollutants into the environment, but whether the rate of release is environmentally acceptable. S/S technologies for waste treatment have been in use for only a few decades, so the number and duration of studies on field disposed S/S are very limited. Decisions about the acceptability of particular S/S products must based on the available short term field data, laboratory data, and models of leaching behaviour.
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CHAPTER
TWENTY
ONE
BIOREMEDIATION
21.1
INTRODUCTION
Bioremediation refers to the decontamination or restoration of a mixture of organic waste and soil by means of enhancing the chemical degradation or other activities of soil organisms. In the past, organic residues were piled, moistened, and allowed to undergo biological decomposition -- a process known as composting of organic wastes. The technology has been expanded to include the treatment of food wastes, agricultural wastes, and wastewater. More recently, bioremediation concept has been applied to the treatment of hazardous wastes, remediation of polluted soils and groundwater. Microorganisms are the primary agents in bioremediation. They will transform almost any material at some given concentration. As discussed in Chapter 4, soils contain microorganisms which either biodegrade or mediate the following: (1) degradation of carbonaceous wastes, (2) transformation of cyanide to mineral nitrogen compounds to inert N2 gas, (3) initiation of metal ion oxidation-reduction, and (4) production of: (a) CO2 with subsequent formation of weakly ionized carbonic acid, (b) simple organic acids, hydrogen, and carbon monoxide, (c) small or large molecular species upon which trace pollutants may be adsorbed, (d) complex organic compounds which may react with waste pollutants, and (e) small-sized organic debris which can infiltrate small pore spaces and pollute surface and groundwater. It is critical that a bioremediation system be designed with an understanding of the biological processes involved. Understanding the transformation pathways allows evaluation of the extent of biodegradation, the accumulation of intermediate products, and the requirements that must be fulfilled for successful biological treatment system. In this regard, this chapter is concerned with the basic principles that govern the biological treatment processes.
21.2
SOIL BIOMASS
21.2.1 Definition
Soil biomass is defined as the total mass (dry weight) of living organisms in soils (Paul and Clark, 1989; Brady, 1990). This definition accounts for all the living fauna andflora, each of which consists of a macro and micro that contribute toward formation and accumulation of soil organic matter in soils. Macro-fauna are the higher animals (e.g., earthworms). Micro-fauna are small worms, normally of length 1-2 mm. They consume dead soil organic matter and microorganisms such as bacteria and algea (Jones, 1983; Brady, 1990). Macro-flora are the higher plant, including the 559
560
BIOREMEDIATION
cultivated crops. Micro-flora (bacteria, fungi, actinomycetes and algae) are the lower plants, or the microscopic small plants. They are the most abundant of all the living organisms in soils. They are usually present near feeder roots of plants and play a vital role in many dynamic microbial reactions. The population of microorganisms decreases rapidly with soil depth. 21.2.2 Bacteria
Bacteria are the simplest form of plant life in soil, and are composed of single cells, usually 5 gm in size. Bacterial cells are composed of 80% water and 20% dry material of which 90% is organic and 10% inorganic. An approximate formula for the organic fraction is CsH702N(Sawyer et al., 1991). As indicated by the formula, about 53% by weight of the organic fraction is carbon. The organic fraction can be represented by the formula C60H87023N12p when phosphorus is present. The inorganic portion consists of P205 (50%), 803 (15%), Na20 (11%), CaO (9%), MgO (8%), K20 (6%), and Fe203 (1%). All these elements are necessary for microbial growth. The bacterial cell composition is shown in Figure 21.1. Bacteria lack nuclear membranes, and their nucleoplasm is not separated from the cytoplasm. The cytoplasmic membrane is a 40 to 80 ~, thick membrane that contains a phospholipid bilayer with proteins embedded within the bilayer. The phospholipid bilayer is made of hydrophobic fatty acids oriented towards the inside of the bilayer and hydrophilic glycerol moieties oriented towards the outside of the bilayer.
Ribosomes
DNA Cell wall
Flagellum
Cytoplasmic membranes
Figure 21.1. Composition of bacterial cells.
Cell wall is composed of a mucopolysaccharide called peptidoglycan or murein (glycan strands cross linked by peptide chains). Peptidoglycan is composed of N-acetylglucosamine and Nacetylmuramic acid and amino acids. A cell wall stain, called the Gram stain differentiates between Gram-negative and Gram-positive bacteria, on the basis of cell wall composition. Peptidoglycan layers are thicker in Gram-positive than in Gram-negative bacteria. In addition to peptidoglycan, Gram-positive bacteria contain techoic acids, which are made of an alcohol and a phosphate group. Microbial cells can move by means of flagella. Bacteria display various flagellar arrangements ranging from monotrichous (polar flagellum, as shown in Figure 21.1) and lophotrichous (bundle of flagella at one end of the cell) to peritrichous (several flagella distributed
SOIL BIOMASS
561
around the cell). The flagellum is composed of protein called flagellin and is anchored by a hook to a basal body located in the cell envelope. Flagellum enables cell to move towards food, light, or oxygen. Ribosomes contain endoplasmic reticulum which is involved in protein synthesis. Endoplasmic reticulum is a system of folded membranes attached to cell membrane. Deoxyribonucleic acid (DNA) is a double stranded molecule that is made of several millions of units called nucleotides. Each nucleotide is made of a 5-carbon sugar (deoxyribose), a phosphate group, and a nitrogen-containing base linked to the carbon-5 and carbon-1 of the deoxyribose molecule, respectively. The nucleotides on a strand are linked together via a phosphodiester bridge. Bacteria are called prokaryotic (Paul and Clark, 1989). The majority of bacteria species are heterotrophic or organotrophic. These bacteria obtain their carbon from an organic substrate (e.g., toxic organic chemicals). On the other hand, autotrophic or lithotrophic bacteria are those which use inorganic matter, such as NH4+, Fe 2§ SO42", and/or CO 2 for food and energy. 21.2.3 Fungi
Fungi are also elementary forms of plant life in soils. In contrast with bacteria, the individual cells of fungi have a nucleus. Hence, they are eukaryotic organism. The cells may be linked together into filaments called hyphae which, collectively, are called mycelium (Paul and Clark, 1989). Fungi lack chlorophyll, the green substance essential for photosynthesis. Because of the absence of chlorophyll, fungi have developed into saprophytes, pathogens or parasites in nature. Saprophytes are plants which obtain their food from non living organic material. They contribute to nutrient cycling in soils. Pathogens are fungi that cause diseases in plants and animals. Parasites are organisms living in close association with another organism, deriving its nourishment from the host, and harming the host organism in the process. Many types of fungi are used in the food and pharmaceutical industries. Production of antibiotics and other chemicals are additional examples of the usefulness of fungi. Together with bacteria, they share the responsibility of decomposing organic chemicals in soils, breaking it down into simpler forms.
21.3
MICROBIAL M E T A B O L I S M AND G R O W T H
21.3.1 Microbial Metabolism
Metabolism is the sum of biochemical transformation that include interrelated catabolic and anabolic reactions. Catabolic reactions are exergonic; they release energy derived from organic and inorganic compounds. Aerobic respiration is an example of the catabolic reactions. Respiration is an energy generating process that involves the transfer of electrons through the electron transport system. The substrate is oxidized, 02 being used as the terminal electron acceptor. The electron donor may be an organic compound (e.g., oxidation of glucose by heterotrophic microorganisms) or an inorganic compound (e.g., oxidation of H 2, Fe(II), NH4 or S~ by chemoautotrophic microorganisms). Anabolic reactions (i.e., biosynthetic) are endergonic; they use the energy and chemical intermediates provided by catabolic reactions for biosynthesis of new molecules, cell maintainance,
562
BIOREMEDIATION
and growth. Photosynthesis is a process that converts light energy into chemical energy, using CO2 as a carbon source and light as an energy source. Light is absorbed by chlorophyll molecules, which are found in algae, plants and photosynthetic bacteria. Oxygenic photosynthesis consists of two types of reactions: (1) light reactions which result essentially in the conversion of light energy into chemical energy, and (2) dark reactions in which the energy produced in type 1 reaction is used to produce CO2. 21.3.2 Microbial Growth
The basic requirements for microbial growth are: (1) energy source, (2) carbon for the synthesis of new microbial cells, (3) inorganic elements (nutrients) such as nitrogen, phosphorus, sulfur, potassium, calcium, and magnesium, and (4) organic elements (growth factors) such as amino acids, purines, pyrimidines, and vitamins. A carbon or energy source is usually referred to as substrate.
Carbon Source Carbon is a basic building block of life. Organic matter and carbon dioxide are the most common sources of cell carbon for microorganisms. For most organisms, the bulk of net energyyielding or energy-consuming metabolic forces involves changes in the oxidation state of carbon. For example, when algae fix carbon dioxide as a carbohydrate, [CH20 ], carbon changes from the +4 to 0 oxidation state. CO 2 + H 20 -* [CH20]
+ 0 2
[21.1]
When algae die, bacterial decomposition results in the reverse reaction; energy is released and oxygen is consumed. In aerobic conditions, the main energy-yielding reaction of bacteria is oxidation of organic matter, expressed as:
02 + [CH20] -. CO 2 + H 20
[21.2]
This type of reaction provides bacteria and other microbes with the energy necessary for the growth and reproduction. Under reducing conditions (anaerobic), the oxygen content of the original organic matter is decreased, leaving relatively higher carbon content. Based on the source of cell carbon, microorganisms are classified into: (1) heterotrophs for which organic carbon is the carbon source, and (2) autotrophs for which carbon dioxide is the carbon source.
Energy Source The required energy may be supplied by light or chemical oxidation reaction. Based on the source of energy, microorganisms are classified into phototrophs and chemotrophs. These types are discussed below.
SOIL BIOMASS
563
Phototrophs: Phototroph microorganisms use light as the energy source. They are subdivided into: (a) heterotrophs, e.g., sulfur bacteria, and (b) autotrophs, e.g., algae and photosynthetic bacteria. Chemotrophs: Chemotroph microorganisms obtain their energy from oxidation of inorganic and organic compounds. They are subdivided into chemoautotrophs and chemohetrotrophs. Chemoautotrophs use carbon dioxide as a carbon source (carbon fixation) and derive their energy from oxidation of inorganic compounds such as NH4, NO2, H2S, Fe 2+, or H 2. Most of them are aerobic. Nitrifying bacteria are widely distributed in soils and water. They oxidize ammonium to nitrate. Sulfur-oxidizing bacteria use hydrogen sulfide (H2S), elemental sulfur (S~ or thiosulfate ($2032) as energy sources. They are capable of growth in very acidic environments (pH 2 or less). Iron bacteria such as Thiobacillusferrooxidans are acidophilic. They derive energy from oxidation of Fe 2+to Fe3+, and are, also, capable of oxidizing sulfur. Other iron bacteria oxidize ferrous iron at neutral pH (e.g., Sphaerotilus natans, Leptothrix ochracea, Crenothrix, Chlonothrix, and Gallionella ferruginea). Hydrogen bacteria (e.g., Hydrogenomonas) use H 2 as energy source and CO 2 as carbon source. Hydrogen oxidation is catalyzed by a hydrogenase enzyme. These bacteria are facultative chemoautotrophs, since they can grow also in the presence of organic compounds. Chemohetrotrophs are the most common nutritional group among microorganisms. They include protozoa, fungi, and most bacteria. The energy source is derived from oxidation of organic matter. Organic compounds serve both as energy source and carbon source.
Nutrients The principal inorganic nutrients needed by microorganisms are nitrogen, phosphorus, sulfur, potassium, calcium, magnesium, iron, sodium, and chloride. Minor nutrients of importance include zinc, manganese, selenium, copper, cobalt, molybdenum, nickel, vanadium, and tungsten.
Growth Factors In addition to inorganic nutrients, organic nutrients may also be needed by some organisms. The major organic nutrients (growth factors) are amino acids, purines, pyrimidines, and vitamins.
Enzymes Enzymes are large protein molecules composed primarily of amino acids twisted into complex shapes by peptide links and hydrogen bonding. They do not undergo structural changes following their participation in chemical reactions and, thus, can be used repeatedly. They lower the activation energy and increase the rate of biochemical reactions. Their role in metabolism is extremely complex. The substrate combines with active site of the enzyme molecule to form an enzyme-substrate complex (ES). A new product (P) is formed and the unchanged enzyme (E) is ready to react again with the substrate. The complexes (ES) allow the substrate to pass through the cell wall. Once inside the cell, intracellular enzymes (E) will further complex with the substrate to catalyse other reactions necessary to obtain energy and to build new cellular material. Different enzymes act in a sequential manner to transform organic pollutants to successively simpler compounds (e.g., hydrocarbons to alcohols).
564 21.4
BIOREMEDIATION MICROBIAL REACTIONS
As discussed in the preceding sections, soil microbial population is an array of organisms, with diverse enzyme systems capable of deriving energy from metabolism of organic and inorganic compounds. The toxic chemicals reaching the soil surface, as a result of disposal products or accidental spills, are subjected to many biological reactions which can affect their persistence and mobility in soils. 21.4.1 Biomass Growth
The growth of biomass depends upon the available energy and nutrient supply. Chemicals applied to soil may become a source of energy for the soil biomass. When a microbial population is offered an energy source, it rapidly increases in number and activity to utilize the energy (Keeney, 1983).
Table 21.1: Physico-chemical properties of selected carbon compounds Molecular Molecular Solubility Compound formula weight (mg/1) Naphthalene 2-Methyl Naphthalene 2-Naphthol
C10H8
128.12
CI0HTCH 3 CloH7OH
142.13 144.12
31.7
25 750
Table 21.2: Mineral salts medium constituents and their concentrations Concentration Salt (mmol/1) Salt NaH2PO4.4 H20 K2HPO4 (NH4)4SO 4
MgSO4.7H20 Co(NO3)2.6H20 AIK(SO4)2.12H20
6.4 12.99 8.33 0.395 0.001 0.001
log K~w 3.36 4.11 2.84
Concentration (mmol/1)
CuSO 4
0.001
ZnSO4.7H20 MnSO4.H20 FeSO4.7H20 Na2MoO4.2H20 Ca(NO3)2.4H20
0.01 0.01 0.01 0.002 0.01
Hibbeln (1996) conducted a study on the growth of Pseudomonas putida bacteria using naphthalene, 2-methyl naphthalene, and 2-naphthol as the sole carbon source. The results are shown in Figures 21.2(a) and (b). The physico-chemical properties of the carbon compounds are given in Table 21.2. A mineral salts medium, with composition given in Table 21.2, was used for all experiments. The mineral salts medium provides a buffered system able to compensate for the hydrogen ion concentration released during the degradation of the organic compounds. The Figures
MICROBIAL REACTIONS
565
depict the rate of Pseudomonas putida growth for 25 and 130 ppm concentrations of the three carbon compounds.
Figure 21.2 (a). Growth of Pseudomonas putida in mineral salts medium amended with 25 ppm concentration of naphthalene, 2-methyl naphthalene, and 2-naphthol.
Figure 21.2 (b). Growth of Pseudomonas putida in mineral salts medium amended with 130 ppm concentration of naphthalene, 2-methyl naphthalene, and 2-naphthol.
566
BIOREMEDIATION
The results indicate that Pseudomonas putida can grow, with naphthalene, 2-methyl naphthalene, and 2-naphthol as the sole carbon source. Similar results were reported by Guerin and Boyd (1991), and Cane and Williams (1982). The concentration of 130 ppm for naphthalene and 2methyl naphthalene was sufficiently above their respective solubilities of 31 and 26 mg/1, and resulted in their precipitation. The presence of precipitates clearly did not restrict rapid biomass increase, as shown in Figure 21.2(b). The concentration of 130 ppm for 2-naphthol appeared to prevent growth and was assumed to be toxic to this organism at elevated concentrations.
21.4.2 Aerobic Respiration Aerobic respiration, e.g., oxidation, occurs when adequate amounts of air (02) are present. In aerobic respiration, O2 acts as the electron acceptor. Aerobic bacteria reduce oxygen by the addition of protons (H § to form H20. Therefore, with adequate aeration, organic chemicals decompose into CO2 and H20 (Eq. [21.2]), and release the inorganic species. The carbon dioxide produced by aerobic decomposition of organic chemicals will react with soil water to form carbonic acid, thus increasing the soil acidity. An increase in soil acidity increases the dissolution of soil minerals, as discussed in Chapters 5 and 15. Such a dissolution of primary minerals is important in soil formation, soil fertility and plant nutrition. For example, calcium carbonate, CaCO3, the main component of lime stone, is insoluble in pure water H20. However, in water containing CO2, calcium carbonate is made soluble by its conversion into calcium bicarbonate, Ca(t-ICO3)2. The reaction, called carbonation, can be written as: 920
[21.3]
+ C O 2 --, 9 2 C 0 3
H2CO 3 + CaCO 3 -~ Ca(HC03) 2
[21.4]
The rate of formation of calcium bicarbonate depends on the partial pressure of the carbon dioxide. The solubility of calcium bicarbonate increases as the partial pressure of carbon dioxide increases. In addition, dissolved carbon dioxide also increases the solubility of sulfate-bearing minerals such as apatite. The residue from coal mines contains large amounts of solid iron sulfide, FeS 2 -- an energy source for autotrophic bacteria. In the presence of oxygen, autotrophic bacteria (e.g., Ferrobacillus or Gallionella) accelerate the oxidation process and produce sulfuric acid, as per the following reaction: 2FeS 2 + 2H20 +
702
-~
2Fe 2+ + 4H +
+
4S04
[21.5]
The extremely high acidity created by this reaction is toxic and has many undesirable effects on soil and the environment. The iron released remains soluble in the acid environment, and is rapidly oxidized into Fe 3+. In a slightly acidic to neutral environment, ferric iron precipitates as Fe(OH)3. These processes can be described by the following reactions:
MICROBIAL TRANSFORMATION OF INORGANIC COMPOUNDS 3 F e 2§ + 0 2
+
4H +
F e 3+ + 3 9 2 0
~
F e 3+ +
2H20 + e-
.-. Fe(OI-Z)3 +
3H +
567 [21.6]
[21.7]
The Fe(OH)3 produced is often deposited as an amorphous colloidal material. 21.4.3 Anaerobic Metabolism In anaerobic conditions, where aeration is poor, microbial decomposition of organic chemicals is very slow or inhibited. Anaerobic bacteria are those which capable of metabolism in the absence of oxygen. When 02 is deficient in soil air, the anaerobic bacteria use the bonded oxygen (e.g., 02 in nitrate, sulfate and carbon dioxide). The end products are H2, H2S, and CH4. In anaerobic reactions, CO2, NO3-, and SO42 act as electron acceptor (Stevenson, 1986). In anaerobic condition, nitrate tends to be reduced into nitrite, NO2-. This reaction, called nitrate reduction, can be written as follow:
2NO 3-
+ CH20-,
2N02- + H20
+ CO 2
[21.8]
In Eq. [21.8], nitrate acts as an electron acceptor. The nitrite formed is, however, toxic and if allowed to build up to sufficiently high concentrations, may inhibit further bacterial growth. Nitrate reduction may ultimately lead to denitrification, a process by which the nitrogen compound is reduced to N 2 gas, according to the reaction:
4NO 3-
21.5
+
5CH20 + 4H+-. 2N2 + 5C02 + H20
[21.9]
MICROBIAL TRANSFORMATION OF INORGANIC COMPOUNDS
21.5.1 Nitrogen Transformation Microbial facilitated reactions of nitrogen compounds include: (1) nitrification where ammonia (NH3) is oxidized to nitrate (NO3-), and (2) denitrification (reduction o f N O 3 and NO 2 to N2), which releases N2 to the atmosphere. These processes are discussed below.
Nitrification Nitrification is the conversion of ammonia into nitrate. It is known to occur in two steps (Alexander, 1965; Brady, 1990). In the first step, Nitrosomonas bacteria convert ammonia into
568
BIOREMEDIATION
nitrate: 2NH4+ + 302 -* 2 N O 2- + 2 H 2 0
+ 4H + + e n e r g y
[21.10]
It is immediately followed by the second step, in which the Nitrobacter bacteria transform the nitrite into nitrate: 2 N O 2-
+
0 2 ~
2 N O 3- + energy
[21.11]
Both of the nitrogen bacteria are aerobic. Each can utilize oxidizable inorganic materials as electron donors in oxidation reactions to yield needed energy for metabolic processes. NO3, produced in the second step as seen in Eq. [21.11 ], is the most soluble form ofN 2 in soils. Nitrate is an anion, and will not be adsorbed by negatively charged clay minerals, hence will not accumulate in soils. The high mobility of NO3 contributes to surface and groundwater pollution. Nitrification is affected by a number of soil conditions, such as aeration, temperature, moisture content, pH, and C/N ratios of the organic compounds.
Table 21.3: Bacteria capable of denitrification (Firestone, 1982) Microbe 9A l c a l i g e n s 9A g r o b a c t e r i u m 9A z o s p i r i l l u m 9 Bacillus 9 Falvobacterium 9 Halobacterium 9 Hyphomicrobium 9Paracoccus 9Propionibacterium 9Pseudomonas 9Rhizobium 9Rhodopseudomonas 9 Thiobacillus
Characteristics of the species Commonly isolated from soils Some specific plant pathogens Capable ofN 2 fixation, commonly associated with grasses Thermophilic denitrifiers Denitrifying species isolated Requires high salt concentration for growth Grows on one-carbon substrates Capable of both lithotrophic and hetrotrophic growth Fermentors denitrifiers Commonly isolated from soils Capable ofN 2 fixation Photosynthetic Generally grow as chemoautotrophs
Denitrification
Denitrification is the conversion of nitrates into N20 and N2 gas. Sometimes called enzymatic denitrification, it is a reduction process that is mediated by anaerobic bacteria in the absence of 02 (Cooper and Smith, 1963; Paul and Clark, 1989; Brady, 1990). A general list of bacteria capable of
MICROBIAL TRANSFORMATION OF INORGANIC COMPOUNDS
569
denitrification is given in Table 21.3 (Firestone, 1982). The sequence of reduction is shown in Figure 21.3, and may be presented as follows (Copper and Smith, 1963): NO 3Nitrate
<5
NO 2- <5 N2 0 <5 Nitrite Nitrous oxide
N2 Nitrogen gas
[21.12]
The release of nitrous oxide, N20, gas into the air contributes to the destruction of the ozone, 03, layer which protects the earth from the sun's ultraviolet radiation.
Figure 21.3. Products formed during denitrification in Melville loam soil at pH 7.8 (Cooper and Smith, 1963).
21.5.2 Phosphorus Transformations Phosphorus in soils exists in organic and inorganic forms. Humus, manure, and other types of non-humified organic matter are the major sources of organic phosphorus in soils. Some of the compounds in soil organic matter considered potential sources of phosphorus are phospholipids, nucleic acids and inositol phosphates. Microbial phosphates are produced by most organotrophic members of the soil microorganism population, such as bacteria actionmycetis, fungi, and protozoa nematodes. Microbial mineralization of organic phosphorus is strongly influenced by environmental parameters such as pH, temperature and soil moisture content. The products formed in the various stages of a microbially mediated transformation will have different properties, and their transport in soil media will be affected by the type of speciation, as shown in Figure 21.4. Inorganic phosphorus is derived mostly from the apatite minerals, which are accessory minerals in all types of rocks.
570
BIOREMEDIATION
I Waste Disposal I Microbial Phosphorus (Actionmycetis, Fungi, and ProtozoaNematodes)
# Mobile inorganic
Phosphorus form and mobility Mobile organic
Medium mobileorganic
Resistant organic
%
Aggregate organic
Type of extract
~ esin
NaHCO a
NaOH
HCI04
NaOHj
Figure 21.4. Mobility of products formed in various stages of a microbially mediated transformation.
In solution, phosphorus is present as either the primary, H3P04-, or secondary, HP042-, orthophosphate ion. The concentration of these ions in the soil solution depends on the pH. In acid soils, H2PO4 will be more dominant than HP042. At pH 6-7, both forms are equally represented in soil solution. At pH greater than 7, HP042 will be dominant and P043 exist in small quantity. The stability of these ions can be explained by the p K a values of orthophosphoric acid, H3PO 4 (Tan, 1994). Orthophosphoric acid undergoes the following dissociation reactions, each characterized by a specific pK, value: H3PO 4 "-" H '
+ H2P04
H2P04-,-, H + + HP042HPO42-,-, H + + PO43-
PKal -
2.17
PKa2 -
7.31
[21.13]
p K 3 - 12.36
where p K a value is the pH at which the compound dissociates into equal concentrations of anions and the original material. Soil pH or water pH is usually above 2.17 and way below 12.36. Therefore, the first and second steps of dissociation will occur in soils at natural conditions while the last step will occur only if the soil pH is above 12.36. This indicates that H2P04 and HP042 are the dominant species
MICROBIAL TRANSFORMATION OF INORGANIC COMPOUNDS
571
of phosphorus in soil solution. The effect of pH on phosphate adsorption on geothite is shown in Figure 21.5. The general trend for adsorption decreases as the pH increases, with bends in the curve near the pK values. This occurs because the amount of phosphate available for exchange and the amount of proton donor capable of neutralizing liberated hydroxide are maximum at the pK values.
Figure 21.5. Effect of pH on phosphate adsorption on geothite surface.
Phosphate adsorption is considered to occur by several mechanisms. Oxides and hydrous oxides of iron and aluminum tend to adsorb much larger quantities of phosphate than silicate minerals (Uhara and Gillman, 1981). At slightly acidic pH, phosphate is thought to be nonspecifically adsorbed on hydrous Fe 3+, A1 oxides, and clay minerals that exhibit a positive charge. The negatively charged surface of minerals such as hydrous manganese oxide at near neutral pH may become positive through the exchange ofH § for divalent cations such as Ba, Ca and Mg (Kawashima et al., 1986). Phosphate fixation, that is, formation of complexes or insoluble metal-phosphate compounds, such as variscite (A1PO4.2H20) and strengite (FePO4.2H20), contributes to phosphate adsorption by soil surfaces (Tan, 1994). 21.5.3 Metal Transformations
Metal transformations include oxidation-reduction of inorganic forms and conversion of metals to organic complex species, and the reverse conversion of organic to inorganic forms. Microbially mediated oxidation-reduction, described in Table 21.4 (Fuller and Warrick, 1985), is the most typical pathway for metal transformation.
572
BIOREMEDIATION
T a b l e 21.4: S o m e m i c r o b i a l t r a n s f o r m a t i o n o f i n o r g a n i c s u b s t a n c e s Element As
Microorganism 9 F. ferroxidant 9 Heterotrophic bacteria 9 M. lactilyticus
P h y s i o l o g i c a l activity As2S 3 oxidized to AsO3 3, AsO43" AsO33 oxidized to AsO43 AsO4 3" reduced to AsO3 3"
Cd
9 Desulfovibro
C d C O 3 + $042 -+- 8H + + 8 e + CdS + 4H20 + C032-
Cu
9 T. ferroxidans 9 F. ferroxidans 9 Desulfovibrio 9 M. latilyticus
CuzS + 4 H 2 0 --~ 2Cu 2+ + 6H + + HzSO 4 + 1 0 e CuS + 4H20 --~ Cu 2+ + 6H + + H2SO 4 + 8 e Cu 2+ and SO42 reduced to CuS; CuzS Cu(OH)2 + H + + e + C u O H + H20
Fe
9 T. ferroxidans, Ferrobacillus spp., and Gallionella. 9 Leptothrix ochracea, Sphaerotilus, Protoza, and algae, 9 M. lactilyticus, and B. circulans, 9 Desulfovibrio
Ni
9 T. ferroxidans 9 Desulfovibrio
Fe 2+ --~ Fe 3+ + e
Adsorption and precipitation
Fe 3+ + e + Fe z+ Fe 3+ + SO42 + 8H + + 9 e --~ FeS + 4H20 NiS + 4H20-~ Ni 2+ + 8H § + SO42 + 8 e NiCO 3 + $042 nt- 8H + + 8 e -~ NiS + 4H20 + C032 Ni(OH)2 + SO42 + 10H + + 8 e -~ NiS + 6H20
9 Thiobactericeae, Chlorobacteriaceae, Beggiatoaceae, and Leucothrix 9 Bacteria, actinomycetes, 9 All microorganisms
Polysulfides reduced to thiosulfate and sulfide S o + 2H § + 2 e -~ HzS
Se
9 M. selenicus 9 M. lactilyticus 9 Neurospora
HzSe + 4H20 ~ SeO42 + 10H + + 8 e Se ~ + H + + 2 e + S e l l HSeO3 + 5H + + 4 e -~ Se ~ + 3H/O
Zn
9 T. ferroxidans
ZnS + 4H20 --~ Zn 2+ + 8H + + $042 + 8 e
Metal chelates
9 Heterotrophic microorganisms
Oxidation o f chelating agent with precipitation o f metal.
HzS + S O+ 2H + + 2e, H2S + H 2 0 + SO42- + 10H + + 8 e
U n d e r acidic conditions, metallic iron (Fe ~ readily oxidizes to the f e r r o u s state (Fe 2§ but at p H v a l u e s g r e a t e r t h a n 5 it is c h e m i c a l l y o x i d i z e d to Fe 3+. T h i o b a c i l l u s f e r r o x i d a n t m e d i a t e s this r e a c t i o n in an acidic e n v i r o n m e n t . B a c t e r i a l o x i d a t i o n o f M n 2+ is m i c r o b i o l o g i c a l l y m e d i a t e d o n l y
MICROBIAL TRANSFORMATION OF ORGANIC COMPOUNDS
573
in neutral and acidic environments. A number of soil bacteria and fungi can oxidize manganese ions. Mercury does not remain in metallic form in an anaerobic soil environment. Microorganisms transform metallic mercury to methyl mercury (CH3-Hg+) and dimethyl mercury (CH3-Hg+-CH3), which are volatile and adsorbable on soil organic matter and soil organisms.
21.6
MICROBIAL TRANSFORMATION OF ORGANIC COMPOUNDS
Organic compounds can be biotransformed by microorganisms in several ways. The best known and understood method involves the use of the organic compound as aprimary substrate, or source of energy and carbon for microbial growth. Primary substrates can typically be mineralized to inorganic constituents. Organic compounds can also be transformed via secondary utilization, in which microorganisms do not derive sufficient energy or carbon for net growth from compound oxidation. There are several types of secondary utilization. Microorganisms may not derive energy or carbon because the concentration of the target organic is too low. Organic compounds that cannot serve as carbon or energy sources for bacteria may also be transformed via cometabolism, which occurs only when a primary substrate is present to support growth. Finally, organic compounds can be transformed while serving as electron acceptors and are reduced in the process. Such reactions may or may not be cometabolic as well. Thus, transformation reactions may be mediated by microorganisms (biotic) as well as by chemical processes (abiotic). In general, biotic transformation reactions are faster than abiotic reactions, but this is not always the case. The basic processes involved in the microbially mediated transformation of toxic organic molecules are discussed below.
21.6.1 Biodegradation Biodegradation is a process in which a toxic organic molecule serves as substrate for microbial growth. In this case, the organic molecules are used by one or more interacting microorganisms and metabolized into CO2 and inorganic molecules. In this way, the microorganisms obtain their requirements for growth and the toxic organic molecules are completely decomposed without losing metabolities in the soil environment. There are two general types ofbiodegradation, namely, mineralization and biotransformation. Mineralization occurs when an organic compound is converted by living organisms to mineral (nonorganic) end products. Energy is produced during mineralization. In the following example, reported by Hibbeln (1996), the Pseudomonas putida bacteria used the carbon substrate for microbial growth while the organic chemicals were mineralized. [~4C] naphthalene, [~4C] 2-methyl naphthalene, and [~4C] 2-naphthol were used in the study. The concentration of each organic compound was 25 ppm. Two series of experiments were conducted. In the first series, mineral salt mixture (MSM) was amended with each organic compound. The composition of MSM is given in Table 21.1. The MSM provides a buffered system which is able to compensate for the hydrogen ion concentration released during the degradation of an organic compound. The mineralization results, of the first series, are shown in Figure 21.6(a). In the second series, 2% kaolinite clay and MSM were amended with each organic compound. The mineralization results are shown in Figure 21.6(b). The Figures depict the % mineralized of the three carbon compounds with and without the presence of kaolinite.
574
BIOREMEDIATION
Figure 21.6(a). PAH mineralization by Pseudomonas putida in mineral salt solution amended with 25 ppm naphthalene, 2-methyl naphthalene, and 2-naphthol.
Figure 21.6(b). PAH mineralization by Pseudomonas putida in 2% kaolinite and mineral salt solution amended with 25 ppm naphthalene, 2-methyl naphthalene, and 2-naphthol.
The results, shown in Figure 21.6(a), indicate that Pseudomonas putida is able to mineralize over 50% of naphthalene, 2-methyl naphthalene, and over 35% of 2-naphthol. In the presence of kaolinite (shown in Figure 21.6(b)), however, the rate of mineralization as well as the total
MICROBIAL TRANSFORMATION OF ORGANIC COMPOUNDS
575
mineralization decreased. The reason is the reduced availability of phosphate trace elements since they are adsorbed onto the kaolinite surface. B i o t r a n s f o r m a t i o n occurs when a parent organic compound is not completely mineralized, and a portion is converted into other organic compounds. An example is the reductive dechlorination of trichloroethylene to produce dichloroethylene.
CClzCHCI
+ H + + 2e-~
CHCICHCI
+ CI-
[21.14]
Some general rules can be made about microbial transformation of organic compounds. While there are many exceptions, such general rules are useful for understanding the environmental fate of organic compounds (Sawyer, 1994): Simple carbohydrates and amino acids are very biodegradable; (1) (2) Fats and oils may be difficult to degrade because of solubility limitations; Hydrocarbons are more difficult to oxidize than alcohols, aldehydes, or acids; (3) Ketones are more difficult to degrade than aldehydes; (4) Ethers are difficult to degrade; (5) (6) Tertiary and quaternary carbons and nitrogens are much more difficult to degrade than the primary or secondary counter parts; (7) Hydrolysis of esters, amides, and carbamates is generally fast and easily carried out by microorganisms; and (8) Addition of a chlorine atom or a nitro group to a benzene ring increases its resistance to biodegradation. Microbiologists have isolated a vast array of microorganisms that have the ability to degrade organic toxicants. A list of some of these microbial species is shown in Table 21.5 (Kumaran and Shivaraman, 1988). Since several xenobiotics are aromatic molecules, the microorganisms must be capable of cleaving the aromatic ring in these compounds. The biochemical pathways have been elucidated for many of the compounds. The fission of the aromatic ring is mediated by enzymes called monooxygenases, which include molecular oxygen into the ring prior to cleavage. 21.6.2 Cometabolic Transformation
Cometabolic transformations include the degradation of toxic organic molecules by microorganisms which grow at the expense of a substrate other than the toxic organic one, without the use of the latter as an energy source. This process, in which enzymes involved in catalyzing the initial reaction are lacking in the substrate, may lead to the accumulation of intermediate products which in many cases are more toxic than the parent organic molecules. Generally, microorganisms in soil completely biodegrade an organic molecule in soil pore water to inorganic constituents through mineralization. During the process, the microorganisms convert some of the carbon to cell constituents. The assimilation results in an increase in the number of biomass of microorganisms. The simplified transformation is:
576
BIOREMEDIATION
biodegradation Organic molecule
CO 2 + 920
+
[21.15]
water / soil inorganic minerals + ENERGY-~ microbial cells + (biomass increase)
Table 21.5: Some microorganisms involved in the biodegradation of organic compounds. Organic pollutants
Organism
Benzoates and related compounds
Arthrobacter, Bacillius spp., Micrococcus, Moraxella, Mycobacterium, P. putida, and P. fluorescens.
Hydrocarbons
Escherichia coli, P. putida, P. aeruginosa, and Candida.
Pesticides DDT Linurin 2,4-D 2,4,5-T Parathion
P. aeruginosa. B. sphaericus. Arthrobacter and P. cepacia. P. cepacia. Pseudomonas spp. and E. coli; P. stutzeri and P. aeruginosa.
Phenolic compounds
Achromobacter, Alcaligenes, Acinetobacter, Arthrobacter, Azotobacter, Bacillus cereus, Flavobacterium, Pseudomonas putida, P. aeruginosa, and Nocardia. Candida tropicalis, Debaromyces subglobosus, and Trichosporon cutaneoum. Aspergillus, Penicillium, and Neurospora.
Surfactants
Alcaligenes, Achromobacter, Bacillus, Citrobacter, Clostridium resinae, Corynebacterium, Flavobacterium, Nocardia, Pseudomonas, Candida, and Cladosporium.
In Eq. [21.15], the organic substrate is mineralized in a growth-linked process. Chlorinated hydrocarbons such as DDT, aldrin, and heptachlor are believed to be subject to cometabolic transformation rather than mineralization. In this situation, microorganisms use these chemicals as nutrients for energy but not to sustain growth. Accordingly, the microbial populations utilize another
MICROBIAL TRANSFORMATION OF ORGANIC COMPOUNDS
577
cell-supporting substrate for growth (Alexander, 1981). This phenomena is characterized by a slow rate of substrate modification. A schematic representation of the cometabolic transformation and mineralization processes is given in Figure 21.7.
T
1
Cometabolized Chemical
t-
O
r
o
c-
m t-
[3.
Cometabolizing_
0 13_
O
o
population
Time
"~
Figure 21.7. Illustration of population changes and metabolism of a chemical modified by mineralizing and cometabolizing soil populations.
21.6.3 Polymerization Polymerization is a process in which toxic organic molecules undergo microbially mediated transformation by oxidative coupling reactions. Microbially mediated polymerization leads to the incorporation of xenobiotic chemicals, which are foreign to natural biota, into soil organic matter. Polymerization can be illustrated by considering phenol oxidation by kaolinite and montmorillonite (Larson and Hufnal, 1980). The proposed phenol oxidation mechanism is that the presence of ions such as Zn, A1, Fe and Cu on the surface or within the clay structure can transfer an electron to an adsorbed oxygen molecule. The oxygen molecule (02) along with a proton (H) from the silicate layer are released from the surface as a hydroperoxyl radical (~ The hydroperoxyl radical is a powerful oxidizing agent. In the presence of phenols (Ar) the hydroperoxyl radical acts as an electron acceptor and oxidation occurs. The newly formed product is called phenoxy radical (Ar.). Phenoxy radicals are coupled together via oxidative coupling reaction to form phenol dimers and oligomers --- formation of polymers. The polymerization process can be illustrated by the following three steps: Step 1" Step 2: Step 3:
~ 03 SiOH + 02
+
Ar + .OOH Aro + Ar~
+ +
SiO4-+ .OOH Arphenol dimers and oligomers
578
BIOREMEDIATION
where (-- 03 SiOH) represents the silica sheet of clay, (-- SiO4) is a deprotonated silicate complex, (oOOH) is a hydroperoxyl radical, (Ar) is phenol molecule, and (Aro) is phenoxy radical molecule. A schematic representation of phenol oxidation and dimer formation is shown in Figure 21.8.
(1)
OH
Clay mineral acting as an electron acceptor (2)
OH
Phenol molecule releasing an electron
OH
(
OH
OH
Phenoxy radical OH
r
Phenoxy radicals coupling together
F o r m a t i o n of a
phenol dimer
Figure 21.8. Schematic diagram showing phenol oxidation by clay (Desjardins, 1996).
Polymerization can be further elaborated in view of the study by Desjardins (1996) to determine the ability of saturated Na § Ca 2+-, and Fe z+- montmorillonite clays to oxidize 2,6 dimethyl phenol (C6H4OHCH3) at various soil pH conditions. The experimental results, shown in Figure 21.9, depict the relationship between the soil pH and the relative mass ion abundance (RMIA) for 2,6 dimethyl phenol dimer (C~4H~402). The experimental results indicate that Na § Ca2+-, and Fe3+- montmorillonite clays can oxidize 2,6 dimethyl phenol to dimers. The presence of two slightly activating groups (-CH3) and a strongly activating group (-OH) makes the ring reactive enough to donate electrons to the clay's Lewis acid sites (i.e., exchangeable cations, or incomplete coordinated aluminum within the structural layer lattice). The phenolic polymerization pattern as a function of soil pH follows a "U" shaped curve due to the changes in clay structure with pH. At pH less than 4, clay soil is acidic and becomes an efficient catalyst for the oxidation of organic compounds. Partial dissolution of aluminum occurs, leading to an increase in aluminum concentrations at the edges of the clay mineral contributing to an increase in soil catalysis (oxidation capacity). At neutral pH, phenolic oxidation process is most likely caused by the unaltered aluminum concentration, an electron transfer to structural Fe3§ Fe203, and high valence exchangeable cations (Solomon and Hawthorne, 1983). A high pH value is expected to dissolve both alumina and silica sheets and, hence, form aluminum, silica, and iron complexes. These complexes are active electron acceptors which contribute to the increased ability
MICROBIAL TRANSFORMATION OF ORGANIC COMPOUNDS
579
of clays to oxidize phenol at high pH. Microbially mediated polymerization leads to an increase in hydrophobic bonding between the newly formed compounds and soil organic matter.
Figure 21.9. Relative mass ion abundance (RMIA) of 2,6 dimethyl phenol (C6H4OHCH 3 ) for saturated Na § Ca 2+-, and Fe 3§ montmorillonite clays.
21.6.4 Microbial Accumulation
Microbial accumulation is defined as the transformation of an element in a microbial tissue from inorganic to organic form, rendering the element not readily available to other organisms. The incorporation of elements (including heavy metals) into biological tissues results in their fixation. The rate of accumulation depends on the properties of the microorganisms and toxic organic chemicals. Tobin (1986) investigated the uptake of Pb, Cu, Zn, and Cd by Rhizopus arrhizus biomass at 25 ~C and pH 4. Heavy metals, in nitrate form, were mixed with Rhizopus arrhizus biomass in batch test. The chemical composition of the biomass was 11% phosphate, 40% protein, 44% chitosan, and 5% carbohydrates. Many of these constituents contain functional groups which will contribute to heavy metal uptake. Such functional groups include the chitosan nitrogen-containing group, the carboxyl and amino groups, peptide bonds of proteins, and the phosphate groups. The biomass was obtained by culturing Rhizopus arrhizus strains in a liquid medium consisting of 0.5% peptone, 0.5% neopeptone, 0.1% sucrose, 0.1% KH2PO4, 0.1% NaNO3, and 0.05% MgSO4.7H20, for 70 hours at 25 ~ C with continuous agitation at 200 rpm on a rotary shaker. The concentrations indicated above are in terms of weight/volume. The adsorption isotherm results are shown in Figure 21.10. The results indicate that the biomass has high affinity for Pb adsorption than Cd, Cu, and Zn.
580
BIOREMEDIATION
At low input heavy metal concentration, the order of biomass affinity for heavy metal adsorption is Pb > Cd > Cu > Zn while at high concentration the order is Pb > Zn > Cu = Cd.
Figure 21.10. Adsorption isotherm of lead, copper, zinc, and cadmium by Rhizopus arrhizus biomass.
Cation uptake by the biomass occurs through a process which is primarily a reversible association of the cations with various functional groups within the biomass. Phosphate and carboxyl moieties appear to be the principal groups involved. These primary interactions may be augmented by association with hydroxyl and other groups. Electrostatic attraction to negatively charged functional groups may further augment cation uptake. Most carboxylate groups would be protonated at pH 4. The protons could be displaced relatively easily by the metal cations according to the reaction: -
COOH
+
M2*~
-
COOM
+
+
H ~
[21.16]
Most of the phosphate groups present as mono-ionic with negative charge above pH 3. Phosphates are recognized as potent ligands for metal and could contribute appreciably to both the electrostatic and coordination binding of metals from solution. Sulfhydral groups of the protein fraction of the biomass constitute another metal binding site. Hydroxyl groups are weak bases and would form only weak bonds with metals. Metal uptake by Rhizopus arrhizus cell wall can be described by a following two-step mechanism comprising of an initial binding reaction between the soluble metal and wall reactive sites, followed by the formation of inorganic crystaloid deposits. As discussed, the wall reactive sites are carboxyl, phosphate, and hydroxyl groups.
BIOTRANSFORMATION REACTIONS
581
21.6.5 Nonenzymatic Transformation Nonenzymatic transformation is an indirect process due to microbial activity. It occurs as a result of microbially-induced changes in environmental parameters such pH and oxidation-reduction. These reactions are referred to as biotransformation reactions which discussed below.
21.7
BIOTRANSFORMATION REACTIONS
The most important biotransformation reactions are hydrolysis, and oxidation- reduction. These are discussed in the following sections.
21.7.1 Hydrolysis Hydrolysis is the most common substitution reaction catalyzed by microorganisms. In this process, water serves as a nucleophile and attacks an organic bond. Other nucleophiles are OH, NO3 , S Q 2-, HS, HCO3, and HPO42. There is no change in the oxidation state of the organic chemical during these transformations. Organic chemicals that have been hydrolyzed include halogenated aliphatics, esters, amides, carbamates, and phosphoric esters. Hydrolysis of halogenated aliphatics, for example, can be demonstrated by (Sawyer et al., 1994): R - C H z X + H 2 0 - . R - C H 2 0 H + H++ X R - C H z X + H S - -~ R-CH2SH + X -
[21.17]
21.7.2 Oxidation Oxidation involves the transfer of electrons from a reduced substance, termed the electron donor, to an oxidizing material, termed the electron acceptor. Generally, we think of the electron donor as being the "food" for the organism. Microbial oxidation is a process where electrons are released through enzyme-catalyzed reactions. An electron acceptor, such as oxygen, nitrate and sulfate, is then required to complete the process. Some organic compounds may serve as primary substrate and be mineralized. Such compounds include benzene, toluene, phenol, chlorobenzene, dichloromethane, nitrotoluene, and hydrocarbon. The most widely used microbial transformation involves the oxidation of organic compounds by the introduction of a hydroxyl group derived from molecular oxygen. Such reactions are usually catalyzed by nonspecific monooxygenase and dioxygenase enzymes that insert either one or both atoms of molecular oxygen into the substrate. Such enzymes catalyze the first step in the mineralization of the relatively inert hydrocarbons. Microbial cells grown on hydrocarbons, such as methane and toluene, can oxidize a wide range of organic compounds. Hydrocarbons are oxidized by certain bacteria under aerobic conditions. The oxidation proceeds through several steps. The first step is very slow biologically and involves conversion to alcohols, with attack occurring on carbon atoms as:
582
BIOREMEDIATION 2CH3CH2CH 3 Hydrocarbon
+ 0 2
2CH3CH2CH20H Alcohol
--~
bacteria
[21.18]
Through additional oxidative steps, microorganisms convert the hydrocarbon to carbon dioxide and water, and derive energy in the process:
C H 3C H 2C H 3 + 5 0
bacteria -~-
2
[21.19]
3C02 + 4H20 + energy
21.7.3 Reduction
Through reduction reactions, xenobiotic (synthetic) compounds may serve as electron acceptors. The electron can originate from a wide variety of organic and inorganic electron donors, which may serve as primary substrates. Generally, oxygen, nitrate, sulfate, and carbon dioxide serve as terminal electron acceptors. However, halogenated compounds such as trichloroethylene (CHCICC12) can be reduced to produce dichloroethylene (CHC1CHC1), vinyl chloride (CHC1CH2) and eventually ethene (CH2 CH2), as shown in Figure 21.11. Generally, the more chlorine atoms a compound has, the stronger its tendency to serve as an electron acceptor.
H§ H
N /c~c
Cl
/ N
CI-
CI
H
~ H
N /c~c
CI
/ \
H
H H+
D ich loroe th yle ne H
\
/C~CN
Cl
f
CI-
Vinyl chloride
CI
H
Cl H§
Tric h Io ro e th yle n e
N
/C~C Cl
H
/
\
H
H
Ethene
Figure 21.11. Anearobic reductive dehalogenation of trichloroethylene to produce dichloroethylene, vinyl chloride, and ethene.
FACTORS AFFECTING MICROBIAL TRANSFORMATION 21.8
583
FACTORS AFFECTING MICROBIAL TRANSFORMATION
Several factors influence the microbial transformation of organic compounds in soils. These factors, which can be categorized as chemical and environmental, are discussed in the following sections. 21.8.1 Chemical Factors
The factors that chemically influence the microbial transformation of organic compounds are: (1) chemical structure of the compound, (2) soil organic matter content, (3) soil pH, (4) other compounds or ions present, (5) concentration of the added compounds, (6) acclimatization period, and (7) type of clay mineral. Chemical Structure The rate of microbial transformation of organic compounds in soils depends to a large extent on the chemical structure of the compound. Since the transformation is mediated by enzymes, differences in the transformation rate could be attributed to: (1) levels of enzymes capable of catalysing the reaction, and (2) differences between the substrate structure and the enzyme. The chemical structure parameters of importance are: (1) type of functional group, (2) position of functional group, and (3) number of functional groups. These parameters are discussed below.
Figure 21.12. Effect of nature of functional groups in the aromatic ring on microbial transformation of substituted benzene in soil suspension.
584
BIOREMEDIATION
(1) Type of Functional Group: The effect of functional group type on the transformation of substituted benzene in a soil suspension is shown in Figure 22.12 (Alexander and Aleem, 1961). The Figure indicates that benzene rings bearing nitro (NOR) and sulfo (SO3H) groups are relatively resistant to microbial transformation. Based on this result, the microbial transformation selectivity order is: COOH
> OH > NH 2 > OCH 3 > S03H
> NO 2
[21.20]
(2) Position of Functional Group: The effect of position of functional group (Figure 21.13a) on the rate of microbial transformation of N-substituted naphthalenes (1-amino- naphthalene, 2amino-naphthalene, and 1-amino-2-methyl-naphthalene) is shown in Figure 21.13b (A1-Bashir, 1994).
Figure 21.13a. Position of the functional group in N-substituted naphthalenes.
Figure 21.13b. Transformation rate of N-substituted naphthalenes as a function of the adsorbed concentration in clay loam soil.
FACTORS AFFECTING MICROBIAL TRANSFORMATION
585
Figure 21.13b depicts the relationship between the rate of transformation and soil adsorbed concentrations. The soil used was natural soil characterized as clayey loam with 3.96% organic matter, 1.6% organic carbon, pH 7.27, and cation exchange capacity of 19.94 meq/100 g soil. The results indicate that the rate of transformation is dependent on the position of the functional group. The NH 2 functional group, for example, is positioned at two different positions in 1-aminonaphthalene and 2-amino-naphthalene, resulting in a higher rate of transformation for 1-aminonaphthalene than for 2-amino-naphthalene. (3) Number of Functional Groups: The effect of the number of functional groups can easily be seen in Figure 21.13b by observing the rate of transformation of 1-amino-naphthalene and 1amino-2-methyl-naphthalene. 1-amino-naphthalene has only one functional group (NH2) whereas 1-amino-2-methyl-naphthalene has two (NH2 and CH3). The results clearly indicate that as the number of functional groups increases, the rate of transformation decreases.
Soil Organic Matter The effect of soil organic matter on the transformation of organic pollutants in soils is well documented. In some instances, soil organic matter hastened microbial transformation rate while in others decreased it. This is due to adsorption of organic pollutants by soil organic matter, hence rendering them less vulnerable to enzymatic attack. The properties of organic matter that are relevant to adsorption are: (1) the proportions of humic acid, fulvic acid, and humin, (2) the presence of active groups such as carboxyl, hydroxyl, carbonyl, and amino, and (3) high cation exchange capacity and surface area. Detailed adsorption mechanisms are discussed in Chapter 5.
Figure 21.14. Effect of soil organic matter content on transformation rate of quinoline as a function of the adsorbed concentration in soils.
586
BIOREMEDIATION
Such conflicting results could be attributed to the difficulty in determining whether soil organic matter facilitates microbial transformation by supplying a nutrient source or an energy source for the microbes, or acts as a cometabolite. In general, when organic pollutants are adsorbed by soil organic matter, they may not be metabolized or transformed by microbes if there are other sources of energy. However, when other sources of energy are available, along with a significant soil organic matter, cometabolism could occur to a substantial degree. A1-Bashir (1994) investigated the rate of microbial transformation of quinoline in two natural clay soils with varying amounts of soil organic matter. The results, shown in Figure 21.14, clearly indicate that the rate of microbial transformation decreases as the soil organic matter content increases.
Soil pH As shown in Figure 21.15, soil pH has an important impact on microbial transformation in soils. Microbial transformations at neutral pH are more favourable than acidic and alkaline conditions. In general, bacterial growth is favoured in near-neutral conditions. However, fungi show higher tolerance for pH variations (Valentine and Schnoor, 1987). Soil pH may also affect the amount of sorption of organic pollutants by varying the supply of hydrogen ions that compete for adsorption sites, and by influencing the charge of variable charge minerals (e.g., kaolinite and oxide minerals).
Figure 21.15. Effect of pH on microbial transformation of 1-amino-naphthalene and 1-nitronaphthalene in soil after 200 days.
FACTORS AFFECTING MICROBIAL TRANSFORMATION
587
Other compounds or ions present In the presence of other compounds or ions in the soil, microbial transformation rates of organic pollutants may be enhanced, depressed, or unaltered. For instance, glycerol inhibits the transformation of glyphosate (N-[phosphonomethyl]-glycine) in soils (Tate and Alexander, 1974) while transformation of 2,6-dichloro-benzamide is enhanced by ethanol, benzamide and glucose (Fournier and Salle, 1975). Smith (1979) found that the presence of dicamba (3,6-Dichloro-2methoxy-benzoic acid) and dichloroprop (2-[2,4-Dichlorophenoxy]-propionic acid) has no effect on the transformation rate of 2,4-D (2,4-Dichlorophenoxy acetic acid) in soil. Microorganisms require certain metallic elements for growth and function. These include, the bulk elements Na, K, Mg, and trace elements such as Mn, Fe, Cu, Zn and Mo. All heavy metals are capable of inhibiting microbial activity (Barkay et al., 1992). Metal ions are unique in that over a rather narrow concentration range, their status can change from an essential growth-promoting element to a toxin.
Figure 21.16. Naphthalene mineralization in 2% kaolinite and mineral salt solution from an initial concentration of 130 ppm by Pseudomonas putida.
Hibbeln (1996) investigated the mineralization of [14C] naphthalene by Pseudomonas putida bacteria. Naphthalene at a concentration of 130 ppm was placed in a mixture containing 2% kaolinite, mineral salt solution, and cadmium nitrates. The mineral salt solution provides a buffered system able to compensate for the hydrogen ion concentration released during the degradation of the organic compounds. The composition of the mineral salt solution is given in Table 21.2. The experimental results are shown in Figures 21.16. Hibbeln (1996) concluded that the results indicate that cadmium has no observable impact on the mineralization activity of Pseudomonas putida. The lack of cadmium impact can be attributed to cadmium complexation with phosphate present in the mineral salt solution. The slight increase in the total percent mineralization, in the presence of a final
588
BIOREMEDIATION
cadmium concentration of 2.12 mmole/1, may have been an indirect effect. Those microorganisms amended with a high cadmium concentration experienced a slight drop in pH, creating a more favourable pH range for growth.
Concentration of the added compounds In heavy metal polluted sites, it is reported that fungal populations are increased (Frostegg~rd et al., 1993). Fungi tend to exhibit greater tolerance toward heavy metals (Hattori, 1992), and acidic conditions (Farago and Mehara, 1993; Doelman and Haanstra, 1984). The predominance of fungal growth under acidic or high metal conditions does not always occur, being significantly influenced by such factors as soil type and heavy metal type. Some metals, e.g., Ag § and Hg 2+, are known to inhibit fungal growth (Cornfield, 1977). Several studies have shown that systems exposed to elevated heavy metal concentrations show a superiority towards Gram-negative bacteria (Barkay et al., 1985; Farago and Mehara, 1993; Doelman and Haanstra, 1979). Gram-negative bacteria have shown to be more metal- tolerant than Gram-positive bacteria in soils with low levels of metal pollution (Silver et al., 1982; Frosteggtrd et al., 1993). In the study by Capone et al. (1983), it was found that the chlorides of rig, Pb, Ni, Cd, and Cd completely inhibited sulfate-reducing bacteria.
Acclimatization period The observed lag period, in laboratory experiments, between the introduction of organic pollutants into the system and the initiation of transformation may reflect such factors as: (1) time for enzyme activation to occur, (2) time for genetic mutation to occur, (3) preferential utilization of other organic compounds, (4) the time necessary for the small microbial population to increase in size to degrade observable quantities of the organic compounds, and (5) preferential utilization of the mineralizable fraction of the soil organic matter over organic pollutants (Mihelcic and Luthy, 1988). Pre-exposure of the system to the same or other compounds (e.g., benzene, PAH) will result in a sharp decrease or elimination of the acclimatization period. It was reported (Bauer and Capone, 1988) that the initial mineralization rate and the total amounts of anthracene and naphthalene mineralized were a direct function of the pre-exposure concentration, with the highest rates occurring at the greatest exposure concentration.
Type of clay minerals The type of clay mineral is directly related to the adsorption capacity of clays. As the cation exchange capacity increases, so does the adsorption capacity. Adsorption, in turn, affects the microbial transformation of organic pollutants. As adsorption increases, the availability of organic pollutants for microorganisms decreases because: (1) There is a physical barrier, or the enzyme involved is unable to form a substrate-enzyme complex with the adsorbed pollutants; (2) The sorption site is not in close enough proximity to a microorganism or enzyme; and (3) The adsorbed pollutants are not concentrated in an area where growth of the microbes occur.
FACTORS AFFECTING MICROBIAL TRANSFORMATION
589
21.8.2 Environmental Factors
The environmental factors that affect the microbial transformation of organic pollutants are: (1) moisture, (2) temperature, (3) aeration, and (4) depth of application. These factors are discussed below.
Moisture The optimum soil moisture content for maximum microbial transformation usually occurs at the upper limit of soil moisture, defined as the level at which the amount of water occupying the soil pore species does not interfere with soil oxygen exchange with the atmosphere. Usually this is just slightly below the normal field capacity. Field capacity is defined as the amount of water remaining in a well-drained soil when the downward flow velocity into unsaturated soil has become small. In a measurable quantity, this can be determined as the amount of water that the soil can hold at 0.1 to 0.33 atm. Since soil pore spaces are occupied by air and water, the soil is considered aerobic with oxidation occurring via aerobic microorganisms. When soil water fully occupies the pore spaces, anaerobic conditions determine the microbial transformation processes. Microbial transformation of organic pollutants is affected by moisture in one of three ways: (1) Increasing microbial transformation rate with increasing moisture up to the level just below the saturation; (2) Decreasing microbial transformation rate with increasing moisture; and (3) Increasing microbial transformation rate in saturated soil for compounds acted upon by obligate anaerobes (bacteria that can only use electron acceptors other than oxygen). For example, methane-producing bacteria are obligately anaerobic. Temperature Temperature affects microbial transformation rate of organic pollutants in two ways: (1) by hastening abiotic reactions, and (2) through increased microbial activity. The first has no maximum temperature limit whereas the second is generally effective up to 45 ~ C, beyond which microbial activity declines and then stops. For example, Bauer and Capone (1988) found that mineralization of anthracene and naphthalene increased up to 4.6 times when temperature was increased from 10 to 30 ~ C in aerobic sediment slurries. Aeration Microbes are sensitive to oxygen stress, their sensitivity depending on whether they are aerobic or anaerobic microbes. In general, some organic pollutants transform faster under aerobic conditions while others transform more readily under anaerobic conditions. For example, benthiocarb degrades equally well either in unsaturated or saturated soils whereas chlorinated hydrocarbons are reported to transform faster in saturated soils. Some organic compounds containing benzene rings transform faster under anaerobic conditions, and hasten the breaking of the C-C bonds in the ring. Healy and Young (1978) reported that during anaerobic fermentation of phenol, the aromatic ring is cleaved, forming CO2 and CH4 as follows: C6960 + 4 H 2 0 - ~
-25
CO 2 + --~ 7 CH4
[21.211
590
BIOREMEDIATION Depth of application
Depth of placement has been observed to affect the microbial transformation of organic pollutants. Deep placement generally reduces transformation rate compared to shallow placement. This is most likely due to decreased microbial and soil organic matter content at lower depths. However, the possibility of photo-transformation and volatilization for shallow or surface applications must not be overlooked.
21.9
K I N E T I C S OF M I C R O B I A L T R A N S F O R M A T I O N
The rate of growth of microbial cells can be defined by the following relationship: [21.221
rg : S g c m
where rg is rate of bacterial growth (mass/volumextime), S~ is specific growth rate (1/time), c m is concentration of microorganism (mass/volume). The substrate limited growth can be defined using the following expression proposed by Monod (1949): CS
[21.23]
Sg - Sg m Csh + r
where Sgmis maximum specific growth rate (1/time), c, is concentration of growth limiting substrate in solution (mass/volume), csh, also known as half-velocity constant, is substrate concentration at one-half the maximum growth rate (mass/volume). Substituting Eq. [21.23] into Eq. [21.22], the growth rate can be defined as: C m Cs rg = Sg m
[21.24]
Csh + Cs
During the reaction process, a portion of the substrate is converted to new cells and a portion is oxidized to inorganic and organic end products. The relationship between the rate of substrate utilization and the rate of growth is given by:
rg
mnc
rs
[21.25]
mcs
where mnc is the mass of the newly formed cells, me,,.is the mass of the consumed substrate, r s is substrate utilization rate (mass/volume• Substituting Eq. [21.24] into Eq. [21.25], the rate of substrate utilization may be expressed
SUMMARY AND CONCLUDING REMARKS
591
as"
Cm Cs )
[21.261
Csh + Cs
By defining rsmax , the maximum rate of substrate utilization per unit mass of microorganisms, as
nc ]
and substituting in Eq. [21.26], the rate of substrate utilization can then be defined as:
Fs
=
-
[21.28]
Fsmax Csh
+ Cs
We have to emphasize that direct application of Eq. [21.28] is not expected since microbial populations and substrate availabilities are quite variable. In addition, the effect of soil adsorption on biotransformation rate is not considered in Eq. [21.28]. Laboratory experiments are required to determine the exact role of biotransformation in the distribution of a compound at a specific site. These tests determine the biotransformation rate constants which are usually based on the total chemical concentration of the soil and, therefore, include sorption effects. First order rate constants for soils can be determined from batch experiments from the slope of the relationship: In
CS
-
-
kt
[21.29]
Cso
where
is the initial substrate concentration, k is the first order rate constant, and t is time. The substrate half-life, t ~ / 2 , can then be calculated using the following relation:
%
tl/2
21.10
-
0.693 k
[21.30]
SUMMARY AND CONCLUDING REMARKS
Despite the recent sharp increase in the number of studies on bioremediation of soils polluted with organic chemicals, many questions have yet to be answered. In particular, more in-depth studies on the microbial transformations under denitrification and anaerobic conditions are specifically
592
BIOREMEDIATION
required as highly polluted soil systems are more likely to be characterized by these conditions. Biotransformation of high molecular weight organic pollutants commonly occurs by cometabolism rather than mineralization. The complex transformation pathways increase the risk that toxic as well as non-toxic intermediates will be formed. Identification of metabolites and their turnover times should be an important aspect of any study of the microbial transformation of organic pollutants. The large number of biological and environmental factors that determine the rate of biotransformation makes a precise modelling of the process difficult.
CHAPTER
TWENTY
TWO
CASE S T U D Y
22.1
HISTORY OF THE HYPOTHETICAL SITE
The hypothetical canal was constructed to facilitate navigation between various cities. Over the years, various industries were established on the banks of the canal. Presently, at a number of places, the presence of pollution bears witness to a history of industrial occupation. For many years, discharges from the sanitary sewer systems of towns bordering the canal, and industrial emissions, have combined to contribute to the bacteriological and chemical pollution of the waters and sediments of the canal. A number of years ago, public access to the canal was prohibited, pollution having reached a level judged dangerous to human health and the environment. At that time, certain users of the canal were observed to have contracted skin infections, presumably caused by the presence of pathogenic bacteria in the water. Shortly afterwards, studies were undertaken to determine the nature of the problem. A few years later, the decision was made to cleanup the site.
22.2
WATER QUALITY
The canal is fed by waters from a large river. The water quality of the river is generally better than that of the canal. For most of the parameters analysed at each station along the canal, the average values are either below the State water quality criteria for recreational activities and the protection of aquatic life (MENVIQ, 1990), or below the detection limit, as shown in Table 22.1. Only the concentrations of total phosphorus and of faecal coliforms occasionally exceed the quality criteria for the practice of recreational activities with primary contact (swimming, windsurfing, etc). Chromium levels exceeding the criteria for the protection of aquatic life have been found in the water. The average concentrations of copper, lead and total phosphorus exceed the State criteria in the downstream of the canal. The observed values are, however, only slightly above the criteria and, consequently, should not affect most aquatic organisms. Water in the upstream of the canal contain concentrations of cadmium, lead and copper equal to or slightly above the criteria for protection of aquatic life. Although they occasionally reach relatively high values, mineral oils and grease and phosphorus vary randomly throughout the canal. Moreover, the faecal coliforms are higher in the upstream of the canal (203/100 ml) than in the downstream (96/100 ml). The quantity of suspended solids reach 21 mg/1 in the upstream and 8 mg/1 in the downstream.
593
594
CASE STUDY
Table 22.1" Water quality in the river, upstream and downstream of the canal
Parameter Arsenic Cadmium Chromium Copper Cyanide Mercury Nickel Lead Zinc Nitrogen TOC Coliform Conductivity SM DO pH Phosphorus PCBs PAHs MO&G Phenol
Recreational activities cleanup criteria (mg/1) 0.05 0.01 0.05 1.0 0.2 0.001 0.25 0.05 5.0
Aquatic life cleanup criteria (mg/1) 0.05 0.0002 0.002 0.002 0.005 0.000006 0.025 0.001 0.03 1.2
200.0
5-9 0.03
> 6.5 6.5-9 0.03 0.000001
0.001
0.005
River (mg/1) < 0.001 < 0.001 0.004 0.003 < 0.005 < 0.0002 < 0.002 < 0.002 <0.05 <0.1 7.1 55.0 77.5 <5.0 12.5 7.8 <0.05 < 0.0001 < 0.0002 0.28 < 0.002
Upstream of the canal (mg/1)
Downstream of the canal (mg/1)
< 0.001 < 0.001 0.004 0.003 < 0.005 < 0.0002 < 0.002 <0.05 <0.05 <0.1 7.5 203.0 71.0 21.0 12.0 7.8 0.09 < 0.0001 < 0.0002
< 0.001 < 0.001 < 0.003 0.0025 < 0.005 < 0.0002 < 0.002 < 0.002 <0.05 <0.1 8.2 96.0 72.0 8.0 13.1 7.7 0.1 < 0.0001 < 0.0002 0.26 0.003
< 0.002
TOC = total organic carbon, SM = suspended matter, DO = dissolved oxygen, PCB = poly chloronated biphenols, PAH = poly aromatic hydrocarbons, conductivity (micro S/cm), MO&G = mineral oil and grease, coliform (n/100 ml), cleanup criteria as stated by MENVIQ (1990).
22.3
SEDIMENT QUALITY
The industrial activity developed along the banks of the canal has left its mark in the form of a layer of contaminated sediments. In the upstream, most of the pollutants have concentrations below the harmful threshold (Level 3) defined by the State criteria (MENVIQ, 1990), as shown in Table 22.2. Nevertheless, despite the fact that occasionally the concentrations of certain pollutants are high, the degree of pollution of the sediments in the upstream is relatively low. Consequently, the sediments in the upstream do not appear to be sufficiently polluted to justify corrective measures in the short term.
SEDIMENT QUALITY
595
Table 22.2: Sediment quality in the u 9stream of the canal
Parameter Arsenic Cadmium Chromium Copper Mercury Nickel Lead Zinc PCBs MO&G TOC (%) PAHs Nitrogen Prosphorus Coliform
Average concentration (mg/1)
Cleanup criteria Level 1 (mg/1)
Cleanup criteria Level 2 (rag/l)
2.8 2.1 61.0 62.0 0.51 39.0 142.0 240.0 0.12 1407.0 5.2
3 0.2 55 28 0.05 35 23 100 0.02 1000
7 0.9 55 28 0.2 35 42 150 0.2 2000
Cleanup criteria Level 3 (mg/1) 17 3 100 86 1
61 170 540 1
1
10
600
2000
Cleanup criteria Level 3 (mg/1)
1
Table 22.3" Sediment quality in the downstream of the canal
Parameter Arsenic Cadmium Chromium Copper Mercury Nickel Lead Zinc PCBs MO&G TOC (%) PAHs Nitrogen Prosphorus Coliform
Average concentration (mg/1)
Cleanup criteria Level 1 (mg/1)
Cleanup criteria Level 2 (mg/1)
5.3 2.4 56.0 84.0
3 0.2 55 28 0.05 35 23 100 0.02 1000
7 0.9 55 28 0.2 35 42 150 0.2 2000
1.72
53.0 304.0 1340.0 0.76 6.92 9.9 50.2 2579.0 1516.0 60.0
17 3 100 86 1
61 170 540 1
1
10
600
2000
1
Levels 1, 2, and 3 correspond to threshold concentrations for no effect, weak effect, and harmful effect, respectively (MENVIQ, 1990).
596
CASE STUDY
The results show that the sediments in the downstream of the canal are rather polluted, as shown in Table 22.3. The concentrations of several of the variables analysed exceed the threshold for harmful effects (Level 3) (MENVIQ, 1990) at various sampling sections. Moreover, analysis of boring samples indicated that the average thickness of the polluted sediment is 260 mm, which results in a contaminated sediment volume of about 120,000 m 3.
22.4
DEVELOPMENT OF CLEANUP ALTERNATIVES
(1) (2) (3) (4) (5) (6)
The following cleanup alternatives have been identified for the hypothetical case study: Storage of sediments on land; On-site storage of sediments on the canal bed; Sediment encapsulation within the canal area; Off-site solidification/stabilization; On-site solidification/stabilization; and Extraction-based treatment of polluted sediments.
Alternative 1: Storage of Sediments on Land
The exterior storage alternative for canal sediment comprises four major steps: (1)
(2)
(3)
(4)
Excavation: The only practical method of removing contaminated sediment from the bed of the canal is dredging. Hydraulic and pneumatic dredges are the best techniques from the technical and environmental viewpoints. Volume Reduction: Reduction of mud (water and sediments) volume could be achieved via separation based on particle size and dehydration. Transportation: Transportation modes depend principally on the characteristics of the dredged materials, the stage at which volume reduction is carried out, and the type of treatment. Pipelines could be used for liquid sediment transport, and trucks for transport of dehydrated sediments. External Storage of Polluted Sediment: The principal storage alternatives are storage on land in existing cells which are managed by various waste management companies (e.g., Cintec Environmental Inc., Quebec; Laidlaw Environmental Services, Ontario, and Chemical Waste Management (CWM) Inc., New York State), and storage on land in new cells which require building maximum security cells according to Provincial and State regulations. Cost estimates show an investment of about CanS 28 million for storage at Cintec, about CanS 22 million for the CWM site and about CanS 5 million for a new cell to contain 120,000 m 3 of polluted sediment.
Alternative 2" On-Site Storage of Sediment on the Canal Bed
The on-site method of polluted sediment stabilization would consist of storing the sediment on the canal bed, covering it with a permeable geotextile membrane, then compacting it by overlaying a bed of crushed stone on top of the membrane. The estimated cost to carry out this option is CanS 6 million.
DEVELOPMENT OF CLEANUP ALTERNATIVES
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Alternative 3: Sediment Encapsulation Within the Canal Area
Encapsulation involves storage of polluted sediment in watertight holding basins constructed within the canal itself. Polluted sediments would be excavated by means of suction flow-reversing dredges and transferred to the holding sections by means of pipelines. The holding sections would act as sedimentation-settling zones, allowing only water that meets the established quality standards to return to the canal. Each watertight section would be subdivided into two compartments for a progressive treatment of the water-sediment mixture. The upstream compartment would receive the water-sediment mixture from the dredging activity for primary decantation, which could be accelerated by adding coagulating agents. It would be isolated from the canal and the downstream compartment by means of watertight geomembranes. The bed and walls of the canal and the lower half of the dike of the downstream compartment would be isolated by watertight geomembranes while the upper half of the dike would be covered by a geomembrane filter. Following treatment, water could be returned to the canal through an outlet located in the downstream part of the compartment, or through the geomembrane filter. When the sediment has completely settled, the two compartments would be covered by a semi-permeable membrane, allowing the escape of gas produced by the decomposition of organic matter in the sediment. A protective layer of sand or any type of clean material would be used as a cover. It will, also, compact the sediment beneath the membrane. The space occupied by the compartments would reduce the width of the canal from 50 m to 20 m, which would still leave sufficient space for the practice of aquatic activities. The total estimated cost for implementing this option is about CanS 10 million. Alternative 4: Off-Site Solidification/Stabilization
Solidification/stabilization, as discussed in Chapter 20, is a treatment technique already in use in various parts of the world. The technique is suitable for treatment of heavy metals and oils and greases at low concentrations. The highly alkaline pH (approx. 12) of treated material prevents any re-colonization by microorganisms. In addition, pathogenic organisms are killed during this treatment. Off-site treatment requires prior dredging of polluted sediments. A sediment storage zone and a treatment zone are then setup on the banks of the canal. Pretreatment could be necessary for material preparation before adding the solidifying agents. Pretreatment may consist of filtering to remove large particles, breaking up aggregates in order to homogenize the material, or addition of additives to improve the sediment texture. The necessity for pretreatment, together with its nature, are evaluated following characterization of the sediments to be treated. Next, chemical reagents are added to the polluted sediments in a mixer, or possibly an ordinary cement truck or mixers (reservoirs equipped with agitators) specially installed on-site during operations. Concentration of reagents to be added to the polluted sediments are evaluated based on sediment characteristics. Water is also added to the mixture, for hydration and chemical reactions. After mixing, sediments and reagents form a product which has the consistency of liquid cement. It is then placed in cells where it may undergo a maturation phase. The period of maturation varies from 24 hours to several months, depending on the reagents added, and the future use of the treated sediments. Several schemes for final disposal may be considered for the product of off-site solidification/stabilization treatment. The treated material could be sent to landfills, where it could
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be used as covering material. It could also be used as embankment material or for general construction purposes. Whatever the final use may be, the treated material should be in conformity with regulations and guidelines for solid waste disposal practice. The unit cost for off-site treatment varies between CanS50 and CanS100 per m 3 of polluted sediments. Hence, treatment of 120,000 m 3 of sediments would result in total costs varying from CanS 6 million to CanS 12 million. Alternative 5: On-Site Solidification/Stabilization
On-site solidification/stabilization is accomplished by adding directly to sediment (i.e., in their location in the canal) a chemical reagent, the proportion of which will vary according to the degree of contamination. Cement is added next in order to produce a highly resistant material with low hydraulic conductivity. The maturation and solidification period of the sediment/reagent mixture occurs underwater within several hours. Since treatment occurs underwater, without prior dredging of sediment, no aqueous effluent is generated. The work zone for on-site solidification/stabilization is the canal itself. Equipment used includes the drilling and mixing system, a crane, tanks that contain reagents, computerized equipment for process control and generators, all of which are located on a barge. The treated material can be left in place or be used to reinforce the sidewalls of the canal provided that it conforms with the solid waste disposal regulations. The unit cost of on-site treatment varies between CanS 100 and CanS 200 per m 3 of polluted sediments. Thus, the treatment of 120,000 m 3 of sediment would result in total costs varying between CanS 12 million and CanS 24 million. Alternative 6" Extraction-based Treatment
Extraction-based treatment comprises several stages. First, the sediments are washed (physically extracted) in order to separate the fine particles in which pollutants are concentrated. Next, metals contained in the fine portion are chemically extracted and recuperated via solvent extraction techniques, as discussed in Chapter 17. Finally, organic pollutants are either treated via surfactant extraction (Chapter 18), bioremediated (Chapter 21), landfilled (Chapter 15) or vitrified, etc. Due to the nature of pollutants, organic and inorganic, it is necessary to follow a consistent sequence of washing, metal extraction and treatment of the organic pollutants.
Sediment Washing Sediment washing is a physical operation, consisting of separating the fine particles from the coarser particles of the sediment. The need for particle separation is based on the hypothesis that pollutants are attached to the finer particles of the sediment. Pollutants, also, attach to soil organic matter, notably due to its high adsorption capacity, as discussed in Chapter 5. Coarse particles, i.e., more than approx. 5 to 10 mm diameter, are assumed not to be contaminated. Following an initial separation, subsequent stages of particle separation are carried out, namely, separation of particles by centrifugal force, metal extraction (e.g., electromagnetic and electrostatic processes), etc. Sediment washing results in coarse particles of clean sediment, and fine particles in which metals and organic pollutants are concentrated. The process also generates polluted water which must be treated before discharge. Air purifiers such as active carbon filters or biofilters should be used. Once particles of diameter exceeding 5-10 mm are removed, the sediment may be subjected
COMPARATIVE ANALYSIS OF CLEANUP ALTERNATIVES
599
to a second type of treatment, i.e., high pressure washing. This type of washing keeps sediments in contact with water under high pressure, gradually forcing separation of fines from coarse particles. The larger washed particles are considered clean whereas the finer particles contain both organic and inorganic pollutants. High pressure washing permits volatilization of certain volatile pollutants, which either remain in the gaseous phase or dissolve in the aqueous phase. Air and water purification systems should, therefore, be in place during this type of treatment. Metal Extraction Following the washing stage, the fine particles of sediment, containing concentrated organic and inorganic pollutants, are subjected to a second treatment stage, namely, metal extraction. The principle of metal extraction, as discussed in Chapter 17, consists of first separating metals present in the fine particle s by means of solvents and chelating agents. Next, chelates are separated from the fine particles, which contain only organic chemicals. Metals are then recovered from chelates by using an acid solution to break the chemical bonds between the chelating agents and metals. Then, chelates are regenerated and re-utilized at the upstream of the extraction process while metals are recuperated for recycling purposes by means of electroplating. Treatment of Organic Chemicals The last stage of sediment treatment is the treatment of organic chemicals (e.g., PCB, PAHs) that are concentrated in fine particles and water. Several options are available, including surfactantbased washing (Chapter 18), biodegradation (Chapter 21), and thermal destruction. Incineration is a proven technology, but it is subject to some social resistance. When this technology is used, sediment loses its structure and cannot be reused for any purpose other than land disposal. Biodegradation enables regeneration of sediment and produces only aqueous effluent as residue. Surfactant-based washing or biological treatment of the residue allows recuperation of water for reuse in the process. Nonetheless, the presence of PCBs hinders conventional methods of biodegradation. In this case, dechlorination would be necessary; it could be achieved by irradiation of contaminated particles with ultraviolet rays, which provoke the removal of chlorates from organic molecules. Next, normal decontamination in which chlorine atoms react with an alkaline metal (e.g., Na) to form salts (e.g., NaC1) could be carried out. Finally, it is possible to landfill, without treatment, organically polluted fine particles. Due to the nature of the contamination, disposal would take place in storage cells or in landfills. In this case, the sediment resource is not reusable. Costs of complete sediment treatment would be between CanS 250 and CanS 399 per m 3 for a total cost varying between CanS 30 and CanS 43 million for the treatment of 120,000 m 3 o f polluted sediments.
22.5
COMPARATIVE ANALYSIS OF CLEANUP ALTERNATIVES
In the present analysis, environmental, technical and economic criteria are used, as discussed in Chapter 2. Each criterion is divided into various options and each option is divided into three classes according to their importance. The method consists of rating each possible option according to the criterion. The chosen option is the one that accumulates the most first ratings. If two or more options have the same number of first ratings, a count of second rating is taken. If necessary, the count is continued (third rating, etc.) until the option with the best performance is determined.
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The criteria used in this comparative analysis are: (1) permanent environmental, (2) temporary environment, (3) technical, (4) technical and economical, (5) economic, (6) nondiscriminatory, and (7) non-pertinent. These criteria are discussed below. 22.5.1 Permanent Environmental Criteria
Option 1: Elimination of Pollution in Canal Sediment This option is used to evaluate whether an option is able to decontaminate the canal sediments. Treatment options eliminate pollution. In containment and fixation options, pollutants are still present, although they are immobilized. The criterion reflects a concern that the sediment pollution problem should, if possible, be permanently resolved. This means that options which ensure permanent sediment cleanup should be preferred over others. Possible ratings are shown in Table 22.4.
Table 22.4: Possible ratings of options of permanent environmental criteria for the comparative analysis Option
Rating
Justification
1
1st 2nd
Complete treatment and elimination of pollutants in sediments. Confinement or fixation of polluted sediments.
2
1st 2nd
Sediments remain in the study zone. Part or all the sediments is transferred outside the study zone.
3
1st 2nd
No changes in the layout of the canal. Changes in the layout of the canal.
4
1st 2nd 3rd
No pollution risk; contaminants are no longer in the study zone. Low risk of pollution; contaminants confined or fixed in the canal in a completely closed compartment. Moderate risk of pollution; semi-closed compartment.
1st 2nd 3rd
Substratum similar to natural state. Substratum composed of small stones (added material). Cement substratum.
Option 2: Sediment Management in Study Zone This option expresses the desire to resolve the contamination problem in the study zone. Transfer of contamination elsewhere could cause social opposition (Not in My Back Yard, NIMBY, syndrome) which could compromise the feasibility of the project. Possible ratings are shown in Table 22.4.
COMPARATIVE ANALYSIS OF CLEANUP ALTERNATIVES
601
Option 3: Heritage This option evaluates changes in the layout of the canal, since it is preferable to preserve the historic or present layout of the canal. Possible ratings are shown in Table 22.4. Option 4:
Risks of Polluting Groundwater and Surface Water after Completion of Treatment
This option covers risks of contamination of the canal and groundwater in the study zone following decontamination operations. Risk of future pollution is directly linked to the presence of pollutants and the possibility that they may again be found in the water. Possible ratings are shown in Table 22.4. Option 5: SubstratumCharacteristics Following Completion of Work Once sediment has been treated or removed, the final characteristics of the substratum at the bottom of the canal will influence recolonization of the canal by flora and fauna. Possible ratings are shown in Table 22.4.
Table 22.5: Possible ratings of options of the temporary environmental criteria for the comparative analysis Option
Rating
6
1st
2nd 3rd
7
1st 2nd 3rd
8
1st 2nd 3rd
Justification Majority of work taking place outside study zone; minimal installations, little storage required, work completed within a year or less. Work carried out in study zone; installations on-site, including storage, work completed within a year or less. Work carried out in study zone; installations on site, including storage, work completed in more than a year. Minimal trucking activity (movement, stopping, loading). High level of trucking activity (movement, stopping, loading and transport of material from outside). Very high level of trucking activity (movement, stopping, loading and transport of sediment to outside study zone). Little or no drop in water level, minimal dredging and mechanical agitation. Little or no drop in water level; dredging. Major drop in water level.
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CASE STUDY
22.5.2 Temporary Environmental Criteria
Option 6: Inconvenience Caused by Presence of Construction Site This option covers all kinds of inconvenience which may impact the public during work period. Inconvenience includes dust, operational noise arising from the cleanup process, odours, and deterioration of the landscape. Inconvenience connected to closing or diversion of the cycle path in several places, due to the presence of a construction site, is also evaluated, as is the duration of such inconvenience. Inconveniences linked to the presence of a construction site beside the canal will principally affect the population living near the canal, or those frequenting the canal. Possible ratings are shown in Table 22.5. Option 7: Inconvenience due to Trucking Activity This option covers such matters as perturbation to circulation and pedestrian security. It takes into account the need to access the canal. Inconvenience from trucking activity will affect the population living near the canal, and will also cause repercussions for several kilometres around the work site. Possible ratings are shown in Table 22.5. Option 8: Impact of Works on Flora and Fauna This option is used to evaluate negative impact (injury, mortality) caused by carrying out a decontamination option, such as that arising from lowering the water level or from turbidity due to dredging or hydraulic agitation. Possible ratings are shown in Table 22.5. 22.5.3 Technical Criteria
Option 9: Works Linked to Substructures after Completion This option evaluates the possibilities of installing new substructures (e.g., subterranean cables), as well as the ease of carrying out and maintaining operations on substructures already in place, after work completion. These substructures include servitudes, subterranean installations, and installations underwater or on the banks of the canal. Possible ratings are shown in Table 22.6.
Table 22.6: Possible ratings of options of the technical criteria for the comparative analysis Option
Rating
9
1st 2nd 3rd
10
1st 2nd 3rd
Justification Canal bottom similar to natural state. Canal bottom similar to natural state but compartments present on canal banks. Characteristics of canal bottom altered (e.g., cement or geomembrane). Proven technology for similar treatment. Proven technology for a different treatment. Technology not commercially proven (at pilot or pre-industrial stage).
COMPARATIVE ANALYSIS OF CLEANUP ALTERNATIVES
603
Option 10: Technical Feasibility This option covers technological risks linked principally to the stage of development or applicability of technologies selected for decontamination of the canal. Possible ratings are shown in Table 22.6. 22.5.4 Technical and Economic Criteria
Option 11: Follow-up Measures in Study Zone and Associated Costs This option evaluates the amount of follow-up which will be necessary after completion of decontamination in the study zone. Follow-up consists of ensuring, on a long term basis, the effectiveness of the restoration. Costs involved are proportional to the amount of follow-up required. This criterion also covers the ease with which appropriate follow-up measures can be carried out. Possible ratings are shown in Table 22.7.
Table 22.7: Possible ratings of options of the technical and economic criteria for the comparative analysis Option
Rating
11
1st 2nd 3rd
12
1st 2nd 3rd
Justification No follow-up required, pollutants are removed from study zone. Control is required, but easily carried out, i.e., small number of sampling points is needed for appropriate follow-up. Control is required and is difficult to carry out; large number of sampling points is required for appropriate follow-up. CanS 10 million or less. CanS 10 - 30 million. over CanS 30 million.
22.5.5 Economic Criteria
Option 12: Total Cost This option evaluates the total cost of the project, with regard to each option, for treatment of 120,000 m 3 of polluted sediments. Possible ratings are shown in Table 22.7. 22.5.6 Non-Discriminatory Criteria
The following options are deemed non-discriminatory, i.e., do not differentiate between the various options, since their influence on all the options is similar.
Option 13: Contaminant Risks Downstream from the Study Zone The six treatment alternatives under consideration are not discriminated by this option. All
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CASE STUDY
the alternatives, except on-site solidification/stabilization, require dredging of polluted sediment, which results in resuspension of contaminated sediment. However, measures can be provided to reduce this effect. With regard to on-site solidification/stabilization, protective measures are included in the design. Furthermore, the treatment work area could be sealed off by preventing water flow from upstream, thereby preventing downstream pollution. Hence, this option cannot be used to discriminate one treatment alternative from the other.
Option 14: Potential for Recreational, Faunic and Landscape Redevelopment This option is considered non-discriminatory because all treatment alternatives considered will result in re-opening of the canal to the public. The potential result will be an improvement of the environment, leading to potential recreational, faunic and landscape redevelopment, and the effect will be similar no matter which treatment alternative is selected. Option 15: Impact of Works on the Local Economy Construction work of any type will have a direct impact on local economy. For example, perturbation of traffic and closing of access roads will adversely affect some local businesses, although an increase in workforce could result in increased business for certain local establishments (e.g., restaurants, etc. ). In addition, regarding employment, the call for bids could specify that a percentage of local personnel must be hired for the project, no matter which option is selected. Therefore, this option is considered non-discriminatory. Option 16:
Impact on Property Values and the Socio-economic Structure
Speculation stimulated by the decontamination of the canal, no matter which treatment option is utilized, will have similar effects on real estate market. 22.5.7 Non-Pertinent Criteria
Option 17:
Industrial Activity during Work Period
Some of the options involve a reduction, major or otherwise, in water level in the canal during part of the work. This drop in water level would cut off supply to some industries along the side of the canal. However, it must be stipulated in the call for bids that water supply to various industries be assured throughout the work period.
Option 18: Stability of Canal Walls If the canal is drained for a long period, measures should be taken to prevent the side walls from caving-in (e.g., technical inspection during installation of metal reinforcement to the wall structure). These requirements would be stipulated in the call for bids, thus making preventive or protective measures of this nature mandatory. Option 19: TimeTable The actual timetable will be known only after tenders are received following call for bids. Once again, a time limit for work completion could be laid down in the call for bids. The expected time required to carry out the work, according to each option, is included in option 6.
COMPARATIVE ANALYSIS OF CLEANUP ALTERNATIVES
605
22.5.8 Importance of Rating Rating allows a relative weight to be attached to each of the options, those options deemed more important being given a heavier weight. In this analysis, three classes of criteria are defined, the first and third representing the most and least important, respectively. Since the choice of option depends on the number of top ratings accumulated, the importance of the several options is taken into account by diminishing the ratings given to the options of the less importance. The revised rating given to an option after the importance of the class has been allowed for will be called the weighted rating. For example, a second rating, for a top-rated option, will have the same weight as a first rating for a second class rated option. Both will be given a second-weight rating. As the maximum number of ratings (class of importance) possible is three, a maximum of five corresponding ranks will be possible, as indicated in Table 22.8 (Holmes, 1971).
Table 22.8: Corresponding ranks (Holmes, 1971) Class of importance (CI)
Relative rank of option (RRO) 1st [ 2nd 3rd
I 1 2
I [
2 3
3
I
4
I The class of importance was determined from a perspective of sustainable development. The class of importance assigned to each of the twelve options is shown in Table 22.9. It is seen that out of the twelve options used in the comparative analysis, five were given first class rating, four a second class and three a third class rating.
22.5.9 Remedial Alternative Selection Evaluation of each of the six treatment alternatives according to the twelve defined options is discussed previously. Four partial Tables, corresponding to the four subcategories of criteria, i.e., permanent environmental (Table 22.4), temporary environmental (Table 22.5), technical (Table 22.6), and technical and economical (Table 22.7), are developed. Ranking of the six treatment alternatives is shown in Table 22.10. Evaluation of performance of the various options is achieved by calculating the number of first ratings accumulated by each option. Should two or more options be awarded the same number of first ratings, the number of second ratings, and if necessary third ratings, will be calculated.
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CASE STUDY
Table 22.9: Class of importance for various options
Rational
Option
Class
1
2
All options conform to the main objective of the project, i.e., "as far as possible the choice of option should favour permanent treatment of contaminated sediments."
2
1
This option expresses the State responsibility.
3
3
The public has expressed a desire to preserve the present layout of the canal.
4
2
Risks of future contamination of canal waters are linked to the presence of pollutants, which could seep to subsurface water in places where the hydraulic gradient is oriented from the canal towards inclined layers. Nonetheless, a follow-up after completion of work would enable problems to be discovered and corrective action to be taken, if necessary.
5
3
Changes of Canal's substrate is possible; the rate and diversity of the substrate changes may, however, be influenced by the type of substrata in place.
6
1
This option is a matter of social acceptability. Alteration of the surrounding conditions usually means increased noise, dust, and unpleasant odours. This may create negative impacts to the local population. The project could be contested by the population owing to inconveniences during work in progress. However, some measures are possible for reducing the nuisance (e.g., appropriate working hours, communication plan, etc.).
7
1
This option is also a matter of social acceptability. Inconveniences related to urban transportation will directly affect the population, business and adjacent industries. Protests could result from these inconveniences.
8
3
This option is important during progress of work. It is temporary. No rare or endangered species are to be found in and around the study zone. Changes are possible even though the canal does not constitute an important environment for wild flora and fauna.
COMPARATIVE ANALYSIS OF CLEANUP ALTERNATIVES
607
Table 22.9 (cont.): Class of importance for various options Option
Class
Rational
9
2
Access to or installation of new substructures on the canal bottom following decontamination will be complicated by a substratum different from the natural one, especially if the new substratum in intended to be watertight. The presence of compartments on the banks could interfere with some future work. Nonetheless, some modifications or adjustments could be made to overcome these difficulties.
10
1
Technical feasibility of the various decontamination options implies the notion of technical risks associated with relatively undeveloped or commercially unproved technologies. Attainment of the project objectives depends directly on the decontamination technology selected.
11
2
Follow-up after work completion requires long term effort and investment, which must be added to costs already incurred.
12
1
The total cost of the project is very important, as it has been determined that the status quo does not involve any public health risk and does not require immediate action. Thus, the public could object if a very costly project was undertaken. On the other hand, if public health was threatened, little importance would have been attached to the total cost of the project, the priority being protection of the population.
The final ratings of the six treatment alternatives are shown in Table 22.11. The encapsulation alternative is awarded the highest classification, this being the alternative with the largest number of first ratings. The storage on land alternative is in the second position, followed by chemical extraction. The fourth, fifth and sixth positions are given to on-site solidification/stabilization, storage with geomembrane (FML) and off-site solidification/stabilization. All methods of comparative analysis have limitations. These limitations are principally attributable to the number and classification of options. Thus, the number of options selected in each category, as well as the classification, are determinant choices which remain subjective, in as much as they are decided upon by a task force comprising a restricted number of specialists. On the other hand, analysis of results of the comparative analysis and most particularly analysis of the strong and weak points of the chosen option make it possible to partially overcome the limits of the method and validate its results.
m 0 co
CASE STUDY
Table 22.10: Ranking of the various treatment alternatives
Storage on land
Storage with FMT.
Encapsulation
On-site SIS
Off-site
Treatment
R
RRO
R
RRO
R
RRO
R
RRO
R
RRO
R
RRO
SIS
1 2 3 4 5
2 1 3 2 3
2nd 2nd 1 st 1st 1st
3 2 3 2 3
2nd 1st 1st 3rd 2nd
3 1 3 4 4
2nd 1st 2nd 2nd 1st
3 1 4 3 3
2nd 2nd 1st 1st 1st
3 2 3 2 3
2nd 1st 1st 3rd 3rd
3 1 3 4 5
1st 1st 1st 1st 1st
2 1 3 2 3
6 7 8
1 1 3
1st 3rd 2nd
1 3 4
2nd 2nd 3rd
2 2 5
3rd 2nd 3rd
3 2 5
2nd 3rd 2nd
2 3 4
2nd 1st 1st
2 1 3
3rd 1st 2nd
3 1 4
9 10 11
2 1 2
1st 1st 1St
2 1 2
2nd 3rd 3rd
3 3 4
2nd 1st 2nd
3 3
1st 2nd 1St
2 2 2
3rd 2nd 3rd
4 2 4
1st 3rd 1st
2 3 2
12
1
2nd
2
1st
1
1st
1
3rd
3
2nd
2
3rd
3
1
R = Rating; RRO = Relative rank of option, which is assigned based on Table 22.8.
SUMMARY AND CONCLUDING REMARKS
609
Table 22.11: Final ranking of the various treatment alternatives Relative rank of option
Storage on land
Storage with FML
Encapsulation
On-site S/S
Off-site S/S
Treatment
Ranking
From the results shown in Table 22.11, the encapsulation option has the highest rating and is therefore, the cleanup alternative selected. Scientific analysis of the individual tables make it possible to point out the strengths and weaknesses of this option. First, it can be noted that encapsulation has a second rating for performance among the permanent environmental criteria. Indeed, sediment management is carried out within the study zone. In addition, the natural granular substratum will remain in place at the end of the project. Encapsulation is a storage process, but the fact that sediment is confined in completely closed compartments limits future risk of water contamination in the canal and groundwater. In terms of the temporary environmental criteria, encapsulation is ranked last. This constitutes the major weakness of the selected option. For technical performance, encapsulation alternative is rated the second. It is a proven confinement technique which is in commercial use for similar projects. Finally, encapsulation is rated first for economic performance, since this treatment alternative has the least cost.
22.6
S U M M A R Y AND C O N C L U D I N G R E M A R K S
It is apparent that if sustainable development is to be achieved, and the requirements of the environmental impact assessment process satisfied, environmental and social considerations as well as technical and economic criteria must be considered from the very beginning of the planning process. This will require leadership on the part of the engineering profession, since it has traditionally dominated the planning process, as well as an evolution towards a more multidisciplinary approach to design that involves the input of biologists, sociologists, anthropologists, etc., at the earliest stages. The case study presented in this chapter is utilized to develop and evaluate remediation scenarios applicable to a specific uncontrolled hazardous waste site. The goal of the case study is to integrate a set of technologies that are capable of eliminating the site' s risks to the degree required by pertinent government standards. The set of remedial options must satisfy the sustainable development criteria, discussed in Chapter 2. The method employed is systematic, and demands multiple inputs from various disciplines.
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GLOSSARY A Abiotic: Nonliving factors or physical factors; the abiotic elements of an ecosystem constitute its climate, geological, and soil components. Absorbed dose: The amount of chemical substance actually entering an exposed organism via the lungs (for inhalation exposures), the gastrointestinal tract (for ingestion exposures), and/or the skin (for dermal exposures). It represents the amount penetrating the exchange boundaries of the organism after contact. It is calculated from the intake and absorption efficiency, expressed in mg/kg-day. Absorption: (1) Penetration of a substance into the body of another. (2) transformation into other forms suffered by radiant energy passing through a material substances. Absorption coefficient: A measure of the amount of normally incident radiant energy absorbed through a unit distance of the absorbing medium. factor: The percent or fraction of a chemical in contact with an organism that becomes absorbed into the receptor. Accelerator: A chemical additive. It is used to promote the dissolution of calcium cations and anions from cement, thus enhancing cement hydration. Acceptable daily intake (ADI): An estimate of the maximum amount of a chemical ( in mg/kg body weight/day) to which a potential receptor can be exposed on a daily basis over an extended period of time --- usually a lifetime --- without suffering a deleterious effect, or without anticipating an adverse effect. risk: A risk level generally deemed by society to be acceptable. Acid: (1) An acid is a hydrogen-containing substance which dissociates in solution to produce one or more hydrogen ions. (2) The Br6nsted concept states that an acid is any compound which can furnish a proton. (3) G.N. Lewis defined an acid as any thing which can attach itself to something with an unshared pair of electrons. deposition: Acid rain; a form of pollution depletion in which pollutants are transferred from the atmosphere to soil or water; often referred to as atmospheric self-cleaning. mine drainage: The seepage of sulfuric acid solutions (pH 2- 4.5) from mines and their wastes which have been placed on the surface; these solutions result from the interaction of water/precipitation with sulfide minerals exposed by mining. rain: Atmospheric precipitation with pH values less than about 5.6, the acidity being due to inorganic acids such as nitric and sulfuric that are formed when oxides of nitrogen and sulfur are emitted into the atmosphere. soil: A soil with a pH value less than 7. Acidity: The capacity of an acid to neutralize a base such as a hydroxyl ion (OH-). Actinomycetes: A group of organisms intermediate between the bacteria and the true fungi that usually produce a characteristic branched mycelium. Includes many but not all, organisms belonging to the order of Actinomycetales. Action level (AL): The level of a chemical in selected media of concern above which there are
611
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GLOSSARY --- A
potential adverse health and/or environmental effects. It represents the contaminant concentration above which some corrective action (e.g., monitoring or remediation) is required by regulation. Activated sludge: The biologically active sediment produced by the repeated aeration and settling of sewage and/or organic wastes. Activity coefficient: A factor which when multiplied by the molecular concentration yields the active mass. Acute exposure: A single large exposure or dose to a chemical, generally occurring over a short period (usually 24 to 96 hours). toxicity: The development of symptoms of poisoning or the occurrence of adverse health effects after exposure to a single dose or multiple doses of a chemical within a short period of time. Adhesion: Molecular attraction that holds the surface of two substances (e.g., water and soil particles) in contact. Adsorption: The attraction of ions or compounds to the surface of a solid. Soil colloids adsorb large amounts of ions and molecules. complex: The group of organic and inorganic substances in soil capable of adsorbing ions and molecules. isotherm: The relationship between the adsorbed concentration and the equilibrium concentrations at fixed temperature and applied pressure. Adveetion: The process by which pollutants are transported along with the flowing fluid or solvent in response to a hydraulic gradient. Aeration: The process by which air in the soil is replaced by air from the atmosphere. In a wellaerated soil, the soil air is similar in composition to the atmosphere above the soil. Poorly aerated soils usually contain more carbon dioxide and correspondingly less oxygen than the atmosphere above the soil. Aerobic: (1) Having molecular oxygen as a part of the environment. (2) Growing only in the presence of molecular oxygen, as aerobic organisms. (3) Occurring only in the presence of molecular oxygen (aerobic decomposition as an example). Aerosol: A colloidal system in which gas, frequently air, is the continuous medium and particles of solids (usually less than 100 microns in size) or liquid are dispersed in it. Aesthetics: The qualities of a site that when adversely affected can result in noticeable and disagreeable perceptions by the senses. These include sight (e.g., visibly stained soil or a film on water), taste (in water, fish, or agricultural products), and odour (in air, water, or soil). Aggregate: Many soil particles held in a single mass or cluster such as a clod and crumb. Air-dry: (1) The state of dryness of a soil at equilibrium with the moisture content in the surrounding atmosphere. The actual moisture content will depend upon the relative humidity and the temperature of the surrounding atmosphere. (2) To allow to reach equilibrium in moisture content with the surrounding atmosphere. Air pollution: The discharge of toxic gases and particulate matter introduced into the atmosphere, principally as a result of human activity. porosity: The proportion of the bulk volume of soil that is filled with air at any given time or under a given condition, such as a specified moisture potential. sparging: An in-situ technique. Air is injected under pressure below the water table to
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enhance biodegradation rate. Algae: Microscopic organisms that subsist on inorganic nutrients and produce organic matter from carbon dioxide by photosynthesis. Aliphatie compounds: Organic compound in which the carbon atoms exist as either straight or branched chains; examples include pentane, hexane, and octane. Alkalinity: The capacity of a base to neutralize the hydrogen ion (H+). Alkali soil: A soil that contains sufficient alkali (sodium). Anophane: An aluminosilicate mineral that has an amorphous or poorly crystalline structure and is commonly found in soils developed from volcanic ash. Alluvial soil: A soil developing from recently deposited alluvium. Alluvium: A general term for all material deposited or in transit by streams, including gravel, sand, silt, clay, and all variations and mixtures of these. Aluminosilieate: Compounds containing aluminum, silicon, and oxygen as main constituents. Amino acids: Nitrogen-containing organic acids that couple together to form proteins. Each acid molecule contains one or more amino groups (-NH2) and at least one carboxyl group (-COOH). In addition, some amino acids contain sulfur. Ammonification: The biochemical process whereby ammoniacal nitrogen is released from nitrogen-containing organic compounds. Ammonium fixation: The entrapment of ammonium ions by the mineral or organic fraction of the soil in forms that are insoluble in water and at least temporarily nonexchangeable. Amorphous material: Noncrystalline constituents of soils. Ampere: See Coulomb. Anaerobic: (1) Without molecular oxygen. (2) Living or functioning in the absence or air or free oxygen. Angstrom (,~,): A unit of length. ,a, = 1x 10~~ m. Anion: Negatively charged ion; during electrolysis it is attracted to the positively charged anode. exchange capacity: The sum of exchangeable anions that a soil can adsorb. Expressed as centimoles of charge per kilogram of soil. Antagonism: The interference or inhibition of the effects of one chemical substance by the action of other chemicals. Anode: The electrode at which oxidation occurs in a cell. It is also the electrode toward which anions travel due to the electrical potential. In spontaneous cells the anode is considered negative. In non spontaneous or electrolytic cells, the anode is considered positive. Anoxic: see anaerobic. Anthropogenic: Created by human activities. acids: Those acids that are the result of human activities. stress: The effect of human activity on other organisms. Approach: The philosophy and procedures used by a regulatory agency to establish criteria. The components of an approach can include the types of information considered, the goal of setting criteria (for example, protecting human health and the environment), the relative priorities to various types of information, and the way (s) that information are combined to set the criteria. Aquasphere: Water system. Aquiclude: A rock formation that is too impermeable or unfractured to yield groundwater.
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Aquifer: A subsurface water-bearing formation that yields water to wells. Arid climate: Climate in regions that lack sufficient moisture for crop production without irrigation. In cool regions annual precipitation is usually less than 25 cm. It may be as high as 50 cm in tropical regions. Natural vegetation is desert shrubs. Arithmetic mean: The arithmetic mean of a number of observations is the simple average. It is commonly referred to simply as the mean. Aromatic compounds: Organic compounds that contain carbon molecular ring structure (e.g., benzene, toluene, ethylbenzene, and xylene). Asbestos: A group of fibrous silicate minerals, typically those of the serpentine group, of which there are three main types: (1) blue asbestos, or crocidolite, (2) white asbestos, and (3) brown asbestos. Asphalt: A product of petroleum refining used for road construction. Assessment criteria: Concentrations of substances in soil and groundwater which can be used to assess site conditions in terms of the potential need for remediation. Where conditions do not exceed assessment criteria, there is no need for further investigation or remediation. As such, the assessment criteria are analogous to the de minimus, "trigger" and "threshold" criteria that some agencies have established. Atmosphere: A series of envelopes in the form of imperfect spherical shells of various materials that are bound to earth by gravitational force. It consists of gases, vapours, and suspended matter. The composition of the lower layers of the atmosphere is assumed for purposes of most engineering calculations as 76.8% nitrogen and 23.2% oxygen by weight; 79.1% nitrogen and 20.9% oxygen by volume. Atom: The smallest particle of an element which can enter into a chemical combination. All chemical compounds are formed of atoms, the difference between compounds being attributable to the nature, number, and arrangement of their constituent atoms. Atomic weight: The weight of an atom according to a scale of atomic weight units, valued as onetwelfth the mass of the carbon atom (C ~2= 12). Attenuation: Any decrease in the amount or concentration of a pollutant in an environmental matrix as it moves in time and space. It is the reduction or removal of pollutant constituents by a combination of physical, chemical, and/or biological factors acting upon the polluted media. Atterberg limits: Water contents of fine-grained soils at different states of consistency. Attraction: The electrostatic force between ions of the unlike charge (sign). Autotroph: An organism capable of utilizing carbon dioxide or carbonates as the sole source of carbon and obtaining energy for life processes from the oxidation of inorganic elements or compounds such as iron, sulfur, hydrogen, ammonium, and nitrites, or from radiant energy. Contrast with heterotroph. Available nutrient: The portion of any element or compound in the soil that can be readily absorbed and assimilated by growing plants. water: The portion of water in a soil that can be readily absorbed by plant roots. The amount of water released between the field capacity and the permanent wilting point. Average daily dose (ADD): The average dose calculated for the duration of exposure, and used to estimate risks for chronic non-carcinogenic effects of environmental pollutants. It is defined by: ADD (mg/kg-day)= {pollutant concentrationxcontact rate}/body weight.
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Avogardo constant (NA): The number of molecules contained in one mole or gram-molecular weight of a substance. N A = 6.022x 1023 mole 1.
B Background concentration" The concentration of a chemical substance occurring in a media removed from the influence of industrial activity at a specific site and in an area considered to be relatively unaffected by industrial activity. Bacteria" Microscopic, unicellular plant cells bounded by a membrane-wall complex and containing a variety of inclusions. Depending on the species and cultural conditions, bacteria occur as individual cells or in clumps or chains of sister cells. The average length lies within the range of 2 to 5 micrometers, although some are as small as 0.2 micrometer, or as large as 100 micrometer. Bacteria are classified on the basis of their requirements of free atmospheric oxygen. Those which require atmospheric oxygen are called aerobic (air-living); those which cannot live in the presence of atmospheric oxygen are called anaerobic; those which do well with oxygen, but can get along without it are termedfacultative anaerobes. Bar: A unit of pressure equal to one million dynes per square centimeter (106 dynes/cm2). 1 bar = 0.987 atm = 1000 mbars = 29.53 in of mercury. Bases: (1) A base is a substance which dissociates in solution to produce one or more hydroxyl ions. (2) Br6nsted concept states that a base is any compound which can accept a proton. (3) G.N. Lewis defines a base as anything which has an unshared pair of electrons. Bedrock: The solid rock underlying soils. Benchmark risk: A threshold level of risk, typically prescribed by regulation, above which corrective measures will certainly have to be implemented to mitigate the risks. Bench terrace" An embankment constructed across sloping fields with a steep drop on the downslope side. Bioaccumulation: The accumulation of pollutants in the organism, in the fatty tissues, in a concentration exceeding that which is found in the organism's environment. Bioassay: Measuring the effect(s) of environmental exposures by intentional exposure of living organisms to a chemical. Bioaugmentation: A process in which specially selected bacteria cultures, that are predisposed to metabolize some target compound(s), are added to impacted media, along with the nutrient materials, to encourage degradation of the pollutants of concern. Bioavailability: (1) The fraction of the total chemical that is available for uptake by biota from the encompassing environment such as water, sediment, soil, suspended matter. (2) The percentage of external exposure which enters the metabolism and is, thus, available at a possible site. (3) The fraction of total contaminant in the interstitial water and on the sediment particles that is available for bioaccumulation. (4) Easily available to plants or microorganisms. It is determined by exchanging heavy metals in soils by neutral salts. Biochemical oxygen demand (BOD)" The degree of oxygen consumption by microbial oxidation of pollutants in the water; also known as biological oxygen demands. Bioconcentration: Concentration of a chemical by ingestion into the organism. factor (BCF)" The ratio of the concentration of a substance in an organism to the
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concentration in the ambient medium at steady state. Biodegradable: Subject to degradation by biochemical processes. Biodegradation: The conversion of waste materials by biological processes to simple inorganic molecules and, to a certain extent, to biological materials. Biological half-life: The time required for half of a quantity of radioactive material absorbed by a living tissue or organism to be naturally eliminated. Biomagnification: Concentration of a chemical byfood supply into the organism. Biomass: The total mass of living material of a specified type (e.g., microbial biomass) in a given environment (e.g., in a cubic meter of soil). Bioreactors: Reactors for the conversion of biomass. Bioremediation: The decontamination or restoration of polluted or degraded soils by means of enhancing the chemical degradation or other activities of soil organisms. Biosphere: The portion of the earth on or in which life exists. Biota: Living organisms. Biotic: Actions of other organisms. Biotransformation: The conversion of a substance through metabolization, thereby causing an alteration to the substance by biochemical processes in an organism. Bioventing: A technique for aerating in-situ polluted soils to simulate biological activity. Bitumen: A naturally occurring material that is petroleum-like in nature, but that is immobile at room temperature. Boltzmann constant: The ratio of the universal gas constant to Avogardo number; equal 1.38054x 1016 erg/~ Bond number: It is used to assess the impact of buoyancy forces on the mobilization of nonaqueous phase liquid in porous media. Mathematically, it is the ratio of gravitational force to surface tension force. Boyle's law for gases: At a constant temperature, the volume of a given quantity of any gas varies inversely as the pressure to which the gas is subjected. For ideal gas, changing from pressure P and volume V to pressure P' and volume V' without change of temperature: PxV = P'xV'. Bottom ash: Ash that occurs at the bottom of a combustor. Br/insted acid: A chemical species which can act as a source of protons. base: A chemical species which can accept protons. Buffering capacity: It is a measure of the ability of a material to resist changes in its pH. Bulk density: The mass of dry soil per unit of bulk volume, including the air space. The bulk volume is determined before drying to constant weight at 105~
C Calcareous soil: Soil containing sufficient calcium carbonate (often with magnesium carbonate). Cancer: A disease characterized by malignant, uncontrolled invasive growth of body tissue cells. It refers to the development of a malignant tumour or abnormal formation of tissue. potency factor: Health effect information factor commonly used to evaluate health hazard potentials for carcinogens. It is usually represented by the cancer slope factor. Capillary fringe: Above the zone of saturation in the ground, capillary pores may exist which,
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if filled with water, form a zone of fringe of moisture higher than the true water table. This is the capillary fringe or zone of capillarity. number: It is used to assess the impact of viscous forces on the mobilization of non-aqueous phase liquid in porous media. Mathematically, it is the ratio of viscous force to surface tension force. water: The water held in the capillary or small pores of a soil, usually with a tension greater than 60 cm of water. See also moisture potential. Capping: Process to cover the wastes, prevent infiltration of excessive amounts of surface water, and prevent release of waste to overlying soil and the atmosphere. Capture zone: The portion of the aquifer that yields water to the well due to pumping. Carbon cycle: The natural circulation of carbon in the biosphere. Carbonation: It refers to the reaction of lime with carbon dioxide to form relatively weak cementing agents. It is the conversion of calcium carbonate into bicarbonate. Carcinogen: A chemical or substance capable of producing cancer in living organisms. Carcinogenesis: Development of cancer cells within an organism. Carcinogenic: Capable of causing, and tending to produce or incite cancer in living organisms. Carcinogenicity: The ability of a chemical to cause cancer in a living organism. Catabolism: Breaking down complex molecules. Energy is released during the process. Catalysis: The process of changing the velocity of a chemical reaction by the presence of a substance that remains apparently chemically unaffected throughout the reaction. In acid catalysis, a proton is transferred from the catalyst to the reactant while in base catalysis, a proton is transferred from the reactant to the catalyst. Catalytic cracking: A process in which heavy gas oils are converted into products of higher volatility by contracting the higher molecular weight hydrocarbon with the hot catalyst. oxidation" A chemical conversion process used predominantly for destruction of volatile organic compounds and carbon monooxide. Cathode: The electrode at which reduction occurs. It is the negative electrode in a cell which current is being forced, but it is the positive pole in a battery. In a vacuum, the cathode is the electrode from which electrons are liberated. Cation: A positively charged ion; during electrolysis it is attracted to the negatively charged cathode. exchange: The interchange between a cation in solution and another cation on the surface of any surface-active material such as clay or organic matter. exchange capacity (CEC): The sum of exchangeable cations that a soil can adsorb. Expressed as centimoles of charge per kilogram of soil. Cemented: Indurated; having a hard, brittle consistency because the particles are held together by cementing substances such as humus, calcium carbonate, or the oxides of silicon, iron, and aluminum. Centipoise: A standard unit of viscosity; equal to 0.01 poise. Water at 20~ has a viscosity of 1.002 centipoise or 0.01002 poise. Chelate: See coordination compounds. Chelating agents: Complex-forming agents having the ability to solubilize heavy metals. Chemical waste: Any solid, liquid, or gaseous material discharged from a process and that may pose substantial hazards to human health and the environment.
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Chlorite: A 2:l:l-type layer-structured silicate mineral having 2:1 layers alternating with a magnesium-dominated octahedral sheet. Chlorofluorocarbons: Organic compounds containing chlorine and fluorine. Chlorophyll: The green substance essential for photosynthesis. Chronic: Of long term duration. daily intake (CDI): The receptor exposure, expressed in mg/kg-day, averaged over a long period of time. exposure: The long term, usually low-level exposure to chemicals, i.e., the repeated exposure or doses to a chemical over a long period of time. It may cause latent damage that does not appear until a later period in time. toxicity: The occurrence of symptoms, diseases, or other adverse health effects that develop and persist over time, after exposure to a single dose or multiple doses of a chemical delivered over a relatively long period of time. Clay: (1) A soil separate consisting of particles less than 0.002 mm in equivalent diameter. (2) a soil textural class containing greater than 40% clay, less than 45% sand, and less than 40% silt. mineral: Naturally occurring inorganic material (usually crystalline) found in soils and other earthy deposits, the particles being clay size, that is, less than 0.002 mm in diameter. Clean-up: See remediation. Clod: A compact, coherent mass of soil produced artificially, usually by such human activities as plowing and digging, especially when these operations are performed on soils that are either too wet or too dry. Coal: An organic rock. Coarse texture: The texture exhibited by sands, loamy sands, and sandy loams (except very fine sandy loam). Cobblestone: Rounded or partially rounded rock or mineral fragments 7.5-25 cm in diameter. Cohesion: Holding together: force holding a solid or liquid together, owing to attraction between molecules. Decrease with rise in temperature. Coil: The term applies to one or more turns of a conductor when wound as a definite unit of an electrical circuit. Coke: Solid product produced by the carbonization of coal; also produced from petroleum during thermal processes. Colloid soil: Organic or inorganic matter with very small particle size and a correspondingly large surface area per unit of mass. Combustible liquid: A liquid with a flash point in excess of 37.8~ but below 93.3~ Cometabolic transformations: The degradation of toxic organic molecules by microorganisms which grow at the expense of a substrate other than the toxic organic one, without the use of the latter as an energy source. Complex: See coordinate bond and coordination compounds. Compost: Organic residues, or a mixture of organic residues and soil, that have been piled, moistened, and allowed to undergo biological decomposition. Composting: Biological decomposition of organic waste. Conductance: The reciprocal of resistance. It is measured by the ratio of the current flowing through a conductor to the difference of potential between its ends. The practical unit of
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conductance, the mho, the conductance of a body through which one ampere of current flows when the potential difference is one volt. The conductance of a body in mho is the reciprocal of the value of its resistance in ohms. Conduction: The transfer of heat by physical contact between two or more objects. Conductivity (electrical): The quantity of electricity transferred across unit area per unit potential gradient per unit time. Reciprocal of resistivity. See specific conductance. Cone of depression: The shape of the depressed water table "due to water extraction by pumping. The surface of the cone is referred to as draw-down-curve. The draw down becomes negligible beyond the radius of influence. Consequence: The impacts resulting from a receptor response due to specified exposures, or loading or stress conditions. Consistence: The combination of properties of soil material that determine its resistance to crushing and its ability to be moulded or changed in shape. Such terms as loose, friable, firm, soft, plastic, and sticky describe soil consistence. Constant charge: The net surface charge of mineral particles, the magnitude of which depends only on the chemical and structural composition of the mineral. The charge arises from isomorphous substitution and is not affected by soil pH. Containment: Emplacement of a physical, chemical, or hydraulic barrier to isolate polluted areas. Contaminant: Any chemical substance whose concentration exceeds backgrounds concentrations or which is not naturally occurring in the environment. migration: The movement of a contaminant from its source through other matrices/media such as air, water, or soil. migration pathway: The path taken by the contaminants as they travel from the polluted site through various environmental media. plume: A body of contaminated groundwater or vapour originating from a specific source and spreading out due to influences of factors such as local groundwater conditions or soil vapour flow patterns. It represents the volume of groundwater or vapour that contains the contaminants released from a pollution source. release: The ability of a contaminant to enter into other environmental matrices (e.g., air, water or soil) from its place/point of origin. Convection: The movement of a solute under the influence of a pressure gradient. Coordinate bond: It is established when an anion or a polar molecule donates an electron pair to a metal cation. The electron donor is often referred to as a metal complex. The formation of a coordinate bond is known as complexation or coordination. Coordination: See coordinate bond. compounds: Molecules or ions in which there is an atom (A) to which are attached other atoms (B) or groups (C) to a number in excess of that corresponding to the oxidation number of the atom (A). The atom referred to as (A) is known as the central or nuclear atom, and all other atoms which are directly attached to (A) are known as coordination atoms. Atom (B) and groups (C) are called ligands. A group containing more than one potential coordinating atoms is termed a multidentate ligand, the number of potential coordinating atoms being indicated by the terms unidentate, bidentate, etc. A chelate ligand is a ligand attached to one central atom through two or more coordinating atoms while a bridging group is attached to more than one atom. The whole assembly of one or more central atoms with their attached
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ligands is referred to as a complex, which may be an uncharged molecule or an ion of either polarity. A polynuclear complex is a complex which contains more than one nuclear atom, their number being designated by the terms mononuclear, dinuclear, etc. number: The number of ligands bonded to the central atom in a definite geometry. Corrective action: Action taken to correct a problematic situation. Corrosivity: A characteristic of substances that exhibit extremes of acidity or basicity or a tendency to corrode steel. Cost-effective alternative: The most cost-effective alternative is the lowest cost alternative that is technologically feasible and reliable, and which effectively mitigates and minimizes environmental damage. It generally provides adequate protection of public health, welfare, or the environment. Coulomb (unit of quantity of electricity): The quantity of electricity transported in 1 sec by a current of 1 Ampere. Coulomb law: The force between two point (electric) charges in free space is a pure attraction or repulsion. It is given by: F = {ql qz}/{4~Cor2} --- where ql and q2 are the magnitudes of the charges in coulombs, r is their separation distance in metres and eo is the permittivity of free space, 8.84x 10 .9 farads/metre. Covalent bonds: The strong forces that hold the atoms together in a molecule. For example, when two hydrogen atoms combine to form an H 2 molecule, the bond is often represented as H:H or H-H with the understanding that the two dots or the straight line drawn between the two hydrogen atoms represent a covalent bond. Cover management factor: A measure of the influence of vegetation systems and management variables on soil loss due to erosion. Creep: The term usually is associated with the slow plastic deformation under constant load. Criteria: Generic numerical limits or narrative statements intended as general guidance for the protection, maintenance and improvement of specified uses of soil and water. Critical micelle concentration: The point at which the surfactant molecules become arranged into clusters or aggregates (micelles). Micelles have hydrophobic interiors and hydrophilic exteriors. Crude oil: Petroleum in its natural state before refining. Crumb: A soft, porous, more or less rounded natural unit of structure from 1 to 5 mm in diameter. Crust: A surface layer on soils, ranging in thickness from a few millimeters to perhaps as much as 3 cm, that is much more compact, hard, and brittle, when dry, than the material immediately beneath it. Cryolite: Cryolite, sodium aluminum fluoride, crystallizes in the monoclinic system. Crystal: A homogeneous inorganic substance of definite chemical composition bounded by plane surfaces that form definite angles with each other, thus giving the substance a regular geometrical form. Crystal structure: The orderly arrangement of atoms in a crystalline material. Crystalline rock: A rock consisting of various minerals that have crystallized in place from magma. Crystallization: (1) Crystal formation from solution takes place when the solution is evaporated or cooled below the saturation point. (2) Formation of crystals by sublimation takes place
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when the vapour of a substance is condensed as a solid without passing through the liquid phase. Current (electric): The rate of transfer of electricity. Current unit is the ampere, a transfer of one coulomb per sec.
D Data quality objectives: Qualitative and quantitative statements developed by analysts to specify the quality of data that, at a minimum, is needed and expected from a particular data collection activity (or site characterization activity). It is determined based on the end use of the data to be collected. Debye length: A theoretical length which describes the maximum separation at which a given electron will be influenced by the electric field of a given positive ion. Decay constant ().): A constant relating the instant rate of radioactive decay of a radioactive species to the number of atoms (N) present at a given time (t). Thus, - (0N/c3t) = ~,N. If N Ois the number of atoms present at time zero then N = N Oe ~'t. Decision analysis: A process of systematic evaluation of alternative solutions to a problem where the decision is made under uncertainty. The approach is comprised of a conceptual and systematic procedure for analysing complex sets of alternative solutions in a rational manner in order to improve the overall performance of a decision-making process. framework: A management tool designed to facilitate rational decision making on environmental pollution problem. Decomposition: Chemical breakdown of a compound (e.g., a mineral or organic compound) into simpler compounds, often accomplished with the aid of microorganisms. De-flocculate: (1) To separate the individual components of a compound particles by chemical and/or physical means. (2) To cause the particles of the disperse phase of a colloidal system to become suspended in the dispersion medium. Deforestation: The change or transfer of forest land into no-afforestation purposes. Degradation: The physical, chemical, or biological breakdown of a complex compound into simpler compounds and by products. de Minimus: A legal doctrine dealing with levels associated with insignificant versus significant issues relating to human exposures to chemicals that present very low risk. It is the level below which one need not be concerned. Denitrification: A process by which the nitrogen compound is reduced into nitrogen gas. Deoxyribonucleic acid (DNA): Nucleic acids are compounds in which phosphoric acid is combined with carbohydrates and with bases derived from purine and pyrimidine. When D2-deoxyribose is present in the carbohydrate compounds, deoxyribonucleic acid is formed, generally termed DNA. It has been observed that DNA is a constituent of the chromosomes of the cell nucleus. Since the number of chromosomes of a cell and its daughter cells are equal, the quantity of DNA in the normal cell of any given species or type should be and is remarkably constant. This quantity does not change by starvation or other action or form of stress. DNA controls the transmission of hereditary characteristics from one generation to the next.
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Dermal exposure: Exposure of an organism or receptor through skin absorption. Desalinization" Removal of salts from saline soil, usually by leaching. Desertification: The drop in soil productivity in dry and semi-dry areas around the world. Desorption: The removal of sorbed material from surfaces. Detection limit" The minimum concentration that can be detected by a single measurement with a known confidence level. Instrument detection limit represents the lowest amount that can be distinguished from the normal "noise" of an analytical instrument, i.e., the smallest amount of a chemical detectable by an analytical instrument under ideal conditions. Method detection limit represents the lowest amount that can be distinguished from the normal "noise" of an analytical method, i.e., the smallest amount of a chemical detectable by a prescribed or specified method of analysis. Detergency" It is a measure of the surfactant ability to remove pollutants from solid surfaces and subsequently keep them suspended in the aqueous media. Surfactant detergency is a measure of its cleaning power. Detergency is expressed as the ratio of mass of dispersed and dissolved pollutants to volume of aqueous solution of surfactant. Detoxification: The biological conversion of a toxic substance to a less toxic species, which may still be relatively complex, or biological conversion to an even more complex material. Dialysis" A process for separating components in a liquid stream by using membrane. Dielectric: A material having low electrical conductivity. The principal properties of a dielectric are its dielectric constant (the factor by which the electric field strength in a vacuum exceeds that in the dielectric for the same distribution of charge), and its dielectric strength (the maximum potential gradient it can stand without breaking down). constant: For a given substance, the ratio of the capacity of a condenser with that substance as dielectric to the capacity of the same condenser with a vacuum for dielectric. It is a measure, therefore, of the amount of electrical charge a given substance can withstand at a given electric field strength; it should not be confused with dielectric strength. The dielectric constant (c) is a function of temperature and frequency and is written as a complex quantity e = c' - ie"--- where c' is the part that determines the displacement current and e" the dielectric absorption. For a non-absorbing, non-magnetic material e' is equal to the square of the index of refraction and the relation holds only at the particular frequency where these conditions apply. F = QQ'/e r --- where F is the force of attraction between two charges Q and Q' separated by a distance r in a uniform medium. Diffraction" The phenomena produced by the spreading of waves around and past obstacles which are comparable in size to their wave length. Diffuse ion layer: The net negative charge on clay surface is compensated by cations which are located on the layer surfaces. In the presence of water, these compensating cations have a tendency to diffuse away from the layer surface since their concentration is smaller in the bulk solution. On the other hand, they are attracted electrostatically to the charged layers. The result of these opposing trends is the creation of a distribution of the compensating cations in a diffuse electrical double layer on the exterior layer surfaces of a clay particle. Diffusion: The movement of atoms in a gaseous mixture or of ions in a solution, primarily as a result of their own random motion. coefficient (D): A measure of the rate of diffusion of a property through a medium. D is expressed as the ratio of mass flux to concentration gradient. It has dimensions of a length
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times a velocity.
Dioctahedral sheet: An octahedral sheet of silicate clays in which the sites for the six coordinated metallic atoms are mostly filled with trivalent atoms as A13§
Dipole: (1) A combination of two electrically or magnetically charged particles of opposite sign which are separated by a very small distance. (2) Any system of charges, such as a circulating current, which has the properties that: (a) no forces act on it in a medium field; (b) a torque proportional to sin 0 --- where 0 is the angle between the dipole axis and a uniform field, does act on it; (c) it produces a potential which is proportional to the inverse square of the distance from it. The hydrogen fluoride molecule can be described as a dipole, with the fluorine atom acting as a negative pole and the hydrogen atom as a positive pole. moment: The product of one of the charges of a dipole unit by the distance separting the two dipolar charges. Disintegration: Physical or mechanical breakup or separation of a substance into its component parts (e.g., rock breaking into its mineral components). Disperse: (1) To breakup compound particles, such as aggregated, into the individual component particles. (2) To distribute or suspend fine particles, such as clays, in or throughout a dispersion medium, such as water. Dispersion: (1) The spreading mechanism, due to variations in seepage velocity, is known as dispersion. (2) Statistically, dispersion is scale-dependent; it refers to the deviation from the mean. Dispersion forces: The force of attraction between molecules possessing no permanent dipole. The dispersion energy is given by: U~ = - {3/4 } •215{VoX0~2/r6} --- where h is Plank's constant, Vo a characteristic frequency, r the distance between molecules, and ~ the polarizability. Disposal: The discharge, deposit, injection, or placing of a waste onto, or into, a land facility. Dissolution: Process by which molecules of a gas, solid, or another liquid dissolve in a liquid, thereby becoming completely and uniformly dispersed throughout the liquid's volume. Distribution coefficient: The ratio of the solute concentration in water to that adsorbed by soil. Dormant period: It identifies the period in which the heat liberation rate is relatively low, during cement hydration. Dorn effect: See sedimentation potential. Dose: The amount of a chemical taken in by potential receptors on exposure. It is a measure of the amount of the substance received by the receptor, whether human or animal, as a result of exposure, expressed as an amount of exposure (in mg) per unit body weight of the receptor (in kg). Dose-response: The quantitative relationship between the dose of a chemical and an effect caused by exposure to such substance. curve: A graphical representation of the relationship between the degree of exposure to a chemical substance and the observed or predicted biological effects or response. evaluation: The process of quantitatively evaluating toxicity information and characterizing the relationship between the dose of a chemical administered or received and the incidence of adverse health effects in the exposed population. Double layer: See diffuse ion layer. Drying: Removal of a solvent or water from a chemical substance or soil. Dump site: A site used for the disposal of solid wastes without environmental controls or
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safeguards. Dyne: The force necessary to give acceleration of one centimetre per second per second to one gram of mass.
E Ecological cycles: Cycles involving land systems, water systems, and the atmosphere which are important to life. Ecology: The branch of science related to the study of the relationship of organisms and their environment. Economic sustainability: (1) The amount one consumes during a period and still be as well off at the end of the period. (2) Maintenance of capital in general, both man-made and natural. Ecosystem: (1) Odum defines an ecosystem as "any entity or natural unit that includes living and nonliving parts interacting to produce a stable system in which the exchange of materials between the living and nonliving parts follows circular path." (2) An ecological community, or living unit, considered together with nonliving factors of its environment as a unit. Ecotoxicity assessment: The measurement of effects of environmental toxicants on indigenous populations of organisms. Eddy current: A current induced in a mass of conducting material by a varying magnetic field. Effect: Local effect is defined as the response produced from a chemical contact with an exposed receptor that occurs at the site of first contact. Systemic effect is defined as the response produced from a chemical contact with an exposed receptor that requires absorption and distribution of the chemical and tends to affect the receptor at sites away from the entry point(s). Effluent: Any polluting substance, usually a liquid, that enters the environment via a domestic, industrial, agricultural, or sewage plant outlet. Electric conduction: Movement of ions in solution. conductivity: The capacity of a substance to conduct or transmit electrical current. In soils or water, measured in siemens/metre, and related to dissolved solutes. field strength: The magnitude of the electric field vector. field vector: At a point in an electric field, the force on a stationary positive charge per unit charge. Under conditions in which the ratio of force to charge is not constant, the field vector is defined as the limit of the ratio as the charge approaches zero. This may be measured in newtons per coulomb or in volts per metre. flux density: At a point, the vector whose magnitude is equal to the charge per unit area. potential: The electric potential at a point is the work done in moving a unit positive charge from the datum point to the point in question. The unit of potential is volt. The difference of potential between two points is the work done in moving a point charge from the first point to the second. It is equal to the difference in the values of the potentials at the respective points. Electrochemistry: That branch of science which deals with the interconversion of chemical and electrical energies, i.e., with chemical changes produced by electricity as in electrolysis. Electrochemical: A process in which ionic species in solution move to an electrode of opposite
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electric charge.
Electrode: In an electric circuit, part of which is composed of other than the usual conductor of copper, or other metal, the terminal connecting the conventional conductor and the conducting substances is an electrode. Electrodialysis: An extraction of dialysis that is used to separate the components of an ionic solution by applying an electric current to the solution thereby causing ions to move in preferred directions through the dialysis membrane. Electrokinetic potential: In colloidal systems, the difference in potential between the immovable layer attached to the surface of the dispersed phase and the dispersion medium. Electrolyte: A conducting solution. Electrolytic cell: It consists of a direct current source, electrodes, electrolyte, and electroactive species. Electrolysis: See electrochemical. Electromagnetic: The term electromagnetic is used to describe: (1) the combined electric and magnetic fields that are associated with movement of electrons through conductors, (2) the combined electric and magnetic effects exhibited by and used by equipment, apparatus, and instruments, (3) the radiation that is associated with a periodically varying electric and magnetic field that is travelling at the speed of light, such as light waves, radio waves, x-rays, gamma radiation, etc. Electromotive force: The electric potential difference (emf) between the terminals of any device which is used or may be used as a source of electrical energy, i.e., to supply an electric current. Electroneutrality: If n l is the number of moles of the ionic species i, and z, is the charge of an ion, the condition of electrical neutrality is: ~ z~ n~- 0. Electroosmosis: The movement of liquid with respect to a fixed solid (e.g., a porous diaphragm or a capillary tube) as a result of an applied electric field. Electrophoresis: A study of the migration of colloidal particles under the influence of an electric field. The motion of colloid particles is very similar to that of ions in an electric field. Electroplating: The coating of an object with a thin layer of some metal through electrolytic deposition. Electrostatistics: The branch of electromagnetism that deals with the effect of stationary (as opposed to moving) electric charges. The basic law of electrostatistics is Coulomb law. Emulsion breaking: The settling or aggregation of colloidal-sized emulsions from suspension in a liquid medium. Encapsulation: A process used to coat waste with an impermeable material so that there is no contact between the waste constituents and the surroundings. Endangered species: Any plant or animal species that no longer can be relied on to reproduce itself in numbers ensuring its survival. Endpoint: A biological effect used as index of the impacts of a chemical on an organism. Environment: The combination of physical and chemical (abiotic) factors and living (biotic) factors which influence the lives of individual organisms. Environmental fate: The ultimate and intermediary destinies of a chemical after release or escape into the environment, and following transport through various environmental compartments. It is the movement of a chemical through the environmental by transport in air, water, and
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soil culminating in exposures to living organisms. It represents the disposition of a material in various environmental compartments (e.g., soil, sediment, water, air, biota) as result of transport, transformation, and degradation. impact assessment (EIA): The interpretation of the significance of anticipated environmental changes related to a proposed project. mobility: The ability and/or ways with which a substance can move in the environment. sustainability: It seeks to protect the raw materials, and ensure that the capacity as a waste sink is not exceeded. technology: The application of scientific and engineering principles to the study of the environment with the goal of improving the environment. Enzymes: Large protein molecules composed primarily of amino acids twisted into complex shapes by peptide links and hydrogen bonding. Equilibrium constant: The product of the concentrations (or activities) of the substances produced at equilibrium in a chemical reaction divided by the product of concentrations of the reacting substances, each concentration raised to that power which is the coefficient of the substance in the chemical equation. Erosion: The wearing away of the land surface by running water, wind, ice, or other geological processes. Ettringite: It is an aluminosulfate mineral. It is a prismatic crystal with a hexagonal cross section. Eutrophic: Having concentrations of nutrients optimal for plant or animal growth. Evaportranspiration: The combined loss of water from a given area, and during a specified period of time, by evaporation from the soil surface and by transpiration from plants. Evaporation: A process for concentrating nonvolatile solids in a solution by boiling off the liquid portion of the waste stream. Exchange capacity: The total ionic charge of the adsorption complex active in the adsorption of ions. See also anion exchange capacity; cation exchange capacity. Exchangeable sodium percentage: The extent to which the adsorption complex of a soil is occupied by sodium. It is expressed as: ESP = exchangeable sodium/cation exchange capacity x 100. Explosives: (1) A solid, gas, or liquid material which, when triggered, will release great amount of heat and pressure by way of a very rapid, self-sustaining exothermic decomposition. (2) Chemicals which decompose spontaneously, or by initiation/simulation, with a rapid release of a high amount of energy. Exposure: The situation of receiving a dose of a chemical substance, or coming in contact with a hazard. It represents the contact of an organism with a chemical or physical agent available at the exchange boundary (e.g., lungs, gut, skin) during a specified period of time. assessment: The qualitative estimation, or the measurement, of the dose or amount of a chemical to which potential receptors have been exposed, or could potentially be exposed. It comprises the determination of the magnitude, frequency, duration, route, and extent of exposure to hazards of potential concern. conditions: Factors (e.g., location and time) that may have significant effects on an exposed population's response to a hazard situations. duration: The length of time that a potential receptor is exposed to the pollutants of concern in a defined exposure scenario.
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frequency: The number of times (per year or per event) that a potential receptor would be exposed to site contaminants in a defined exposure scenario. parameters: Variables used in the calculation of intake (e.g., exposure duration, inhalation rate, average body weight). pathway: The course a chemical or physical agent takes from a source to an exposed population or organism. It describes a unique mechanism by which an individual or population is exposed to chemicals or physical agents at or originating from a polluted site. point: A location of potential contact between an organism and a chemical or physical agent. route: The avenue by which an organism contacts a chemical, such as inhalation, ingestion, or dermal contact. scenario: A set of conditions or assumptions about sources, exposure pathways, concentrations of chemicals, and potential receptors that aid in the evaluation and quantification of exposure in a given situation. External surface: The area of surface exposed on the top, bottom, and sides of a clay crystal or micelle. Extracellular enzyme: An enzyme that is capable of acting outside an organism.
F Fabric filters: Filters made from fabric materials and used for removing particulate matter from liquid or gas streams. Facultative organism: An organism capable of both aerobic and anaerobic metabolism. Faraday's constant (F): It is equal to the charge carried by one mole of univalent cations. F 96485 C mole ~. Fauna: The animal life of a region or ecosystem. Feldspar: A group of aluminum silicate minerals commonly found in many types of rock. Ferrihydrite: A dark reddish brown poorly crystalline iron oxide that forms in wet soils. Field capacity: The percentage of water remaining in a soil two or three days after its having been saturated and after free drainage has practically ceased. Field sampling plan: A documentation that defines in detail the sampling and data gathering activities to be used in the investigation of a potentially polluted site. Filtration: The use of porous barriers to collect solids but allows liquid to pass. Fine-grained mica: A silicate clay having a 2:1-type lattice structure with much of the silicon in the tetrahedral sheet having been replaced by aluminum and with considerable interlayer potassium, which binds the layers together and prevents interlayer expansion and swelling, and limits interlayer cation exchange capacity. Fine texture: Consisting of or containing large quantities of the fine fractions, particularly of silt and clay. Fixation: (1) For other than elemental nitrogen: the process or processes in a soil by which certain chemical elements are converted from a soluble or exchangeable form to a much less soluble or to a non-exchangeable form; for example, potassium, ammonium, and phosphate fixation. (2) For elemental nitrogen: process by which gaseous elemental nitrogen is chemically combined with hydrogen to form ammonia.
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GLOSSARY--- F
Fixed hearth incinerators: Incinerators with single or multiple (non-mobile) hearths which are used for the combustion of liquid or solid wastes. Flammable: A substance that will bum readily. liquid: A liquid having a flash point below 37.8~ solid: A solid that can ignite from heat remaining from its manufacture, or which may cause a serious hazard if ignited. Flexible m e m b r a n e liner (FML): See lining systems. Flocculate: To aggregate or clump together individual, tiny soil particles, especially fine clay, into small clumps or floccules. Flora: The sum of all kinds of plants in an area at one time. Flotation: A process for removing solids from liquids using air bubbles to carry the particles to the surface. Flushing: A cleanup process in which the soil is left in-place and the water is pumped into and out of the soil in order to clean it. Flux: The rate of transfer of a pollutant from one domain to another. Fly ash: Particulate matter produced from mineral matter in coal that is converted during combustion to finely divided inorganic material and which emerges from the combustor in the gases. Free product: Chemical constituents that floats on groundwater. radical: An unsaturated molecular fragment in which some of the valence electrons remain free, i.e., do no partake in bonding. Examples are methyl (CH3o) or phenyl (C6H5~ radicals. Frequency: Rate of oscillation; units 1 cycle per sec = 1 Hertz. function: An expression giving the frequency of a variate-value x as a function of x; or, for continuous variate, the frequency in an elemental range dx. Unless the contrary is specified the total frequency is taken to be unity, so that the frequency function represents the proportion of variate-values x. From a more sophisticated standpoint the frequency function is most conveniently regarded as the derivative of the distribution function. The generalization to more than one variate is immediate. If the distribution is regarded as defining the probabilities of occurrence of the value of x, the frequency function is sometimes called the probability densityfunction and the distribution function itself is called the cumulative probability function. Friable: A soil consistency term pertaining to the ease of crumbling of soils. Fugacity: In thermodynamics, a measure of the tendency of a substance to escape by some chemical process from the phase in which it exists. Fugitive dust: Atmospheric dust arising from disturbances of particulate matter. Fugitive dust emissions consist of the release of chemicals from contaminated surface soil into the air, attached to dust particles. Fulvic acid: A term of varied usage but usually referring to the mixture of organic substances remaining in solution upon acidification of a dilute alkali extract from the soil. Fungi: Simple plants that lack a photosynthetic pigment. The individual cells have a nucleus surrounded by a membrane, and they may be linked together in long filaments called hyphae, which may grow together to form a visible body. Fungicide: A chemical used to control fungal action.
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G Gas: A state of matter in which the molecules are practically unrestricted by cohesive forces. A gas has neither definite shape nor volume. constant (R): The constant factor in the equation of state for perfect gases. The universal gas constant is R = 8.3143 joules/~ -- 1.9872 cal/~ Gaseous pollutants: Gases released into the atmosphere that act as primary or secondary pollutants. Geoaccumulation: The tendency of a substance to persist in the soil over a long term. Geosphere: The complex and variable mixture of minerals, organic matter, water, and air that make up soil; also known as land system. Geostatistics: It is a statistical method that is useful in situations where a sample value is affected by its location and its relationship with its neighbours, i.e., spatially correlated data. Gibbs free energy: See Gibbs function. function: A mathematical defined thermodynamic function of state, which is constant during a reversible isothermal process. Also, called Gibbs free energy, thermodynamic potential. Gibbsite: An aluminum trihydroxide mineral most common in highly weathered soils such as oxisols. Geothite: A yellow-brown iron oxide mineral that accounts for the brown colour in many soils. Glassification: Encapsulation of a waste in a glass-like coat; vitrification. Gram-negative: See Gram stain. Gram-positive: See Gram stain. Gram stain: A method of staining microorganisms which enables such organisms to be classified into two main groups, those which retain the stain being described as Gram-positive, and those from which the stain is decolorized being described as Gram-negative. The organisms are first stained with either gentian violet, or its analogue, crystal violet, and then treated with a solution of iodine. An organic solvent, usually alcohol, is then applied, which washes out the stain from Gram-negative organisms, leaving Gram-positive organisms with the violet stain unaffected. A counter stain of some contrasting colour is then applied to demonstrate the Gram-negative organisms. Gram-positive organisms include staphylococci, streptococci, pneumococci; among the Gram-negative are gonococci, meningococci, Bacillus coli, and the
salmonella. Granular structure: Soil structure in which the individual grains are grouped into spherical aggregates with indistinct sides. Highly porous granules are commonly called crumbs. Gravitational water: Water that moves into, through, or out of the soil under the influence of gravity. Greenhouse effect: The entrapment of heat by upper atmosphere gases such as carbon dioxide, water vapour, and methane just as glass traps heat for a greenhouse. Increase in the quantities of these gases in the atmosphere will likely result in global warming that may have serious consequences for humankind. gases: Gasses that contribute to the greenhouse effect. Groundwater: Subsurface water in the zone of saturation that is free to move under the influence of gravity, often horizontally to stream channels. Grouting: It is a modification technique used to: (1) strengthen the ground, (2) reduce hydraulic conductivity of fill voids, (3) solidify polluted soils.
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H Half-life: The time taken for half of the atoms in a given amount of radioactive material to decay. See biological half-life. Half-velocity constant: Substrate concentration at one-half the maximum growth rate. Hazard: The inherent adverse effect that a chemical or other object poses. It is that innate character which has the potential for creating adverse and/or undesirable consequences. It defines the chance that a particular substance will have an adverse effect on human health or the environment in a particular set of circumstances which creates an exposure to that substance. assessment: The evaluation of system performance and associated consequences over a range of operating and/or failure conditions. It involves gathering and evaluating data on types of injury or consequences that may be produced by a hazardous situations or substance. identification: The systematic identification of potential accident, upset conditions, etc. It is the recognition that a hazard exists and the definition of its characteristics. The process involves determining whether exposure to an agent can cause an increase in the incidence of a particular adverse health effect in receptors of interest. index (HI): The sum of several hazard quotients for multiple substances and/or multiple exposure pathways. quotient (HQ): The ratio of a single substance exposure level for a specified time period to a reference dose of that substance derived from a similar exposure period. ranking system (HRS): A scoring system used by the US EPA to assess the relative threat associated with actual or potential releases of hazardous substances at contaminated sites. The HRS is the primary way of determining whether a site is to be included on the National Priorities List (NPL). Hazardous substance: Any substance that can cause harm to human health or the environment whenever excessive exposure occurs. waste: A solid waste, or combination of solid wastes, which because of the quantity, concentration, or physical, chemical, or infectious characteristics may (1) cause or significantly contribute to an increase in mortality or an increase in serious irreversible, or incapacitating reversible, illness or (2) pose a substantial present or potential hazard to human health or the environment when improperly treated, stored, transported, or disposed of or otherwise managed. Heap leaching: A technology for ex-situ treatment of polluted soils. In the process, the polluted soil is excavated and mounted on a pad. The metals of interest are removed by passing extraction fluid through the polluted soil using a liquid distribution system. The extraction fluid is collected in a solution pit and processed to remove the metal of interest. Heavy metals: Metals with particle densities greater than 5 Mg/m 3. Hematite: A red iron oxide mineral that contributes red colour to many soils. H e n r y ' s constant: An indicator of the stripability of a volatile pollutants from the soil moisture phase. It is the ratio of the vapour pressure of a pollutant to its mole fraction in soil moisture. law: The mass of a slightly soluble gas that dissolves in a finite mass of a liquid at a given temperature is very nearly directly proportional to the partial pressure of the gas. This holds for gases which do not unite chemically with the solvent.
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Histogram: It is a graphical representation of a grouped frequency distribution. Rectangles are formed by using the class interval as the base and the frequency of the class as the height.
Herbicide: A chemical that kills plants or inhibits their growth; intended for weed control. Heterotroph: An organism capable of deriving energy for life processes only from the decomposition of organic compounds and incapable of using inorganic compounds as sole sources of energy or for organic synthesis. Contrast with autotroph. Human equivalent dose: A dose which, when administered to humans, produces effects comparable to that produced by a dose in experimental animals. health risk: The likelihood (or probability) that a given exposure or series of exposures to a hazardous substance will cause adverse health impacts on individual receptors experiencing the exposures. Humic acid: A mixture of variable and indefinite composition of dark organic substances, precipitated upon acidification of a dilute alkali extract from soil. substances: A series of complex, relatively high molecular weight, brown-to black-coloured organic substances that make up 60 to 80% of the soil organic matter and are generally quite resistant to ready microbial attack. Humidity: Mass of water vapour present in unit volume of the atmosphere, usually measured as grams per cubic metre. It may also be expressed in terms of the actual pressure of the water vapour present. Itumification: The processes involved in the decomposition of organic matter and leading to the formation of humus. Humin: The fraction of soil organic matter that is not dissolved upon extraction of the soil with dilute alkali. Humus: The more or less stable fraction of the soil organic matter remaining after major portions of added plant and animal residues have decomposed. Usually it is dark in colour. Hydration: Chemical union between an ion or compound and one or more water molecules, the reaction being stimulated by the attraction of the ion/compound for either the hydrogen or the unshared electrons of the oxygen in the water. Hydraulic conductivity: An expression of the readiness with which a liquid flows through a soil in response to a given potential gradient. Hydrocarbon: Organic chemicals/compounds, such as benzene, that contain atoms of both hydrogen and carbon. Hydrodynamic dispersion coefficient: The sum of diffusion and dispersion coefficients. Hydrogen bond: The chemical bond between a hydrogen atom in one molecule and a highly electro-negative atom such as oxygen or nitrogen in another polar molecule. Hydrologic cycle: The cycle of water movement from the atmosphere to the earth and back to the atmosphere through various stages or processes, as precipitation, interception, runoff, infiltration, percolation, storage, evaporation, and transpiration. Hydrolysis: The reaction between water and a compound (commonly a salt). The hydroxyl from the water combines with the anion from the compound undergoing hydrolysis to form a base; the hydrogen ion from the water combines with the cation from the compound to form an acid. Hydronium: A hydrated hydrogen ion (H30+), the form of the hydrogen ion usually found in an aqueous system.
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Hydrophilic: Having greater affinity for water, or water-loving. compounds: Substances that tend to become dissolved in water. effect: The reduction of unfavourable interactions between the ionic head group of the surfactant and the non-polar solvent molecules. Hydrophobic: Tending not to combine with water, or having less affinity for water. compounds: substances that tend to avoid being dissolved in water and are more attracted to nonpolar liquids (e.g., oils). effect: The drive to minimize the unfavourable interactions between water and the hydrophobic tail of the surfactant molecule. Hydrosphere: Water in various forms; total mass of free water in solid or liquid state on the surface of the earth, i.e., Oceans, lakes, rivers, glaciers, and ice. Hydrous mica: See fine-grained mica. Hygroscopic coefficient: The amount of moisture in a dry soil when it is in equilibrium with some standard relative humidity near a saturated atmosphere (about 98%), expressed in terms of percentage on the basis of oven-dry soil. Hydrosphere: The discontinuous envelope of water, both fresh and salt, which covers a major portion of the lithosphere.
Ideal gas: A gas which conforms to Boyle's law and has zero heat of free expansion. Igneous rock: Rock formed from the cooling and solidification of magma and that has not been changed appreciably since its formation.
Ignitability: Characteristic of liquids whose vapour are likely to ignite in the presence of an ignition source; also characteristic of non-liquids that may catch fire from friction or contact with water and that burn vigorously. lllite: See fine-grained mica. Immobilization: The conversion of an element from the inorganic to the organic form in microbial tissues or in plant tissues, thus rendering the element not readily available to other organisms or to plants. Impact assessment: See environmental impact assessment. Impervious: Resistant to penetration by fluids. Incineration: The controlled combustion of materials in an enclosed area. In-situ immoblization: A process used to convert waste constituents to insoluble or immobile forms that will not leach from a disposal site. treatment: A process that can be applied to wastes in a disposal site by direct application of treatment processes. Individual excess lifetime cancer risk: An upper bound estimate of the increased cancer risk, expressed as a probability, that an individual receptor could expect from exposure over a lifetime. It is a statistical concept and is not necessarily dependent on the average residency time in an area. Induction: The production of an electric charge or magnetic field in a substance by the approach or proximity of an electrified body, a magnet, or any other source of an electric or magnetic
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field. The term induction implies that there is a relatively non-magnetized medium between the body in which the electric or magnetic field is induced and the electrified body, or other source of the electric or magnetic field. field: This is the magnetic field set up around a conductor by the current in the conductor. If the field changes, it causes a back electromotive force to be induced in the conductor. The magnetic fields present around dc and low-frequency circuits are considered to be of this sort since the error in the assumption is entirely negligible. All the energy stored in this field is returned to the circuit when the current flow is stopped. Infiltration: The downward entry of water into the soil. Ingestion: An exposure type whereby chemical substances enter the body through the mouth, and into the gastrointestinal system. Inhalation: The intake of a substance by receptors through the respiratory tract system. Intake: The amount of material inhaled, ingested, or dermally absorbed during a specified time period. It is a measure of exposure, expressed in mg/kg-day. Inner-sphere surface complex: The complex that is obtained when no water molecule is interposed between the surface functional group and the ion or molecule it binds. Inoculation: The process of introducing pure or mixed cultures of microorganisms into natural or artificial culture media. Insecticide: A chemical that interferes in the life cycle of certain insects to control insect populations; the chemical may also kill insects. Intercalation: Penetration of an organic molecule into the inner spacing of a structure. Interlayer: Materials between layers within a given crystal, including cations, hydrated cations, organic molecules, and hydroxide groups or sheets. Internal surface: The area of surface exposed within a clay crystal or micelle between the individual crystal layers. Ion exchange: A means of removing cations or anions from solution onto a solid resin. Ionic bond: The electrostatic forces that hold the ions together, as between Na § and CI to form ordinary table salt NaC1. charge: Either the total charge carried by an ion or the charge carried by an ion which has unit charge. migration: The movement of charged particles of an electrolyte toward the electrodes under the influence of the electric current. mobility: The ratio of the average drift velocity of an ion in solution to the electric field. It is expressed as the ratio of ion conductance to Faraday constant. potential: The ratio of the charge on an ion to its radius. strength: A measure of the effectiveness of the forces restricting the freedom of ions in an electrolyte. It is defined as one-half the sum of the terms obtained by multiplying the total concentration of each ion by the square of its valence, i.e., g = { 1/2 } ~ ci zi2, where g is the ionic strength, c is ionic concentration, and z is valence. Ionization: A process which results in the formation of ions. Such processes occur in water, liquid ammonia, and certain other solvents when polar compounds (such as, acids, bases, or salts) are dissolved in them. Dissociation of the compounds occurs, with the formation of positively- and negatively-charged ions. The charges on individual ions being due to the gain or loss of one or more electrons from the outermost orbits of one or more of their atoms.
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potential: It is the energy per unit charge, for a particular kind of atoms, necessary to remove an electron from the atom to infinite distance. It is numerically equal to the work done in removing the electron from the atom, expressed in electron-volts. Ion pair: This term is used to denote: (1) A positive ion and a negative ion or electron, having charges of the same magnitude, and formed a neutral atom or molecule by the action of radiation or by any other agency that supplies energy. (2) As postulated in the Debye-HiJckel theory, in concentrated solutions of strong electrolytes (two or more), ions may occasionally approach each other so closely that they may form pairs (or groups) without entering into permanent chemical combination. Ions: Atoms, group of atoms, or compounds that are electrically charged as a result of the loss of electrons (cations) or the gain of electrons (anions). Immiscible solvent: A liquid that dissolves or extracts a substance from solution in another solvent without itself being very soluble in that the other solvent is termed an immiscible solvent. Interstratification: Mixing of silicate layers within the structural framework of a given silicate clay. Isomer: One of two or more molecules having the same atomic composition and molecular weight, but differing in geometrical configuration. Isomorphous substitution: The replacement of one atom by another of similar size in a clay crystal lattice without disrupting or changing the crystal structure of the mineral.
J Joule: The SI energy unit defined as a force of 1 newton applied over a distance of 1 metre; 1 joule = 0.239 calorie.
K Kaolinite: An aluminosilicate mineral of the 1:1 crystal lattice group; that is, consisting of single silicon tetrahedral sheets alternating with single aluminum octahedral sheets. Kelvin: The Kelvin, the unit of thermodynamic temperature, is the fraction 1/273.16 of the thermodynamic temperature of the triple point of water. The triple point of water, the fundamental reference point, is 273.16 ~ K. Krafft temperature: A characteristic temperature beyond which the solubility of non-aqueous organic chemicals will undergo a sharp increase. Kriging: It is an interpolation procedure. It is used to estimate the expected value of specially distributed variable from adjacent values.
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L Labile: A substance that is readily transformed by microorganisms or is readily available for uptake by plants. Land: An area of the earth's surface; the characteristics of which embrace all reasonably stable, or predictably cyclic, attributes of the biosphere vertically above and below this area including those of the atmosphere, the soil and the underlying geology, the hydrogeology, the plant and animal populations and the results of past and present human activity, to the extent that these attributes exert a significant influence on present and future uses of the land by man. farming: Spreading of waste on land for solid phase bioremediation; see bioremediation. Landfill: A site where waste is placed in or on land (often in separate trenches depending upon the waste and to prevent contact of reactive waste materials) and which should be lined to prevent leakage and runoff of the polluted surface water. Latent period: The time between the initial induction of a health effect from first exposure to a chemical and the manifestation or detection of actual health effects. Layer: A combination in silicate clays of tetrahedral and octahedral sheets in a 1:1, 2:1, or 2:1:1 combination. Leachability index: An index for comparing the relative mobility of pollutants. Leachate: Aqueous liquid generated when water percolates or trickles through waste materials or contaminated sites and collects components of those wastes. Leaching usually occurs at landfills as a result of infiltration of rainwater or snowmelt, and may result in hazardous chemicals entering soils, surface water, or groundwater. Leaching: The removal of materials in solution from the soil by percolating waters. Lewis acid: A chemical species which can accept an electron pair from a base. base: A chemical species which can donate an electron pair. Lifetime average daily dose (LADD): The exposure, expressed as mass of a substance contacted and absorbed per unit body weight per unit time, averaged over a lifetime. It is usually used to calculate carcinogenic risks. It takes into account the fact that, whereas carcinogenic risk values are determined with an assumptions of lifetime exposure, actual exposures may be for a shorter period of time. exposure: The total amount of exposure to a substance that a human would be subjected to in a lifetime. risk: Risk which results from life time exposure to a chemical substance. Ligand: See coordination compounds. Lignin: The complex organic constituent of woody fibers in plant tissue that, along with cellulose, cements the cells together and provides strength. Lignins resist microbial attack and after some modification may become part of the soil organic matter. Lime: Calcium oxide. Limestone: A sedimentary rock composed primarily of calcite. Lining systems: They are used to prevent a potential pollutant in a waste from migrating to the surface water, groundwater and food chain. Liner materials could vary in chemical composition from compacted soils to highly crystalline polymeric materials which are chemically resistant and have very low hydraulic conductivity.
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Liquid: A state of matter in which the molecules are relatively free to change their positions with respect to each other, but restricted by cohesive forces so as to maintain a relatively fixed volume. limit: The water content corresponding to the arbitrary limit between liquid and plastic states of consistency of a soil. Lithosphere: The minerals in the earth's crust. Loam: The textural-class name for soil having a moderate amount of sand, silt, and clay. Loam soils contain 7-27% clay, 28-50% silt, and 23-52% sand. Loamy: Intermediate in texture and properties between fine-textured and coarse-textured soils. Includes all textural classes with the words loam or loamy as a part of the class name, such as clay loam or loamy sand. Lowest observed adverse effect level (LOAEL): The chemical dose rate causing statistically or biologically significant increases in frequency or severity of adverse effects between the exposed and control groups. It is the lowest dose level, expressed in mg/kg body weight/day, at which adverse effects are noted in the exposed population. observed effect level (LOEL): The lowest exposure or dose level of a substance at which effects are observed in the exposed population. The effects may or may not be serious. Low-level waste: A type of radioactive waste depending upon the amount of radioactivity.
M Macronutrient: A chemical element necessary in large amounts (usually 50 mg/kg in the plant) for the growth of plants. Includes C, H, O, N, P, K, ca, Mg, and S. Macropores: Large soil pores, generally having a diameter greater than 0.06 mm, from which water drains readily by gravity. Mass flow: Movement of ions with the flow of water. Maximum daily dose (MDD): The maximum dose calculated for the duration of receptor exposure, and used to estimate risks for subchronic or acute noncarcinogenic effects of environmental pollutants. contaminant level (MCL): A legally enforceable maximum chemical concentration standard that is allowable in drinking water. contaminant level goal (MCLG): A nonenforceable health goal for public drinking water systems. Mean square difference: The mean square difference of n sample values x about a point a is defined as ~ ( x - a)2/n. Used as a measure of dispersion, a is usually chosen to be the sample mean, or the sample median about which the mean deviation is a minimum. Median: The median is the data value that is larger than half of the data set and smaller than the other half. Median lethal dose (LDs0): The dose at which 50% of the organisms remained alive, with reference to uptake of a chemical substance. lethal concentration (LCs0): The dose at which 50% of the organisms remained alive, with reference to inhalation of a chemical substance. Metabolism: The sum of all biological transformations.
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Metal complex: See coordinate bond. Methane: An odourless, colourless gas commonly produced under anaerobic conditions. When released to the atmosphere, methane contributes to global warming. Micas: Primary aluminosilicate minerals in which two silica tetrahedral sheets alternate with one alumina/magnesia octahedral sheet with entrapped potassium atoms fitting between sheets. They separate readily into visible sheets or flakes. Micelles: See critical micelle concentration. Microbe: A microscopic or ultramicroscopic organism (e.g., bacterium or virus) Microbial accumulation: The transformation of an element in a microbial tissue from inorganic to organic form, rendering the element not readily available to other organisms. Microfauna: That part of animal population which consists of individuals too small to be clearly distinguished without the use of a microscope. Includes protozoans and nematodes. Microflora: That part of the plant population which consists of individuals too small to be clearly distinguished without the use of a microscope. Includes actinomycetes, algae, bacteria, and fungi. Micronutrient: A chemical element necessary in only extremely small amounts ( less than 50 mg/kg in the plant) for the growth of plants. Examples are B, C1, Cu, Fe, Mn, and Zn. (Micro refers to the amount used rather than its essentiality). Microorganisms: Algae, bacteria, and fungi. Micropores: Relatively small soil pores, generally found within structural aggregates, and having a diameter less than 0.06 mm. Contrast to macropore. Migration: See ionic migration. Mineralization: The conversion of an element from an organic form to an inorganic state as a result of microbial decomposition. Minerals: Naturally occurring inorganic solids with well-defined crystalline structures. Mitigation: The process of reducing or alleviating a problem situation. Mobility: The migration behaviour of a substance within or between the soil, water, and air. Modelling: The use of mathematical equations to simulate and predict real events and processes. Moisture potential: See soil water potential. Molar solubilization ratio (MSR): The number of moles or organic compound per unit increase in the micellar surfactant concentration. Molal solution: It contains one mol/1000 g of solvent. Molar solution: It contains one mol or g mol wt of the solute in 1 litre of solution. Mole: The amount of a substance which contains as many elementary entities as there are atoms in 0.012 kg of carbon 12. Molecular weight: The sum of the atomic weights of all the atoms in a molecule. Molecule: A molecule is the smallest particle of any substance which can exist free and still exhibit all of the chemical properties of the substance. A molecule is a local assembly of atomic nuclei and electrons in a state of dynamic stability. Molecules are formed by the association of individual atoms. Monitoring: The measurement of concentrations of chemicals in environmental media or in tissues of humans and other biological receptors/organisms. Monomer: A single molecule, or a substance consisting of single molecules. Montmorillonite: An aluminosilicate clay mineral in the smectite group with a 2:1 expanding
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crystal lattice, with two silicon tetrahedral sheets enclosing an aluminum octahedral sheet. Isomorphous substitution of magnesium for some of the aluminum has occurred in the octahedral sheet. Considerable expansion may be caused by water moving between silica sheets of contiguous layers. Mulch: Any material such as straw, sawdust, leaves, plastic film, and loose soil that is spread upon the surface of the soil to protect the soil and plant roots from the effects of raindrops, soil crusting, freezing, evaporation, etc. Municipal solid waste: The heterogeneous mass of throw-away from residential and commercial sources, open areas and treatment plants.
N Naphtha: The volatile fraction of petroleum which is used as a solvent or as a precursor to gasoline.
Neurotoxicity: Hazard effects that are poisonous to the nerve cells. Neutralization: A process for reducing the acidity or alkalinity of a waste stream by mixing acids and bases to produce a neutral solution.
Nitrification: The biochemical oxidation of ammonium to nitrate, predominantly by autotrophic bacteria.
Nitrogen fixation: The biological conversion of elemental nitrogen (N2) to organic combinations or forms readily utilize in biological processes. Nonenzymatic transformation: An indirect process due to microbially-induced changes in environmental parameters such as pH and oxidation-reduction. Nonhumic substances: The portion of soil organic matter comprised of relatively low molecular weight organic substances; mostly identifiable biomolecules. Nonpolar bond: When two identical atoms joined together, as is the case in H 2 and F2. Nonpolar solvent: A solvent whose constituent molecule do not possess permanent dipole moments and which do not form ionized solutions. Non-specific adsorption: It is referred to adsorption in the diffuse ion layer. No observed adverse effect level (NOAEL): The chemical intakes at which there are no statistically or biologically significant increases in frequency or severity of adverse effects between the exposed and control groups. It is the highest level at which a chemical causes no observable adverse effect in the species being tested or the exposed population. effect level (NOEL): The dose rate of chemical at which there are no statistically or biologically significant increases in frequency or severity of any effects between the exposed and control groups. It is the highest level at which a chemical causes no observable changes in the species being tested or exposed populations under investigation.
O Obligate anaerobes: Bacteria that can only use electron acceptors other than oxygen (methane-
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producing bacteria). Octahedral sheet: Sheet of horizontal linked, octahedral-shaped, units that serve as the basic structural components of silicate minerals. Each unit consists of a central, six-coordinated metallic atom (e.g., A1, Mg, or Fe) surrounded by six hydroxyl groups that, in turn, are linked with other nearby metal atoms, thereby serving as inter-unit linkages that hold the sheet together. Octanol-water partition coefficient (kow) : The ratio of the amount of organic chemical in octanol to that in water. Off-site: Areas outside the boundaries or limits of a presumed polluted site. Ohm (unit of electric resistance): The electric resistance between two points of a conductor when a constant difference of potential of 1 volt, applied between these two points, produces in this conductor a current of 1 ampere; this conductor not being the source of any electromotive force. Ohm's law: It states that when electricity flows through an electrical conductor, the current density is proportional to the field strength: I = kE. The constant k, known as conductance, is the reciprocal of the resistance R; so ! = E/R. On-site: The boundaries or limits of a presumed polluted site. Organic carbon content (%): This reflects the amount of organic matter present, and generally correlates with the tendency of chemicals to accumulate in the soil or sediment. soil: A soil that contains at least 20% organic matter (by weight) if the clay content is low and at least 30% if the clay content is as high as 60%. Osmotic pressure: Pressure exerted in living bodies as a result of unequal concentrations of salts on both sides of a cell wall or membranes. Water moves from the area having the lower salt concentration through the membrane into the area having the height salt concentration and, therefore, exerts additional pressure on the side with higher salt concentration. Osmotic potential: See soil water potential. Outer-sphere surface complex: The complex that is obtained when at least one water molecule is interposed between the surface functional group and the ion or molecule it binds. Outersphere surface complexes involve electrostatic bonding mechanisms. Oxidation: The loss of electrons by a substance; therefore a gain in positive valence charge, and in some cases, the chemical combination with oxygen gas. Ozone: A highly reactive gas comprising triatomic oxygen (03) formed by recombination of oxygen in the presence of ultraviolet radiation.
P Parasites: Organisms living in close association with another organism, deriving its nourishment from the host, and harming the host organism in the process. Particulate matter: Particles in the atmosphere or on a gas stream that may be organic or inorganic and originate from a wide variety of sources and processes. Pathogens: Type of fungi that cause diseases in plants and animals. Pathway: Any specific route which environmental pollutants take in order to travel away from the source and to reach potential receptors or individuals.
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Permanent charge: see constant charge. Permeability: The ease with which gases or liquids pass through a bulk mass of soil. Permissible exposure limit (PEL): A maximum (legally enforceable) allowable level for a chemical in workplace air. Persistence: The degree to which a chemical substance remains in its original form in the environment. Pesticides: Plant protection agents or crop protection products. pH: The negative logarithm of the hydrogen ion activity (concentration) of a soil. The degree of acidity (or alkalinity) of a soil as determined by means of a glass or other suitable electrode or indicator at a specified moisture content or soil/water ratio, and expressed in terms of the pH scale. PH-dependent charge: That portion of the total charge of the soil particles that is affected by, and varies with, changes of the pH scale. Phosphate fixation: Formation of complex or insoluble metal-phosphate compounds. Photosynthesis: A process that converts light energy into chemical energy. Phytotoxic substances: Chemicals that are toxic to plants. Pica: The behaviour in toddlers and children (usually under age 6 years) involving the intentional eating/mouthing of large quantities of dirt and other objects. PKa: The pH value at which a compound dissociates into equal concentrations of anions and the original material. Plank's constant (h): A universal constant of nature which relates the energy of a quantum of radiation to the frequency of the oscillator which emitted it. It has the dimensions of action (energy• Expressed by E = h v - - - where E is the energy of quantum and v is its frequency. Its numerical value is 6.626176(36)• 1027 erg sec. Plastic limit: The water content corresponding to an arbitrary limit between the plastic and semisolid states of consistency of a soil. Plastic soil: A soil capable of being moulded or deformed continuously and permanently, by relatively moderate pressure, into various shapes. Point of zero charge: The pH value of a solution in equilibrium with a particle whose net charge, from all sources, is zero. Polar bonds: Bonds in which the electron distribution is unsymmetrical. Polar solvent: It consists of polar molecules, that is, molecules that exert local electrical forces. In such solvents, acids, bases, and salts, that is, electrolytes, in general, dissociate into ions and form electrically conducting solutions. Pollutants: See contaminant. Pollution: The introduction into the land, water, and air system of a chemical(s) that are not indigenous to those systems or the introduction into the land, water, and air systems of indigenous chemicals in greater-than-natural amounts. Polymeric materials: Polyvinyl chloride (PVC), chlorosulfonated polyethylene (CSPE), high density polyethylene (HDPE). Polymerization: This term is used to designate a reaction in which a complex molecule of high molecular weight is formed from a number of simpler molecules of the same or different sorts. Polynuclear complex: See coordination compounds.
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Population at risk (PAR): A population subgroup that is more susceptible to hazard or chemical exposures. It represents that group which is more sensitive to a hazard or chemical than is the general population. excess cancer burden: An upper bound estimate of the increase in cancer cases in a population as a result of exposure to a carcinogen. Pore size distribution: The volume of the various sizes of pores in a soil, expressed as percent of the bulk volume (soil plus pore space). Porosity: The volume percentage of the total soil bulk not occupied by solid particles. Portland cement: It is made by heating together lime stone and clay or some other sources of silica at about 2700 ~ F, forming a mass called clinker. A small amount of gypsum is added and the clinker is ground to a fine powder. Potency: A measure of the relative toxicity of a chemical. Potential, electric, (~): At any point, it is measured by the work necessary to bring unit positive charge from an infinite distance. The potential at a point due to a charge q at a distance r in a medium whose dielectric constant is c is, ~) = q/{c• Power: The time rate at which work is done. Units of power, the watt, one joule (ten million ergs) per ser the kilowatt is equal to 1000 watts. If an amount of work W is done in time t, the power or rate of doing work is P - W/t. Power will be obtained in watts if W is expressed in joules (107 ergs) and t in sec. Power developed by a direct current: The power in watts developed by an electric current flowing in a conductor is P =ExI = RxI 2 --- where E is the difference of potential at its terminals in volts, R its resistance in ohms, and I the current in amperes. The work done in joules in a time t sec is W = E•215 = R•215 Pozzolan: A material that contains silica or silica and alumina. It has little or no cementation value itself, but under certain conditions can react with lime to produce cementitious material. Pozzolanic reaction: It refers to the reaction of hydrous silica and alumina with calcium ions to form insoluble compounds. ppb(parts per billion): An amount of substance in a billion parts of another material. Also expressed by lag/kg or ~tg/1. ppm(parts per million): An amount of substance in a million parts of another material. Also expressed by mg/kg or mg/1. Precipitation: See solubility product. Primary mineral: It has not been altered chemically since deposition and crystallization from molten lava. pollutants: Pollutants that are emitted directly from the sources. waste treatment: Preparation for further treatment, although it can result in the removal of by-products and reduction of the quantity and hazard of the waste. Probability: The likelihood of an event occurring. density function: See frequency function. distribution: A quantity possesses a probability distribution if it takes any one of a set of values with different probabilities. Formally, this is the same as the frequency function, where the relative frequencies are interpreted as probabilities. Protein: Any of a group of nitrogen-containing organic compounds formed by the polymerization
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of a large number of amino acid molecules and that, upon hydrolysis, yield these amino acids. They are essential parts of living matter and are one of the essential food substances of animals. Pump and treat: Extraction of polluted groundwater from an aquifer with subsequent treatment at the surface and disposal or re-injection.
Q Quality assurance(QA): A system of activities designed to assure that the quality control system is performing adequately. It consists of the management of investigation data to assure that they meet the data quality objectives. This commonly includes designing appropriate protocols, ensuring they are carried out, and independently testing data quality. control (QC): A system of specific efforts designed to test and control the quality of data obtained in an investigation. It consists of the management of activities involved in the collection and analysis of data to assure they meet the data quality objectives. It is the system of activities required to provide information as to whether the quality assurance system is performing adequately. Activities include following the sampling protocols, and routinely checking calibration of laboratory equipment. Quantitative: Description of a situation presented in exact numerical terms. Quantum: Unit quantity of energy postulated in the quantum theory. The photons is a quantum of the electromagnetic field, and in nuclear field theories, the meson is considered to be the quantum of the nuclear field.
R Radio frequency heating: A process of increasing soil temperature to promote vaporization of pollutants.
Radius of influence: See cone of depression. R a n d o m variable: A variable which can take any one of a given set of values with assigned probability. In statistics, a particular value of random value is often referred to as a variatevalue, and sometimes variate is used as synonymous with random variable. Range: For stationary function, the length scale at which the sill is obtained describes the scale at which two measurements of the variable become practically uncorrelated. This length scale is known as range, correlation length or radius of neighbourhood. Reaction: The degree of acidity or alkalinity of a soil, usually expressed as a pH value. Reactive substances: Chemicals that undergo rapid or violent reaction under certain conditions. waste: Waste unstable under ambient conditions. Receptor: An object or location that is affected by a pollutant. Recycling: The use or reuse of chemical waste as an effective substitute for commercial products or an ingredient or feedstock in an industrial process. Redox potential: The electrical potential (measured in volts or millivolts) of a system due to the
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tendency of the substances in it to give up or acquire electrons.
Reduction: The gain of electrons, and therefore the loss of positive valence charge by a substance. In some cases, a loss of oxygen or a gain of hydrogen is also involved.
Reference concentration (RfC): A concentration of a chemical substance in an environmental medium to which exposure can occur over a prolonged period without an expected adverse effect. The medium in this case is usually air, with the concentration expressed in mg of chemical per m 3 of air. dose (RID): The maximum amount of a chemical that the human body can absorb without experiencing chronic health effects, expressed in mg of chemical per kg body weight per day. It is the estimate of lifetime daily exposure of a noncarcinogenic substance for the general human population (including sensitive receptors) which appears to be without an appreciable risk of deleterious effects, consistent with the threshold concept. Refraction: The bending of a light or energy wave as it passes through material with varying wave velocities. Regulated waste: Waste that have been singled out for regulatory control. Relative humidity: (1) The ratio of the quantity of water vapour present in the atmosphere to the quantity which would saturate at the existing temperature. (2) The ratio of the pressure of water vapour present to the pressure of saturated water vapour at the same temperature. Remedial action: Those actions consistent with a permanent remedy in the event of a release of a hazardous substance into the environment, meant to prevent or minimize such releases so that they do not migrate to cause substantial danger to present or future public health or welfare or the environment. alternative: An action considered in the feasibility study, that is intended to reduce or eliminate significant risks to human health and/or the environment at a polluted site. investigation (RI): The field investigation of hazardous waste sites to determine pathways, nature, and extent of pollution, as well as to identify preliminary alternative remedial actions. It addresses data collection and site characterization to identify and assess threats or potential threats to human health and the environment posed by a site. Remediation: The management of a contaminant at a site so as to prevent, minimize, or mitigate damage to human health or the environment. Remediation is a broader term than cleanup in that remediation options can include physical actions such as removal, destruction, and containment, as well as the use of institutional controls such as zoning designations or orders. criteria: Concentrations of substances in soil or groundwater which are intended as general guidance to protect and maintain specified uses of soil and water at contaminated sites. At concentrations greater than these criteria, the need for remediation is indicated. Remediation criteria can vary according to land use (e.g., agricultural, residential/park land, and commercial/industrial). Representative sample: A sample that is assumed not to be significantly different than the population of samples available. Repulsion: The electrostatic force between ions of the like charge (sign). Residual risk: The risk of adverse consequences that remains after corrective actions have been implemented. Resistivity: In electricity, a characteristic proportionality factor equal to the resistance of a centimetre cube of a substance to the passage of an electric current perpendicular to two
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GLOSSARY--- S parallel faces. Also called specific resistance.
Respiration: An energy generating process. It involves the transfer of electrons through the electron transport system.
Response: The reaction of a body or organ to a chemical substance or other physical, chemical, or biological agent.
Retardation factor: A measure of how much slower a solute migrates than water. Risk: (1) The probability or likelihood of an adverse consequence from a hazardous situation or hazard. (2) The potential for the realization of undesirable adverse consequences from impending events. It is a measure of the probability and severity of an adverse effect to health, property, or the environment. acceptance: The willingness of an individual, group, or society to accept a specific level of risk in order to obtain some gain or benefit. appraisal: The assessment of whether existing or potential biologic receptors are presently, or may in the future, be at risk of adverse effects as a result of exposures to pollutants originating at a polluted site. assessment: Estimation of the degree of adverse effects due to release of specific wastes from a particular site. control: The process to manage risks associated with a hazard situation. management: Application of the best available technologies for protecting human health and the environment. reduction: the action of lowering the probability of occurrence and/or the value of a risk consequence, thereby reducing the magnitude of the risk. specific dose (RSD): An estimate of the daily dose of a carcinogen which, over a lifetime, will result in an incidence of cancer equal to a given risk level. It is the dose associated with a specified risk level. Rock: The material that forms the essential part of the earth's solid crust, including loose incoherent masses such as sand and gravel, as well as solid masses of granite and limestone. Runoff: The portion of the precipitation on an area that is discharged from the area through stream channels. That which is lost without entering the soil is called surface runoff and that which enters the soil before reaching the stream is called groundwater runoff or seepage flow from groundwater.
S Salt: Any substance which yields ions, other than hydrogen and hydroxyl ion. A salt is obtained by displacing the hydrogen of an acid by a metal.
Sample blank: Blanks are samples considered to be the same as the environmental samples of interest except with regard to one factor whose influence on the samples is being evaluated. Blanks are used to ensure that pollutant concentrations actually reflect site conditions, and are not artifacts of the sample handling processes. The blanks consist of laboratory-prepared sample bottles of distilled or de-ionized water that accompany the empty sample bottles to the field as well as the samples returning to the laboratory, and are not opened until both the blanks and the actual site samples are analysed.
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duplicate: Two samples taken from the same source at the same time and analysed under identical conditions. Sand: A soil particle between 0.05 and 2.0 mm in diameter; a soil textural class. Sandstone: A sedimentary rock formed by compaction and cementation of sand grains; can be classified according to the mineral composition of the sand and cement. Saprophytes: Plants which obtain their food from non living organic material. Saturation extract: The solution extracted from a saturated soil paste. percentage: The water content of a saturated soil paste, expressed as a dry weight percentage. Screening: A process for removing particles from waste streams and used to protect downstream pretreatment processes. Secondary mineral: A mineral resulting from the decomposition of a primary mineral or from the re-precipitation of the products of decomposition of a primary mineral. See also primary minerals. pollutants: Pollutants produced by interaction of primary pollutants with another chemical or by dissociation of a primary pollutant or other effects within a particular ecosystem. Sedimentation potential: An electric field is created when charged particles move relative to stationary liquid. The movement of the particles could be under gravitational or centrifugal fields. This phenomenon is sometimes called the Dorn effect. Sedimentary rock: A rock formed from materials deposited from suspension or precipitated from solution and usually being more or less consolidated. The principal sedimentary rocks are sandstones, shales, and limestones. Selective sequential extraction (SSE): A technique used to selectively extract heavy metals from soil constituents (e.g., oxides, hydroxides, carbonates, soil organic matter). Semiarid: Term applied to regions or climates where moisture is more plentiful than in arid regions but still definitely limits the growth of most crop plants. Natural vegetation in uncultivated areas is short grasses. Semivariogram: It is a graph that is most commonly used in applied geostatistics to describe the special correlation between variables. It contains information about the scale of fluctuation of a variable. A measure of the similarity between variables (co-variance between two variables) for distance h is obtained. This is repeated for all samples that are h distance apart and the average squared difference divided by double the number of samples obtained. This similarity measure is called y(h). Theses are plotted on an x-y plot with the x-axis being the distance h, and 7(h) on the y-axis. Sensitive receptor: Individual in a population who is particularly susceptible to health impacts due to exposure to a chemical pollutant. Sewage effluent: The liquid part of sewage or wastewater, it is usually treated to remove some portion of the dissolved organic compounds and nutrients present from the original sewage. sludge: Settled sewage solids combined with varying amounts of water and dissolved materials, removed from sewage by screening, sedimentation, chemical precipitation, or bacterial digestion. Also called biosolids. Silica/alumina ratio: The molecules of silicon dioxide (SiO2) per molecule of aluminum oxide (A1203) in clay minerals or in soils. Sill: The value at which the semivariogram stabilizes. The sill is approximately equal to the
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variance of the data. Siloxane surface: The plane of oxygen atoms on the surface of a 2:1 layer silicate. This plane is characterized by the distorted hexagonal symmetry with its constituent oxygen atom. Silt: A soil consisting of particles between 0.05 and 0.002 mm in equivalent diameter. A soil textural class. Site assessment: Process used to identify toxic substances that may present at a site and to present site-specific characteristics that influence the migration of pollutants. categorization: A classification of sites to reflect the uniqueness of each site. Skewness: That property of frequency distribution involving its symmetry or asymmetry. Skewness is usually measures by the departure from zero of the third moment about the mean, standardized by division by o 3, where o is the standard deviation. Slag: A product of smelting, containing mostly silicates. Slope factor (SF): A plausible upper bound estimate of the probability of a response per unit intake of a chemical over a lifetime. It is used to estimate an upper bound probability of an individual developing cancer as a result of a lifetime of exposure to a particular level of a carcinogen. Slurry reactors: Tanks where wastes, nutrients and microorganisms are placed. Smectite: A group of silicate clays having a 2:1-type lattice structure with sufficient isomorphous substitution in either or both the tetrahedral and octahedral sheets to give a high interlayer negative charge and high cation exchange capacity and to permit significant interlayer expansion and consequent shrinking and swelling of the clay. Montmorillonite, beidellite, and saponite are in the smectite group. Sodium adsorption ratio: It is a ratio between sodium concentration and the square root of half the sum of calcium and magnesium concentrations. It is generally used to measure the soil dispersion potential. Soil: (1) A dynamic natural body composed of mineral and organic materials and living forms in which plants grow. (2) The collection of natural bodies occupying parts of the earth's surface that support plants and that have properties due to the integrated effect of climate and living matter acting upon parent material, as conditioned by relief, over periods of time. air: The soil atmosphere; the gaseous phase of the soil, being that volume not occupied by soil or liquid. alkalinity: The degree or intensity of alkalinity of a soil, expressed by a value greater than 7 on the pH scale. enrichment factor (SEF): The ratio of the concentration of chemical species in soil to those in crustal rock. erodibility factor: A measure of a soil's inherent susceptibility to erosion. erosion: The phenomenon in which the top soil is lost by wind and rain. flushing: A technology for in-situ extraction of toxic chemicals from polluted soil. An extraction fluid is applied to the un excavated undisturbed polluted soil. At the base of the polluted zone, the flushing fluid is recovered using subsurface drainage pipes, trenches, wells or well points. genesis: The mode of origin of the soil, with special reference to the processes or soilforming factors responsible for the development of the soil from the unconsolidated parent material.
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heaping: Piling wastes in heaps several feet high on an asphalt or concrete pad. moisture potential: See soil water potential. organic matter: The organic fraction of the soil that includes plant and animal residues at various stages of decomposition, cells and tissues of soil organisms, and substances synthesized by the soil population. Commonly determined as the amount of organic material contained in a soil sample passed through a 2-mm sieve. salinity: The amount of soluble salts in a soil, expressed in terms of percentage, milligrams per kilogram, parts per million (ppm), or other convenient ratios. solution: The aqueous liquid phase of the soil and its solutes consisting of ions dissociated from the surface of the soil particles and of other soluble materials. structure: The combination or arrangement of primary soil particles into secondary particle, units, or peds. The secondary units are characterized and classified on the basis of size, shape, and degree of distinctness into classes, types, and grades, respectively. vapour extraction: An in-situ technique for extracting vapours from polluted soils. water potential: A measure of the difference between the free energy state of soil water and that of pure water. Technically it is defined as "that amount of work that must be done per unit quantity of pure water in order to transport reversibly and isothermally an infinitesimal quantity of water from a pool of pure water, at a specified elevation and at atmospheric pressure, to the soil water, at the point under consideration. "The total potential consists of the following potentials: matric potential: The portion of the total soil water potential due to the attractive forces between water and soil solids as represented through adsorption and capillarity. It will always be negative. osmotic potential: The portion of the total soil water potential due to the presence of solutes in soil water. It will generally be negative. gravitational potential: The portion of the total soil water potential due to the differences in elevation of the reference pool of pure water and that of the soil water. Since the soil water elevation is usually chosen to be higher than that of the reference pool, the gravitational potential is usually positive. Solid waste: Garbage refuse, sludge from a waste treatment plant, water supply treatment plant, or air pollution control facility and other discarded material, including solid, liquid, semisolid, or contained gaseous material resulting from industrial, commercial, mining, and agricultural operation, and from community activities. Solidification: A term used to describe operations which improve the physical and handling characteristics of the waste. Solubility: A property of a substance by virtue of which it forms mixtures with other substances which are chemically and physically homogeneous throughout. The degree of solubility is the concentration of a solute in a saturated solution at any given temperature. product: A value dependent upon the temperature and the solvent, characteristic of electrolytes. It is the product of the concentrations of ions in a saturated solution and defines the degree of solubility of the substrate. When the product of the ion concentrations exceeds the solubility product, precipitation commonly results. Solubilization: The increased aqueous solubility of organic compounds beyond critical micelle concentration. It is the ratio of mass of dissolved pollutants to volume of aqueous solution
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GLOSSARY--- S of surfactant.
Solute: The constituent of a solution which is considered to be dissolved in the other, the solvent. The solvent is usually present in larger amount than the solute.
Solution mining: An ex-situ soil washing procedure in which polluted soil is excavated and placed in an agitated vessel with an extracting solution. When the soil is sufficiently clean, the solids are separated from the liquid, and the liquid is processed or reused. Solvent: The term solvent generally denotes a liquid which dissolves another compound to form a homogeneous liquid mixture in one phase. extraction: A process for extracting pollutants from soils by mixing soil suspension with a solvent that will extract the adsorbed pollutants onto the soil particles. See chelating agents. Source reduction: The reduction or elimination of a chemical waste at the source. Specific adsorption: Inner-sphere surface complexation is termed specific adsorption. Adsorption in this case occurs in the siloxane surface. See Inner-sphere surface complexation and siloxane surface. conductance: The specific conductance, or conductivity, of a conductor of electricity is the conductance of the material between opposite sides of a cube, 1 cm in each direction. The unit of specific conductance is ohm 1 crn ~ or mho crn -~. surface: The solid particle surface area per unit mass or volume of the solid particles. Spontaneous ignition: Ignition of a fuel, such as coal, under normal atmospheric conditions; usually induced by climatic conditions. Stabilization: The addition of chemicals or materials to a waste to ensure that the waste is maintained in its least soluble or least toxic form. Stationary function: A function is considered stationary if it consists of small-scale fluctuations (compared to the size of the domain) about some well-defined mean value. Statistics: It is a specialization that is generally concerned with the analysis and interpretation of uncertainty resulting from limited sampling of the variable of interest. Steam stripping: A process of injecting steam at 150 to 200 ~ C into the soil to increase the volatilization rate and solubility of pollutants. Stern layer: The layer between the surface of the clay particle and the first layer of adsorbed counter-ions. Across Stern layer, the potential distribution is linear. Stratosphere: Atmospheric layer directly above the troposphere. Streaming potential: When a pressure difference is applied between the ends of a porous material, a flow of liquid in induced through the porous material. As this flow carries the charge of the double layer with it, a potential difference arises between the ends of the porous material. Stripping: A means of separating volatile components from less volatile ones in a liquid mixture by the partitioning of the more volatile materials to a gas phase of air or stream. Structural analysis: Structural analysis is the selection and fitting of mathematical expressions for the required first two moments of the regionalised variable. Subchronic: Relating to intermediate duration, usually used to describe studies or exposure levels spanning 5 to 90 days duration. daily intake (SDI): The exposure, expressed in mg/kg-day, averaged over a portion of a lifetime. exposure: The short term, high level exposure to chemicals, i.e., the maximum exposure or
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doses to a chemical over a portion of a life time.
Surface acidity: The efficiency of a surface to act as an acid catalyst. impoundment: Use of natural depressions, engineered depressions, or diked areas for treatment, storage, or disposal of chemical water. tension: The tendency of one fluid to spread on or adhere to a solid surface in the presence of other immiscible fluids. It is a measure of surface wettability. water: Water in lakes, streams, and reservoirs. Surfactants: Surface active agents. Surfactant molecules have two distinct regions: hydrophobic (water disliking) and hydrophilic (water liking). A surfactant will accumulate at an oil-water interface with its hydrophobic tail in the oil phase and its hydrophilic head in the water phase. Surfactants are typically classified, according to the nature of their head group, as cationic, anionic, nonionic and amphoteric. Cationic surfactants are small groups of softening and coating agents with a positively charged polar solubilizing group -- usually an amino or a quaternary nitrogen functional group. Anionic surfactants normally have a sulphonatesulphate or phosphate functional group as their solubilizing group. Nonionic surfactants do not have a charge on their solubilizing group. They normally contain a polyoxyethylene group which acts as the solubilizing group. extraction: A process for extracting pollutants from soils by mixing soil suspension with surfactants that will extract the immiscible organic pollutants. See surfactants. Sustainable development: It is the development that meets the needs of the present without compromising the ability of future generations to meet their own needs. Synergism: An interaction of two or more chemicals that results in an effect that is greater than the sum of their effects taken independently. It is the effects from a combination of two or more events, efforts, or substances that are greater than that would be expected from adding the individual effects. Systemic: Relating to whole body, rather than individual parts of exposed individual or receptor.
T Terrace: A raised, more or less level or horizontal strip of earth usually constructed on or nearly on a contour and designed to prevent accelerated erosion by diverting water from undesirable channels of concentration; sometimes called diversion terrace. Tetrahedral sheet: Sheet of horizontally linked, tetrahedral-shaped units that serve as one of the basic structural components of silicate (clay) minerals. Each unit consists of a central fourcoordinated atom (e.g., Si, A1, Fe) surrounded by four oxygen atoms that, in turn, are linked with other nearby atoms (e.g., Si, A1, Fe), thereby serving as inter-unit linkages to hold the sheet together. Thermoplastic processes: Blending the waste with melted asphalt, polyethylene, or other thermoplastic additives. Thermosetting processes: Mixing the waste with reactive monomers. Threshold: The lowest dose or exposure of a chemical at which a specified measurable effect is observed and below which such effect is not observed. dose: The minimum exposure dose of a chemical that will cause toxic effects.
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GLOSSARY--- U
limit: A chemical concentration above which adverse health and/or environmental effects may occur. Through-solution mechanism: It describes the formation of a crystalline product at points remote from the dissolving phase, during cement hydration. Tolerance limit: The level or concentration of a chemical residue in media of concern above which adverse health effects are possible, and above which corrective action should therefore be undertaken. Topo-chemical mechanism: It describes the kinetics of growth and the orientation of the crystal in the product layer. Toxic: Harmful, or deleterious with respect to the effects produced by exposure to a chemical substance. wastes: Chemicals discharged from domestic and industrial sources that are harmful or fatal when ingested or absorbed. Toxicant: Any synthetic or natural chemical with an ability to produce adverse health effects. It is a poisonous pollutant that may injure an exposed organism. Toxicity: Defined in terms of standard extraction procedure followed by chemical analysis for specific substances. assessment: Evaluation of the toxicity of a chemical based on all available human and animal data. It is the characterization of the toxicological properties and effects of a chemical substance, with special emphasis on the establishment of dose-response characteristics. characteristic leaching procedure (TCLP): A test designed to determine the mobility of both organic and inorganic pollutants present in leachate. Trace elements: Those elements that occur at very low levels in a given system. Transpiration: The process by which water enters the plant's leaves from the atmosphere. Transport number: The fraction of the total current carried by one particular charge carrier. Treatment: Any method, technique, or process that changes the physical, chemical, or biological character of any chemical waste so as to neutralize the effects of the constituents of the waste. Trioctahedral: An octahedral sheet of silicate clays in which the sites for the six coordinated metallic atoms are mostly filled with divalent cations such as Mg 2+. Triple point: The thermodynamic state at which three phases of a substance exist in equilibrium. The triple point of water occurs at a saturation vapour pressure of 6.11 millibar and at a temperature of 273.16 ~ K.
U Uncertainty: The lack of confidence in the estimate of a variable's magnitude or probability of occurrence.
factor (UF): It is used to provide a margin of error when extrapolating from experimental animals to estimate human health risks.
Universal soil loss equation: An equation for predicting the average annual soil loss per unit area per year.
Unsaturated flow: The movement of water in a soil that is not filled to capacity with water.
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zone: The portion of the soil between the ground surface and the water table and includes the capillary fringe.
V Vadose zone: Unsaturated zone; the region in which the water in soil is below field capacity. Valence: The number of individual chemical bonds or linkages with which a particular atom can attach itself to other atom. van tier Waals forces: The weak forces that contribute to intermolecular bonding. There are two principal van der Waals forces. The most important force at short range is the repulsion between electrons in the filled orbitals of atoms on neighbouring molecules. The van der Waals repulsion energy is expressed as: E R = b e -at - - - where b and a are constants for two interacting atoms. The second van der Waals force is the attraction that results when electrons in the occupied orbitals of the interacting atoms synchronize their motion to avoid one another as much as possible. The energy resulting from this attractive force is known as the London energy, which varies inversely with the sixth power of the separation between atoms: London energy = - d / r 6 - - - where d is a constant and r is the distance between atoms. The total energy of van der Waals interactions is the sum of the attractive and repulsive energies. Vapour: The words vapour and gas are often used interchangeably. V a p o u r is more frequently used for a substance which, though present in the gaseous phase, generally exists as a liquid or solid at room temperature. G a s is more frequently used for a substance that generally exists in the gaseous phase at room temperature. Thus, one would speak of iodine or carbon tetrachloride vapours and of oxygen gas. pressure: The vapour pressure of a substance (solid or liquid) is the pressure exerted by its vapour when in equilibrium with the substance. Variable charge: See pH-dependent charge. Variance: In statistics, the variance of a population is the second moment about the mean, that is to say, the average of the square of deviations from the mean. It is the most commonly used measure of dispersion. In mechanics, the term refers to the number of degrees of freedom of the system. In physics and chemistry, the variance is the number of degrees of freedom of a system, or the degrees of freedom themselves. Variation (coefficient of): The coefficient of variation of a distribution isdefined as the standard deviation divided by the mean. Vat leaching: A procedure used to separate clay fractions based on their particle sizes. Vermiculite: A 2:l-type silicate clay usually formed from mica that has a high net negative charge stemming mostly from extensive isomorphous substitution of aluminum for silicon in the tetrahedral sheet. Viscosity: A property of fluids which appears as a dissipative resistance to flow. It is used as an index of the stability of liquids and calculated using Stokes' law. Vitrification: A phase conversion process in which a waste is melted at high temperature to form an impermeable mass; also known as galssification. Volatile: (1) Having a low boiling or subliming temperature at ordinary pressure. (2) Having a
652
GLOSSARY--- W high vapour pressure, as ether, naphthalene, benzene, or methyl chloride.
W Waste management: An organized system for waste handling leading to elimination or disposal in ways that protect the environment.
Water pollution: Any change in natural waters that may impair their further use, caused by the introduction of organic or inorganic substances.
table: The upper surface of groundwater or that level below which the soil is saturated with water. table, perched: The surface of a local zone of saturation held above the main body of groundwater by an impermeable layer of stratum, usually clay, and separated from the main body of groundwater by an unsaturated zone. Weathering: All physical and chemical changes produced in rocks, at or near the earth's surface, by atmospheric agents. Wetland: An area of land that has hydric soil and hydrophytic vegetation, typically flooded for part of the year, and forming a transition zone between aquatic and terrestrial systems. Wilting point: The moisture content of soil, on an oven dry basis, at which plants wilt and fail to recover their turgidity when placed in a dark humid atmosphere.
Z Zeta potential: See electrokinetic potential. Zero point of charge: The pH at which a mineral has no charge, or has equal amounts of negative and positive charges.
Zone of aeration: Unsaturated zone.
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INDEX
aeration, 60, 345 aerial photographs, 146 aerobic, 60, 172, 174, 394, 428, 589 biodegradation, 428 microbes, 589 aggregates, 470 aggregation, 79, 309, 319, 357, 394, 470, 491 number, 470, 491 agricultural, 4 pollution, 4 air, 1, 2, 6, 9, 11, 22, 28, 33, 47, 123, 136, 412 diffusivity, 412 permeability, 420 pollution, 6, 9 algea, 559 alkaline soils, 346, 445 alkalinity, 172, 556 amine, 83 amorphous, 59, 80, 125 amphibole, 64 reaction, 60 amphoteric, 467 groups, 467 anaerobic, 60, 172, 395,567, 589 bacteria, 567 biodegradation, 278 microbes, 589 animals, 8, 124, 345 amino acid, 111 anion, 61,105, 113,317 exchange capacity, 105, 113 exclusion, 317 anionic, 467 groups, 467 surfactants, 467 anode, 502 anodic reaction, 502 anthropogenic, 6, 11, 59 aquatic, 23, 125
A abiotic, 48, 126, 129, 209, 573 processes, 209 acceptable daily intake, 270 accumulation, 125 acid, 6, 59, 446, 473 buffering capacity, 59 dissociation constant, 473 leachate, 335 mine drainage, 37, 360 precipitation, 6 rain, 6 soils, 446 acid-base reactions, 172, 374 acidic, 346, 451 groups, 451 soils, 346 activated carbon adsorption, 406 active barrier system, 301 activity, 2 acute, 33, 51,130, 135 effect, 33,130, 132 mortality, 130 oral LD50, 138 tests, 137 toxicity, 446 adsorbate, 85, 89, 95 adsorbent, 85, 97 adsorption, 61, 65, 66, 80, 85,125,230, 374, 391 characteristics, 230 complex, 85 isotherm, 93, 96, 215 adsorption-desorption, 172, 304, 311 adsorptive, 85, 97 advection, 209, 210, 375 advection- dispersion, 219 adverse effect, 130, 267, 274, 346 adverse impact, 13 693
694
GEOENVIRONMENTAL ENGINEERING
systems, 125 aqueous, 467 solubilization, 467 arid climates, 336 artesian pressure, 283 assessment, 13 assimilative capacity, 22 pollution, 18, 22 atmosphere, 1, 2, 7, 8, 11,123,406 atmospheric pressure, 409 attraction, 61,308 reaction, 60 auger, 164 hollow-stem, 164 solid-stem, 164 autotrophs, 562
B background, 272 bacteria, 8, 345,559, 560 barrier, 59 system, 59 basal spacing, 75 base dissociation constant, 473 basic groups, 451 batch equilibrium test, 230 bentonite grout, 169 bioaccumulation, 23, 44, 126, 209 concentration, 139 effect, 124 bioavailability, 1, 11, 59, 66, 79, 123, 124, 126, 445,463 bioconcentration, 113, 133, 135, 139, 139 factor, 113, 133,135, 139 biodegradation, 113,135, 172, 209, 559,573 biological, 7, 11, 9, 28, 59, 130, 378 activity, 59 environment, 28 gas treatment systems, 406 processes, 374, 382, 394 systems, 130 biomagnification, 139
biomass, 67, 79, 125,428 biosphere, 1, 8, 9, 11,123 biota, 123, 143 biotic, 48,209, 573 barrier, 360 processes, 209, 216 biotransformation, 139, 573 rate, 132 bond, 67, 75 covalent, 68 electrostatic, 67 hydrogen, 68, 75 ionic, 67 van der Waals, 68 Bond number, 483 breakthrough, 230 brucite sheet, 71, 77 buffering capacity, 106, 374 bulk, 85, 211 diffusion, 211 solution, 85
C calcite, 396 calcium silicate hydrate, 534 calibration, 209 capillary barrier, 347 number, 483 capture zone, 300, 303,407 carbon, 6, 7, 10, 60 dioxide, 6, 10, 60 carbonate, 125, 127, 396 carbonation, 566 carboxyl, 83 carcinogenic, 33,270 carrying capacity, 22, 27 cathode, 502 cathodic reaction, 502 cation, 61, 85, 91 adsorption, 91 exchange, 63, 75-77, 83, 85, 106, 113,
INDEX 391, 411,477 cationic, 467 groups, 467 surfactants, 467 cement, 169, 320, 319, 530 -bentonite cutoff walls, 320 -bentonite method, 319 grout, 169 CFC, 7, 8, 10 charge, 500 chelate ring, 110 chelating agent, 449, 451 chelation, 110, 374 chemical, 3, 5, 11, 28 environment, 28 extraction, 433 industry, 5 potential, 505 processes, 209, 378, 382 chemical grouts, 321 acrylamide, 321 aminoplasts, 322 lignochrome, 322 phenoplasts, 322 silicate, 321 chemisorption, 445 chemotrophs, 563 chlorinated polyethylene, 399 chlorite, 74 chlorosulfonated polyethylene, 399 chronic, 33,130 daily intakes, 269 effect, 33, 44,130 exposures, 269 tests, 137 toxicity, 44, 135,446 clay, 63, 70, 77, 80, 85 structure, 70 surface, 85 clay barrier layer, 350 cleaning power, 471 clinker, 530 clod size, 354 closure, 336, 337 cloud point, 472
695 clusters, 353,470 co-disposal, 57 cometabolic transformations, 575 cometabolism, 573 compaction energy, 354 complementary error function, 220 complex formation, 374 complex ion, 389 complexation, 86, 90, 110, 172, 449, 587 compliance criteria, 304 compositional factors, 378 composting, 559 concentrate channel, 524 concentration, 22 gradient, 211 conceptual model, 209 conditional constants, 457 conductance, 158 conduction, 499 cone of depression, 282, 284, 296 confined aquifer, 282 constant, 80 surface charge, 80, 97, 520 surface potential, 80, 520 containment, 307 continuity equation, 215 coordinate bond, 449 coordination number, 61, 69, 110, 449 co-precipitation, 440 correlation length, 190 coprecipitation, 66 corrosive, 33 corrosivity, 45 counter-ion, 107 covalent, 86 bond, 86, 90, 111,390 convection, 508 cover, 54 cover management factor, 339 cracks, 376 cradle-to-grave, 49, 55, 58 creep, 383 crystallization, 535 cumulative distribution, 186 current, 156
696
GEOENVIRONMENTAL ENGINEERING
current density, 500
D decision process, 210 decontamination, 59 deep wells, 290 deforestation, 6, 9, 10 degradation, 2, 4, 11, 79, 124, 126, 129 de-nitrification, 346, 567 deprotonation, 80 dermal, 137 dermal contact, 130 desiccation, 344, 355,375,376 desertification, 7, 10 desorption, 65 detergency, 468, 471 dialysate channel, 524 dielectric constant, 61, 98, 104, 107, 108, 120, 318,360,511 chemical, 117 differential settlement, 383 diffuse ion, 88, 97, 308 layer, 308, 512 swarm, 88, 97 diffusion, 11, 83, 123,209, 211,304, 361, 375,414, 535,554 coefficient, 212, 509 dioxin, 5 direct contact, 123,266 dispersed, 410 dispersion, 11, 82, 83, 123,209, 213,467 dispersivity, 214 displacement grouting, 324 disposal, 5, 145 facilities, 145 dissociation, 75, 81 reactions, 570 dissolution, 89, 114, 386, 535 reaction, 89 dissolved, 3, 61 materials, 61 distribution coefficient, 112, 215, 231
dolomite, 396 domains, 353 domestic pollution, 4 dormant period, 533 Dom effect, 516 dose-response analysis, 266 dose-response ratio, 138 drainage, 54 draw-down, 282, 284 curve, 282 drilling, 165 air-rotary, 165 cable-tool, 165
E ecological, 18, 21 economics, 18 systems, 18, 21 economic, 8, 17, 18, 20, 27 criteria, 599 development, 17, 20 impact, 8 objectives, 27 sustainability, 18 system, 18 ecosystem, 2, 3, 6, 8, 9, 11, 18, 21, 24, 27, 47, 59, 83, 124, 125, 130, 132 approach, 24 engineer, 2 ecotoxicological, 124, 131 edge surfaces, 97 edge-to-edge, 309 edge-to-face, 309, 312 effect, 131, 140 parameters, 131 electric, 90, 97, 98, 104, 111,430, 489, 500 current, 500 double layer, 97, 111 field, 90, 430 potential, 98, 104, 489, 500 electrical, 63, 149 charge, 63
INDEX conductivity, 149, 154, 500, 501 potential, 107, 158 resistivity surveys, 149 electrode, 502, 504 potential, 504 electrodialysis, 523 electrokinetic, 328, 499, 510 electrolysis, 328 electrolyte, 106, 502 concentration, 106 electrolytic cell, 502 electromagnetic conductivity, 149 electromotive force, 507 electron, 581 acceptor, 581 donor, 581 electron-pair, 110 acceptor, 110 donor, 110 electronic conduction, 503 electroosmosis, 328, 511 electrophoresis, 328, 511 electroplating industry, 37 electrostatic, 88 attraction, 91 bond, 88, 111 environmental, 1, 2, 4, 12-15, 23, 24, 27-29, 33, 123, 144, 263,599 assessment, 13, 24 attributes, 29 conservation, 12 criteria, 599 degradation, 1, 13, 20, 274 education, 12 ethic, 12, 14 factors, 378 impact, 13, 24, 28, 147 loads, 12 management, 12, 17 mobility, 273 objectives, 27 pollution, 2, 5, 14 protection, 13, 14, 263 regulation, 33, 144 risk, 33
697 sustainability, 27 system, 23 enzymes, 563 equilibrium constants, 450 erosion, 7, 54, 336, 339 erosivity index, 339 error function, 220 estimation error, 198 ettringite, 531,538, 549 evapotranspiration, 347 exchangeable, 127 expected value, 191 exponential model, 192 exposure, 123, 124,126, 138, 144, 209, 263, 268 parameters, 144 extraction systems, 289
F fabric, 352 face-to-face, 309, 312 fate, 123,134, 140, 144,209, 263 parameters, 131, 144 pollutants, 263 fauna, 9, 113, 559 feldspar, 64 field, 181 blank, 181 capacity, 62, 341,347, 589 samples, 181 fixation, 346 flat layer, 97 flexible, 227, 229, 368, 399 membranes, 368, 399 wall, 229 wall permeameter, 227 flocculated, 410 flocculation, 308 flora, 9, 113,559 flow index, 380 fluid phase, 61 fraction of organic carbon, 215
698
GEOENVIRONMENTAL ENGINEERING
freeze-thaw, 336, 357 frequency of occurrences, 186 functional groups, 80, 82, 451,584 fungi, 561 funnel-and-gate system, 333
G gas, 59, 406, 419 flow, 406, 419 geoaccumulation, 59, 129 geochemical, 172 processes, 172 geocomposites, 349 geoenvironmental, 1, 2, 10, 15 geogrids, 349 geology, 1 geomembrane, 53,597 geomembrane barrier layer, 349 geonets, 349 geophone, 160 geosphere, 1, 11, 59 geophysical, 147 techniques, 147 geosynthetic, 347, 349 material, 347 geotextiles, 349 gibbsite sheet, 71 global, 4, 6, 10, 12, 14 environment, 10, 12, 14 population, 4 warming, 6, 10 Gram-negative bacteria, 588 Gram-positive bacteria, 588 gravitational water, 341 gravity drains, 289 greenhouse, 8, 9 effect, 8 gases, 9 ground penetrating radar, 149 groundwater, 1, 15, 33, 52, 53, 134, 136, 143, 166, 209, 265,288, 307, 405 disposal, 57
extraction, 307 flow directions, 146 flow system, 147 management, 1 path-line analysis, 288 pollution, 148 pumping, 281 recovery well, 406 transport, 209 grouting, 321 growth factors, 562 gypsum, 64
H half-life, 135,475,591 hazard, 33, 45, 129 index, 270 potential, 129, 131 hazardous, 2, 5, 6, 10, 15, 33, 143 liquid, 15 sites, 143 solid, 15 waste, 2, 5, 6, 10, 15, 33, 37, 44, 50, 143 waste landfill, 53 health, 2, 15,266 effects, 266 risks, 274 heap leaching, 435 heavy metals, 3, 37, 42, 124, 169 Henry's law constant, 132,405, 412 heterotrophs, 562 high density polyethylene, 399, 400 histogram, 186 human, 1, 2, 4, 10, 11, 21, 23, 28, 33, 124, 267 activity, 1, 2, 10 environment, 28 exposure, 267 health, 4, 15, 33, 46, 47, 274 humic, 60, 129, 482 substances, 129, 482
INDEX humid, 67, 336 climates, 336 humus, 79, 113 hydration, 61 shell, 61 hydraulic, 54, 61, 70, 82, 85, 146, 209, 210, 283,308, 315,352, 373,408, 485, 550 barriers, 289, 308, 362 conductivity, 54, 57, 61, 70, 82, 85, 146, 147, 227, 283,315,352, 373, 408,410, 485,550 gradient, 147, 209, 210, 283 management system, 289 volume defects, 376 hydrocarbon, 42 hydrodynamic dispersion coefficient, 214, 217 hydrogen, 7, 114 bonding, 114 hydrogeology, 1 hydrolysis, 132, 132, 386, 450, 474, 471,551, 581 rate, 132 hydrophile-lipophile balance (HLB) number, 468 hydrophilic, 466, 486 effect, 487 moiety, 466 hydrophobic, 112, 133,215,425,466, 486 bonding, 425,579 effect, 112, 133,487 interaction, 215 moiety, 466 hydrosphere, 1, 11,123 hydroxide, 127 hydroxyl, 71, 86, 397 groups, 86, 397
ignitability, 45 ignitable, 33 illite, 72, 76, 114
699
immiscible pollutants, 210 immobilization, 209 impact, 143 assessment, 143 impermeable cutoff walls, 307 incineration, 5, 40, 54 inclusion, 66 indirect contact, 123 industrial waste, 33 infiltration barrier, 337 information content, 185 ingestion, 124, 267 inhalation, 130 injection systems, 289 injection wells, 290 inner-sphere surface complex, 85, 95,390 inorganic chemical effect, 315 in-situ treatment methods, 281 interaction, 465 intercalation, 77, 553 interfacial tension, 467, 468,484 ion, 67, 542 exchange, 542 valence, 67 ion exchange membranes, 523 ionic, 61, 69, 73, 86, 467 bond, 73, 86 bonding, 67 charge, 61 conduction, 503 potential, 90 radius, 69, 89 strength, 451 ionization, 132 isoelectric point, 82 isoelectric weathering, 105 isomorphic substitution, 76 isomorphous substitution, 64, 73, 75, 76, 80, 86, 97
J jet grouting, 327
700
GEOENVIRONMENTAL ENGINEERING
K kaolinite, 64, 72, 114 Krafft temperature, 472 kriging, 198
L Lagrange multiplier, 199 land, 1, 2, 6, 7, 10, 11, 15, 21, 28, 33, 47, 50, 59, 61,143,266 development, 21 disposal, 15, 33, 50, 51, 56, 59, 61 environment, 10, 59 management, 15 sustainability, 21 use, 143, 144, 275 landfill, 22, 40, 52, 57 layer charge, 73 leachability index, 136, 555 leachate, 22, 35, 335,371 collection, 53 composition, 35 concentration, 57 corrosion potential, 160 plume, 22, 170 removal, 53 system, 53 leaching test, 136, 230, 553 Lewis acid, 82 sites, 578 ligand, 69, 110, 389, 449 ligand exchange, 114 linear model, 195 liner system, 53, 371 lipophilic tail, 466 liquid, 37, 59, 414 limit, 318 phase, 59 phase diffusion, 414 waste, 37, 38 lithosphere, 123 long-range forces, 391
low hydraulic conductivity barriers, 281
M macro-pores, 344 marine, 3, 10 oxygen, 3 pollution, 3, 10 matrix hydraulic conductivity, 375 mean function, 191 mean square difference, 187 mean square estimation error, 199 median, 187 median lethal, 138 concentration, 138 dose, 138 metabolic pathways, 559 metabolism, 124, 126 metal complex, 449 methane, 7 mica, 64, 67 micelle, 470 microbial, 129, 174, 562, 569, 579 accumulation, 579 activity, 174 growth, 562 mineralization, 569 microorganisms, 174, 395,559 micro-pores, 344 migration, 501 mine tailings, 33 mineral, 1, 59, 64 fraction, 64 primary, 64 secondary, 64 mineralization, 124, 135, 346 mining wastes, 36 miscible, 210 pollutants, 210 mixed-layer, 78 mobility, 60, 124, 134, 136, 445 pollutants, 60 mobility index, 137, 414
INDEX moieties, 466 molar absorptivity, 132 molar ratio, 64 molecular weight, 391 monitoring well, 147, 166, 170 installation, 147 materials, 147 mono-landfill, 57 montmorillonite, 67, 114 mortality, 130, 137 multi-dentate, 449 municipal solid waste, 33, 34, 35 mutagenic, 33
N natural, 2, 9, 14, 18, 20, 26 attenuation, 276 capital, 20, 26 ecosystems, 2 environment, 9, 20 resources, 14, 18 nitrification, 346, 567 nitrogen, 7 non-aqueous phase liquids, 169 nonionic surfactants, 467 non-specific adsorption, 88 nonstationary, 190 normal distribution, 201 nugget effect model, 194 numerical criteria, 268 nutrient, 3, 61,344, 345,562
O obligate anaerobes, 589 observational method, 144 octahedral sheet, 72 octanol-water partition coefficient, 112, 119, 132, 133,136, 360, 391,412 ogive, 186 oil pollution, 2
701
oleophilic, 467 olivine, 64 on-site, 6, 596 disposal, 6 storage, 596 oral, 137 organic, 3, 36, 94, 124, 129, 425 acids, 463 carbon content, 112 carbon fraction, 425 carbon partition coefficient, 215 chemical, 117, 124 chemical effect, 317 compounds, 36 ligand, 94 matter, 1, 125, 342 substances, 5 organism, 1, 9, 23 organochlorine, 5 osmotic velocity, 514 outer-sphere surface complex, 88, 393 oxidation, 60, 360, 551,581 oxidation rate, 132 oxidation-reduction, 135, 172, 374, 559 oxide, 65, 80 oxygen, 7, 362 flux, 362 ozone, 2, 7, 8, 10, 400 layer, 2, 8, 10 depletion, 8
P palygorskite, 74, 78 partial pressure, 60 particulate grouts, 321 bentonite grouts, 322 matter, 9, 38 Portland cement, 322 passive barriers, 307 pathway, 12, 263,265 PCB, 5 peds, 353
702
GEOENVIRONMENTAL ENGINEERING
penetration grouting, 324 permeable reactive walls, 307 persistence, 123, 134, 477 pesticide, 5, 10, 36, 125, 130 waste, 36 pH, 66, 91, 172 pH-dependent, 77, 394 pH-dependent charge, 80, 82 phenolic, 83 phosphate fixation, 128 photochemical degradation, 477 photolysis, 132 rate, 132 reaction quantum yield, 132 phototoxicity, 45, 126, 346 phototrophs, 563 physical, 11, 28, 209, 378, 382 change, 28 environment, 28 processes, 209, 378, 382 pickling solutions, 38 plants, 8, 124 plasticity index, 377 polarization, 90 pollutant, 1, 22, 52, 59, 61, 85, 123, 143 characteristics, 147 migration, 147 mobility, 85 pathways, 123,143 retention, 59 transport, 85, 143 polluted soil, 60 Pollution, 2, 5, 9, 10, 12, 36 impact, 36 prevention, 12 polymerization, 135,577 polyvinyl chloride, 399 pore, 1, 59, 232 fluid, 1 space, 59 volume, 232 porosity, 212, 480, 556 Portland cement, 530 portlandite, 549 potassium fixation, 86
potential, 61 matric, 61 osmotic, 61 pressure, 61 soil water, 61 potential determining, 80, 97, 104, 311, 312 anions, 312 ions, 80, 97, 104, 311 power model, 195 pozzolan, 529 pozzolanic reaction, 541 precipitation, 8, 89, 172, 332, 374, 388, 440 impact, 8 reaction, 89 precipitation-dissolution, 172, 304 reactions, 304 prediction, 210 pressure ridge, 289 primary, 64, 97, 541 bonds, 97 cementing product, 541 minerals, 64 probability density function, 188 protonation, 80, 112, 450 pulsed pumping, 300 pumping wells, 292 pyroxene, 64
Q quality assurance/quality control, 176 quartz, 64
R radio frequency heating, 430 radius of influence, 282, 408,409, 417 radius of neighbourhood, 190 rainfall intensity, 33 9 seasonal distribution, 339 random function, 191
INDEX range, 190, 191 rate constant, 216 reactive, 33,329 barriers, 329 reactivity, 45 receptors, 143, 147, 263 recharge basins, 290 redox, 66, 89, 129, 520 potential, 66, 129, 520 reaction, 89 reduction, 60, 551 refraction, 160 regionalised or field variable, 191 regulation, 5, 33 remedial, 2, 15, 143 268 action, 143, 170 measures, 15 options, 143 planning, 268 techniques, 143 remediation, 209, 270, 407 goals, 270 strategies, 209 technique, 407 repulsion, 308, 384 resistivity, 155 resource, 20, 22 non-renewable, 20 renewable, 20, 22 retardation, 209, 216, 231,230 coefficient, 216, 231 parameter, 230 retention mechanisms, 441 rigid wall permeameter, 227 risk, 263 assessment, 267, 269 management, 263,268 root time method, 243
S safe disposal, 15 salinity, 172
703
salt, 3,296 water intrusion, 296 sand, 63 Schlumberger spacing, 156 seams, 401 secondary, 64, 541 cementitious products, 541 minerals, 64 sediment, 125 encapsulation, 597 washing, 598 sedimentation potential, 511 seepage velocity, 210 seismic surveys, 149 selective sequential extraction, 91,127, 437 selectivity, 90 self-diffusion, 211 self-healing material, 364 self-sealing, 366 semivariogram, 188, 191 sensitivity, 209 sepiolite, 74, 78 shallow wells, 289 short-range forces, 391 shrinkage, 77, 355 silica tetrahedron, 64 silicate mineral, 125 sill, 191,203 siloxane, 86, 116, 393 cavity, 86, 116, 393 surface, 86 silt, 63 site, 143,209, 268 assessment, 209 characterization, 143,268 climate, 146 investigation, 218 topography, 146 skewness coefficient, 188 slope, 385 failure, 385 stability, 385 -steepness factor, 339 slurry walls, 308 smectite, 64
704
GEOENVIRONMENTAL ENGINEERING
social, 8, 18, 24, 27 capital, 20 justice, 24 objectives, 27 sustainability, 20 socioeconomic, 14, 132 soft acids, 390 soil, 1, 2, 5, 7, 33, 36, 51, 57, 59-61, 70, 79, 83, 85, 89, 124, 125, 136, 143,230, 263,320, 345,374, 378,405 acidity, 60 adsorption capacity, 70 adsorption coefficients, 413 aeration, 61 aggregate, 113 air, 59 attenuation, 230 -bentonite cutoff walls, 320 -bentonite method, 319 buffering capacity, 70, 374 carbonate, 59, 92 carrying capacity, 33 characteristics, 125 compaction, 61 composition, 125,380 erosion, 2, 7, 345 erodibility, 339 flushing, 435 hydraulic conductivity, 114 liner, 53 loss equation, 338 macro-pores, 411 material, 36, 378 micro-pores, 411 mineralogy, 410 moisture, 61,420 moisture content, 410, 425 moisture storage, 347 organic carbon, 425 organic content, 413 organic matter,.59, 60, 79, 83, 112, 129, 411 organisms, 130 pollutant, 85 solution, 61, 85
porosity, 410 structure, 79, 344, 378, 410 surfaces, 85 system, 59 temperature, 408 washing, 435 water, 60, 61, 85 venting, 405 soil-water, 132 organic carbon, 132 partition coefficient, 132 solid, 5, 37, 63, 66, 85 phase, 63, 66, 85 solution, 66 surface, 85 waste, 5, 37 waste management, 5 solidification, 529 solidification/stabilization, 596 solidified waste, 554 solubility product, 474 solubilization, 470 ratio, 494 solute concentration, 218 solution mining, 435 sludge, 38 specific, 86,149, 173,283, 515 adsorption, 86 conductance, 149, 173, 515 discharge, 283 spherical model, 194 split spoon sampler, 166 stability constant, 450, 456 stabilization, 529 standard deviation, 188 standard penetration test, 166 stationary, 190 stationary model, 191 steady state, 239 steam stripping, 420, 430 Stern layer, 107, 512 storage coefficient, 283 streaming potential, 511 structure, 85 structural analysis, 192
INDEX sub-lethal Effects, 130 subacute, 135 subsurface, 281,308 barriers, 308 control systems, 281 drains, 281 sulfide, 125 surface, 65, 90 acidity, 477 area, 64, 76, 77, 80, 83, 106 charge, 65, 80, 91, 104, 311 charge density, 97, 477 complexation, 90 functional group, 90 potential, 104 surface runoff, 347 tension, 468 water, 143 surfactant extraction, 465 support practice factor, 339 sustainable, 13, 15, 17, 18, 21, 22, 23, 26, 30, 263 development, 12, 13, 15, 17, 18, 21, 22, 23, 26, 30, 263,278 solutions, 21 yield, 22 swell, 77, 383
T tailing and rebound phenomena, 302 target concentration, 270 TCLP, 47, 51 technical criteria, 599 temperature, 8 teratogenic, 33 tetrahedral sheet, 72 thaumasite, 549 thermal destruction, 406 thermoplastic processes, 530 thermosetting processes, 530 through-solution mechanism, 536 time-lag method, 241
705
tobermorite, 534, 537 topo-chemical, 536 topographic factor, 339 tortuosity, 212, 556 toxic, 5, 123,266 chemicals, 123 effects, 267 material, 266 waste, 361 toxicity, 45,477 toxicological, 123,267 effects, 267 parameters, 132 trace elements, 66 transformation, 124 transient states, 239 transmissivity, 283,287 transport, 11, 82, 124, 239, 263,501 parameters, 239 pollutants, 263 processes, 11,124 number, 501 treatment, 10 tricalcium aluminate, 537
U unbiased estimates, 185 unconfined aquifer, 283 uni-dentate, 449 unsaturated soils, 341 zone, 209, 428 uptake, 268
V vadose zone, 209, 405 valence, 107 van der Waals interactions, 114 vapour, 60, 405,406, 465,476 density, 410
706
GEOENVIRONMENTAL ENGINEERING
extraction process, 60 extraction wells, 406 flow, 406 flow path, 408 phase, 405,476 phase diffusion, 414 phase pressure, 410 pressure, 132, 405,465,476 treatment unit, 406 velocity, 410 viscosity, 410 variable charge, 105, 106, 397, 521 mineral, 521 soil, 105,397 surface, 106 variance, 188 vat leaching, 435 vegetation, 146, 338 vermiculite, 64, 67, 77 volatile organic, 405, 414 compounds, 405 pollutants, 414 volatilization, 60, 209, 476 voltage, 156 volume, 375,382 change, 382 defect hydraulic conductivity, 375
W waste, 1, 2, 5, 10, 15, 20, 33, 34, 59, 83, 85, 209, 266, 336, 371,373 concentration, 49 containment, 83, 85,373 disposal, 5, 49 durability, 556 emission, 20 generation, 50 impoundment, 371 incineration, 49 leachate, 85 management, 5, 10, 15, 33, 34, 336 material, 15, 33,266
minimization, 50 policy, 55 recycle, 50 recycling, 33 reduction, 17, 33, 34, 49 regulated, 34 retention, 59 reuse, 33 rock, 37 separation, 49 source, 34 storage facility, 209 stream, 38 toxic, 34 treatment, 15, 33, 49 unregulated, 34 water, 15, 37, 42, 51 wastewater sludges, 365 water, 2, 3, 7, 10, 11, 18, 28, 33, 47, 63, 86, 123, 136, 412 balance method, 347 bridging, 113 capillary, 63 diffusivity, 412 flux, 210 free, 63 -holding capability, 341, 391 hygroscopic, 63 molecule, 86 pollution, 18 quality, 10 solubility, 132, 134 vapour, 7 weathering, 64, 67 stages, 67 well, 164, 166, 172, 176, 289 casing, 164 materials, 164 screen, 164, 166 point systems, 289 purging, 172, 176 Wenner spacing, 156 wet-dry, 385 wettability, 468 wilting point, 341,348
INDEX
Y yield stress, 314
Z zero point of charge, 82, 521 zeta potential, 311, 512
707
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