Environmental Contamination (Studies in Environmental Science)

ENVIR 0NME NTAL C 0NTAMINAT10N A selection ofpaperspresented at the 5th international Conference on Environmental Con...

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ENVIR 0NME NTAL C 0NTAMINAT10N

A selection ofpaperspresented at the 5th international Conference on Environmental Contamination, Morges, Switzerland, 29 September- 1 October 1992

Studies in Environmental Science 55

ENVIRONMENTAL C 0NTA MINAT I0N

Edited by

J.-P. Vernet Institute E-A. Forel University of Geneva Versoix, Switzerland

ELSEVIER Amsterdam

-

London

-

N e w York - Tokyo 1993

ELSEVIER SCIENCE PUBLISHERS B.V Molenwerf 1 P.O. Box211,1000AEAmsterdam,The Netherlands

L i b r a r y o f Congress C a t a l o g i n g - i n - P u b l i c a t

on D a t a

I n t e r n a t i o n a l C o n f e r e n c e on E n v i r o n m e n t a l C o n t a m i n a i o n ( 5 t h : 1 9 9 2 : Morges. S w i t z e r l a n d ) Environmental contamination : a s e l e c t i o n o f papers presented a t t h e 5 t h I n t e r n a t i o n a l C o n f e r e n c e on E n v i r o n m e n t a l C o n t a m i n a t i o n , M o r g e s , S w i t z e r l a n d , 29 S e p t e m b e r - 1 O c t o b e r 1992 / e d i t e d b y J . - P . Vernet. p. cm. -- ( S t u d i e s i n e n v i r o n m e n t a l s c i e n c e ; 55) I n c l u d e s 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 i n d e x . ISBN 0-444-89868-9 I. V e r n e t . J . - P . ( J e a n - P i e r r e ) 1. P o l l u t i o n - - C o n g r e s s e s . 11. T i t l e . 111. S e r i e s . TD172.5.1545 1992 363.77--dc20 93-3600 1 CIP ISBN: 0-444-89868-9

0 1993 Elsevier Science Publishers B.V. All rights reserved. No part of this publication may be reproduced, stored in a retrieval system or transmitted i n any form or by any means, electronic, mechanical, photocopying, recording or otherwise, without the prior written permission of the publisher, Elsevier Science Publishers B.V., Copyright & Permissions Department, P.O. Box 521,1000 A M Amsterdam,The Netherlands. Special regulations for readers in the USA - This publication has been registered with the Copyright Clearance Center Inc. (CCC),Salem, Massachusetts. Information can be obtained f r o m theCCCaboutconditionsunderwhich photocopiesofpartsofthispublication may be made inthe USA. All other copyright questions, includ-ing photocopying outside of the USA, should be referred to the copyright owner, Elsevier Science Publishers B.V., unless otherwise specified. No responsibility isassumed bythe publisher forany injury and/ordamageto personsor 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. This book is printed on acid-free paper Printed i n The Netherlands

Studies in Environmental Science Other volumes in this series 1 2 3 4 5 6

7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33

Atmospheric Pollution 1978 edited by M.M. Benarie Air Pollution Reference Measurement Methods and Systems edited by T. Schneider, H.W. de Koning and L.J. Brasser Biogeochemical Cycling of Mineral-Forming Elements edited by P.A. Trudinger and D.J. Swaine Potential Industrial Carcinogens and Mutagens by L. Fishbein Industrial Waste Management by S.E. Jmrgensen Trade and Environment: A Theoretical Enquiry b y H. Siebert, J. Eichberger, R . Gronych and R. Pethig Field Worker Exposure during Pesticide Application edited by W.F. Tordoir and E.A.H. van Heemstra-Lequin Atmospheric Pollution 1980 edited by M.M. Benarie Energetics and Technology of Biological Elimination of Wastes edited by G. Milazzo Bioengineering, Thermal Physiology and Comfort edited by K. Cena and J.A. Clark Atmospheric Chemistry. Fundamental Aspects by E. Meszaros Water Supply and Health edited by H. van Lelyveld and B.C.J. Zoeteman Man under Vibration. Suffering and Protection edited by G. Bianchi, K.V. Frolov and A. Oledzki Principles of Environmental Science and Technology by S.E. Jmrgensen and I. Johnsen Disposal of Radioactive Wastes by Z. Dlou hy Mankind and Energy edited by A. Blanc-Lapierre Quality of Groundwater edited by W. van Duijvenbooden, P. Glasbergen and H. van Lelyveld Education and Safe Handling in Pesticide Application edited by E.A.H. van Heemstra-Lequin and W.F. Tordoir Physicochemical Methods for Water and Wastewater Treatment edited by L. Pawlowski Atmospheric Pollution 1982 edited by M.M. Benarie Air Pollution b y Nitrogen Oxides edited by T.Schneider and L. Grant Environmental Radioanalysis by H.A. Das, A. Faanhof and H.A. van der Sloot Chemistryfor Protection of the Environment edited by L. Pawlowski, A.J. Verdier and W.J. Lacy Determination and Assessment of Pesticide Exposure edited by M. Siewierski The Biosphere: Problems and Solutions edited by T.N. Veziroglu Chemical Events in t h e Atmosphere and their Impact o n t h e Environment edited by G.B. Marini-Bettolo Fluoride Research 1985 edited by H. Tsunoda and Ming-Ho Yu Algal Biofouling edited by L.V. Evans and K.D. Hoagland Chemistry for Protection of the Environment 1985 edited by L. Pawlowski, G. Alaerts and W.J. Lacy Acidification and its Policy Implications edited by T. Schneider Teratogens: Chemicals which Cause Birth Defects edited by V. Kolb Meyers Pesticide Chemistry by G. Matolcsy, M . Nadasy and Y. Andriska Principles of Environmental Science and Technology (second revised edition) by S.E. Jmrgensen and I. Johnsen

34 35 36 37 38 39 40 41 42 43 44 45 46 47 48

49 50 51 52 53 54

Chemistry for Protection of the Environment 1987 edited by L. Pawlowski, E. Mentasti, W.J. Lacy and C. Sarzanini Atmospheric Ozone Research and its Policy Implications edited by T. Schneider, S.D. Lee, G.J.R. Wolters and L.D. Grant Valuation Methods and Policy Making in Environmental Economics edited by H. Folmer and E.van lerland Asbestos in Natural Environment by H. Schreier How t o Conquer Air Pollution. A Japanese Experience edited by H. Nishimura Aquatic Bioenvironmental Studies: The Hanford Experience, 1944-1984 by C.D. Becker Radon in the Environment by M. Wilkening Evaluation of Environmental Data for Regulatory and Impact Assessment by S.Ramamoorthy and E. Baddaloo Environmental Biotechnology edited by A. Blazej and V. Privarova Applied Isotope Hydrogeology by F.J. Pearson Jr., W. Balderer, H.H. Loosli, B.E. Lehmann, A. Matter, Tj. Peters, H. Schmassmann and A. Gautschi Highway Pollution edited by R.S. Hamilton and R.M. Harrison Freight Transport and the Environment edited by M. Kroon, R. Smit and J. van Ham Acidification Research in The Netherlands edited by G.J. Heij and T. Schneider Handbook of Radioactive Contamination and Decontamination by J. Severa and J. Bar Waste Materials in Construction edited b y J.J.J.M. Goumans, H.A. van der Sloot andTh.G. Aalbers Statistical Methods in Water Resources by D.R. Helsel and R.M. Hirsch Acidification Research: Evaluation and Policy Applications edited by T. Schneider Biotechniquesfor Air Pollution Abatement and Odour Control Policies edited byA.J. Dragt and J. van Ham Environmental Science Theory. Concepts and Methods i n a One-World, Problem-Oriented Paradigm by W.T. de Groot Chemistry and Biology of Water, Air and Soil. Environmental Aspects edited by J. Tolgyessy The Removal of Nitrogen Compounds from Wastewater by B. Halling-S~rensenand S.E. J ~ r g e n s e n

vii

P R E F A C E

It is becoming increasingly necessary that we face the problems ofglobal environmental contamination, taking into account interactions between the different contaminants and natural environments such as the atmosphere, soil, water and sediment, as well as flora and fauna. With fauna, the threat comes from man, his survival and his unhealthy need to refuse change and evolution. For him global warming is an aggression, even though it is intrinsic to the quaternary period that drastic and repeated changes in climate occur without any human interference. Those same variations have already allowed the moderate climate of the regions in the northern hemisphere to go through alternating glacial and interglacial periods. A new sociological dimension should thus be added to the simple observation of the natural environment and its contamination by man. This new dimension has resulted in making every gesture towards the environment a highly politicized act, as the environment has become a primordial dimension of our society, almost a religion for a growing number of Westerners. Scientific research is hindercd rather than helped by these political interferences.

In this field, fundamental and aiplied research tend to become mixed up. This sometimes goes too far, as in the case of atmospheric pollution. Before scientists elucidate the real weight of human activity on global warming, a dangerous amalgam has been made, giving man responsibility for climatic evolution and the hasty implementation ofenergy taxes. Such taxes are not, as declared, incitations to energy saving but an opportunity for political circles to increase state revenue - the heavy taxes on fuel have demonstrated the failure ofsuch a policy well. The environmental problems have become necessary for the society that generates them as a source ofincome. The growing importance of ecology has attracted men hungry for power, who can redirect our system’s dynamics to the benefit oftheir own ambitions. Others have a fundamentally different vision since, for them, ecology has become a new means ofslowing down our society’sdevelopment, or even to destroy it, causing a retrograde step. Finally, in an early developmental phase, industrial circles found themselves held back by this excessive ecological intervention. Later, a “green”industry developed in western countries that satisfied the new values and contented industrials for a while. Future development will take place in countries less concerned about the, often irrational, requirements of their ecological and political circles that penalize the industry in the western countries. For the above reasons, scientists will have to leave their ivory tower and communicate with the political and economical circles if this ecological drifting is to be limited. A worldwide extension of norms is necessary if we are to have a successful world economy. It is certain,

nevertheless that some measures are wasteful in terms of finance and energy.

...

Vlll

In the light of the above, consideration must be given to scientists from different fields. Economists, chemists, geologists, biologists, pedologists and sanitary engineers can now voice their opinions.

The first field considered in this work is atmospheric pollution. This has become a major problem of our time, and could soon develop into a scandal through the overestimation of the impact of green house gases such as ozone, NOxor SO, Which allows their use as political instruments. Another growing field of importance covers the fundamental questions posed by the problem of soils and their interaction with waste from human activities, for example, sterile mining waste, sewage sludge, and contaminated dredged material. How do these different elements react and how tolerant can our system be towards them? Continental aquatic systems are still matters ofprime importance, and pose such questions as what part do sediments play in the evolution of eutrophication and how do bacteria affect drinking water or bathing? A volume such as this, containing a selection of the most important work presented at the 5th

International Conference on Environmental Contamination, reflects some aspects of the present state ofresearch and our society’s dominant preoccupations. It can be noted that atmospheric pollution and the use by agriculture of wastes such as sewage sludge and contaminated dredged material, provide a major concern. We cannot conclude, however, without seizing this opportunity of synthesizing our knowledge on elements as important as mercury or selenium. Through their efforts and expertise, the different authors have contributed to the success of this book. The Editor would like to thank all the authors for their important contributions and Elsevier Science Publishers, Amsterdam, for their most appreciated assistance in bringing this to fruition.

Jean-Pierre Vernet

1 July 1993

F.-A. Fore1 Institute University ofGeneva Switzerland

IX

CONTENTS Preface

vii

List of Contributors

xiii

Chapter 1

Aspects of environmental contamination Ethical aspects ofenvironmental protection. B. B ~ G E N M E I E R .

Chapter 2

1

Atmospheric pollution Effects of air pollutants on man, animals, plants and buildings: mechanisms and dose-response effects.

S. HIPPELI and

E.F.

ELSTNER.

13

Effects ofair pollution on the condition of sessiIe oak forests in Hungary.

I. Mf?SZAROS, I. M6DY and M. MARSCHALL.

23

Long term effects of acid deposition: Implications on the performance

of high level nuclear waste repositories. J. NEBOT and J. BRUNO.

35

On character of ejection of radionuclides out of earth surface. I.V. MELIKHOV, 2. VUKOVIC and V. SIPKA. Chapter 3

65

Soils and contamination from mining uses Trace elements dynamics in soils and aquifers of western Switzerland.

0.ATEM, J.C. VEDY and A. PARRIAUX.

19

Leaching behaviour of granulated nonferrous metal slags. F.M.G. TACK, P.H. MASSCHELEYN and M.G. VERLOO.

103

Environmental impact of mining activities on the Hermioni area, Greece.

S.P. VARNAVAS,

KRITSOTAKIS.

A.G. PANAGOS

and

K.G.

119

x

Chapter4

Human impacts on soils by wastes and contaminated dredged material uses Beneficial and toxic effects of chromium in plants: solution culture, pot and field studies. J. BARCEL6, Ch. POSCHENRIEDER, M.D. V k Q v E z and B. GUNSE.

147

Interactive effects of the application of different Cd forms and an acidibing agent on plant available metals and postharvest soil extractability. R. NOGALES, D. HERVLS, J. SOT0 and F. GALLARDO-LARA.

173

Evolution of heavy metal species in leachates and in the solid phase during composting of municipal solid wastes. P. PRUDENT, C. MASSIANI and 0. THOMAS.

187

Effect ofseveral industrial wastes on soil respiratory activity. E. MARTI, R. CRUmAS, M.A. GARAU, E. DE MIGUEL and M.T. FELIP~.

217

Long-term evaluation of plants and animals colonizing contaminated dredged material placed in upland and wetland environments. D.L. BRANDON, C.R. LEE, J.W. SIMMERS, J.G. SKOGERBOE and G.S. WILHELM.

23 1

Contaminated aquatic sediments and waste sites as toxic chemical time bombs. U. FORSTNER.

259

The importance of biological testing in the assessment of metal contamination and site remediation. C.R. LEE, J.W. SIMMERS, D.L. BRANDON, L.J. ONEIL, M.J. CULLINANE and J.M. ROBERTSON. Chapter 5

293

Inland waters Recovery from eutrophic to oligotrophic states in lakes: role of sediments. D. SPAN, V. COPPEE, J. DOMINIK, G. BALVAY, F. BERTHIER, C. MARTIN and J.-P. VERNET.

303

xi

Dynamics of the autochthonous and contaminant bacterial colonization of lakes (lake of Cadagno and lake of Lugano as model systems). R. PEDUZZI, A. DEMARTA and M. TONOLLLZ.

323

The role of the bacterial community in the radionuclide transfers in freshwater ecosystems. F. HAMBUCKERS-BENIN, A. HAMBUCKERS and J. REMACLE.

337

Effects ofplants on the accumulation of Zn, Pb, Cu and Cd in sediments of the Tagus estuary salt marshes, Portugal. I. CAGADOR, C. VALE and F. CATARINO.

355

Studies on heavy metals of periphyton and its host plant / Phragmites australis (Cav.) Trien ex Steudel / in shallow lakes. G. LAKATOS.

365

Origin and pathways of Cadmium contamination in the Gironde Y. estuary, Garonne river and tributaries. J.M. JOUA"EAU, LAPAQUELLERlE and

Chapter 6

C.LATOUCHE.

373

Synthesisand methods Mercury pollution and cycling in aquatic systems. F.M. D I W .

391

Analysis of Selenium. M. SAGER.

403

Selenium occurence and ecology. M.SAGER.

459

AUTIIOR INDEX

475

SUBJECT INDEX

477

This Page Intentionally Left Blank

...

XI11

LIST OF CONTRIBUTORS

AlTEIA 0.

ATE-Pedology, Swiss Federal Institute ofTechnology (EPF), Lausanne, Switzerland.

BALVAY G.

Institut National de la Recherche Agronomique, Thonon lcs Bains, France.

BARCELO J.

Laboratorio de Fisiologia Vegetal, Facultad de Ciencias, Universidad Aut6noma de Barcelona, Bellatema, Spain.

BERTHIER F.

Syndicat Intercommunal do Lac d'Annecy (SILA), CranGevrier, France.

BRANDON D. L.

Environmental Laboratory, US Army Engineer Waterways Experiment Station, Vicksburg, Mississippi, U.S.A.

BRUNO J.

MBT Tecnologia Ambiental, Parc Tecnolbgic del Valles, Cerdanyola, Spain.

BURGENMEIER J.

Department of Economics, University ofGeneva, Geneva, Switzerland.

CACADOR I.

Departamento de Biologia Vegetal, Universidade dc Lisboa, Lisboa, Portugal.

CATARINO F.

Departamento de Biologia Vegetal, Universidade de Lisboa, Lisboa, Portugal.

COPPEE V.

Institute F.-A. Forel, University ofGeneva, Versoix, Switzerland.

CRURAS R.

Laboratori d'Edafologia, Facultat de Farmacia, Universitat de Barcelona, Barcelona, Spain.

CULLINANE M. J.


DEMARTA A.

Laboratory ofmicrobial Ecology, University ofGeneva, and Cantonal Institute of Bacteriology, Lugano, Switzerland.

D'ITRI F. M.

Institute of Water Research, Michigan State University, East Lansing, Michigan, U.S.A.

DOMXNXK J.


ELSTNER E. F.

Institut fur Botanik und Mikrobiologie, Biochemisches Labor, Technische Universitat Miinchen, Miinchen, Germany.

FELIPO M.T.

Laboratori dEdafologia, Facultat de Farmlcia, Universitat de Barcelona, Barcelona, Spain.

FURSTNER u.

Technische Universitat Hamburg-IIarburg, Arbeitsbereich Umweltschutztechnik, Hamburg, Germany.

xiv

GALLARDO-LARA F.

Estacion Experimental del Zaidin, Granada, Spain.

GARAU M.A.

Laboratori d’Eddologia, Facultat de Farmacia, Universitat de Barcelona, Barcelona, Spain.

GUNSE B.

Laboratorio de Fisiologia Vegetal, Facultad de Ciencias, Universidad Autonoma de Barcelona, Bellaterra, Spain.

HAMBUCKERS A.

Microbial Ecology, Department of Botany, University of Liege, Liege, Belgium.

HAMBUCKERS-BERHIN F.

Microbial Ecology, Department of Botany, University of Litge, Liege, Belgium.

H E R V ~ SD.


HIPPELI S.

Institut fir Botanik und Mikrobiologie, Biochemisches Labor, Technische Universitat Miinchen, Miinchen, Germany.

JOUANNEAU J. M.

DCpartement de GCologie et OcCanographie, UniversitC de Bordeaux I, Talence, France.

KRITSOTAKIS K. G.

Department of Mineralogy,University of Mainz, Mainz, Germany.

LAKATOS G.

Department of Ecology, Lajos Kossuth University, Debrecen, Hungary.

LAPPAQUELLERIE Y.

Dkpartement de Gkologie et OcCanographie,UniversitC de Bordeaux I, Talence, France.

LATOUCHE C.

DCpartement de GCologie et OcCanographie,UniversitC de Bordeaux I, Talence, France.

LEE C.R.


MARSCHALL M.

Botanical Department, Lajos Kossuth University, Debrecen, Hungary.

MARTI E.

Laboratori dEdafologia, Facultat de Farmacia, Universitat de Barcelona, Barcelona, Spain.

MARTIN C.

Institute F.-A. Forel, University ofGeneva, Versoix, Switzerland

MASSCHELEYN P. H.

Laboratory for Analytical Chemistry and Agrochemistry, University ofGent, Gent, Belgium.

MASSIANI C.

UniversitC de Provence, Laboratoire de Chimie et Environnement, Marseille, France.

MELIKIIOV I. V

Lomonosov State University, Moskow, Russia.

M k Z h O S I.


DEMIGUEL E.

Laboratori dEdafologia, Facultat de Farmacia, Universitat de Barcelona, Barcelona, Spain.

xv

MODY I.


NEBOT J.

MBT Tecnologia Ambiental, Parc Tecnolbgic del Vallts, Cerdanyola, Spain.

NOGALES R.


O’NEIL L. J.

Environmental Laboratory, US A m y Engineer Waterways Experiment Station, Vicksburg, Mississippi, U.S.A.

PANAGOS A. G.

Department ofApplied Geology, Technical University of Athens, Athens, Greece.

PARRIAUX A.

GEOLEP, Swiss Federal Institute ofTechnology (EPF), Lausanne, Suisse.

PEDUZZI R.

Laboratory of microbial Ecology, University ofGeneva, and Cantonal Institute of Bacteriology, Lugano, Switzerland.

POSCHENRIEDER Ch.

Laboratorio de Fisiologia Vegetal, Facultad de Ciencias, Universidad Autonoma de Barcelona, Bellaterra, Spain.

PRUDENT P.

Universite de Provence, Laboratoire de Chimie et Environnement, Marseille, France.

REMACLE J.

Microbial Ecology, Department ofBotany, University of Litge, Liege, Belgium.

ROBERTSON J. M.

Ware & Freidenrich, Attorneys at Law, Palo Alto, California, U.S.A.

SAGER M.

Landwirtschaftlich-Chemische Bundesanstalt Wien, Wien, Austria.

SIMMERS J. W.


v.

Institute of Nuclear Sciences VinEa, Belgrade, Yugoslavia.

SIPKA

SKOGERBOE J. G .


SOT0 J.


SPAN D.


TACK F. M. G.


THOMAS 0.

Universite de Provence, Laboratoire de Chimie et Environnement, Marseille, France.

TONOLLA M.

Laboratory of microbial Ecology, University of Geneva, and Cantonal Institute ofBacteriology, Lugano, Switzerland.

VALE C.

Instituto Nacional de InvestigaCIo das Pascas, Lisboa, Portugal.

xvi

VARNAVAS S. P.

Department of Geology, University of Patras, Patras, Greece.

V ~ Q U E ZM. D.

Laboratorio de Fisiologia Vegetal, Facultad de Ciencias, Universidad Autonoma de Barcelona, Bellatem, Spain.

d D Y J.-C.

IATE-Pedology,Swiss Federal Institute ofTechnology(EPF), Lausanne, Switzerland.

VERLOO M.G.


VERNET J.-P.


WKOVIC

2.

WILHELM G . S.

Institute of Nuclear Sciences Vin&, Belgrade, Yugoslavia. The Morton Arboretum, Lisle, Illinois, U.S.A.

CHAPTER 1 Aspects of environmental contamination


1

ETHICAL ASPECTS OF ENVIRONMENTAL P R O T E ~ I O N

B. Biirgenmeier' * Department of Economics, University of Geneva, 102 Bd Carl-Vogt, 1211

Geneva 11, Switzerland

INTRODUCTION I n the actual public debate about how the environment can be most effectively protected, there is increasing emphasis on economic instruments (Tietenberg, 1989). The pre-eminence of economics in social relations, is also extended to environment protection. The best results are supposedly achieved through optimum allocation of productive resources by means of the market mechanism. According to this view, scarcity can best be managed by market forces, backed u p by measures which are likewise market-oriented. This has consequently become the preferred environmental strategy i n the OECD countries (OECD, 1991a). The measures taken in order to protect the environment so far went to the opposite way. Direct control has taken over market incentives. The antagonism between norms and incentives still overshadows actual policy recommendations and takes it as granted tha t environment protection should be exclusively submitted to economic constraints. However, the need for a n ethical yardstick is clearly expressed (Etzioni, 1991). Why does this claim persist? This contribution seeks to give a n answer to this question. I t is divided into three parts. The first discusses the distinction between the positive and normative aspects of economics. The second part recalls the ethical content of the market as sole reference of the functionning of our society. Finally, the third part examines the issue of environmental protection in the light of both market and policy failures.

VALUE JUDGMENTS AND ECONOMIC SCIENCE The current school of thought claims to take ethics into account by portraying economics as both positive and normative. However, this interpretation is inadequate. The positive aspect of economics, reflecting a n approach common i n the natural sciences, is based on observable facts. I t highlights causal relationships

2

and seeks to analyse economic mechanisms on the basis of available information. However, the nature of this information, which is frequently unquantifiable and historically unique, has led to an emphasis on purely deductive methods, which form the basis for many of the conclusions in mainstream economics. This current of thought is typified by the neoclassical school, which takes as its main premise the much-disputed hypothesis of economic rationality. Human behaviour is deemed to be rational when consumers maximize their satisfaction and producers their profits, subject t o constraints such as differing incomes and production costs respectively. Optimal strategies for both consumers and producers can best be determined by the mechanism of market forces. This leads t o the important conclusion that maximum economic welfare is the result of individual strategies, and that society does not exist as a separate entity but is merely the sum total of individuals behaving rationally. The key issue here is efficiency. The normative aspect of economics involves value judgements. It is therefore based on a subjective portrayal of the facts and considers the economy as it should be, using criteria which can only be identified by the collective decision-making mechanism of political choice. Economic theory acknowledges that each economic agent is perfectly free t o express value judgements, and thus there may conceivably be as many different points of view as there are agents. If one point of view is to prevail, it can only be the majority view, identified as such by a democratic process. Here the key issue is equity, embodied in laws and regulations which represent an institutional codification of society’s moral and ethical judgements. This rough outline of the positive and normative aspects of economics provides a conceptual link between efficiency and economics on the one hand, and equity and politics on the other. Seen from this point of view, the market economy is indissociable from democracy - neither is conceivable without the other. Accordingly, strict application of the principles of the market economy and political democracy should eliminate the need for a separate ethical approach to economics (Novak, 1987); ethical issues can then be taken into account simply by sustaining both of these collective decision-making mechanisms. Despite this essential conclusion, which is central t o the functioning of Western industrial society, recent years have seen a rapidly growing need to introduce ethical considerations into economics. This is because both the boundary between the normative and positive aspects of economics and the distinction between efficiency and equity are unclear, for the following reasons (Biirgenmeier, 1992): - The distinction between the normative and positive aspects of economics and their linking with politics and economics respectively are, of course, gross oversimplifications. Political logic displays certain rational traits, and value judgements play an undoubted part in economic behaviour. A great deal of rhetoric - some would say ideological bias - has gone into presenting positivism as the only rational standard for human economic behaviour. Not only is the

3

definition of economic rationality a tautology, but human behaviour and institutions quite definitely influence one another. The definition of economic behaviour is tautological in that a deductive approach is used to obtain results compatible with the economic rationality on which much of economic policy is based. This is not to say that the hypothesis of rational economic behaviour is wrong, or to deny that it has profoundly influenced our understanding of how society functions. However, it does mean that economic models based on this hypothesis are bound to treat social issues as mere problems of constrained optimization. Not only is this likely to result in circular reasoning, but inherent in the market model are value judgements which are not separate from the economic sphere but specific to it. We can then no longer relegate the normative aspects of economics to the political sphere and treat economics as a scientific discipline in the pure positivist tradition. Were ethcs nothing more than the expression of value judgements, there would still be a place for it in economics. As for the influence of economic behaviour on institutions and vice versa, it must be acknowledged that institutional criticism of the market model has not succeeded in diverting the mainstream of economic thought. Mainstream economics is eager t o prove the existence of economic laws which are independent of place and time. In contrast, a n apparent common thread in the institutional criticism, and the basis for socio-political approaches to the economy, is the idea that economics is culturally determined and cannot be viewed in isolation from Western cultural history. Once again, were ethics nothing more than a cultural phenomenon, there would still be a place for it in economics. - The notion of the market as an efficient allocator of factors of production is tending to be extended to areas other than economics. Ceaseless intellectual effort has gone into attempts to apply economic reasoning t o such fields as law, politics, sociology and medicine, on the principle that efficiency must be achieved before there can be any question of equity. Such attempts suggest a wish to see production constraints take priority over all other social, o r indeed ethical, considerations.

THE ETHICAL REFERENCE OF THE MARKET The market is merely one of a number of collective decision-making mechanisms. The others include the democratic process and the civil service. Comparing these mechanisms, one can observe a tendency for decisions to be reached not by market forces or by the democratic process, but by private or public bureaucracies. This shift towards a more administrative form of collective

4

decision-making has caused a change of attitude. Not only are there new groups of voters, but society’s image of the economy has changed, transferring the emphasis from efficiency to equity. The market mechanism has begun to make way for more political decision-making mechanisms in which problems of organization have supplanted problems of trade. This development is made easier by the neoclassical misapprehension of the role of the State, and by the problems of redistribution of income and wealth that are inherent in the functioning of any market economy. Neoclassical theory sees growth as the answer to problems of equity, and suggests that optimal growth will be achieved by the operation of individualism in a harshly competitive market. Politics, on the other hand, proposes to deal with problems of equity by creating conditions conducive to the development of solidarity. In a society characterized by numerous market failures and a lack of solidarity, the emergence of bureaucratic mechanisms and the need for an ethical frame of reference are thus inevitable. There is thus a place in economics for an ethical component which reflects particular value judgements or a particular culture and is backed up by the various collective decision-making mechanisms society has established. All that is necessary is to acknowledge that economic theory is essentially normative and that it is merely one part of the social whole. Should ethics then be assigned a universal value and also be approached in positivist terms? An example of such an approach is the demonstration of the universal nature of human rights. If ethics is indeed an inherent feature of human nature, then an extra dimension needs to be added to the hypothesis of rational behaviour by economic agents. Failing this, a positivist approach t o values will exclude them from the scope of economics, relegating ethical standards t o the field of theology or political philosophy. This must be avoided a t all costs. Instead of attempting as mainstream economic theory does - t o show that economic rationality is the only predominant human behaviour that can be generalized in a manner which is independent of history, culture and institutional development, one may quite legitimately suppose that ethics is a feature which is inherent in and exclusive t o man. Such an approach would treat ethics independently of economics and would place its study within the field of philosophy. This would be the consequence of excluding all value judgements from economic theory, and would lead to the creation of a specialized branch of study for ethics. This would be in keeping with the way in which the transmission of knowledge is currently organized, namely via an increasingly impenetrable web of specialization. In order t o ensure ethics a lasting place in economic reasoning, then, the hypothesis of economic rationality should not be rejected (since the strategies consumers and producers use to maximize their utility and profits respectively reflect essential motives in human behaviour), but it should be developed further. Economic agents act emotionally as well as rationally. Emotional intensity thus adds a moral dimension t o economic and social actions and is a t variance with scientific interpretations based on the principle of causality,

5

which reduce emotionality to the level of what has been termed "hallucinatory reality". To quote Claude Olievenstein (1988, p. 191): "What is unspoken, particularly with regard to morality or ethics, cannot be written off as hallucinatory - human memory is too powerful for that. We thus have the astonishing situation in which part of reality is kept, as it were, under cover. Things which exist but cannot be fitted into a logical interpretation are either censored or, at best, contemptuously relegated to the nether world of the social sciences...". Attempting to rethink contemporary economic theory along these lines doubtless involves a n unprecedented intellectual challenge; yet such a n approach may well be the only way to help economics out of the positivist impasse it has got into. The methodological obstacles are enormous. Economics has only managed to bypass the normative issues by adopting a n abstract conceptual approach to society, based on deductive reasoning. For a complete view of social reality, economics must open itself up t o other disciplines - yet such a n interdisciplinary approach is inevitably more inductive. Despite the fact th at all problems are essentially interdisciplinary, knowledge has hitherto progressed through specialization. Our education system can scarcely be said to have encouraged diversity, and it is worth recalling Plato's story (Blumenberg, 1987) in which the philosopher Thales fell down a well while gazing a t the heavens. A Thracian servant girl made fun of him for having his head in the clouds and failing to see things th a t were right under his nose (and feet). Although nowadays Thales is famous and the servant girl is not, both of them had a point. An ethical approach to economic problems necessitates a n inductive review of contemporary economic theory in the light of knowledge acquired in other disciplines. Such is the challenge posed by contemporary problems, and in particular by environmental pollution, th at there is really no other choice.

PROTECTION OF THE ENVIRONMENT AND THE FAILURE OF COLLECTIVE DECISION-MAKING MECHANISMS Since the market is merely one of a number of collective decision-making mechanisms, environmental policy must be seen on various levels, including not only the economic, but also the political and administrative spheres. Accordingly, i t is prone to the failures observed in each of these spheres, and must come t o terms with the various collective decision-making mechanisms, which may supplement, oppose or reinforce one another. Thus the implementation of a n environment policy is a f a r more normative process than economic theories on the internalization of external costs are prepared to admit.

6

The following table sets out the various defects inherent in each of the collective decision-making mechanisms. The economic sphere is represented by the market and the political sphere by democracy, while the State is represented purely by its administrative functions. Yet the functions of the state are more complex than this, and impinge on the environment in numerous ways. We need only recall that the State is not only itself a producer and consumer of goods and services, but also intervenes in environmental economics as the supervisor of all manner of constraints. It thus guides individual choices, selects alternatives and regulates - a t times repressively - the social costs associated with environmental damage. It is therefore an oversimplification to reduce the State to its purely administrative dimension; however, this happens to be the discretionary space used for all collective decisions which the other mechanisms have failed to reach, and the table clearly shows that, when faced with the problem of implementing environmental policies, society must reconcile the failures that occur in the various collective decision-making mechanisms. These failures have some common features and therefore lend themselves to a tabular presentation, which will be discussed from an ethical point of view.

The main failures observed in the various collective decision-making mechanisms

Market

Democracy

Civil service and legal authorities

Redistribution of income and wealth

Intensity of preferences is not expressed

Increase o f public spending

1 Imperfect competition

Results may be logically inconsistent (Condorcet's paradox)

4

Public goods

Increase in discretionary space available to civil service

5 No incentive to keep oneself informed

7

3

2

6

Civil service seen as a public good 9

This table indicates, column by column, the main failures observed in the various collective decision-making mechanisms (Frey, 19811, but expressly omits collective decisions made by special interest groups such as umbrella organizations and occupational associations or trade unions. Since the latter operate simultaneously in all three of the spheres mentioned, it is not felt necessary to assign them a separate collective decision-making process. If the table is read horizontally, it is possible - admittedly, with some effort - to discern links between the various collective decision-making mechanisms. The first objection to the market economy (1)was so considerable that it encouraged the growth of an alternative economic system based on central planning. This first market failure thus not only profoundly shaped the history of social emancipation in the countries which espoused the market economy, but was also to result in unprecedented disillusionment in the countries which adopted the central-planning model. Accordingly, no workable environmental policy can be conceived of outside this frame of reference. However, the implementation of any such policy in our society remains subject - even though such matters lie within the "unspoken" realm (Olievenstein, 1988) - to the issue of redistribution of income and wealth, a problem which market mechanisms fail to deal with. Variations in market prices and in relative prices between market sectors (between "non-polluting" and environmentally harmful goods) lead to instances of market exclusion which are scarcely acceptable from the point of view of equity. Mainstream economists (Arrow, 1977) readily admit this, but insist on maintaining a clear distinction between positive and normative reasoning. They claim that economics is there to analyse how the market works, and that it is up to politics to settle matters of social justice. However, if we turn t o the collective decision-making mechanism known as democracy, we are forced t o admit that this is no answer (2). The democratic principle of "one man, one vote" does not allow intensity of preferences to be expressed. There is little point in calling for internalization of external effects if, in budgetary terms, one is not prepared to accept the resulting price rises. This leads to the pursuit of alliances and agreements which enable such objectives to be achieved: in other words, environmental protection for oneself, but at other people's expense. The consequences of this paradox are not only that a minority succeeds in imposing its will by democratic means, but that problems of environmental pollution are often settled by appealing to a third collective decision-making mechanism, namely the civil service (3). Economists who seek to avoid value judgements by appealing to the democratic mechanism thus help to strengthen the role of the civil service, whose measures (such as technical regulations) supposedly run counter to market forces (Baumol and Oates, 1979). In order to soften the blow, public (State) expenditure tends to increase (4). Accordingly, the State's share in changes in GNP is following the same trend (OECD, 1991b), suggesting - perhaps fallaciously - a n increase in the non-commercial sector of the economy. The top line of the table makes it clear that a positivist approach t o the economy cannot be separated from a normative approach expressed in political

8

terms. Such separation entails costs which, in turn, have economic consequences. The increase in public spending can thus be seen as a logical consequence of the separation between the economic and political spheres. Let us assume for the time being that economic models of how the economy runs represent an ideal (like any model). In that case, Todorov's comment is surely relevant (1991, p. 109): "What is ideal is only effective if it remains in touch with what is real; however, this does not mean that it should be dragged down to an "accessible" level, but rather that it should not become separated from the realm ofinquiry. Instead of neutral scientists and technicians on the one hand and moralists who disregard human reality on the other, we would then have researchers who are aware of the ethical dimension of their research, and men of action who keep abreast of achievements in the field of inquiry." This conclusion - that only acknowledgement of the ethical dimension can reconcile the positive and normative aspects of policies in the field of environmental economics - is further reinforced by the second line of the above table. A market economy modified by environmental incentives can only function if a number of hypotheses are satisfied; one such hypothesis is that there is a large number of market agents, without which competition cannot be guaranteed as an effective allocating mechanism (4). In practice, the frequent occurrence of imperfect competition is evidence of a second market failure. This lack of competition means that environmental incentives remain ineffective. Not only does the competitive oligopoly model produce a 'stand-off on prices which thus - within certain limits - remain unresponsive to incentives of this kind, but duopolies of the Nash-Cournot type also lead to situations which are less than optimal. Each situation which justifies technical regulations to protect the environment then becomes a situation in which strategic games ( 5 ) are acted out in the democratic sphere, thus helping t d increase the discretionary space available to the civil service departments (6) whose duty it is t o implement the regulations and monitor their effectiveness. Finally, the third line of the table reminds us that the economic sphere not only displays price flexibility, but also includes a significant non-commercial sector. According to economic theGry, this sector comprises public goods, whose prices cannot be determined by market forces (7). The environment is part of this, although it has an intrinsic value and thus enjoys a special status. If we choose to consider it external to the market, we return to what has become a classic pattern: the idea that the market and the environment on the one hand, and the economy and the State on the other, are diametrically opposed. In that case, the environment only acquires an economic value once it is internalized into a market. This process necessitates monetarizing the environment so that it can be analysed by a method which has been known t o us for nearly 40 years, namely the internalization of external effects. Thus the environment undergoes the same mental portrayal as the whole of society: the predominance of economic considerations over all others, even ethical ones.

9

The notion of the environment as a public good has one further consequence: a dissociation between collective costs and individual benefits. This is used to justify State intervention. Since it is in no-one’s interest to bear the external costs, the State must do so in order to remedy this third market failure. This lack of individual incentive is again encountered in the democratic collective decision-making mechanism. Since the aim of voting is to reflect the main underswells in society, a single vote in isolation is meaningless. Accordingly, there is no obligation on individual voters to keep themselves informed (8). If voters do not believe their votes will be decisive, they will be likely to vote without bothering t o find out exactly what they are voting on. As a result, collective preferences regarding environmental pollution will not be correctly expressed. This has two major consequences. Firstly, political propaganda attempts t o take advantage of the lack ofinformation. Transparency - a vital condition for any collective decision-making mechanism - is limited. Add to this a considerable lack of certainty in o u r scientific knowledge regarding complex environmental processes, and the result is inevitably that voters will be seriously ill-informed. In the absence of incentives to verify the information disseminated concerning the environment, the results of democratic voting procedures may well be biased. The second consequence concerns the very nature of social organization. If economic rationality is bounded by lack of information, markets cannot function optimally. Finding their capability limited, agents then seek to evade the mechanism of competition and t o protect themselves by means of alliances among the various special interest groups. Individual interests are no longer expressed through the market, but through the highly organized network which society has become. The administrative collective decision-making mechanism takes advantage of this lack of transparency to re-justify the increase in its own discretionary power. Yet, in turn, it is subject to the same influences. The civil service, being a public service (91, possesses the very characteristics which previously led to market failure. As a result, it too is influenced by the interplay of alliances and clashes between the various pressure groups, and thus becomes a n integral part of the organizational network of society. In the field of environmental policy, the civil service therefore tends, by virtue of the regulations which it issues, to relinquish its independent, symbolic function as the embodiment of the public interest and, instead, to reflect the failures that have affected the other collective decision-making mechanisms. Once we see the problem of policy in these terms, therefore, consensus will be more costly and governments tend to gratify established pressure groups. However, established alliances in matters of economic policy must increasingly acknowledge the presence of new pressure groups which as yet have no institutional basis. The established alliances have acted as a channel for public opinion (which governments do not always perceive correctly) and have thereby helped the latter in determining their policy. A t the same time, governments have relied on such alliances to spread information, especially in

10

order to popularize policies. The emergence of new and often short-lived pressure groups has therefore proved most disconcerting. As long as such groups are without an institutional basis, governments lack a policy guide, and are therefore tempted to hedge their decisions with extra precautions. As a result, expert reports proliferate and refuge is sought in environmental impact statements. The end result of this refusal to discuss values is an increasingly pettifogging, bureaucratic approach to environmental problems.

CONCLUSION Social sciences which study the decision-making process show the importance of values in the implementation of effective environmental policies. The underlying reference becames crucial. The value system on which public policy is based is bound to change. In order t o legitimate any public action in order to protect the environment, the ethical references will change too. This raises the issue of what society's perception of the state actually is. The need to protect the environment emphasizes the profound distinction between general and individual interest and between the public and private spheres. Marking out the boundaries between these spheres remains the essential task of politics and as to be exposed to the criterion of transparency. But communication and information systems which are supposed to keep transparent the boudaries between private and public interests are not a public but an economic good: Information is produced according t o economic principles. It is therefore not astonishing that environmental protection has to be submitted to the same logic too. In this perspective, ethics is only seen as the expression of moral philosophy. As a matter of fact, ethics has to become integrated part of ecological economics.

REFERENCES Arrow, K.J. (1977), "The Organization of Economic Activity, Issues pertinent to the Choice of Market versus non-market Allocation",in Public Expenditure and Policy Analysis, R.H. Haveman and J. Margolis, eds., Rand MacNally College Publishing Company, Chicago. Baumol, W.J. and Oates, W.E. (1979),Economics, Environmental Policy and the Quality of Life, Prentice Hall, Inc., Englewood Cliffs, N.J. Blumenberg, H. (1987), Das Lachen der Thrakerin. Eine Urgeschichte der Theorie, Suhrkamp, Frankfurt am Main.

Biirgenmeier, B. (1992), Socio-Economics: an Interdisciplinarv Approach: Ethics, Institutions and Markets, Kluwer Academic Publishers, Boston, Dordrecht, London. Etzioni, A., Lawrence, P.R. (19911, Socio-Economics: Towards a New Synthesis, M.E. Sharpe, Inc., h o n k , New York, London. Frey, B. (1981), Theorie demokratischer Wirtschaftspolitik, Franz Vahlen Verlag, Munich. Novak, M. (1987), Une Qthique 6conomique, Les valeurs de 1'Cconomie de marchb, Les Qditions du Cerf, Institut La Bobtie, Paris (translated from the English "The Spirit of Democratic Capitalism", 1982). OECD (1991a), "Recent Developments in the Use of Economic Instruments for Environmental Protection in OECD Countries", OECD Environment Monomaph, No. 11. OECD (1991b), OECD Economic Studies, International Comparisons, Paris. Olievenstein, C. (1988), Le non-dit des Qmotions, Editions Odile Jacob, Paris. Thomas Jr., W.L. e t al. (1956), Man's Role in Chankng the Face of the Earth, University of Chicago Press, Chicago. Tietenberg, T.H. (1989), "Marketable Permits in the U.S.: A Decade of Experience", in Karl W. Roskamp, ed., Public Finance and the Performance of Enterprises, Wayne State University Press, Detroit. Todorov, T. (1991), Les morales de l'histoire, Bernard Grasset, Paris.


CHAPTER 2 Atmospheric pollution


13

Effects of air pollutants on man, animals, plants and buildings: mechanisms and doseresponse effects S . Hippeli and E.F.Elstner

Institut fur Botanik und Mikrobiologie, Biochemisches Labor, Technische Universiat Miinchen, Arcisstr. 21, 8000 Miinchen 2, Germany

ABSTRACT Air pollution has become a major public concern since the beginning of industrialization, including motor car exhaust since the past three decades. Besides direct effects on living organisms, effects on buildings as well as on climate have to be considered. For living organisms, SO NO2, ozone, certain hydrocarbons and particles may have toxic effects whileqor buildings acids and their anhydrides (SO , N02) exhibit destructive power. Epidemiological studies together with clinical triafs and experiments in exposition chambers, including biochemical model reactions, contribute to our knowledge about potential dangers and increase the understanding of corresponding mechanisms and dose-response effects. Comparism of the almost daily appearing threatening reports in the press with the digest of more than 800 relevant original scientific publications allows the statement that impacts of ozone and NOx on health and performance of plants and animals are widely overestimated and appear to be used as political instruments. In contrast, combinations of SO2 with soot particles may represent an underestimated toxic potential. New experimental results clearly demonstrate the generally detoxifying power of the 3-way catalytic converter of Otto motor engines.

INTRODUCTION Air pollution by industry, traffic and agriculture is a growing threat all over the world. In the last few years especially automobil exhaust has moved into the focus of public and political discussions. Besides SO2 from coal-burning power plants, NO and hydrocarbons from gasoline-driven engines and petrol industry, together with their secondary products such as ozone and different peroxides have been identified as responsible for adverse effects on man, plants and buildings. Airborne soot particles stemming both from industrial incinerators and diesel engines envisage growing attention as potential carcinogens and inducers of respiratory diseases. The combinations of SO2 and soot from coal burning have been epidmiologically shown to be responsible for thousands of deaths during several severe smog episodes both in London and in New York not too long ago. In this review a digest of epidemiological data, clinical studies, measurements in exposition chambers as well as ex vivo or in vitro experiments on the effects of trace gases and particulate matter on animals, plants and buildings is reported. More than 800 original reports, revies and official statements have been incorporated into this documentation [I] thus allowing certain conclusions as to the relevance of continuously appearing reports on the increasing danger of air pollution.

14

EFFECTS OF AIR POLLUTANTS

Dose responses and reaction mechanisms Most trace gases or particles in the air are not directly toxic per se in ambient concentrations. Toxic effects on living cells are in most cases obtained indirectly via the formation of peroxides and other reactive oxygen species (ROS) [2, 31. This holds for N02, ozone, particles and partially also for S02. Since peroxides and other ROS are strongly under metabolic control in the living aerobic cell, destructive reactions are only observed after regulatory processes or reconstructive potentials have become limiting or inactivated; these complex processes connecting oxidative damage to biomolecules with cellular antioxidant properties and repair processes have recently been reviewed [4-61. General biochemical reactivities of potentially toxic trace gases are summarized in table 1 .

Table 1 Biochemical reactions of potentially toxic trace gases (nogtr = not of general toxicological relevance) ~

~~

~~~

~

binds to heme groups thus reversibly blocking respiration; only a problem in high concentrations; ("garage effect", suicides); nogtr acts as free radical; binds to heme groups; is rapidly oxidized to NOg; is physiologically "known" as a regula tor of blood vessel tonus; nogtr acts as free radica1,forming nitro compounds and peroxides; precursor of strong acids; is physiologically "known" but of potential toxicity strong oxidant forming peroxides; splits double bonds and oxidizes amino acids in proteins; toxic; (LA-type smog indicator) precursor of strong acids;forme free radicals and rapidly activates peroxides producing extremely reactive alkoxyl- or hydroxyl- radicals; (London-type smog indicator) contain aromatic hydrocarbons, nitroaromats, quinones and chelated transition metals thus initiating toxic redox cycling; catalyse the autoxidation of SH-groups and aecorbate; oxidize SO2 and NO? yielding sulfuric and nitric acids; (London-type smog indicator).

15

Biochemical effects may not always directly reflect physiological impacts since complex metabolic interactions may mask primary molecular events and thus interfere with the expected straight forward symptom expression. In concentrations up to the highest daily means, all the above mentioned gases are metabolized by several general mechanisms or constitutive pathways. Detoxification pathways in plants and animals may follow different routes. Some physiological effects of important air pollutants are listed in table 2.

Table 2 Physiological and ecological effects of relevant air pollutants effects on respiratory parameters. acts as plant fert i 1 izer ; 03

influences membrane equilibria (transport blocker); effects on respiratory parameters; may reduce resistance against viral infections; may reduce plant growth;

so2

irritates eyes and airways; nectrotizes plant tissues; monocausal agent of forrest die-back ( "Waldsterben") ;

soot

mutagenic,carcinogenic and allergizing;

par-

ticles

+ acute toxicity at high concentrations; soot provoke leukocytes (alveolar macrophages) but parimpair their bactericidal properties; potential ticlss inducer of chronic respiratory dieseases; SO2

Clinical and epidemiological studies and field observations Except epidmiological studies and field observations, clinical studies or chamber- and model-reactions have mostly been performed with overdoses of pollutants or doses in the upper region of occasionally observable gas- or particle-concentrations. It is thus not surprising that toxic effects or metabolic limitations (growth, development, performance) have been reported. Data from clinical and epidemiological studies under relevant conditions clearly indicate that a) gaseous components such as NO2 or ozone have no acute or chronic effects besides provoking regulatory systems (respiratory parameters: decrease of air flow in the range of 4-7%) or growth retardations (up to 10%)in some sensitive plants; b) air borne particles, especially those from diesel engine driven cars (trucks), in cooperation with S 0 2 , may most probably be involved in induction of sensitivities towards certain allergens (pollen, house dust mites) and also in the asthmatic complex;

16

c) acid anhydrides such as SO2 and NO2 in cooperation with soot particles as

carriers and catalysts are assumed to be mainly responsible for stone weathering and thus the decay of buildings. Tables 3 and 4 show some dose-response-effectsrelations for NO2 and ozone, as the key trace gases for photodynamic smog situations. Table 3 Dose-response data for NO2 Concentrations up to 850 pg/m 3 (1/2 h means) may be measured in urban areas Effects on no effects in healthy persons in concentrations up to 1900 pg/m3; asthmatics, especially under man and an ima1s strain, may recognize respiratory restrictions under 950 pg/m3; increase of respiratory infections at concentrations over 5000 pg/m3 were reported for animals; no indication for lung cancer after chronic exposition Effects on no negative effects under 4000 pg/m3; cooperation plants with NH4+ as fertilizer; under acid conditions soil losses of Mg++ and Ca++-ions; changes in ecological equilibria due to altered competitive behaviours, especially in N-intolerant ecotypes such as boglands

Table 4 Dose-response data for ozone Concentrations up to 450 pg/m3 (1/2 h means) may be measured close to, but outside of urban or industrial centers; Effects on sensitive persons (not identical to asthmatics!) man and may undergo reversibel respiratory restrictions anima 1s (ca. 4-8% reduction at 200-400 p g / m 3 ) ; during strain respiratory functions change already at concentrations above 160 pg/m3; adaptations after repeated expositions have been observed;Headache, coughing and throat irritations were individually reported at concentrations above 200 pg/m3; in mice, increased susceptibilities for viral infections ( > 160 pg/m 3 ) were observed; no indication for carcinogenic potentials

L

Effects on losses in crop yields and growth reductions up to plants 10% in sensitive plants ( 7 h up to 70 pg/m 3 ) were reported.

17

Motor exhaust condensates Certain components of motor car exhaust undergo condensation or dissolve as soon as they get in contact with cold or water containing surfaces. In order to compare the condensates of different motor concepts (Diesel, Otto-engine without catalytic converter (Bok) or Otto engine equipped with a regulated 3-way catalysator (Kat)) we [7] installed a condensation trap onto the exhaust tube of the vehicle on the test platform (figure 1) and followed the official FTP75 test procedure (federal test program).

Figure 1. Arrangement for trapping aqueous motor exhaust condensates during the "federal test program" (FTP75)

The water soluble condensates of this motor operation cycle, representing a mean city tour, were tested for their activities to modify biological molecules such as a. low molecular weight antioxidants (ascorbate, cysteine and glutathione) and proteins, such as BSA; b. crocin, a water soluble model substance representing polyene structures (typical for biomenbranes or poly unsaturated fatty acids) and

c. enzymes (glycerinaldehyde-3-phosphatedehydrogenase or xanthine oxidase).

Ascorbate oxidation. A strong stimulation of ascorbate oxidation, measured as oxygen uptake, is observed in the presence of BoK, while in the presence of Diesel only weak oxidation and, in the presence of KAT, no oxidation is observed . Fig. 2 shows the oxidation of ascorbate, treated with 100 p1 of the different condensates.

18

Data calculated as oxygen uptake/100 pl are compared to the same data calculated on the basis of the "FTP-cycle". (During the "FTP-cycle" the three motortypes tested emit different volumes of condensates due to their different fuel consumption (diesel) or the catalyzed water consumption (KAT). The corresponding volumes of the aqueous condensates were as follows: BoK: 647 ml/l FTP-cycle KAT: 1001 ml/l FTP-cycle Diesel: 439 ml/ 1 FTP-cycle )

0

Ia

5c

150

.*

e

2:

d

100

X u U

a

0

BoK

&KAT

0

Figure 2. Oxygen consumption during ascorbate oxidation, calculated on the basis of volume (100 pI) and on the basis of the FTP cycle.

Oxidation of sulfhydryl compounds. SH-groups play an important role in metabolic redox chemistry and generally also exhibit antioxidative properties in living cells. As corresponding model compounds we analyzed the effects of the different condensates on the SH-status of cysteine, glutathione and serum albumine. Cysteine oxidation is strongly enhanced by both BoK and Diesel whereas KAT has no influence. Glutathione oxidation is only slightly enhanced by Diesel and more strongly by BoK while SH-groups in serum albumine are exclusively oxidized by BoK. Fig. 3 shows the amount of oxidized SH-groups in cysteine calculated on the basis of the used volume (100 pl) and calculated on the basis of the "FTP-cycle". The oxidizing activity of the diesel-condensate is less pronounced under the condition of the "FTPcycle" due to the lower fuel consumption and thus condensate production.

19

-

100

I

Ir.

=:

9

=: x

X

v

v

m

P

z

y1

n

2

k

100 c. a

0

50

i%

mI

50

I v)

U N u ."

U

.-w

'0

9

0

0

0

0

Figure 3. Loss of free SH groups in cysteine.

Influence on enzymic activities. Glycerinaldehyde-3-phosphatedehydrogenase (GAP-DH) is a key enzyme in the glykolytic pathway and thus of predominating importance for intermediary metabolism. This SH-containing enzyme is sensitive towards oxidation and is rapidly inactivated by peroxide [8]. Increasing amounts of BoK or Diesel increasingly inhibit enzyme activities; KATcondensates are much less active. In comparison with GAP-DH, Xanthine oxidase (XOD) is an enzyme which is not oxygen sensitive; in contrast, it produces reactive oxygen species. Addition of BoK to the XOD test system results in a 68% inhibition of oxygen uptake after 30 min incubation; Diesel yields 20%inhibition under the same conditions while KAT is without any effect. Tab. 5 represents relative activities of the individual aqueous condensates as far as their influence on the enzymes GAP-DH and XOD is concerned. These data were calculated on the basis of comparable condensate volumes as well as on "FTP-cycle". The inhibitory activity of the BoK-condensate was set as 1 for each of the enzymes. The other activities were calculated in relation to this standard value. The KATcondensate showed the lowest, the BoK-condensate the strongest inhibitory capacity.

20

Table 5 Calculation of the relative desactivation factors of GAP-DH and xanthine oxidase in the presence of corresponding amounts of aqueous condensates from different motor concepts

QAP-DH

XOD

100 pl

FTP

100 p1

FTP

BoK

1.00

1.50

1.00

1.50

Kat

0.11

0.25

0.00

0.00

Diesel

1.31

1.31

0.29

0.29

The results clearly document the loss of destructive (oxidative) power of Otto-engine exhaust after passing the regulated 3-way catalytic converter.

Particles and cooperative effects. Soot particles have been shown to contain carcinogenic properties in certain animal models [9] epidemiological studies concerning indoor coal burning in China [lo].

0' 0

50 100 Diesel s o o t particles (vg)

Figure 4.Enhancement of chemiluminescence of butylhydro-peroxide in the presence of 0.5 m M sulfite and increasing amounts of diesel soot particles.

21

There is also growing evidence that respiratory diseases and allergic reactions may be induced and/or enhanced by particle inhalation. Model experiments have shown that diesel soot particles, enhanced by aqueous SO solutions (bisulfite), exhibit a considerable destructuive potential concerning vita? biomolecules such as SHcompounds, polyenes and certain enzymic properties [l 11. Specially the peroxide activating potential has to be addressed in this context since activated leucocytes, which excrete peroxide, exhibit a change in their physiological properties (decrease of superoxide formation and increase of phagocytosis) after soot particle contact (incubation) in the presence of bisulfite [ 121. As shown in figure 4, light emission from the decay of a peroxide (t-butyl hydroperoxide) in the presence of bisulfite is enhanced by diesel particles where the detection limit of particles is in the range of 5-10 pg/ml. The effects on biomolecules can impressively be documented by the oxidative destruction of the biological dye crocin (a water soluble carotenoid from Crocus spec.) which can be suppressed by the addition of superoxide dismutase indicating the importance of the superoxide radical anion in this bleaching reaction (fig. 5 ) .

Crocin Crocin/Sulf i t Crocin/SulFit/Diesel

soat/SOD

Crocin/Sulfit/Diesel

soot

x

0

0

5

10

15

20 25 30

Time (min)

Figure 5. Bleaching of the biological carotenoid "crocin" by 400 pg diesel soot particles and 0.5 mM sulfite: effect of superoxide dismutase (SOD)

CONCLUSIONS From our studies, partially documented in this report, we wish to draw the following conclusions:

22

1) Impacts of ozone and NOx during typical summer "smog" episodes on humans and plants are widely overestimated 2) Diesel soot particles in cooperation with SO2 immissions have to be considered as inducers of chronic respiratory diseases and/or allergic sensibilization 3) It has been documented with the aid of several independent indicator reactions that the regulated catalytic converter prevents the emission of condensable factors initiating oxidative destructions of biomolecules; the responsible oxidants may also be involved in the provocation of inflammatory processes in the respiratory tract and in the "corrosion" of plant cuticles finally yielding loss of structural resistances against certain fungal pathogens.

REFERENCES 1 Elstner EF, Hippeli S . BI Wissenschaftsverlag (in preparation). 2 Menzel DB. In: Pryor WA, ed. Free radicals in biology, vol 2. Academic, New York 1976, 181-202.

3 Mudd JB. In: Pryor WA, ed. Free radicals in biology, vol 2. Academic, New York 1976, 159-180 4 Hippeli S, Elstner EF. In: Sies H., ed. Oxidative Stress Academic, New York 1991, 3-55. 5 Elstner EF, ORwald W. Free Rad Res Comms 1991: 12/13: 795-807. 6 Elstner EF. Der Sauerstoff - Biochemie, Biologie, Medizin, BI-Wissenschaftsverlag. Mannheim, Wien Zurich 1990, 53Opp.

7 Blaurock B, Hippeli S , Metz M, Elstner EF. Arch Toxicol (in press). 8 Hyslop PA, Hinshaw DB, Halsey WA Jr,Schraufstiitter IU et al. J Biol Chem 1988: 263: 1665-1675 9 Heinrich U , Pott F, Rittinghausen S . In: Ishinishi N et al. eds. Carcinogenic and mutagenic effects of diesel engine exhaust. Elsevier, Amsterdam 1986, 441-455. 10 Mumford JL, He XZ, Chapman RS, Cao SR et al. Science 1987: 235: 217-220. 11 Hippeli S , Elstner EF. Z Naturforsch 1989: 44c: 514-523. 12 Hippeli S , Elstner EF. Free Rad Res Comms 1990: 11: 29-35.

23

Effects of air pollution on the condition of sessile oak forests in Hungary I. MCszBros, I. Mody, and M. Marschall Botanical Department, Lajos Kossuth University, Debrecen, Hungary H-4010

INTRODUCTION It is widely recognized that many European forests suffer from the new-type decline and show various signs of degradation [ 1,2,3,4]. There is considerable evidence that the phenomenon of forest decline in Europe cannot be explained by a single cause but it is considered a complex multifactoral process involving the natural and anthropogenic stresses. Industrial air pollution has been implicated as an important cause which may alter directly and indirect1 the healthy state of forests [I, 51. Beside inducing serious direct damages in the tree oliage they also result in the deterioriation of many soil processes and thereby influence the conditions of plant mineral nutrition. Air pollution can cause soil acidification and in this way affects the availability of nutrients and microbiological activity [4, 61. This may lead to a reduction in the rates of biochemical processes and a considerable decrease in the overall biological activity of the soil environment. As a consequence of the direct and indirects impacts of air pollution trees can be weakened and predisposed to injuries from naturally occurring stresses as summer drought [ 1, 4, 7, 8, 91. During the past two decades an accelerated deterioration of natural forests has also been observed in Hungary which has affected mainly the natural stands of sessile oak (Quercus petraea) on the mountainous and hilly regions [ l , 6, 10, 111. The decline of sessile oak appeared first in the North Hungarian Central Range, but from the middle of the 1980s this new dieback process has been observed in the western Transdanubian part of the country, too [ 121. The most serious decline of sessile oak stands has appeared in the industrial districts where the percentage of the decline of Quercus petraea may exceed 45% [ 11. Repeated surveys indicated that significant soil acidification took place in the Hungarian forest stands during the past 25 years [ 1, 131. This paper presents the results of studies aiming to reveal the alteration of the condition of sessile oak stands as consequence of imission loads. Three different aspects of oak forest degradation were investigated. Firstly emphasis was laid on the soil compartment and its main chemical properties and microbial activity were examined. Secondly the responses of underground vegetation layers were studied with respect to the species composition and the leaf NR activity of some dominant shrub species. The third aspect of this work was the investigation of physiological state of healthy and declining trees of Quercus petraea in forest stands with different imission loads [4, 8, 9, 141.

r

OUTLINE OF STUDY AND METHODS

Five years ago we started extensive comparative studies o f sessile oak forests in the nothern industrial region of the country [l, 111.

24

S i t e 1 (Kazincbniclka)

A

S i t e 2 (Fonagysag ) Site 3 ( S i k f A k u t )

Three forest sites were selected for the present experiments situated in the eastern part of the Hungarian Central Range (Fig. 1). Site I (Kazincbarcika) is in close vicinity of a chemical plant which emits a variety of pollutants including large amounts of nitrogen dioxide, chlorine and ammonia. The main products of this chemical plant involve different N and P fertilizers [15]. Site 2 (Fhnagysrlg) and Site 3 (Sikfiikut) are located a long way from the industry under less pollution influences. Soil and plant samplings were performed during the summer months. Soil samples were taken from the upper 10 cm layer under the litter. Soil unolyses involved the measurements of pH (in suspension with 1 M KCI), the available inorganic nitrogen forms (NH4N, NO N) and phosphorus and the ex&angable K, Mg and Ca using standard methods of soil chemistry [ 161.

Figure 1. Location of sampling areas There were also measurements on the levels of organic micro-pollutants (chlorinated pesticides and polychlorinated biphenyls) in the soil with gas chromatography after extraction by hexane. T o describe the soil biological activity at sites with different pollution degrees the microbial carbon dioxide ( C 0 2 ) production and phosphatase activity [4, 17, 18, 191 were measured. Microbial CO release was measured in fresh soil samples (at 50% WCmax) by determining the C d 2 release during 12 day-incubation at 27 ' C . During the experiment C 0 2 was trapped in 2n NaOH solution and estimated after measurement of the unneutralised alkali. Soil phosphatase activity, with some modifications, was determined at the original pH value of the soil according to KrBmer and Erdei [22]. The soil samples for phosphatase assay were air-dried, disodium phenylphosphate was used as substrate, and the liberated phenol was measured cholorimetrically. Since the underground vegetation may indicate well the levels of environmental loads [9] the herb and moss layers were also surveyed in ten 4x4 m plots in forests Site 1 and Site 2. All species appearing in each plot were recorded. While the surveys of mosses covering the bark of trees and the soil surface focused only o n the presence/absence data in case of herbs the individuals were also counted. Leczfnifrute reductuse activity was used as a possible indicator of the external nitrogen input into the forest sites [21, 22, 231. The activity of the enzyme was measured in leaves of some dominant shrub species with in vivo methods as described by Pizelle and Thiery

25

[24]. Leaf samples of approx. 0.2 g were cut into pieces and then were vxiium infiltrated for 30 min in buffer containing 0.02M K N O solved in O.1M NaK-phosphate, pH 7.5. Afterwards the samples were incubated for in the dark at 30 "C. Triton X-100 was added to the incubation medium to prevcnt i t from becoming turbid i n the presence of the acidic sulphanilamide solution used for measuring NO?. The nitrite formed by NRA was assa ed cholorirnetrically after adding 1 % (w/v) sulphanilaniide in 3 M HCI and NED (N-I-naphtyl-ethylenediamine-~liliydrochlori~le)solution to the 0.01%

fh

6,")

sample solution. For the estimation of the physiological state of sessile oak at the selected sites apparently healthy and declining trees were chosen o n the basis of the visible symtoms in their foliage. As in numerous studies the measurement of kuf pliotosyntheiic pignlerzt compositoiz was used as an assay for air pollutant effects.[25, 26, 271. Photosynthetic pigment - especially chorophyll - formation is known to be very sensitive t o almost any factor which disturbs metabolic processes, for example light intensity, temperature, mineral nutrient status, water stress and waterlogging as well as many pathovens, and last but not least air pollutants [26, 28, 29, 301. For the investigation of the pigment composition leaves were collected from the upper (shade) and lower (sun) parts of healthy and declining sessile oak trees. The extraction and determination of pigments were carried o u l following Czuhajowska and Przybylski [3I]. Concentrations of chlorophylls (a and 13) a n d carotenoids were determined in the same extract made with 100 o/o acetone. The absorbance of extracts were measured at wavelengths of 410.5, 044 and 662 nm. Care was taken i n measuring the turbidity at 750nm [32]. Concentrations of pigments were calculated by Holm's formulae [33]also used by Watts and Eley [33].

RESULTS AND DISCUSSION Chemical properties of soils The soil of each sampling site helongs to the brown forest soil type. The results of the soil chemical and biological analysis are summarized in Table 1, Figure 3, 3, 4, and 5. Regarding all the analyzed soil parameters there are suhstantial differences between the investigated areas.

Table 1 Mean values of soil chemical analysis (0-20 cm)

PH (KCV NH4-N (nig/lOOg) N03-N (mg/lOOg) P (mg/1OOg) K ( m d 1OOg) Mg (mgllOOg) Ca (md1OOg)

Site I

Site 2

Site 3

2.95 3.72

3.70 1.52 0.00 0.33 12.23 9.33 114.23

5.18 1.20 0.10 0.59 7.94 5.30 293.8 1

3.96

1.85 8.19

5.20 59.20

26

On the basis of pH values it can be stated that the degree of the soil acidity is the highest in the forest stand close to the industrial point-sourcewhich may be attributed to the acid load caused by the emitted NO and SO .The soil acidity tends to decrease with the distance from the industry (Table 1f In correlation with the low soil pH at site 1near the industry the base saturation of the soil is reduced and it is especially poor in calcium. Inorgmc nitrogen forms (Table 1) and phosphorus exist in elevated uantities in the soil of the polluted area (Site 1) that can be the result of the obviously 'gher external de osition of these elements. %he level of the total tested organic chemicals, chlorinated pesticides and PCBs (Fig. 2) indicates also external load into the soil of the forest near the industrial lant where the contamination by their total quantity amounts to 178%, compared wit the more distant site (Fig. 3)

R

K

200

/L

150-

100

50 -

total pesticides

0Slte 2

-

Pas

0

total

Oestlcides

PCBs

Slte 1

(undisturbed) (disturbed)

Figure 2. Concentrations of the main pesticides and PCBs in the soil

Fipre 3. Degree of the load of organic mcropollutants into the polluted area as expressed in the percentage of values at unpolluted site

As it has already been reported [34] among the chlorinated pesticides hexachlorcyclohexanshave been shown to accumulate most significantly in the soil of the polluted site.

Microbial activity of soils The biological activity of soil - the capacity of soil microorganisms to perform the biolo 'cal transformations of organic matter and thereby determining the soil fertility and t e processes of the nutrient cycles react very sensitively to the disturbances of the habitat [4, 171. As resented in Figure 4 and 5 the activity of soil microflora has shown differences accor&y to the sites and has reflected well the alterations in chemical properties and external oads near and further from the industry (Table 1).

I?

-

27

The microbial CO roduction, one of the frequently used measures of the soil biological activity [4, 35f h ave been shown to be of very low intensity in the industrial area, near Kazincbarcika where a considerable soil acidity was also established (Fig. 4). This might be caused by the effects of acidification since the naturally occurinp microor anisms were driven out due to the pH conditions changed [4,6,13]. Moreover it takes a lpong time while acidophilic communities can be developed and replace the role of existing species [4]. mg g-1 day-1

O

Z

i

5

g-1 h-1

1

1.5

0.1

"

1

1

Site 1 (disturbed)

"

Site 3 (undisturbed)

Figure 4. C 0 2 production of soil (0-20 cm)

Site 1 (disturbed)

Site 3 (undisturbed)

Figure 5. Soil phosphatase activity (0-20 cm)

Many reactions of soil matter transformations may be catalyzed by enzymes existing outside the microorganisms and plant root s stems, therefore soil enzymes may also indicate the biological activity of soils [ 17, 181. is generally accepted that hosphorus i s taken up by plant roots as inor anic hosphate. The large proportion of however, in many soils is organically boun , so t e rmneralization rate of its organic fractions is of great importance. Soil phosphatase lays a major role in this mineralization process [17]. The phosphatase activity of soil &ig. 5 ) has shown good positive correlation with the chan es of pH and microbial CO production of the soil at the sites investigated. It was sixfoh lower in samples from h e industrial area [19]. This together with the low insensi of microbial C02 production reflect slow decomposition processes of organic materi s in the soil of the polluted area [4,61. Lowerin of the rate of decomposition may lead to considerable accumulation of litter at pollute sites [ 11. Our work suggests that the level of microbial CO roduction and the phosphatase enzyme activity can be used as good indicators for the3eration in the biological activity of soils due to the external anthropogenicloads.

C K

s

it

$

li

Species composition of herb and moss layers In the disturbed forest site near the industry the herb layer is poor in species, it consists of only 11 species as compared to that of the control site (Site 2) where the number of the appearing species is much higher (31 species). The herb layer surveys were evaluated on the basis of the nature conservation values introduced by Simon [36]. It was stated that the native herbaceous species (values 1-6) are repressed and the

28

B

roportion of disturbance tolerant and weed s ecies (values 7-10) is higher than in the Forest Site 1 than in Site 2, further from the in ustry (Fig. 6). Similar floristic alterations in the herb layer of disturbed forest sites have been presented by MolnAr e t al. [lS]. Table 2 Species composition of in forest stands near (Site 1) and furt er (Site 2) from the emission source

U"

60

life strategy

Sites 1 2

40

20

a

Slte 1

(disturbed)

Slte 2

(undtstur bed)

Lophocolea heterophylla P + + Hypnumcupressiforme P + + Brachytheciumrutabulum P + + B. velutinum P + + Plagiothecium denticulatum P + + Amblystegium serpens P + + Pylaisia polyantha P + + P + + Atrichum undulatum Isothecium mnyurum P Plagiomnium affine Ls + Bryum flaccidum C P perennial species LS:long-life shuttle species

+ +

Figure 6. Results of herb layer surveys

Ccolonist

There have been also important changes in the individual number of species nearer the in try. The herb layer of Site 2 is rich in individuals which have amounted in total to 109 m- .How er, in Site 1in close vicinity to the industry the individual number has been only 3 m- . The resence of mosses [37] also indicates some signs of degradation in the polluted site (Tab, 2). Although there are not large differences regarding the strategic spectrum of the occurring moss species [38,39] the decrease of species richness and disappearance of some species can be observed near the industry.

9

7

Leaf NR activity There were significant differences in the leaf nitrate activity (NRA) of dominant species (Fig. 7) between the sites. In the polluted area (Site 1) the species had much higher NRA (0106-3.81 pmol g-' d.w. than in the forest (Site 3) far from the industry (0.02-0.57 pmol g' d.w.). d e s e findings are in accordance with other results in the literature [22,23,24,40]. It has 4enerally observed that in forests having significant NO3- deposition and nitrification plants are found to have considerable nitrate reductase activity in root or leaf tissue or both, so NRA may therefore be useful as an indication of the soil status including availability of NO3 [22,23]. It is well-known that the absorption of atmospheric NO enhances the NRA of plants [24] and the rate of NR induction is controlled by the 3 0 3 - uptake of the plant [41]. These reactions are generally subjected to the regulation by the same process: induced

-

29

by nitrate flux into the plant [42]. Our measurements indicate that the imission load into the forest near the chemical industrial plant is due to nitrogenous compounds.

Acer campestre Acer tat ar i cum Euonymus europaeus

E u o n y m u s verrucosus

b

3.8 1

Sambucus nigra

0

02

04

06

08

1

umol NO, g-1 (d w )

RE8 Site

Site I

ElSite

2

3

Figure 7. Nitrate reductase activity of leaves Physiological state of sessile oak When the study aimed at revealing the physiological damages of sessile oak in the samplin sites it was found that the sli ht visible discolouration of leaves is connected with a gastic (40-60%) degradation o the photosynthetic pigments (Fig. 8, 9, 10, 11) 1431.

B

ma a-1

/L

H D Sire1

H D Site2

=sun

H D Site3 @shade

Fi re 8. Total igment concentration in eaves of hea thy (H) and damaged (D) trees

H"

P

H D Site 1

I

H D Site 2 =sun

H D Site 3 =shade

Figure 9. Concentration of chlorophyll a in leaves of healthy (H) and damaged (D) trees

30

Figure 10. Concentration of chlorophyll b in leaves of healthy (H) and damaged (D) trees

Figure 11. Concentration of carotenoids in leaves of healthy (H) and darnaged (D) trees

This was also shown in a number of surveys around point sources, the degree of visible injury was related to the loss of chlorophyll attributed to gaseous pollutants [26]. The reduction of photosynthetic pigments has appeared typically in larger extent in the upper part of the damaged trees. In case of sun leaves the deterioration of total pigments (Fig. 8) was very similar in all sites and amounts to 50-55%. Considering the shade leaves, however, there were significant differences in the degradation degree of pigments among the sites. In Site 1 (Kazincbarcika), close to the point pollution source the total pigment content (Fig. 8) and the concentrations of individual components (Fig. 9, 10, 1I ) in the shade leaves of damaged trees has changed slightly compared with the respective crown layer of healthy trees and 80-90% of the pigment content still remained. At the same time in areas located in larger distances from the industry (Site 2:FbnagysBg and Site 3:Sikfo”klit) the degree of total pigment deterioration in the shade leaves of damaged trees is more similar to the sun leaves and it has reached 55%. These results suggest that in close vicinity to the pollution source the direct effects of gaseous pollutants on the foliage of the sessile oak may be more pronounced than the indirect influences [lo, 111. In spite of large differences in the total leaf pigment content between crown layers and sites the pigment ratios changed in relatively smaller degree (Table 3). The content of total carotenoids (Fig. 11) decreased less in the damaged trees than the chloro ylls (Fig. 9, 10) and depending on the crown layer its level was 6 5 8 5 % of that of the heaEhy trees. It is generally accepted that carotenes are the second most abundant pigments in leaf tissues after chlorophylls [26] therefore Arndt suggested [44] the use of beta-carotene concentration as a reliable early indication of injury from several fumigation including SO2, H F and HCI.

31

Table 3 Ratio of the main pigment components

Site 1

healthy trees damaged trees

Site 2


Site 3


sun leaves shade leaves sun leaves shade leaves sun leaves shade leaves sun leaves shade leaves sun leaves shade leaves sun leaves shade leaves

chla/b

chl/car

3.39 2.9 1 3.46 2.98 3.83 3.11 4.3 1 3.39 2.5 1 2.88 2.79 2.85

3.7 1 4.4s 2.8 1 4.34 3.79 4.32 3.22 4.0 1 4.60 4.74 3.40 3.82

Our studies suggest that beside the concentrations of individual pigment constituents their ratios are also very informative concerning the state of the photosynthetic pigment systems and may he widely used for the indication of the changes induced by the disturbing effects involving the air pollutants. As shown in Table 3 in o u r studies there was a decrease in the ratio of chlorophylls to carotenoids with the decline which was much pronounced in the sun leaves.

ACKNOWLEDGEMENTS

This study was performed partly in the framework of the scientific cooperation between the L. Kossuth University, Debrecen, Hungary and the Institute of Chemical Ecology, G S F Neuherberg, Germany. The authors are very grateful to G. L6rinci and I.Gebefugi for their help in performing the analysis of the PCBs and pesticides.

REFERENCES

1 2 3 4

Jakucs P. (1988): Ecological Approach to forest decay in Hungary. AMBIO 16:267-274. Kubikovi I. (1991): Forest dieback i n Chechoslovakia. Vegetatio 93: 101-108. Schulze E.D., Lange O.L., Oren R. (eds.)(lY80): Forest Decline and Air Pollution. Ecological Studies 77:475. Springer Verlag, Berlin. Smith W.M. (19x1): Air pollution and forests. Interactions between air contaminants and forest ecosystems. Springer Verlag, New York.

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5 6 7

8 9 10 11

12 13 14

15

16 17 18 19

20 21 22 23 24

Villanueva V.R. and Santerre A. (1989): On the mechanism of adaptive metbolism of healthy-resistant trees from forest polluted areas. Water, Air, and Soil Pollution 4859-75. Jakucs P., Meszriros I., Papp B.L. and T6th J.A. (1986): Acidification of soil and decay of sessile oak in the Sikf6klit Project area (N-Hungary). Act. Bot. Hung. 32~303-322. Amundson R.G., Alschner R.G., Fellows S., Rubin G., Fincher I., Van Leuken P., Weinstein L.H. (1991): Seasonal changes in the pigments, carbohydrates and growth of red spruce as affected by exposure to ozone for two growing seasons. New Phytol. 1183127-137. Langebartels C., Ernst D., Heller W., Sandermann H. (1991): Reactions of trees to air pollutants. Proceedings of the Ecological Approaches of Environmental Chemicals. 15-17. April, 1991. Debrecen, Hungary. pp.274-285. Lewitt J. (1980): Responses of plants to Environmental Stresses. Academic Press, New York. Meszriros I. (1990): Comparative study of healthy and diseased sessile oak trees. Proceedings of the International Symposium on Forest Decline in Bavaria and Eastern Europe, Passau Mesziros I. (1991): Effects of soil acidification on the physiology of sessile oak trees. Proceedings of the International Symposium on Ecological Approaches of Environmental Chemicals. 15- 17. April, 1991. Debrecen, Hungary. pp.306-313. Pagony H. (1992): VitalitBtsabschwBchung der Traubeneichenbestiinde. Allgemeine Forst Zeitschrift 2:98-99. Stefanovits P. (1986): Some new data on the soil acidification. Magyar TudomBny 31 :339-341. Lange O.L., Hebber U., Schulze E.D., Ziegler H. (1989): Atmospheric Pollutants and Plant Metabolism. In: Shulze E.D., Lange O.L., Oren R. (eds.) Ecological studies 77:238-273. Springer Verlag, Berlin. Molnar E., Szabb M., Szab6 J. (1991): Case study o n the pollution status of soils, vegetation and groundwater in the Sajo valley, Hungary. Proceedings of the International East-West Symposium o n Contaminated Areas in Eastern Europe, 25-27. November, 1991. Gosen, Germany. Papers Session 2125-27. Allen S.E. (ed.) (1974): Chemical analysis of ecological materials. Blackwell Scientific Publications, Oxford. pp. 1 19-234. Burns R.G. (ed.)(1978): Soil Enzymes. Academic Press, New York. Fauvel B., Rouquerol T. (1970): The phosphatase test considered as an index of soil activity and evolution. Review Ecol. Biol. Soil 7:393-406. Meszaros I., M6dy I. (1991): Applicability of microbial C 0 2 production and phosphatase enzyme activity as indicators of pollution loads into brown forest soils. In: Teller A., Mathy P., Jeffers J.N.R. (eds.) Responses of forest ecosystems to environmental changes. Elsevier Applied Science, London and New York. pp.835-836. Kranier M. and Erdei G. (19%): Investigation of the phosphatase activity of soils with the use of disodium phenylphosphate. Agrokem. Talajt. 7:361-366. A d a m M.A., Attiwill P.M. ( 1982): Nitrogen mineralization and nitrate reduction in forests. Soil Biol. Biocheni. 14: 197-202. Hiigberg P., Granstriim A., Johansson T., Lunmark-Thelin A., NBsholm T. (1986): Plant nitrate reductase activity as an indicator of availability of nitrate in forest soils. Can. J. For. Res. Vo1.16, 6:1165-1169. Thoene B., Schroder P., Papen H., Egger A. and Rennenberg H. (1991): Absorption of atmospheric NO by spruce (Picea abies L. Karst.) trees I. NO2 influx and its correlation with nizate reduction. New Phytol. 117575585. Pizelle G., Thiery G. (1986): Reduction of nitrate in the perennial tissues of aerial parts of Alnus glutinosa. Physiol. Plant. 68:347-352.

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25 20

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37 38 30

40 41 42 43 44

Baillon F., Dalschaert X., Grassi S., Geiss F. (1988): Spruce photosynthesis: possibility of early damage diagnosis due to exposure to magnesium o r potassium deficiency. Trees 2: 173-179. Darral N.M., Jiger H.J. (1084): Biochemical diagnostic tests for the effect of air pollution on plants. In: Koziol M.J., Whatley F.R. (eds.) Gaseous Air Pollutants and Plant Metabolism. Buttenvorths. London. pp.333-349. Howell R.K. and Kremer D.F. (1973). The chemistry of physiology of pigmentation in leaves injured by air pollution. J. Env. Qual. 2:434-438. Heher U., Yin Z.H., Dittrich A,, Pfanz H., Lange O.L. (1990): The response of leaves t o potentially acidic gases. In: Ulrich B. (ed.) Proceedings of the International Congress on Forest Decline Research: State of Knowledge and Perspectives. Friedrichshafen, Germany. pp.499-5 16. Rabe E. and Kreeb K.H. (1980). Bioindication of air pollution by chlorophyll destruction in plant leaves. Oikos 34: 163-367. Tschaplinski J., Timothy J., Norby, Richard J. (199 1): Physiological indicators of nitrogen response in ;I short rotation sycamore plantation 1. CO assimilation, photosynthetic pigments and soluble carbohydrates. Physiol. Plant. $2: 117-120. Czuchajowska Z., Przybylski T.( 1978): The seasonal changes of chlorophylls and carotenoids in unpolluted and polluted needles of Pinus silvestris. Bull. Acad. Polon. Sci. Skr. Sci. Biol. XXVI:30L)-37h. Watts D.F., Eley J.H. (1081): Changes in the chlorophyll a:h ratio during autumn coloration of Populus sargentii. Bull. Torrey Bot. Club 108:379-382. Holm G. (1954): Chlorophyll mutations in barley. Acta Agric. Scand. 4:457-47 1. Meszaros I., Liirinci G., Gebefiigi I. (lY91): Accumulation of pesticides and chlorinated hydrocarbons in brown forest soils. Proceedings of the International Symposium on Ecological Approaches of Environmental Chemicals. 15- 17. April, l901. Debrecen. Hungary. pp. 314-318. Lett1 A. (1990): Influence of industrial SO2 emissions on microorganisms of forest soils. Ekol6gia (CFR) VoI.0, 3:315-330. Simon T. (1988): Nature conservation ranks of the Hungarian vascular flora. Abstr. Bot. 12:1-23. Vajda L., Orbin S. ( 1983): Magyarorszrig mohatl6rrijrinak kezikijnyve. Akademiai Kjadh, Budapest. During H.I. ( 1979): Life strategies of bryophytes: a preliminary review. Lindbergia 5:2-18. Orhfin S. (1984): A mohik strategirii 6s TWR ertekei. Separatum Act. Acad. Paed. Agrensis-nova series tom XVII. pp.755-765. Campbell W.H. (1988): Nitrate retluctase and its role in nitrate assimilation in plants. Physiol. Plant. 74:214-219. Sinha S.K., Nicholas D.J.D. (1981): Nitrate reductase. In: Paleg L.G., Aspinall D. (eds.): The physiology and biochemistry of drought resistance in plants. Acadeniic Press, New York. pp.145-171. Stewart G.R. (1Y9l): The comparative ecophysiology of plant nitrogen metabolism. In: Porter J.R., Lawlor D.W. (ecls.): Plant growth: interactions with nutrition and environment. Cambridge University Press, Cambridge. pp.55-8 1. Ricks G.R. and Williams R.J.H. (1975): Effects of atmospheric air pollution on deciduous woodland. Part 3. Effects on photosynthetic pigments of leaves o f Quercus petraea (Mattuschka Leibl.). Env. Poll. 8:97-106. Arndt U. ( 1971): Konzentratioiisanclerungen hei Blattfarbstoffen unter dem Einfluss von Luftveruntreinigungen, Ein Disskussionsbeitrag zur Pignientanalyse. Environmental Pollution 237-48.


35

Long term effects of Acid deposition: Implications on the performance of high level nuclear waste repositories J. Nebot and J. Bruno MBT Tecnologia Ambiental, Parc Tecnologic del Valles, 08290 Cerdanyola, Spain.

1. INTRODUCTION The Earth's global environment can be considered with regard to the mass transfer as an (almost) closed system. The atmosphere's outer layer, the stratosphere, acts as an interface between the system and the surroundings, allowing at the same time energy transfer. The different compartments (atmosphere, geosphere and litosphere) forming our environment, the biosphere, can globally be considered to be in a steady state with respect to the redox reactions or to the proton and electron balances. Obviously, this global stationary state is locally and regionally disturbed. In fact, naturally occurring processes can cause the above-mentioned decoupling between proton producing and proton consuming reactions. One of this natural processes responsible for local proton imbalances is the aggradation of vegetation. In terrestrial ecosystems, accumulation of biomass occurs in inmature forests until the establishment of the definitive population of the stand. In addition, other processes can cause the proton balance to be upset, i.e. the human activities. At the present moment, one of the major causes of these local and/or regional imbalances is the anthropogenic input of protons to the environment originating from the combustion of fossil fuels for energy. The natural cycling of some principal elements, such as carbon, nitrogen and sulfur has been altered by the combustion of fossil fuels. This activity has caused the rate of the oxidation reactions to increase, while the reduction rates remain unaltered. Accordingly the concentration of oxidized compounds of the above mentioned elements has resulted globally increased in the case of CO,, and locally or regionally increased in the case of sulfur and nitrogen oxides. This increase was first noticed in the troposphere due to the smaller volume and the relatively short mixing time of this reservoir. Once in the atmosphere, sulfur and nitrogen oxides are transformed into the corresponding "strong acids", H,SO, and HNO,, which are quickly removed from the atmosphere and deposited in the litosphere. Several processes in the soil, such as denitrification, sulphate reduction and chemical weathering can neutralize the human-enhanced deposition of acids. In the long-term, the quantitatively most important one is the chemical weathering of minerals. Carbonate minerals are widespread around the earth's crust. The carbonate rocks are easily weatherable minerals, exhibiting a fast kinetics of dissolution. Hence, they provide an enormous potential for neutralizing the acid load into the soil. However, in some areas of North America and Fennoscandia, the dominant type of minerals forming the bedrock are crystalline rocks, such as granite, gneiss and quartz,

36

which are only slowly weathered. In these regions, nor the shallow and base-poor soils deriving from the crystalline rocks neither the weathering of the bedrock itself, can keep pace with the acid deposition. This results in a transfer of this acidity to the downstream compartment in the acid flux: the hydrosphere. Within the hydrosphere, the subcompartment first showing the acidification effects is the surface water, namely lakes and streams. Groundwater can also be affected over long exposition to acid deposition. The element responsible for the acid flux to the surface waters is mainly aluminum, which has been exchanged into the soil solution by protons. The first signs of local acidification were early noticed in areas where ore smelting was a current practice. Linne, in 1734, described that around a 500 years-old smelter at Falun (Sweden) no herbs could grow. However, the first major unification of knowledge about acid precipitation was achieved by Oden, a soil scientist of Uppsala (Sweden). In 1961, Oden's studies showed that acid precipitation is a large scale phenomenon and hypothesized that probable ecological consequences of the acid precipitation would be changes in surface water chemistry, decline of fish populations, leaching of toxic metals from soils, decreased forest growth and accelerated damage to materials. Unfortunately, at the present moment, Oden's hypothesis have been widely confirmed and this study itself is aimed at assessing the long-term consequences of the acid deposition on mineral weathering rates and soil erosion rates. The objective of this report is to make a prospective study of the effects of the ongoing acidification of soils, surface waters and groundwater on the host rock of high level waste repositories in granitic formations. Two main impacts on the performance of such repositories can be expected: I/ on the geological stability of the granitic formations through increased rates of chemical weathering and erosion and, 21 on the chemical composition of groundwater. This report will focus on the impact of acid deposition on the rates of weathering and erosion of granitic bedrock. Three different scenarios, based on the atmospheric emissions of CO, and SO,, are explored up to next Ice Age, which is "scheduled" to happen in some 58.000 years (1). A simple conceptual model is developed in order to asses the consequences of the environmental acidification for deep bedrock disposal.

2. SCENARIOS The main cause of acid deposition is the release of SO, and NO, into the atmosphere. The origin of the emissions of SO, is from natural sources, including both, geological activity (i.e. volcanoes and some hot springs) and biogenic activity. Anthropogenic emissions are mainly originated from the combustion of sulfur-containing fossil fuels. This is the quantitatively most important source of SO,. More than 90% of SO, emissions in Eastern Europe and industrial areas of North America are the result of the burning of fossil fuels (2). The anthropogenic total emissions of SO, account approximately for 150 millions tons per year (3). SO, emissions have nearly tripled in Europe since 1900, with the largest increase since the World War II (2).

37

NO, released into the atmosphere is also originated from the combustion of fossil fuels and from the biomass burning. Some 40 millions tons of NO, are yearly emitted into the atmosphere (3). Fossil fuels (coal, oil and gas) supply at the present 88% of global energy requirements and nuclear energy provides most of the rest (4). As fossil fuel combustion is the main source of SO, and NO,, a central scenario related to this process is proposed. It is assumed the ratio of fossil fuel energetic supply to the nuclear one will be constant until depletion of fossil fuels, although this is very uncertain under present conditions. Sulfur dioxide concentrations over Europe in last years were reviewed and projected 100 years ahead by Graedel et al. (3) (Figure 2-1). These projections assume that population and energy consumption will grow and that the firing of coal for energy (a major source of SO,) will increase. Taking into account that the world coal reserves total about 950 billion metric cubic tons (4), at the today's production rates reserves will last for almost 300 years. Hence, in our calculations we have assumed by the year 2300 the atmospheric acidity will peak. Furthermore, three different sub-scenarios are studied depending on how stringent emission controls might be during next centuries: mild control (scenario A), moderate control (scenario B) or severe control (scenario C). These scenarios are based on the same criterium (SO, emissions) as in the RAINS model developed by the IIASA (5). It is considered that the build-up of atmospheric SO, up to depletion of fossil fuels will continue at the rate estimated by Graedel et al. (3) for next 100 years. A linear increase of varying slope is thought to be the most realistic approach to model the evolution of the atmospheric SO, concentrations (Figure 2-l), although from the mathematical point of view the best fitting for the measured and 100 years projected SO, concentrations is an exponential curve. As a consequence, Eq 2-1, 2-2 and 2-3 are used to describe the build-up of atmospheric SO, (ppb) up to the depletion of fossil fuels. SOzAyear = 14 + 0.149*(year-1970) SOZByear = 14 + 0.445*(year-1970) SO,Cyear = 14 + 1.324'(year-1970)

l i i U

2,""

2"10

ZiOa

2150

2 x 0

2250

2x0

Is.,

Figure 2-1. Measured and projected atmospheric concentrations of S 1, (3). The linear fitting is taken in order to model the atmospheric SO, concentrations up to the depletion of fossil fuels (2300).

38

Results of the calculations based on Eq 2-1, 2-2 and 2-3 for the expected sulfur dioxide concentrations in the atmosphere for the explored scenarios by the exhaustion of fossil fuels are shown in Table 2.1.

CONTROL EMISSIONS

RATE OF INCREASE (PPb*Y-l)

scenario A scenario B scenario C

severe moderate mild

0.149 0.445 1.324

[~OzI,,, (PPb) 63 161 45 1

Table 2.1. Rate of increase and expected concentrations of atmospheric SO, over Europe at the depletion of fossil fuels (2300) depending on the explored scenario. It is well known that CO, concentration is also increasing in the atmosphere due to the same processes as SO, and NO, (consumption of fossil fuels) as well as deforestation. This accounts for perhaps half of the build-up since the year 1800 and 20% of current emissions (6). The current average concentration is around 350 ppm, while the background concentration is estimated to be some 290 ppm (6-8). The proposed scenario assumes a moderate increase of CO, during next centuries. When making long-term predictions (more than 300 years) a moderate rate of increase of 0.8 ppm per year (7) is more realistic than the ones assumed in shortterm predictions i.e. 1.3 ppm per year as projected by Wayne (8) or CO, concentration by 2000 to be 21% higher than the 1977 concentration (9). Hence, we have studied in the various scenarios the increase in CO, concentration and consequently, the acidity and the buffering capacity of the C0,-HC03- system. The evolution of atmospheric concentration of CO, (pprn) from present to year 2300 is described by Eq 2-4. CO, yeat = 334 + 0.8*(year-1970) (Eq 2-4) It has to be noticed that the evolution of the NO, concentrations in the atmosphere is not considered in the scenarios. Two main reasons justify this assumption. The first one is the fact that reactions of oxidation of SO, in the atmosphere yield a sulfur residence time of several days; this corresponds to a transport distance of hundreds to a thousand kilometers. The formation of HNO, is more rapid, this results in a shorter travel distance, thus it can be assumed that nitrogen deposition accounts only for local acidification processes which are not considered in our regional model. The second reason is that nitrogen is a growth-limiting nutrient in many terrestrial ecosystems in Europe and North America. Thus, a moderate additional input of nitrogen will usually lead to an increased primary production. A growing forest needs 0.5 to 0.8

39

g.m-2.y-1of nitrogen. This is partly confirmed by the fact that the analytical data from 12 Swedish stations performed by Rodhe et al. (10) showed no further increase of NO,- in precipitation between 1972 and 1984. The scenario developed considers that after depletion of fossil fuels (2300) the remainder of SO, is washed out from the atmosphere until its preindustrial level. The half-life of SO, in the atmosphere is assumed to be 10 days (0.027 years) and its natural background concentration 5 ppb. Assuming a first order kinetics, the removal of SO, from the atmosphere is described by Eq 2-5. InSO, year = InSO, 2300 - (25.3*(year-2300)) (Eq 2-5) The CO, concentration in the atmosphere after depletion of fossil fuels will decrease slowly due to the fact that its average residence time in the atmosphere is estimated to be some 100 years. This is much longer than the average residence time of sulfur dioxide (a few days). The decrease of atmospheric CO, is also considered to follow a first order kinetics, assuming 290 ppm as a final background pre-industrial concentration (this concentration corresponds to 1860 based upon data of Schneider, 1989) and a half-life time of 100 years, the decay of atmospheric CO, after depletion of fossil fuels is described by Eq 2-6 (Table 2.2 shows the half-life time and rate constants used to model the kinetics of removal of atmospheric SO, and CO, after depletion of fossil fuels). InCO, year = InCO, 2300 - (0.007*(year-2300)) (Eq 2-61 After recovery of SO, and CO, background concentration, the scenarios developed are kept to cover a time span of some 58.000 years in which SO, and CO, levels remain constant until next Ice Age.

so* scenario A scenario B scenario C CO,

HALF-LIFE TIME

RATE CONSTANT

(year)

(year-1

0.027

25.3

100

0.007

TIME REQUIRED RESTORE CONDITIONS (year)

0.100 0.137 0.178 106

Table 2.2. Half-life time, rate constants and computed time required to restore background concentrations of atmospheric SO2 and C02 from 2300 on.

40

3. DEFORESTING EFFECTS OF ACID PRECIPITATION Although an increase in forest damage has been recorded in Europe and North America since the early seventies (11,12), no consensus on its causes and mechanisms has yet emerged. Among the hypothesized causal agents are virtually all the pollutants species present in industrialized environments. Our interest focuses on tree damage due to acid deposition, including both wet and dry deposition effects. Basically, the suggested mechanisms of damage can be separated in two main groups: - Direct effect of the gaseous pollutants on the foliage. - Indirect effect on the root system through acidic deposition on soils. The first mechanism is mainly exhibited by gaseous sulfur dioxide. In addition to adsorption on leaf surfaces, SO, enters the leaves through the open stomata and can cause abnormalities in cell structure and altered metabolic activity. In wet surfaces, SO, and dry deposited sulphate particles dissotve in water and the H,SO, produced enhances the erosion of the leaf surface, which protect the leaf from the water losses and prevents leaching of nutrients from the foliage . Furthermore, a mechanism of co-deposition of SO, and NH, in coniferous trees has been suggested (13). This mechanism may cause an enhanced deposition of nitrogen compounds over large areas of Europe. Acid deposition may cause soil changes detrimental to forest vegetation, either by stripping nutrients from the soils or by mobilizing phytotoxic elements. It is still matter of discussion the way that acidic deposition on soils can cause the reported mortality of fine roots of declining trees (12). An increase of aluminum in soil solution was noted as was a calcium deficiency in roots in other declining stands. It has also been suggested that the ratio AI:Ca is responsible for the forest dieback (14 and references therein). It has been described that not only the SO, concentration, but also the period of exposure determines the degree of forest damage. For that reason the threshold concept can not be applied to tree damage caused by SO,. The IUFRO Air Pollution Section, at this 1982 meeting in Oulu, Finland, decided that no safe limit could be set for SO, ambient concentration to protect forest trees (1 5). In view of the variety of pathways to damage the trees as well as that there is no safe threshold, it seems quite reasonable to conclude that the effect of SO, on the forest trees is related to a physiological stress which makes the trees much more sensitive to other factors i.e. diseases and climate (16). The forest decline induced by acid deposition might eventually lead to a large areas of deforested land around Central and Northern Europe. The main consequences of such deforestation processes concerning the geochemical stability of granitic formations are: A/ Medium to poor forest soils are formed into podzols (the predominant soil type in Sweden). Podzolic soils exhibit a pronounced stratification. The litter accumulates on the top of the soil and it is slowly decomposed into humus by bacteria and mainly by fungi. The raw humus layer often exhibits a pH below 4. Some of the processes occurring in forested podzols concerning the acidic flux are:

41

- The regeneration of the cation store in humus layer is supplied by the decomposition of the litter. Tree roots act as a base cation pump through the A, and B horizons of the podzols (17). The base cations originated by weathering in the B horizon are translocated in forested soils to the humus layer, by root uptake, litterfall and litter decomposition. - Complex bacterial populations are supported in forested ecosystems, carrying out among others processes, the reduction of sulphate and denitrification. Both processes are a sink for hydrogen ions. - The acid load into the soil due to nitrogen compounds is considered to be of lesser importance in front of sulphate deposition due to the uptake of such compounds during biomass growth (see 2. Scenarios). As deforestation proceeds, this sink of nitrogen compounds will disappear, thus, an increase in the acidity of the soil solution can be expected. Hence, deforestation could result in a decrease of base saturation status in the top soil and in a decreased capacity of sinking protons and nitrogen compounds. This is likely to cause an increasing rate of acidification on such deforested soils. It has been well established from laboratory studies that mineral weathering rates are dependent on hydrogen ion activity (18 and references therein). A field study comparing two forested basins in Central Europe representing strongly acidified due to high deposition of SO, and less acidified environments of industrial and rural countryside, indicated increased weathering rates of bedrock and depletion of exchangeable cations from soils due to such an acidification (19). Therefore, from lab and field studies, it seems reasonable to conclude that an increase in proton activity in soils due i.e. to deforestation can account for a faster weathering rate. BI It is well known that afforestation is the best way to prevent soil denudation. Denudation is a complex process involving detachment and downslope transport under the influence of both gravity (mass movement) and water in motion (slope erosion) (20). Weathering and soil formation are dependent upon several factors i.e. temperature, particle size, surface effects, biotic effects (21 and references therein) and variation in these factors causes the processes to proceed at different rates. If the rate of soil formation is defined as the rate at which rock is converted into soil, then weathering and soil formation are closely related processes, particularly in granite. On relatively pure limestone, however, a large depth of rock may weather giving only soluble species and leaving only a shallow soil. Soil formation implies a loss of mass due to several processes i.e. dissolution of mobile elements. Although the bulk density of the soil formed is lower than the parent material bulk density (soil formation from bedrock implies an increase in the porosity), there is an overall reduction in the total volume of material, i.e. the volume of soil formed is less than the volume of rock weathered to form it. Hence, the process itself of bedrock weathering and soil formation implies a net surface lowering or landscape reduction, even without taking into account additional processes like denudation. As an example, the rates of weathering and soil formation on granite were studied in two areas of Rhodesia using small watersheds (22,23). The rate of granite weathering was calculated using Barth's equation (24) which links the rate of weathering to the amount of an element removed in solution per unit of time, its concentration on the rock and its concentration on the weathered product. Results of these field studies indicated

42

rates of granite weathering of 15.4 and 11.O mm per 1000 years for the higher and lower rainfall areas, respectively. The soil production rates were 11.0 and 4.1 mm per 1000 years. Thus, the net surface lowering rates were 4.4 and 1.7 mm per 1000 years, even in areas where soil production is high. An estimation of the rate of surface lowering or landscape reduction under deforestation is a key point in this study. The maximum expected surface lowering rate corresponds to a situation where the soil is continuously removed while forming as it could happen in deforested areas. In this case, the rate of surface lowering is the same as the rate of bedrock Weathering. In southern Fennoscandia where thin soils are often found, deforestation could be responsible in short periods of time for total losses of existing soil due to denudation. In addition, no further recovery of soil mantle can be expected in deforested areas. A deep mature soil over the bedrock protects it from weathering due to its buffer capacity. As the soil mantle thickens, the weathering zone is farther removed from the surface being decreased the available surface and consequently the weathering rate. An example of the role played by the soil preventing the chemical weathering of the bedrock is the field study conducted in upland forested areas in New England (Hubbard Brook Forest). As mentioned by Johnson et al. (25) only a moderate rate of weathering has been measured in this area in spite the acidification of their waters. The cause of this slow weathering rate is not clear. However, the maturity of the soil was suggested as a contributing agent. On the other hand, in shallow or non-existing soils, fresh rock surfaces are constantly being created. These surfaces are completely exposed to physical, chemical and biological weathering agents. Among the physical agents are precipitation and changes of temperature (freezing and melting processes) causing rock breakdown. Biological processes involve the direct colonization of bedrock by microflora and microfauna exerting a mechanical effect on rock as well as a chemical one releasing i.e. organic acids. The final result of all these effects is again an increased rate of weathering of the bedrock. The effects of deforestation discussed in A and B cannot be quantified separately. However, as it has been shown by Schnoor (26) weathering rates measured in the field are always 1-2 orders of magnitude lower than the ones measured in the laboratory at the same pH conditions (Figure 3-1). This has been rationalized by taking into account that the hydrological condition in the field is not the same as in the lab. Under ideal laboratory conditions, weathering proceeds much faster than in the field because of the larger availability of wetted surface. -1 1

c -12 m

’?

E

2

13

-14

c a

8

-15 -16

; -7

-6

5

4

3

2

1

lcg Fbwraie I Mass Soil (L day.’ g-’ )

Figure 3-1. Dissolved silica release rate (weathering rate) vs flow-rate:mass ratio (26).

43

A similar phenomena could be thought to be the final stage of deforestation, where maximum surface and porosity are available due to the disruption of the soil and “fresh” surfaces are continuously formed as a result of the increased erosion. Hence, in our model the worst case scenario is considered to be the one causing total deforestation. This leads to a situation where weathering rates are large as the ones measured in laboratory studies.

4. DESCRIPTION OF THE MODEL A conceptual model was designed to study the environmental acidification processes. The total sulfur concentration (as total sulphate) is considered the main driving variable of such processes. A compartment structure (box model) was found to be a good approach to simulate the physically different reservoirs involved in the process of acidification. Such compartments or reservoirs are assumed to be connected by the transfer of sulfur compounds. Four reservoirs were considered to be relevant in the natural flow of sulfur, namely atmosphere (troposphere), surface waters, soil-bedrock and groundwater. A schematic representation of the compartment system is shown in Figure 4-1. A more detailed description of the designed compartments and assumptions is given in next subchapters. The model developed in this report refers to the atmosphere, soil and bedrock compartments The modelling of surface and groundwater compartments is not considered in detail in this work. A regional model of the hydrological conditions of the region is under development. The consequences of the acidification processes in these compartments will be studied when a complete hydrological model for surface and groundwater mixing will be available.

water

Lqran1te

I

I

weathering] I

water

Figure 4-1. Schematic representation of the acid flux through the compartments involved in the developed model.

44

4.1. Atmosphere As mentioned above, the main driving variable for the environmental acidification processes is the total deposition of sulfur. In the developed model the atmospheric compartment is the origin of the acid flux, being sulphate the assumed species responsible for the acidic deposition. Sulfur is emitted as a gas, sulfur dioxide, but it is transformed into fine sulphate particles once in the atmosphere (8,27). A process which, among others, gives sulphate particle is direct SO, photochemical oxidation, schematically shown as: 2S0, + 0, + hv ---> 2S03 where SO, is further hydrolyzed into H,SO,:

SO, + H,O + M ---> H2S04+ M The model assumes all the SO, is finally washed out from the troposphere in the form of H,SO,. No difference is considered in the deposition mechanism (dry or wet) on how the acidity reaches the soil reservoir. The average SO, atmospheric concentration is used as the main variable in order to describe the scenarios. The sulphate concentration is considered to drive the acidification of soils. Hence, a linkage between SO, atmospheric concentrations and sulphate deposition has to be defined. A direct correlation is assumed between SO, atmospheric values and sulphate concentration in precipitation on Southern Sweden. In order to obtain an average value for sulphate concentration in rainwater in this area, we relied on data taken from the European Air Chemistry Network. The area considered comprises from 5 9 N to 60QNlatitude and from 1OQEto 20GE longitude (basically, Central and Southern Sweden). Around this region 44 stations were monitored for rainwater chemical data between 1955 and 1979 by the Swedish University of Agricultural Sciences of Uppsala and the International Meteorological Institute of Stockholm (28). We used selected data sets (see Appendix A) to calculate the average sulphate concentration in rainwater in 1970. Median values of each station have been used to estimate average sulphate concentration in precipitation in the concerned area (28,29). The computed value was assigned to 1970 as the origin of our calculations. The available sulphate values (28) referred to excess sulphate (i.e. the non-seasalt fraction). Since in ocean islands and coastal areas sulphate originating from seaspray contributes very significantly to the sulphate concentration in precipitation samples, the seasalt fraction of sulphate was deducted in this study based on the sodium concentration according to the following formula: [S0,2-]ex= [S042-]- 0.25 “a+] In order to calculate the total sulphate concentration (seasalt + non-seasalt), the sodium average concentration in precipitation has to be known. This value was calculated from the data .of the sodium content in precipitation taken from Soderlund et al. (30) by using median values of each station. The average sodium content in precipitation was calculated to be 47 ~ m o l e . d m - ~ . The computed average excess sulphate content of the rainfall was also 47 ~ m o l e . d m -Hence, ~. the calculated total concentration of sulphate was 58 p m ~ l e . d m - ~ .

45

This concentration was taken as a base level corresponding to 1970 for next calculations. All calculations were made by taking as average for the region considered an annual precipitation of 1000 mm.m-* (31 and references therein). The linear relationship between time and concentration of sulphate (mole.dm-3) in precipitation is described for each scenario by Eq 4-1, 4-2 and 4-3. S0,2-Ayear = 5.8'1 0 ~ +5 0.497*1O-6*(year-l970) (Eq 4-1 1 S0d2-Byear= 5.8*10-5 + 1.403*1O-6*(year-l970) (Eq 4-21 S0d2-Cyear= 5.8'1 0-5+ 5.067'1 O-6*(year-1970) (Eq 4-31 In order to validate the reliability of such equations to predict the total sulphate concentrations of the rainfall, a check against the measured concentrations during the eighties was made. Our results are in good agreement with the measured data of the Co-operative Program for the Monitoring and Evaluation of the Long Range Transmission of Air Pollutants in Europe (EMEP) (32). These data are presented as isolines of volume-weighted average of sulphate in precipitation on Europe (1 978-1982). Southern Scandinavia falls between the 1.5 and 4.5 mg S042--S.dm-3isolines. Assuming the scenario B (moderate control of emissions), the one actually operating in Scandinavia, we estimate a sulphate concentration of 2.3 mg S042--S.dm3. Base cations can neutralize in the atmosphere a substantial amount of acid deposition, this has to be considered in order to calculate the acid load entering the soil compartment. Alkalies are generated in the atmosphere as the carbonates of windblown dust, generally of natural origin, and from seaspray in coastal areas. It is known from precipitation measurements (33) that most of the base cation contain is in the form of calcium and magnesium. It is difficult to quantify base cation deposition. As noticed by Kamari (cited in 34) the ratio base deposition/sulfur deposition is fairly constant. The combined Ca2+ and Mg2+deposition was estimated to neutralize an average of 33% of the sulfuric acid in bulk precipitation, within a range of 12% to 44%. Hence, we assume in the model that one third of the sulfuric acid is readily neutralized in the atmosphere before reaching the soil. The computed total acidity (Htot) of rainwater as a function of SO,2- concentration - ~ each scenario is described by Eq 4-4. and dissolved CO, (H,CO,(aq)) in m ~ l e . d mfor HtOtA,B,Cyear = 1.33*S042-A,B,Cyear+ 2'H2C03(aq)year (Eq 4-4) where H,CO,(aq)year is calculated from CO, Henry's constant as shown in Eq 4-5. H2C03(aq)year= 1.132*10-5+ 2.71 1'10-8'(year-1970) (Eq 4-51 Due to the composition of rainwater, the total and free acidities are almost similar and H+ concentration approximates the Htot (36). The expected pH values of the rain solution in southern Sweden are shown for the three studied scenarios in the period 1970-2300 in Figure 4-2.

46

42

1

1970

20m

2033

2060

2090

21a

219

2180

2210

2240

2270

2300

V ,..

Figure 4-2.Evolution of the rainwater pH up to depletion of fossil fuels (1970-2300). 4.2 Soil-Bedrock The soil is considered in this model as an interface between the atmosphere and the bedrock, for this reason they are not treated as a separate reservoirs. This chapter is devoted to the acidification processes which occur in forested soils. Since agricultural soils are intensively managed with lime and other chemicals, acidification is not expected to proceed in these soils. Furthermore, one of the objectives of this study is to assess the degree of deforestation due to acid deposition. 4.2.1.Processes of acidification in forested soils

Soil acidification has been defined as a decrease in the soil acid neutralizing capacity (alkalinity) accomplished by removal of alkaline earth cationic components or, to a lesser extent, by addition of acidic components. This occurs by an irreversible flux of protons to the soil. The proton sources include: - atmospheric inputs of acidic or potentially acidic substances - net assimilation of cations by vegetation - net mineralization of anions from organic matter - deprotonation of weak acids - oxidation reactions - precipitation of cations as a consequence of secondary phase formation - mineral weathering of anionic components We will only consider the first point which represents the anthropogenic input of protons to the soil system (acid deposition), whilst the rest are taken as background processes which do not upset the global balances. Protons entering or produced in the soil are removed by: - export in drainage water - net mineralization of cations in organic matter - net assimilation of anions by vegetation - protonation of anions - reduction reactions - weathering of cationic components - precipitation of anions Once the acid input to the soil is completely neutralized, cations are removed by vegetation uptake and/or export into drainage water, resulting in soil acidification and in an increase in the alkalinity of percolating water.

47

Different reactions are responsible for the soil buffer capacity. The concept of buffer capacity corresponds to the above mentioned concept of alkalinity or acid neutralizing capacity. It is defined as the total reservoir of basic components in the soil. Ulrich (cited in 37) classified the inorganic buffering reactions in the following buffer ranges: - Carbonate buffer range: the buffering capacity in this range is supplied by CaCO, (Eq 4-6), hence only calcareous soils are within this range. The amount and the rate of dissolution of CaCO, in such soils is usually enough to buffer moderately large amounts of acidic deposition. The soil solution pH within this range is higher than 6.2. CaCO, + H+ = HCO; + Ca2+ (Eq 4-61 - Silicate buffer range: this corresponds to the buffering range of feldspar weathering reactions which can be generalized as described by Eq 4-7. KAISi,O, + 7 H,O + H+ ---> AI(OH),(s) + 3 H,SiO, + K+ (Eq 4-71 These are kinetically slow reactions and cannot be described as an equilibrium buffer capacity. Hence a silicate buffering rate capacity is defined. This corresponds to the rate of reaction 4-7. The pH of the soil solution in this range is between 5.0 and 6.2. - Cation exchange buffer range: base cations such as Ca2+, Mgz+, Na+ and K+ exist in the soil mainly complexed by the organic matter (humic and fulvic substances) and/or bound to clay particles. The buffer capacity given by exchange reactions with humic substances can be described by a general reaction (Eq 4-8) as follows:

HUM-Ca + 2 H+ = HUM-H, + Cazi (Eq 4-61 The buffer capacity given by surface reactions or: clay particles may be generalized in reactions of the type: >SiOK + H += >SOH + K+ (Eq 4-91 where the symbol > denotes a surface process. These two processes are kinetically fast and they constitute the more readily available buffering capacity of non-calcareous soils. The number of reactive sites available is approximated to the Cation Exchange Capacity (CEC). Base saturation (0) is the fraction of CEC consisting in base cations. Buffer capacity of the cation exchange range (OCE) is defined as the product of CEC and 0. If the total acid load overalls the cation exchange buffer capacity, this buffer becomes depleted and silicate weathering becomes the only neutralizing process available. The pH of the soil in this buffer range is between 5.0 and 4.2. - Aluminum hydroxide buffer range: as the alkalinity of the soil solution is depleted and as a consequence of feldspar weathering, AI(III) hydroxide becomes abundant in the soil solution. At the low AI(III) concentrations and pH range of this buffer (pH= 4.2 to 3.0),the main equilibrium is given by :

AI(OH),(S) + 3 H+ = ~ 1 3 ++ 3 H,O for poorly crystallized gibbsite this is a fast reaction.

(Eq 4-10)

48 If this buffer capacity is exhausted, other solid phases dissolution and precipitation reactions can take over. For instance, Fe(OH),(s) dissolution has been considered to buffer the soil solution at very low pH values. 4.2.2 Main assumptions of the model developed

In our effort to model the process of soil acidification in Southern Fennoscandia, we assume, based on data taken from the FAO-UNESCO Soil Map of the World (1974) (in 34), that the type of forest soil predominant in the area to be studied is orthic podzol. The orthic podzol is a non-calcareous soil deriving from granitic or base-poor material. Podzols in Scandinavia are often shallow soils, exhibiting a pronounced stratification. On the top of the soil it is found a humus layer originating from the slow decomposition of litter. The pH of this layer is often below 4 in medium to poor podzols. A bleached horizon (A2), usually grayish in color due to intense leaching, follows below the raw humus layer. Below this layer, there is an enrichment zone, the B-horizon, reddish to brownish. Iron leached from the bleached horizon precipitates here in trivalent form as goethite (17). In order to simplify the calculations, the model assumes that the soil layer is a homogeneous box 50 cm deep. In addition, we deal with yearly variations in soil acidification in order to avoid seasonal fluctuations. This seasonal variations are mainly caused by biological activity and tend to be internally compensated in the ecosystem over the year. Since podzols are non-calcareous soils, it is postulated that the carbonate buffer range has been surpassed. Therefore, the model assumes that at the origin of our calculations (1970) the soils considered here are at the stage at which cation exchange is the dominant buffer reaction. Cation exchange buffer capacity (OCE) includes in our model the ion exchange capacity of the organic substances as well as the cation exchange from silicate interlayer positions by a slow diffusion phenomenum reported in young soils of Scandinavia by Graunstein (cited in 17). It is postulated that the soil exhibits some capacity of recovery in front of acid deposition due to the organic matter, represented in our model by an initial term (OCEO). This recovering capacity is assumed to be linearly depleted with time at the same rate as acid load increases as computed in Eq 4-1 1, 4-12 and 4-1 3. (Eq 4-1 1) DCEOAyear = 2.0 - (7.182’1 O-4*(year-l970)) (Eq 4-1 2 ) BCEOByear = 2.0 - (1.924*10-3*(year-I970)) DCEOCyear = 2.0 - (6.809*1O-3*(year-l970)) (Eq 4-13) Silicate weathering rate (see section 4.2.4) is assumed to be only slightly dependent on the soil solution pH. If we assume that the weathering rate (brSi) is proportional to [H+Io5,a decrease of pH in the soil solution of 2 units results in a ten-told increase in the rate of weathering. In addition, the rate of release of base cations is assumed not to depend on the soil pH. The modifying effect of the forest canopy (forest filtering) on the deposition is also taken into account in the model. There are two ways in which the canopy is modifying precipitation input to soil. One way is dry deposition. From an ecological point of view, any vegetation canopy behaves like a filter or a sink for the fluxes of matter passing along its surface, the filter efficiency being very dependent upon its physical and

chemical properties (38). Hence, forest filtering can cause forested soils to receive more sulfur deposition than adjacent cropland or pastureland. Another way of modifying soil input is by leaching substances previously taken up by the roots and translocated to the upper parts of the tree. Thus, when they get to the soil surface the total precipitation input to soil is a mixture of substances coming from outside the ecosystem, i.e from the atmosphere, and other compounds which are merely completing an internal cycling (38). In addition, it has been shown by lvens (cited in 34) that forest filtering is quantitatively different in coniferous stands or in deciduous stands. The total surface of the needles of coniferous trees is bigger than the total surface of broad-leaved trees. Furthermore, in wintertime deciduous trees loose their leaves, while coniferous needles are kept through the year. Both factors cause the coniferous stands to intercept much more acid deposition than the deciduous stands. It is difficult to quantify the concentrating effect of coniferous canopies. Our model assumes that soil under coniferous trees receives an acid load due to sulfur compounds 1.6 times greater that soil under deciduous trees (35). In our model the net surface lowering is defined as the depth of bedrock weathered minus the depth of soil subsequently formed. A direct correlation between the assumed silicate weathering rate and bedrock surface lowering rate is assumed in order to calculate the net surface lowering. One of the critical assumptions of the model is that once acidification proceeds further to the AI(OH), buffer range, deforestation proceeds for the lifetime of the forest stand (50 years). Deforestation causes the total lost of soil mantle due to the increased effects of erosion and denudation. In these conditions the dissapearence of the soil mantle is irreversible and there is no longer soil formation. As a consequence, due to the disruption of the soil structure, the bedrock surface becomes more available and the weathering rates get closer to the ones measured in ideal laboratory conditions. 4.2.3 Assignment of values to soil parameters

A key point in the development of any regional model is the correct quantification of the parameters involved in the model such as, in our case, cation exchange capacity and base saturation. Ideally, field measurements of the relevant soil parameters should be preliminary to modelling and forecasting. In our study, since direct field measurements of such soil parameters were not accessible, literature values had to be relied upon. Following an extensive literature search it was decided to use the data elaborated by Kauppi et a1.(34) based on data reported in the Appendix of the FAO-UNESCO Soil Map of the World (1974). In our model, initial conditions (1970) are always set to the less favorable conditions. The CEC and 0 values are initially set to 20 moles.m-* and 0.1 respectively, giving a cation exchange buffer capacity of 2.0 mo1es.m-2. Silicate weathering rates were also taken from Kauppi et al. (34 and references therein). These weathering rates were calculated from the data reported in the Geological Map of Europe and the Mediterranean Region. The initially assigned silicate weathering rate is 0.05 mole.m-2.y-'. The weathering rates measured across Europe range from 0.02 to 0.2 mole.m~z.y-i(cited in 34). We assume that the above mentioned effect of soil disruption caused by extensive

50 acidification over the Al(OH), buffer range, increases 1.5 orders of magnitude the silicate weathering rates. This is based on the observations and discussions of section 3 regarding the differences between measured weathering rates in laboratory and field conditions. Under these conditions, the silicate weathering rate (brSiAl) is assumed to be 1.58 moles.m-2.y-1. Lerman (39) reported rates of bedrock surface lowering, elaborated from chemical weathering field measurements (40) in different geological environments, ranging from 0.006 to 0.14 mm.y-'. An intermediate bedrock lowering rate of 0.075 mm.y-' is used in our model. The assumed rate of soil formation from granite is 9.10-3mm.y-' (20). A conversion coefficient (f) of 1.5*10.3 is used in our model to relate silicate weathering rate (mo1e.m-2.y-I)and bedrock surface lowering rate (m.y-') based upon the assigned values to bedrock surface lowering and chemical weathering rates. 4.2.4 Model development

The main features of the developed model can be summarized as follows: - calculation of the annual acid load entering the soil - determination of the dominant buffering range - calculation of the annual soil pH - calculation of the net surface lowering at the corresponding year The model computations are separated into the following steps: 1. The acid load entering the soil per year (HA,B,Cyear) is calculated from total acidity of rainwater (Eq 4-4) assuming an annual precipitation of 1000 mm.m-2. 2. The annual cation exchange buffer capacity (RCEyear) is calculated by subtracting the silicate weathering rate (brSi) from the acid input into the soil. This result is then subtracted from the annual initial buffer capacity of the cation exchange range (DCEO). These calculations can be represented as: (Eq 4-14) OCEyear = DCEOyear - (Hyear - brSi) This is an annual iterative step while soil KE>O (the cation exchange buffer capacity of the soil is enough to neutralize the annual acid load). If OCE=O, it is considered that the soil drops into the AI(OH), buffer range. 3. The soil pH is calculated within the silicate, cation exchange and upper aluminum buffering ranges f D>O) according to a non-linear relationship between base saturation and pH (41) as follows:

(Eq 4-15) PHyear 4.0 + 1.6 * (OCEyear/CEC)o.75 4. The silicate weathering rate (brSiyear) is recalculated as a function of pH in periods where acid load entering the soil is high (1970-2300). The weathering rate constant (ksi) is calculated assuming a 0.5 order dependence on free proton activity. 5. The total amount of silicate weathered (WSi) during any period is calculated by integrating the weathering rate equations within that period. 6. The net surface lowering (NSLow) is calculated by establishing a direct correlation between the amount of silicate weathered (WSi) during the corresponding period, the bedrock surface lowering rate and the soil formation rate (SoF) as follows: NSLowyear = (WSiyear * f) - (SOF* nyears)

(Eq 4-16)

51

where f is the conversion factor from mole.m-2to m deep and nyears is the time span considered. The driving variable of the developed model is acid load (HA, HB, HC) which depends on the SO, and CO, atmospheric concentrations for each scenario considered. Based on the acid load entering the soil compartment three different periods of time can be distinguished (see section 2.): - a period of 330 years, from 1970 to 2300 (estimated exhaustion of fossil fuels) during which the soils are receiving an increasing acid load. - a short period of time (a few months for SO2 and some 100 years for CO2) for the recovery of the background concentration of such gases. The soil Compartment receives a decreasing acid load. - finally, a time span of some 58.000 years up to next Ice Age, where the acid load entering the soil is constant and corresponds to the acidity originated from natural sources. The model computations are performed separately for each of the above mentioned periods of time.

Increasing acid load: 1970-2300 The annual acid load (moles. m-2. year-') for each scenario is calculated according to Eq 4-17 as follows: (Eq 4-17) HA,B,Cyear = a'l 03*S0,2-A,B,Cyear + 2*1O3'H,CO3(aq)year where a= 1.33 in soils supporting deciduous stands and a= 2.13 (1.33'1.6) in soils supporting coniferous stands (see Eq 4-4 and section 4.2.2). The evolution of the annual acid load in forested podzols along this time-span for the three scenarios studied is depicted in Figures 4-3 and 4-4.

Im

an,

2 x

2%

mo

m

year

Figure 4-3. Acid load entering soils supporting deciduous forests from 1970 to depletion of fossil fuels (2300) for the three studied scenarios.

Figure 4-4. Acid load entering soils supporting coniferous forests from 1970 to depletion fossil fuels (2300) for the three studied scenarios.

52

The total acid load received by the soil during the period 1970-2300 depending on the scenario studied and the type of afforestation is shown below in Table 4-1.

Total Acid Load (moles.m-*) scenario B scenario C

scenario A Deciduous stands

72.0

137.8

403.8

Coniferous stands

109.0

214.2

639.8

Table 4-1. Total acid load received by podzols in Southern Sweden depending on the scenario studied and the type of afforestation during the period 1970-2300. The variation of the cation exchange buffer capacity with time is calculated according to Eq 4-18. OCEA,B,Cyear = OCEOA,B,Cyear - (HA,B,Cyear - brSi) (Eq 4-18) The evolution of cation exchange buffer capacity and base saturation of the soil during this period of time depending on the afforestation is shown in Figures 4-5, 4-6, 4-7 and 4-8. I

1w

am

m

891

mo

m

vear

Figure 4-5. Evolution of soil cation exchange buffer capacity indeciduous forests from 1970 to depletion of fossil fuels (2300) for the three studied scenarios.

Figure 4-6. Evolution of soil cation exchange buffer capacity in coniferous forests from 1970 to depletion of fossil fuels (2300) for the three studied scenarios.

53

Figure 4-7. Evolution of the base saturation of soils supporting deciduous stands from 1970 to depletion of fossil fuels (2300) for the three studied scenarios.

Figure 4-8. Evolution of the base saturation of soils supporting coniferous stands from 1970 to depletion of fossil fuels (2300) for the three studied scenarios.

The predicted soil pH values under different conditions of afforestation during the period 1970-2300 are shown in Figures 4-9 and 4-10.

Figure 4-9. Evolution of the pH of soils supporting deciduous forests from 1970 to depletion of fossil fuels (2300) for the three studied scenarios.

Figure 4-10. Evolution of the pH of soils supporting coniferous forests from 1970 to depietion of fossil fuels (2300) for the three studied scenarios.

As mentioned above, the silicate weathering rate was kept constant through the computations. Once the pH of the soil was calculated, a check of the reliability of this assumption was made. The silicate weathering rate constant (ksi) was calculated by assuming that the initial weathering rate (0.05 mole.m-2.y-1)was measured at pH= 4.33. By taking into

54

account a 0.5 order kinetics of the silicate weathering rate (1 8) ksi is calculated as:

- brSi * (10-4.33)-0.5 kSI. (Eq 4-19) Therefore, the dependence of silicate weathering rate upon time (or soil acidity) for each scenario can be written as a rate equation as follows: (Eq 4-20) brSiA,B,Cyear = ksi * (HA,B,Cyear)O.' The variation of the silicate weathering rate with time (or soil acidity) is depicted in Figures 4-1 1 and 4-12. The total amount of silicate weathered (molem2) can be easily calculated by integrating the weathering rate equation (Eq 4-16) within 1970 and 2300, except for the scenario C. In this worst case scenario, the soil cation exchange buffer capacity is exhausted by 2144 in the case of soils supporting deciduous stands, and by 2078 in soils under coniferous forests. The rate equation developed can only be applied within this buffer range and the amount of weathered silicate is computed by this method up to those limits. During next 50 years (see section 4.2.2.) the weathering rate is assumed to remain constant and finally from 2128 (coniferous stands) and 2194 (deciduous stands) the silicate weathering rate is assumed to be the one which applies for the AI(OH), buffer range. Table 4-2 shows the weathered silicate during this time span depending on afforestation and scenario.

Figure 4-11. Variation of the weathering rate with time (or soil acidity) from 1970 to depletion of fossil fuels (2300) on soils supporting deciduous forests depending on the studied scenario.

Figure 4-12. Variation of the weathering rate with time (or soil acidity) from 1970 to depletion of fossil fuels (2300) on soils supporting coniferous forests depending on the studied scenario.

The net surface lowering is calculated for the scenarios A,

B as shown in Eq 4-21.

NSLowA,B1g70-2300= (f * WSiA,B1970-2300) - (SoF (2300-1970)) (Eq 4-21)

55

where f= 1.5'10-3, being the conversion coefficient from silicate weathering rate (mo1e.m2.y-1)to surface lowering rate (m.y-l)(see sections 4.2.2 and 4.2.3).

scenario A

Weathered Silicate (mole.m-2) scenario B scenario C

Constant weathering rate: coniferous and deciduous stands

16.5

16.5

16.5

Weathering rate dependent on the pH: deciduous stands

18.0

19.0

227.1

Weathering rate dependent on the pH: coniferous stands

18.4

(8.7+3.5+214.9)

19.9

281.O (5.6+3.6+271.8)

Table 4-2. Total amount of silicate weathered (in molem2) in forested soils during the period 1970-2300 considering a constant weathering rate (0.05 mole.m-2.y-1)and a weathering rate dependent on the soil pH. In the case of the scenario C, the three terms correspond respectively to the silicate weathered during the period of dominance of the cation exchange buffer range, to the intermediate period during which deforestation proceeds and to the time span where AI(OH), dissolution is the dominant buffer reaction.

The net surface lowering exploring the scenario C is computed according to Eq 422 in three steps depending on the weathering rate: NSLowC1~ ~ o - ~ ~ o o = N S L g70-year+(50'brSiyear)+((2250-year)*brSi~~) OWC~ (Eq 4-22) where year = 2144 in deciduous stands and year = 2078 in coniferous stands. The calculated net surface lowering by 2300 in Swedish podzols depending on the type of afforestation is shown in Table 4-3.

Net Surface Lowering (rn) scenario C

scenario A

scenario 6

Deciduous stands

0.024 (0.022)

0.026 (0.022)

0.34

Coniferous stands

0.024 (0.022)

0.027 (0.022)

0.42

Table 4-3. Expected net surface lowering by 2300 for the studied scenarios depending on the forest stand assuming a weathering rate dependent on soil pH. Figures in brackets denote the expected surface lowering assuming a constant weathering rate.

56

Decreasing acid load: 2300-2406 After exhaustion of fossil fuels SO, atmospheric concentration returns very rapidly to the pre-industrial background level (1-2 months) (see section 4.1). In the developed model it is considered that SO,, as sulphate, accounts during this period for a constant acid load into the soil corresponding to the background sulfur cycling. Eq 2-6 and 4-5 describing the decay of atmospheric CO, from 2300 on and the solubility of CO, in rainwater are used to calculate the annual concentration of dissolved CO, (H,CO,(aq)) according to Eq 4-23. (Eq 4-23) H,CO,(aq)year = H,CO,( aq)2300*e-O007(year-2300) Hence, the acid load received by the soil can be calculated as indicated by Eq 424. (Eq 4-24) Hyear = a * l O3*SO,'-backg + 2*1O3*H&0,(aq)year where a = 1.33 in deforested soils (scenario C) or soils supporting deciduous stands and a = 2.13 in soils supporting coniferous stands. The evolution of the annual acid load entering the soil during this time span is shown in Figure 4-13. The soil cation exchange buffer capacity in the scenarios A and B is computed as in the previous period, by using the Eq 4-25: (Eq 4-25)

I

005,

I

Figure 4-13. Evolution of the acid load entering the soil in the scenarios A and B depending on the type of afforestation during the period of decay of CO, atmospheric concentration (2300-2406).

The variation of the soil cation exchange buffer capacity and base saturation with time within the scenarios A and B is depicted in Figures 4-14 and 4-15. The pH of the soil for the scenarios A and B is again calculated from the above mentioned Reuss' non-linear relationship (Eq 4-15). The evolution of the soil pH for the scenarios A and B during this period is depicted in Figure 4-16. Since it has been previously shown that for the scenarios A and B, only a slight increase on the silicate weathering rate can be expected as a consequence of the increased activity of hydrogen ions in the soil solution, the total amount of silicate

57

weathered during the decay period is calculated assuming the weathering rate to be constant (brSi). In the worst case scenario (C), the amount of weathered silicate is computed by using the weathering rate corresponding to the aluminum buffer range (brSiAl). The net surface lowering corresponding to this period is calculated for the scenarios A and B as was from 1970 to 2300. For the scenario C, it is computed according to Eq 4-26, assuming no formation of soil subsequent to weathering of silicate minerals (see section 4.2.2). (Eq 4-26) NSLowC2300-2406 = wSic2300-2406 The total amount of silicate weathered and the corresponding net surface lowering from 2300 to 2406 for the three explored scenarios is shown in Table 4-4.

Figure 4-14. Evolution of the soil cation exchange buffer capacity in the scenarios A and B depending on the type of afforestation during the period of decay of CO, atmospheric concentration (2300-2406)

Figure 4-16. Evolution of the soil pH in the Figure 4-15. Evolution of soil base scenarios A and B depending on the type saturation in the scenarios A and B depending on the type of afforestation during of afforestation during the period of decay of CO, atmospheric concentration the period of decay of CO, atmospheric concentration (2300-2406). (2300-2406).

58

scenarios A,B

Acid Load (molem2) Weathered Silicate (mole.m-2) Net Surface Lowering (m)

scenario C 3.0 167.5 0.4

3.0 5.3 0.007

Table 4-4. Acid load, amount of weathered silicate and net surface lowering from exhaustion of fossil fuels to the recovery of CO, background concentration for the three studied scenarios.

Constant acid load from natural sources: 2406-60.000 (Next Ice Age) The acid input to the soil-bedrock system during this time-span is assumed to be constant and in our model corresponds to 1860 atmospheric concentrations of SO, and CO,. The annual acid load is computed according to Eq 4-27 giving a result of 0.024 mo1e.m-2.y-1. (Eq 4-27) Hyear = 1.33'1 03*S0,2-backg + 2'1 03*H2C03(aq)backg The annual cation exchange buffer capacity is computed for the scenarios A and B as in the preceding periods according to Eq 4-9. The evolution of the soil cation exchange buffer capacity and soil base saturation in the most favorable scenarios (A and B) are shown in Figures 4-17 and 4-18.

12.

10.

0 5-

8-

severe conlrol

bane

CEBC 6.

..1ur.t,on

0 3.

imdema 4

Ol-

2.

0,

01

m

m

m

m

m

m

m

yes,

Figure 4-17, Evolution of soil cation exchange buffer capacity from the recovery of background atmospheric concentrations of CO, on for the scenarios A and B.

2400

2450

2500

2550

2600

2650

27M

yen

Figure 4-18. Evolution of the base saturation of the soil from the recovery of background atmospheric concentrations of CO, on for the scenarios A and B.

In these favorable scenarios, the soil pH is calculated as in previous periods (Eq 415) and its evolution with time is depicted in Figure 4-19.Table 4-5 shows the amount of

59

weathered silicate and the net surface lowering from recovery of background concentrations up to next Ice Age depending on the scenario considered. The global results of acid load, weathered silicate and net surface lowering (1970 to next Ice Age) of the developed model for each of the studied scenarios are shown in Table 4-6.

Figure 4-19. Evolution of the pH of the soil from the recovery of background atmospheric concentrations of CO, on for the scenarios A and B.

scenarios A,B

Acid Load (mo1e.m-2) Weathered Silicate (mole.m-2) Net Surface Lowering (m)

1.38*103 2.88*103 3.8

scenario C 1.38*103 9.1'1 O4 136.5

Tabley-5. Acid load received by Swedish podzols from 2406 up to next Ice Age. Amount of weathered silicate and corresponding expected surface lowering depending on the scenario explored.

scenario A

scenario

B

scenario C

__

Total Acid Load (mo1e.m-2) deciduous forests 1.46'1 03 1.49*103 coniferous forests Total Weathered Silicate (mole.m-2) 2.9'103 Total Net Surface Lowering (m) 3.8 Soil buffer range by the cation recovery natural levels Cop,SO2 exchange

1.52*103

I .60*103 2.9*103 3.8 cation exchange

1.79*103 2.02*103 9.147 O4 137.1 AVH)3

Table 4-6. Total acid load received by forested podzols from 1970 to next Ice Age depending on the explored scenario. Estimated amount of silicate weathered, expected

60 surface lowering and soil dominant buffer reaction after recovery of background concentrations of atmospheric SO, and CO, .

5. SUMMARY AND CONCLUSIONS As mentioned above, there are two ways in which environmental acidification could have an impact on the planned HLNW repositories. In this report we have focussed on the effects on the geological stability of the bedrock. A central scenario related to the combustion of fossil fuels for energy is explored. Sulfur dioxide emissions resulting from the use of these fuels are considered to be responsible for the acidification processes. It is also taken into account the expected global build-up of CO, concentration in the atmosphere. Three different sub-scenarios are studied depending on how stringent SO, emission controls might be during next centuries: mild control (scenario A), moderate control (scenario B) or severe control (scenario C). A linear increase of varying slope is thought to be the most realistic approach to model the evolution of the atmospheric SO, and CO, concentrations. For the scenarios A and 8 the depth of weathered bedrock is calculated by using the silicate weathering rate corresponding to the cation exchange buffer range. In the worst case scenario (C), the amount of weathered host rock is computed by using the silicate weathering rate corresponding to the aluminum buffer range. The main conclusions of the model calculations performed along the previous chapter are: - Due to the short residence time in the atmosphere of SO,, the time required to restore atmospheric background concentrations is short. This means that any global reduction on the emissions has an immediate effect on the various reservoirs. The behavior of CO, is rather different due to its longer residence time. The appropriate reduction measures would have effect only one century later. - The average atmospheric concentration for the scenario C is respectively 7 and 2.5 times larger than for scenarios A and B. However, the resulting acid toad introduced into the soil up to next Ice Age is only a 20% larger in the worst case scenario (C). This is mainly due to the fact that the anthropogenic perturbation lasts for only 0.5% of the total time covered in these calculations. - There is a well defined threshold for soil acidification, when the cation exchange capacity is depleted. In our calculations this would only happen in the worst case scenario (C). However, for the average case scenario (the one actually operating in Scandinavia), the degree of acidification has a dependence on the type of afforestation. At the end of the "fossil fuel age" (2300), for deciduous forests the pH of the soil is expected to decrease to regional average values around 4.2. In the case of soils under coniferous stands, the pH is expected to decrease down to 4.05. What is more important, the remaining cation exchange buffer capacity ranges from 5% for deciduous forests to only 1Yofor the coniferous ones. These are extremely low values for a buffer capacity to be effective and irreversible damage could be expected even in the average case scenario, particularly for podzolic

61

soils supporting coniferous forests. As a matter of fact, larger sensitivity to acidification has been observed in coniferous forests (12). - In the scenarios A and B the total weathering is only slightly affected by the differences in soil acidity. Over the acid load threshold value (total exhaustion of the cation exchange buffer capacity), the weathering rates are largely increased and so it is the resulting total weathering (case C). This a consequence of a major disruption in the soil structure and the larger availability of wetted surface. - The net surface lowering resulting from scenarios A and B is only around 4 meters up to next Ice Age. This amounts only to 1% of the depth of the repository (500 meters). In the worst case scenario, the computed net surface lowering amounts to some 140 meters (between 25 and 30% of the total repository depth). As we have already discussed, this is the result of a large acidification, extensive deforestation and consequently increased weathering rates. The net result of this scenario would be that the Scandinavian ecosystem would become hardly habitable and the performance of the HLNW repository a lesser problem for the biosphere. However, the increased weathering results in general in a larger penetration of the acidified surface waters into the undisturbed groundwater system. This could possibly have larger implications in the performance of the HLNW repository. These implications will be discussed in a forthcoming report in this series. The main consequence of this study is that an expanded usage of fossil fuels without the pertinent emissions control could affect the already stressed Swedish ecosystem. This has to be kept in mind when balancing different energy alternatives and their related cycles.

6. REFERENCES 1 Ahlbom K, Aikas T, Eriksson, LO. SKBnVO Ice Age Scenario.1990;

SKB Technical report. Semb A. Atmos. Environ. 1978; 12: 455-460. Graedel TE, Crutzen PJ. Scient. Am. 1989;261:(3), 28-36. Gibbons JH, Blair PD, Gwin HL. Scient. Am. 1989; 261 :(3), 86-93. Alcamo J, Shaw R , Hordijk L. (eds.) The rains model of acidification.1989; Kluwer Academic Publishers, Dordrecht. 6 Schneider SH. Scient. Am. 1989;261:(3), 38-47 7 Raiswell RW, Brimblecombe P, Dent DL, Liss PS. Quimica Ambiental. Ediciones Omega S.A., Barcelona:1983. 8 Wayne RP. Chemistry of Atmospheres. Oxford University Press, New York:l985 9 Sekihara K. Possible climatic changes from carbon dioxide increase in the atmosphere. In: J. O'M. Bockris (ed.) Environmental Chemistry. Plenum Press, 1. New York: 1 9 7 7 ; ~285-31 . 10 Rodhe H, Rood MJ. Nature 1986;. 321: 762-764. 11 Jacobson JS. Experimental studies on the phytotoxicity of acidic precipitation .In:T.C. Hutchinson, M. Havas (eds.). NATO Conference on effects of acid precipitation on vegetation and soils .Toronto: 1978;.Plenum Press, p. 151-160.

2 3 4 5

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12 Johnson AH, Siccama TG. Environ. Sci. Technol. 1983;17:(7), 294-305. 13 McLeod AR, Holland MR, Shaw PJA, Sutherland PM, Darrall MN, Skeffington RA. Nature 347, 277-279. 14 Wolt JD. Effects of acidic deposition on the chemical form and bioavailability of soil aluminum and manganese. In: A. Lucier and S.Haynes (eds.) Mechanisms of forest response to acidic deposition. Springer-Verlag, New York:l990 15 Tomlison II GH. Environ. Sci. Technol 1983;17:(6), 246-256. 16 Makela A, Schopp W. Regional-scale SO2 forest impact calculations. In: J. Alcamo, R. Shaw and L. Hordijk (eds.) The Rains model of acidification. Kluwer Academic Publishers, Dordrecht, 1989; p.263- 296. 17 Jacks G, Knutsson G., Maxe L., Fylkner A. Effect of acid rain on soil and groundwater in Sweden. In: 6. Yaron, G. Dagan and J. Goldshmid eds. Pollutants in Porous Media. Springer-Verlag,l984 ; p. 94-1 14. 18 Stumm W, Wollast R. Coordination chemistry of the weathering. Reviews of Geophysics, 1990. 19 Paces TJ. Geol SOCLondon:1986; 143: 673-677. 20 Slaymaker 0. Slope erosion and mass movement in to weathering in geochemical cycles. In: A. Lerman and M. Meybeck (eds.) Physical and chemical weathering in geochemical cycles.1988; NATO AS1 Series. 21 White GN, Feldman SB, Zelazny LW. Nutrient release by mineral weathering. In: A. Lucier and S. Haynes (eds.) Mechanisms of forest response to acidic deposition. Springer-Verlag, New York. 1990 22 Owens LB, Watson JP. Soil Sci. SOC.Am. J. 1979a; 43,160-166. 23 Owens LB, Watson JP. Geology 1979b; 7, 281-284. 24 Barth TFW. Geochim. Cosmochim. Acta 1961; 23, 1-8. 25 Johson NM, Reynolds RC. Science 1972; 117, 514-516. 26 Schnoor JL. Kinetics of chemical weathering: a comparison of laboratory and field weathering rates. In: W. Stumm (ed.) Aquatic chemical kinetics. Wiley & sons Inc. 1990 27 Bricard J. Aerosol production in the atmosphere. In: J. O'M. Bockris (ed.) Environmental Chemistry. Plenum Press, New York: 1977; p. 313-330. 28 Rodhe H, Granat L, Soderlund R. Report CM-64,Dept. of Meteorology of University of Stockholm and International Meteorological Institute of Stockholm.1984a 29 Rodhe H, Granat L. Atmos Environ 1984b; 18 (12), p. 2627-2639. 30 Soderlund R, Granat L. Report CM-54, Dept. of Meteorology of Univ. of Stockholm and lntl. Meteorological Institute of Stockholm.1981; 31 Tanke M, Gulik J v The global climate (Atlas). Mirage Publishing, Amsterdam: 1989 32 Hordijk L, Shaw R, Alcamo J. Background to acidification in Europe. In: J. Alcamo R. Shaw and L. Hordijk (eds.). The Rains model of acidification. Kluwer Academic Publishers, Dordrecht: 1989; p. 31-60. 33 Likens GE, Wright RF, Galloway JN, Butler TJ. Acid rain. Scient. Am. 1979; 241(4), 39-47. 34 Kauppi P, Alcamo J. Modelling soil acidification in Europe. In: J. Alcamo, R. Shaw and L. Hordijk (eds.) The Rains model of acidification. Kluwer Academic Publishers, Dordrecht, 1989; p. 179-221. 35 Kauppi P, Alcamo J. Linkages in the Rains model. In: J. Alcamo, RShaw and L. Hordijk (eds.) The Rains model of acidification. Kluwer Academic Publishers, Dordrecht, 1989; p. 297-317.

63

36 Johson AH, Sigg L. Acidity of rain and fog: conceptual definitions and practical measurements. Chimia 1985; 39, 59-61 37 Berden M, lngvar Nilsson S, Rosen K, Tyler G. Soil Acidification,extent, causes and consequences. Report 3292, National Swedish Environment Protection Board.1987. 38 Mayer R , Ulrich B. Input to soil, especially the influence of vegetation in intercepting and modifying inputs - A review. In: T.C. Hutchinson, M. Havas (eds.). NATO Conference on effects of acid precipitation on vegetation and soils, Toronto, 1978;. Plenum Press, p. 173-182. 39 Lerman A. Weathering rates and major transport processes. An introduction. In: A. Lerman and M. Meybeck (eds.) Physical and chemical weathering in geochemical cycles. NATO AS1 Series. Kluwer Academic Publishers, Dordrecht,l988; p. 1-10, 40 Wright RF. Influence of acid rain on weathering rates. In: A. Lerman and M. Meybeck (eds.) Physical and chemical weathering in geochemical cycles. NATO AS1 Series. Kluwer Academic Publishers, Dordrecht,l988; p. 181-196. 41 Reuss JO. Implications of the calcium-aluminum exchange system for the effect of acid precipitation on soils. J. Environ. Qual.1983; 12:(4), 591-595.

7 . ACKNOWLEDGEMENTS This work has been financed by SKB (Swedish Nuclear Fuel and Waste Management). We are very indebted to Dr. Peter Wikberg for his encouragement and support.

64

APPENDIX A. Selected stations to calculate the average total sulphate content in precipitation in Central and Southern Sweden. Median excess sulphate and sodium concentrations (pmole.dm-3) (Rodhe et al., 1984a and Soderlund et al, 1981). STATION

STATION

s042-

Na+

NUMBER

NAME

median value

median velw

98 1

20 6

16

KVARNTOR

21

FIAHULT

65 7

23 5

23

PLONNiNG

1090

79 6

32

SKURUP

1200

55 1

38

GOTEBORG

160 0

131 0

39

STOCKHOLM

1760

33 4

40

BOHUSMAL

97 2

336 0

42

RYDAKUNG

76 6

124

43

UPPSAIAN

96 8

193

44

KLUNKHYT

92 1

174

46

FARNABRU

91 7

129 14 1

47

BJORSUND

83 4

48

TARNA

90 0

14 5

51

AS

76 6

24 6

80

SJOANGEN

73 0

170

124

ARUP

98 I

23 5

126

GRANAN

81 6

28 0

128

KOMOSSE

72 8

42 7

131

SODERARM

1160

1250

133

GRIMSO

77 4

11 1

134

KALIANDS

69 5

150

135

AKERSHUS

84 7

23 3

1%

TORSO

66 2

10 1

137

VASE

83 0

18 9

65

ON CHARACTER OF EJECTION OF RADIONUCLIDES OUT OF EARTH SURFACE l.V.Melikhova, i.Vukovikb and V.Sipkah 'Lomonosov State University, Moskow bInstitute of Nuclear Sciences VinEa, Belgrade SUMMARY It has been established that the concentration of radionuclides at each point of atmosphere varies deviating from mean values in an order of magnitude or even more [l-41. Concentration variance are so significant that a problem of establishing the regularities of this change is recognized. In order to ascertain those regularities the results of many years rneasurernents of radioactivity of the ground layer of the air, carried out in The Institute of Nuclear Sciences in VinCa near Belgrade, have been analyzed. Those measurements which were done using standard methodology [S], characterize a particular region of Serbia, but may also be typical for any other region. The analysis has shown that the change of content of radionuclide in the ground layer of the air represents random ergodic process characterized by the probability of appearance of different meaning of concentration which change in accordance to the Fokker-Planck equation. Concentration of radionuclides has been changing impulsively showing minor dependance upon meteorological conditions, probably due to the seismic variations of radionuclides ejection froni the earth surface.

CHARACTER OF THE USED DATA The results (data) of radioactivity of the ground layer of the air were analyzed during the period of 1963-1991, at three checkpoints, located in a line towards dominating wind direction. Distance between first and second point was 2 km and between second and third point S km. At each checkpoint ground air was constantly filtered through a tube of 100 cm2 cross section at a height of I m froni the ground totalling 550 m2/24hr. The air was filtered through paper filters which retained more than 80% of radionuclides in molecular or aerosol state [5]. Filter radioactivity was measured either constantly or periodically. During measurements in situ filter was nioved at constant speed through a camera of proportional 2 7r counter [ 5 ] . During periodical ineasurements filter was immobile. Filter was measured at proportional 27r counter or spectrometer with NaI(T1) detector [6]. Each filter was measured several times in order to carry out measurement of contribution of short-lived (with disintegration constant (A > 2 10 sec-') and long-lived radionuclides. Filter radioactivity was compared with standards of'"K, '"Sr and '"U of appropriate activity. During measurements in situ instantaneous concentration was determined according to formula C = ( v x .At,,)-' where: v=500 m3/24h/,x=0.8 efficiency of retain of radionuclides, At 1 month fluctuations B(At), (Tables 2 and 4) basically are connected t o the appearance of the groups of daily impulses and monthly impulses. With the increase of At those fluctuations tend t o average. A t three check-points function (1 1 ) possesses identical character but different meanings of Aj. It is apparent through identity of functions D(B) which are standardized to mean meaning BlAt) at all three check points (Fig.7.).

77 REFERENCES 1 Junge CE, Air Chemistry and Radioactivity, Mainz Academic Press 1963. 2 Stein L, Radiochimica Acta 1983, 32: 163-17 1 . 3 Minato S, J.Nuclear Sci.Technology 1980, 17: 451 -469. 4 Kataoka T, at al. J.Nuclear Sci. Technology 1982, 18: 831 -836. 5 SmiljaniC R, PatiC D, Atmosphere Protection (Belgrade) 1979, 15: 29-32. 6 SmiljaniC R , Sipka V, Radiation Protection Report IBK-VinEa, Belgrade 1990. 7 Jain RK, Urvan LV, Stacey GS, Environmental Impact Analysis, New York: Van Nostarand Reinhold Comp. 1977. 8 Krasovskii AA, Phase Space and Statistic Theory of Dynamic Systems (in Russian)M, Nauka, 1974. 9 Data on Environmental Radioactivity Collected in Italy. Cornitato Nazionale Energia Nucleare, B10/26/62, 1962, B10/03/63, 1963.


CHAPTER 3 contamination from mining uses


19

Trace elements dynamics i n soils and aquifers of western Switzerland. 0. Atteiaa, J.-C. VCdp, A. Parriauxb. BIATE-Pedology, Swiss Federal Institute of Technology (EPF), Ecublens CH1015 Lausanne, Switzerland.

bGEOLEP, Swiss Federal Institute of Technology (EPF), Ecublens CH- 1015 Lausanne, Switzerland.

INTRODUCTION Along with increasing use of trace metals in industrial processes, many areas presently show abnormally high levels of these trace elements. This contamination, reinforced by the spreading of organic substances, poses a significant problem for the quality of drinking water. A better understanding of trace elements transfer in natural media is needed to manage water resources. The soil, and to a larger extent the pedosphere, supports and regulates many abiotic and biotic processes. By acting as a dynamic geoderma or skin of the earth, it protects, for example, the lithosphere from impacts of man’s activities. The chemical composition of rain is altered when it comes in contact with the vegetation cover, the soil surface, and percolates through the body of the soil cover. Soil represents a high capacity buffer medium: due to its porosity, to the presence of mineral and organic components, the pedosphere serves as source, filter and sink of fluxes of matter during atmosphere-pedosphere-hydrosphere interaction. Trace elements studies in earth sciences mainly originate in mining. In this field, trace elements are concentrated in specific minerals or rocks. Interactions between these minerals and solutions have been studied by geochemistry [ll. Soil geochemistry is well studied concerning major elements [21, but few papers deal with the geochemistry of trace elements in soils [3,41. On the other hand, soil scientist began trace elements studies with the advent of soil contamination and acid rain studies [5,6]. Numerous papers during the last two decades deal with pollutants adsorption equilibrium on clay o r organic matter [71. In biogeochemical studies of particular element, mass balance studies try to attach fluxes to biological cycling or weathering [8,9,10]. Some of these studies have proven t o be very useful to understand the biogeochemical cycles, but it remains difficult to have access to the real occurring processes. The key soil and aquifer properties such as pH, organic matter content, clay minerals content, oxidation and reduction processes are pre-eminent factors controlling the chemical behaviour of metals in ecosystems. The soil cover controls the concentrations of metal ions and complexes in the soil solution and

80

thus exerts a major influence on the quality of deep water. The behaviour of trace elements in natural media have been described using the atomic characteristics of the elements and the Goldschmidt rules. Taking into account the actual soil fbnctioning could be helpful in continuing this first classification. In this paper we try to specify some limits of trace elements fluxes in natural, low contaminated, ecosystems. This is done by comparing fluxes and concentrations of elements in rain, snow, soil solution and aquifer outlet. We first describe trace element behaviour at the general level, for the whole ecosystem. The role of soil vs. that of the aquifer, which is a n important task in hydrogeology, is defined using the behaviour of some trace elements. Taking in account concentrations of elements in soils and soil solutions, it is possible to precise the role of some phenomena involved in trace elements fluxes, mainly organic matter adsorption, biological cycling, weathering, soil acidification, etc.

SITES, MATERIALS AND RlETHODS Sites: A network studying aquifer typology is presently implemented in western Switzerland. This network includes a long term survey of 17 sites on a transect representing the principal alpine lithologies [ill. It is completed by a short term network analysing the chemical composition of numerous springs on some specific lithologies. The soil types existing in the network vary significantly; major types are typical of temperate climate zones [121: calcic cambisol, acid cambisol, Podzol (FA0 soil classification). For the soil study, we selected 7 sites in the whole network (fig.1).

Figure 1.Location of the studied sites.

81

The selection of the site is described in another paper [131. Here, we present the results of the three most important sites: Jura, Lutry and Argentihres. The Jura site is located in the Swiss Jura and concerns two calcic cambisol, one under forest and the other under extensive pasture; this site is called JUR. These soils are both thin with high organic matter content and nearly neutral pH. The Lutry site stands 20 km north of the Lausanne city. It is a small watershed under spruce forest with a deep acid cambisol lying on a fairly homogeneous sandy burdigalian molasse; this site will be referred as LRY. The Argentihres site is in France near Chamonix a t 2000 m altitude. The soil is totally covered by rhododendron and few short larches are spread near the site. This soil is a typical podzol developed on a recent gneiss moraine 1141; this site is referred to as ARG. Main site characteristics are summarised in Table 1. Table 1. Main characteristics of the sites. Altitude: altitude of the soil station/ spring: altitude of the spring/ Prec.: precipitation amount/ Int.: interception by forest canopy/ ET: Evapotranspiration/ size: estimated size of the watershed. Site

Altitude Spring Prec.

Int.

ET

Size Underlyingrock

m

m

mdy mdy mdy

LRY

885

880

1200

150

420

ARG JUR

2000

-1200

1700

0

350

?

Gneissicmoraine

1330

970

1600

300

250

-40

MalmLimestone

km2

0.02 Burdigalianmolasse

Soil survey is used to test the homogeneity of the watershed and to define the number of sites to be installed on each watershed according to the major soil units. Only JUR is equipped with two sites, one under forest and another under pasture, these land uses corresponding statistically to different soil types [ 151. Hydrological limits of the watersheds are appreciated with different techniques like structural analysis, tracers and isotopes [16]. The behaviour of trace elements from atmosphere to soil and spring is analysed on the three main sites only. Chemistry of the hundreds of springs of the short term network is used to test the validity of our results.

Material: To collect soil solution, two types of lysimeters (tension or free tension one) have been placed at the contact between the soil and geological layers [131. In the ARG and LRY soils, another lysimeter series is installed below the A2 and B1 layer respectively. The choice of the lysimeter type is made according to soil texture and structure. Only the LRY site is equipped with porous cups. On this

82

site tensiometric measurement allow the determination of the suction to apply to porous cups and the validation of a simple hydraulic model 1171. Chemistry of the extracts obtained with porous cups are compared using two types of lysimeters [ 151. Water balance has been verified using cumulative chloride fluxes analysis in rainfall and soil solution 1151. Bulk precipitation is collected with polyethylene (PE) f h n e l s and throughfall under forest is obtained with three PE gutters. All the materials have been tested and proved no trace elements contamination. Prior t o installation, all materials were rinsed with nitric acid and deionized water. After every sample collection, collectors are washed with distilled water. Spring water is collected directly from springs in PE bottles.

Analysis: Collected water (rain, snow, soil solutions and spring water) is analysed for anions, major cations and trace elements. Before analysis all the samples were filtered at 0.45 pm. Samples are partly stored at -2OOC for anion analysis and partly acidified with 0.5% of suprapure HN03 and stored at 4°C. Major cations are determined with a DCP-AES (ARL,Spectraspan 11). Sulphates are analysed with ICP (Jobin-Yvon 38) in S form and with ion chromatography (Sykam). The results agree well with a difference less than 10 %. Other anions are determined by automated colorimeter technique (Technicon).Trace elements are analysed with ICP-MS [181. For total soil content of major elements, soil is grounded and fused at 1200°C with a melter product (Sr metaborate) and then dissolved with HNO3 2%. For total trace elements content, X-ray fluorescence is performed on pastilles of grounded soil. For extractable elements we used two techniques: (i) Acetate Ammonium-EDTA (NH4EDTA) at pH 4.65 with a 1 : l O soil-solution ratio for the determination of "mobile" elements [ 191 and (ii) oxalate-oxalic acid (Tamm reactant) at pH 3 with a 1:40 soil-solution ratio for the determination of amorphous oxides [20,211. The role of the extraction procedure is analysed elsewhere [El. On the LRY soil samples, simple water extract with a 1 : l O soil to solution ratio has been tested. Fluxes calculation: Precise calculation of water fluxes in soil requires an important device in the field and good computer codes [221. In trace element budget studies fluxes are commonly calculated with precise concentrations of elements but with very poor estimates of water fluxes. Considering that the hydraulic functioning of the lysimeters and rain gauge are consistent around the year, we can consider the average concentration, weighted with volume, as a good estimate of yearly element concentration. To allow comparison of element concentrations at different levels of an ecosystem, evapotranspiration must be taken into account because it modifies concentration of elements in solution. Concentrations of the elements in the atmospheric water were calculated by removing the amount of estimated evapotranspiration from the total amount of precipitation. This calculation is also done between the two lysimeter depths when they exist. No evapotranspiration is assumed between soil bottom and spring. Evapotranspiration calculations are detailed elsewhere [151.

83

To appreciate the storing or releasing capacity of soil for each element, we calculated ratios of input concentrations divided by output ones. If this ratio is higher than 1 the element is accumulated in the system, o r budget is positive, it is the inverse when ratio is Iower than 1.Two of these indicator ratios have been calculated: from atmosphere to spring to test the whole ecosystem response and from atmosphere to soil bottom to test soil effects. Hydrological processes To be able to compare concentrations of elements at different levels of the ecosystem we have assumed that all the spring water comes from the atmosphere and passes through soil. Nevertheless this assumption is rarely entirely verified in mountainous aquifers. Figure 2 presents the major types of circulation of water existing in aquifers. The first case (1) is the one considered in our hypothesis: continuous flow between atmosphere and spring. In mountains, important zones are bare rock (2) and therefore soil effect acts on a small part of the water entering the aquifer. If soil bedrock is different from that of the aquifer (31, these two media can be totally disconnected. In aquifers, water,can flow through discontinuities, such as faults, joints and layering, (4) rather than in the whole rock. In this case water is in contact with a material eventually different from the bulk rock considered as the aquifer. Presently, it is known that variation in flows can modify the chemical characteristics of water such as pH, redox potential, temperature.

Figure 2: Diagram of the flows in soil-aquifer systems in mountain. In real conditions, it is very difficult to quantify the role of each of the previous hydrologic passes. This is why ratios of element concentrations obtained in each medium will be used as indicators, but the discussion will be focused on phenomena.

84

RESULTS Atmospheric deposition Study of atmospheric deposition of trace elements is detailed elsewhere [231. Average values are presented in Table 2;major features can be described as the following: - Bulk precipitation is fairly homogeneous over the whole region for most elements. - Only Cu, Pb, Zn and B clearly originate from human activities. - Fe, Ni, V, Rb and C r show low levels of concentrations and mainly come from earth eolian dust. - Sr and Ba show variable concentrations which are influenced by the existence of local eolian transport from outcrops surrounding the site. Table 2. Median and mean concentrations of trace elements (pgA)in rain and snow on the six sites. pgA B median 3.1 mean

V

Cr

Fe

0.3

0

3

Mn

Ni

Cu

Zn

Rb

Sr

Ba

Pb

6 0 1.9 21 0.5 2 2.6 1.8 5.6 0.45 0.3 8.04 5.63 0.34 2.75 22.3 0.37 2.45 6.8 2.17

The effect of forest on trace element concentration is described in another paper [24]. Concentrations of most heavy metals in throughfall are similar to that of bulk precipitation. Sr is enriched by roughly 30% and Ba by more than 100%; Fe and B are slightly retained by vegetation. Mn and Rb show important biological cycling, throughfall concentrations being an order of magnitude higher than those of precipitation. Hence, we won't focus on variability of atmospheric input in the following discussion,but forest effects will be considered for some elements.

Spring water composition Trace element contents of spring water are presented in fig.3 and 4 for the total network (AQUITYF') and for three different types of aquifers. Presently Evaporites of the Trias, Flyschs of the Niesen nappe, and Cristalline rocks of the Mont Blanc massif have been obtained. Similar studies on carbonates and molasse have been started; see table 3 for a description of the whole network. Strontium is considered as trace element, even if its concentration can reach 8 to 10 mgA because it has been less studied than major elements and its concentration in atmosphere and soil solution is at trace level. For other trace elements, Ba presents the highest concentration (from 10 to 120 pgA) followed by B, Fe and Li which concentrations vary from 0 to 60 MA. Among this group, Ba is more concentrated in Flysch aquifers and Li in Evaporites.

85

Table 3. Spring of the global AQUITYP network (geological substratum followed by spring symbol). LIO Alluvial gravels LUC Aquitanian CHE Malm Jura molasse MAL Delta gravels CHA Burdigalian LRY Triasic BOR molasse poz limestones Fluvio-glacial MOR Niesen Flysch LLI Helvetic Malm SAR gravels limestones Fluvio-glacial DIZ Gurnigel nappe ALL Alpine NOC gravels under THI Evaporitic Trias BLE moraine subAlpine COR Cretaceous TIL Cristalline Mont BOR molasse Jura Blanc

'267 75

+AOUITY P (1 7) 4 Evaporitic (86)

Flvschs 1206)

+ Cristalline (1 10)

50

25

0k-b

Sr

0

L As

I

U

Mo

I

W

Figure 3. Composition of spring water (symbols represent the median, bars the 10 and 90% percentiles, number of springs in parentheses).

V, Co, Rb, Ni, Cu, Cr, Mn and Zn concentrations in spring water are quite low having median values lower than 1pgA, the range of variation is also quite small. Among this group, Evaporite aquifers present statistically higher Rb, Ni, Cu and Mn concentrations, the highest difference existing for Rb. In Cristalline aquifers Rb concentrations show a larger dispersion than in the general network. Concentrations of U in Cristalline massif can reach very important values (up t o 1 mgA) and become the first trace element at some places. High concentrations of W, Mo and As are also noticed in these aquifers.

86

Concentrations of Mo and W are lower than pgA in the other aquifers of the short term network. We do not present here the results of I and Br analysis, even if they exist at concentrations higher than 1 pgA, for different reasons: (i) they have not been analysed everywhere, (ii) the detection limit is higher than for other elements and (iii) their behaviour is different from the other elements as they are anions in water. Numerous other trace metals and rare earth elements have been detected and analysed in the semi-quantitative mode. Their concentrations lie between 0.01 pg/l and 1 pgA. Further analyses are being developed t o quantify the concentrations of these elements in the range of ngA. PLll

P@ 120

l8

-8-

AQUlTYP(17)

++

Evaporitic (86) 16

100

+

Flyschs (206)

80

+

Cristalline (110) 12

?

!) 29.3

23

14 10

60

a

40

2oi 0

I

Ib

lb

Jk n

Zn

Figure 4. Composition of spring water (symbols represent the median, bars the 10 and 90% percentiles, number of springs in parentheses).

Trace elements budgets at the ecosystem level We call budget at the ecosystem level the behaviour of the elements from the inlet t o the outlet of the watershed. In this first approach, soil + aquifer is considered as a big black box. We analyse the ratio, R1, of average weighted concentration of one element in rain and snow divided by its concentration in spring water. Therefore, if the ratio of the element is higher than 1,the element is accumulated in the whole ecosystem, the budget being positive; in the opposite case the element is released during the water course with a negative budget. Values of R1 for each of the sites and for the numerous springs of the short term network are reported in fig.5. We can separate 4 groups: - Mn, Zn, Cu, Pb, and to a lesser extent Fe, are accumulated in the ecosystem (R1=8 to 30).

87

- Ni, Rb, B, V and Ba concentrations are almost unchanged during the ecosystem crossing. - Sr is highly dissolved in every ecosystem. - Cr shows important differences between sites. - U, Li and Co are dissolved in some aquifers. Atmosphere I springwater

1

I.

JUR

0

ARG

.

0

.

0

*

0

0

1 .

LRY AQUITW

0

~

0

0.01

1

.

Zn Mn Fe Pb Cr Cu Rb Ni

B Ba

V

-T/

Sr Co

U

Li

Figure 5. Value of R1 ratio of trace element concentrations in meteoric water to spring water. (error bars for the global network present 10 and 90% percentiles). Difference of ratios between sites is generally important. The use of ratios can give considerable importance to slight differences in spring water concentrations if these concentrations are low. Therefore, these ratios must be used only for description of general behaviour of trace elements.

Soil effect The average concentrations of trace elements in the different sites and different media are gathered in Table 4. Major elements are presented as reference to define the whole soil behaviour. Statistically different groups obtained by an Analysis of Variance (ANOVA)are differentiated. The elements can be divided into three groups: (i) Mn, Zn and Fe that can reach concentrations greater than 50 pgA in some sites, (ii) Ba, Sr, B and Cu whose concentrations are between 1 and 20 pgA and (iii) Ni, Pb, V, Cr and Rb generally lying around 1 pgA. Among these elements Ba, Rb, Pb and Cr do not present any statistical differentiation between sites. No differentiation between

88

sites is valid for all the elements, but trace element content is generally higher at LRY and lower at ARG. Table 4. Average weighted concentrations of elements in soil solutions. Period of March 1990 to November 1991. Concentrations are weighted with volume of water. For trace elements, letters show statistically different groups of data. mg/l pH Ca Si Mg K Na Al C1 N-NO3 S-SO4 ARG 4.88 0.59 1.75 0.11 0.17 0.37 0.34 0.43 0.21 0.46 JUR 7.66 33.7 5.86 0.38 0.16 0.49 0.09 1.42 1.71 0.93 LRY 5.62 2.79 4.94 1.59 0.37 0.93 0.41 1.41 0.33 2.85 pg/l ARG JUR LRY

Mn Zn Fe Ba 4.62a 11.P 69.4b 12.la 1.09a 13.9 17.28 18.8a 158b 130b 20.6a 168

Sr B 1.85a 3.39 11.9b 9.64b 12.7b 12.4b

Cu Ni Rb Pb V Cr 1.48a 1.08a 1.08a 0.88a 0.08a 0.02a 2.42a 0 . 7 9 0.4a 0.25a 0.72b 0.238 5.68b 2.04b 0.69a 0.83a 1.11b 0.21a

Table 5 shows soil solution concentrations of some heavy metals in different ecosystem studies. Table 5. Average concentrations of some heavy metals obtained in other studies: Solling, Spanbeck (Germany): [251 apart for Ni and Cr: [81, Varsjo and Gardsjon (Sweden): [lo], soil layers 1 , 2 , 3 are respectively A layer: 15 cm, B1 layer: 35 cm and B2 layer: 55 cm. PLgfl

Solling Spanbeck Varsjol 2 'I

3

Gardsjonl " 2 "

3

Zn

Cu

Ni

Pb

Cr

520

9

15

2

0.6-1.2

540

9.6

34

4

0.9

1.2

65

2

2.2

1.7

151

1.4

4.1

0.8

38

4.1

1.8

19.5

2.3

55

0.7

2.5

2.5

1.2

75

0.7

3.6

1

1.3

4

89

Concentrations a t the Solling site are higher than those existing in each of our sites. But Solling site is known t o be highly polluted and highly acidified. The two Swedish sites present soil solution concentrations of Cu, Zn and Ni similar to LRY levels and even slightly lower. These concentrations are higher than the one existing in JUR and ARG. Concerning Pb and Cr, concentrations in all our sites are lower than those of Swedish soils. For Cu,KeZZer [261 finds results lying between LRY and JUR values for one cambisol and one podzol in the same region of Switzerland. The studied sites can be qualified as slightly contaminated in heavy metals for ARG and JUR whereas LRY seems t o show intermediate levels. Soil effects on atmospheric water composition is tested by using R2, i.e. the ratio of concentrations in soil solution to those in atmospheric water. This ratio has been plotted in fig.6 with the same representation as for R1 in the previous section. Values of R2 are spread on both sides of 1 for many elements, the exceptions being Cu and Pb which are retained in all soils and Ba, and t o a lesser extent Sr, Ni and Fe, which are released by all soils. Soil solution I Atmosphere 0

0

8

0

-.-

0

0

0

~

w

H

'

H I

0 -

I

Pb Cu Cr Mn Zn V

B Rb Sr Ni Fe Ba

Figure 6 . Value of R2 ratio of trace element concentrations in meteoric water t o soil solution. (Mn and Rb ratios have been calculated using concentrations in precipitation to exclude biological recycling effect). Some elements (V, B) present R2 values close to 1 on all sites. This behaviour suggests that soil has no effect on the concentrations of these elements. By contrast, values of R2 for Fe, Mn and Zn show important differentiation between soil types. Therefore, few trace elements found in soil solutions are influenced by pedogenetic processes.

90

Elements concentration on their way through the ecosystem

6 ,4 Q,

32 0

-

.4 Q,

='2

0 1251 - 100 75

4

.

.

-

Q,

2 2

=L

0

-

- - I

Ni

BP

Tf

A2/B1 Bs/B2 Sp

50

25 0 125 1004

Zn

:75-1

BP

Tf

A2/B1 BdB2 Sp

Figure 7. Average concentrations (pg/l) of trace elements in bulk precipitation (BP), Throughfall (TO, soil solution (A2,Bs, B1, B2) and spring (Sp) for the three main sites.

91

Fig.7 describes the concentrations of each element for the different levels of the ecosystem (atmosphere, soil and spring). Used concentrations are the average concentrations weighted by volume and corrected for evapotranspiration or interception effects as stated in the fluxes calculation section. The ARG spring composition is a n average of water coming from different fractures of one tunnel, therefore it has to be considered with care. Concentrations of Cu and Pb decrease from atmosphere to soil solution, the decrease being dramatic for Pb, concentration in soil top layers is low and conservative until spring. Copper shows smoother decrease in each compartment. For V, Ni and Rb, levels of concentration in top soil are close to atmospheric ones. Between top and deep soil layers, besides Ni concentrations that increase at the LRY site, all elements concentrations remain constant. Ba concentrations always increase between atmosphere and top soil. While this increase of concentration remains at the LRY site in deep soil, concentrations decrease a t the two other sites. Fe levels are quite stable between atmosphere and soil solution and become null in spring water. The major exception to this scheme being the A2 layer of ARG which releases a large amount of Fe compared to other sites. Nevertheless Fe concentrations remain at trace level. A t the JUR and ARG sites, Zn concentrations decrease along the water way and Mn levels are always very low. At the LRY site the scheme is totally different with a net release of Mn and Zn in the B layer. Mn pattern at LRY is different from Zn one: concentrations abruptly increase between precipitation and throughfall and Mn is retained in the top soil. This pattern is characteristic of biological cycling as cited by different authors [24,27,28]. The site showing the more specific features is LRY: not only levels but also vertical evolution of concentrations are different from the other sites. This fact is surprising because the most similar sites, JUR and ARG, are the most different ones from the pedogenesis point of view.

DISCUSSION

Trace elements budgets for the whole ecosystem One can find the sequence of ecosystem budget in different kind of studies concerning the general behaviour of trace elements [291. For instance Tardy [3] worked on the concept of mobility which is a ratio of element concentration in solution to its concentration in solid phase. Hence accumulated elements ( R b 1 ) will show a lower mobility than elements dissolved during water course. Study of solids and solutions in the Massif Central (France) leads to the following mobility sequence: Sr > Ni, Rb, V, Cu, Pb > Ba, Cr > Fe, Mn. Groups and order between groups are the same as in our study except for Cu and Pb. The difference existing for these elements can be explained by the high concentration of these elements in atmospheric deposition that was not taken in account in Tardy study. Nevertheless the validity of the previous scheme is of little practical impIication for many elements because of the important differences existing between sites (fig.3).

92

Trace elements retained by the ecosystem (Cu,Pb) Only Cu and Pb are retained by soils and aquifers. Their concentrations decrease along the entire water course while their total concentration in solid increase from the bottom to the top soil layers. This pattern is specific of elements retained in the topsoil, and has also been observed by different authors [10,251. Concerning Pb, we show elsewhere 1151 that this element is not submitted to biological cycling by forest. This is also concluded by Heinrichs [8].In both acid soils, Pb is almost totally extracted by T a m reactant (cf. Table 6). Moreover, in these soils total concentration of Pb in soil is directly related to the total organic matter content (fig.8). The difference of slope of Pb - organic matter correlation between the two sites is due to higher organic matter content at ARG site than at the LRY one, whereas total amount of Pb is similar (cf. Table 5). In soil solutions of low pH, such as in acid sites, the concentration of dissolved Pb is equal t o the one existing in the carbonated site. In other studies [301, i t can be seen that even in more acidic soils Pb is retained with organic matter in the top of mineral layers. In the studied soils Pb seem to remain bounded to organic matter and only moves when this one moves. This is also described by Tyler [311 and Berggren [32]. It is important t o study origin and movement of organic matter and particularly its relationship with hydrological factors like quick snowmelt on podzols.

60,

1

I

Regression lines

LRY Pb = 1.99 * %O.M. -3.42,

R2 = 0.97 ARG: Pb =O. 47* %O.M. + 9.91, R2 = 0.98

0

20

40 60 % Organic Matter

80

Figure 8. Relationship between total soil content of Pb and organic matter.

93

Table 6. Total soil content (ppm or mg/kg), % extractable with T a m reactant (Ox.)and Acetate-NH4-EDTA (Ac.) of some trace elements. for JUR,f: forest, p: pasture. For JUR and LRY numbers are median depths of the layer, for ARG layers are referenced by names. B: bedrock composition. Mn Ox Ac Pb Ox Ac Cr Ox Ac Ni Ox Ac V OxAc ppm % % pprn % % ppm % % ppm % % ppm % % JURf3 1010 39 19 29 5 5 110 1 0 57 14 1 137 3 0 JURfl5 1030 42 13 24 3 3 118 1 0 63 13 0 146 3 0 91 2 0 JURf25 740 45 6 8 7 4 44 18 0 103 4 0 JURp5 1120 46 0 31 8 0 111 1 0 46 23 0 141 4 0 JURpl5 1130 45 9 16 13 4 110 1 0 46 22 0 134 4 0 JURp25 1210 43 9 21 4 3 112 1 0 46 21 0 142 3 0 J U R B 110 6 26 9 4 LRY4 LRY14 LRY30 LRY5O LRY70 LRY9O LRYllO LRY130 LRY150 LRY B

440 660 940 840 750 680 620 500 790 570

ARGaO ARGal ARGe ARGbh ARGbhs ARGbs ARGB

50 110 190 250

38 53 57 50 52 38 38 29 48

15 2 0 100 340 100 420 1 420

12 4 2 3

2 1 1 1

5 8 1 0

45 30 2

46 10 4 6 5 3 4 5 1. 1.

83 42 100 41 100 51 100 19 100 16 100 15 100 10 88 7

77 100 94 106 99 118 105 121 110 122

51 3 14 20 12 12 13

21 62 100 100 100 39 100 19 100 16 72 15

18 0 0 18 1 0 9 2 0 30 14 1 32 13 2 22 5 0 26

6 5 5 3 4 3 3 3 1

2 0 0 0 0 0 0 0 0

15 100 0 20 100 7 21 98 2 26 52 1 30 54 1 36 24 0 41 18 0 42 16 0 35 14 0 55

44 55 54 5 6 5 9 59 55 55 5 1 53

10 92 73 2 100 100 0 100 100

9 11 6 28 2 1 29 4 0 92 15 1 56 8 0 41 2 0 43

n

n 6 100 17 7

10 7 7 5 5 3 3 3 2

1 0 0 0 0 0 0 0 0

Concerning Cu, Xray fluorescence detection limits (approx. 4 mg/kg) are too high to measure Cu levels in the different soils. Cu concentrations in the soil solution are similar in the soils with pH varying from 4 to 8. This behaviour could be explained by Cu strong complexing with organic matter a s cited by Berggren 1321, B u f f e [33] or KeZZer [261. The average ratio of element in soil solution (pg/l) vs. “mobile” amount (NH4EDTA in mg/l) is 0.42 for Pb and 26 for Cu which shows that Pb is much more strongly bounded to organic matter than cu.

94

Trace elements showing similar behaviour in all soils (B, V,Cr, Ni, Ba) Average B concentrations in soil solution are close to the existing one i n atmospheric deposition and almost not influenced by soil type. B and Ba concentrations, as Si ones, increase in summer and decrease in winter. This evolution can be related linearly to a drought factor on the ARG and JUR sites (fig.9) and water suction at LRY (Not shown). This drought factor can be expressed as the inverse of a rain intensity factor or in month / mm of rain. Concentrations increase when soil receives less water, this increase being higher than pure evaporation effects. Keren [34,35] showed that B concentrations depend on the soil to solution ratio and the clay type existing in the soil. Co-evolution of B and Si concentrations can be expIained by similar chemical properties: they are both metalloids, undissociated weak acids being hydroxides, at the soil solution pH. As the interlayers are generally filled with hydroxides [361, the relationship between concentration and drought could be explained by clay shrinkage which excludes B hydroxides from their interlayers.

40 -

B = 2.31*A + 4.25 R2 = .8

I

Ba = 6.75*A + 9.53,

30-

2 d 60-

I

2 20F9

.d

0

.

10

0

0 0

8 101214 A (month I mm of water) 2

4

6

0

2 4 6 8 1012 A (month I mm of water)

4

Figure 9. Relationship between B and Ba concentrations in soil solution and a drought factor (A) in month/ mm of rainfall. Concentrations of V, Rb,Cr in soil solutions are low and similar in different soils. The role of the soil is difficult to specify because of tiny differences between atmospheric and soil solution concentrations. On the LRY and ARG sites we have calculated the total amount of one element lost in the upper 80 cm (or 50 cm for ARG). Soil concentrations have been corrected for density effects by assuming that Zr is invariant. The depletion of one element is calculated as the difference between corrected soil

95

concentration and bedrock concentration. This depletion is compared t o the annual fluxes lost in soil solution at this depth. This is equivalent to the time t o deplete the soil layer with a constant flux equal to the currently measured one. For major elements, table 7 shows that this simple calculation leads to a time of the same order of magnitude as the soil age (8 t o 10000 years). This can be interpreted as a constant dissolution rate for these elements, this rate being independent of soil pH and organic matter content evolution. Table 7. Calculated time (thousands of years) to deplete the soil layer with the current flux LRY

Al K Si Mg 13.5 13.4 4.6 -

ARG

6.6 3.8 3.5

Rb

Ni

V

Cr

Fe

Ba

6.2

9

-

100

28

1.2 0.01

3.6 4.9

-

13

1.8 0.01 0.01

11.6 24

Zn

Mn 0

Calculated time of depletion for Rb, Ni and V are similar t o the one calculated for major elements. Calculation has not been done at ARG for Ni because Ni lost by weathering can be biologically recycled and adsorbed to organic matter. For V at LRY, fluxes in solution are too close to fluxes in atmospheric deposition regarding the errors made on these measurements. These three elements can be regarded as originating from weathering in soils. The fluxes are different and decrease from Ni to Rb and V. This can be related to the extractable proportion of these elements (Table 6) and to their different solubilities. The low soil content of these elements and the slow speed of the weathering process justifies the low concentrations in solution. In minerals, Rb substitute K with a ratio of 160 to 300 [l] because of its similar radius and charge [37].A t the LRY site, K/Rb equals 228 in bedrock, 180 in soil top layer and 333 in soil solution. For ARG, this ratio equals 166 in bedrock, 105 in the soil top layer and 196 in soil solution. These ratios are fairly close which shows that Rb behaviour is close to K one. The ratio decreases from bottom t o top of the soil and is higher in soil solution than in soil. This proves that Rb is more resistant to weathering than K. All ratios are higher for LRY than for ARG so the composition of soil solution from different soils reflects the difference of mineral composition. Rb concentrations or WRb values are then related to the composition of the original material and to the weathering rate. Comparison for Cr shows very long time for achieving current depletion. C r being in particulate form at this pH [381,soil solution concentrations do not indicate total annual flux. Release of Ba in soil solution has not been related to a particular chemical phenomenon though its presence in feldspars [l] can explain its high concentration in soil solutions. Similarities between B and Ba (fig.9) cannot be explained by classical chemical properties of these elements, which are very different. Ba behaviour needs to be studied more deeply.

96

Elements influenced by pedogenetic processes (Fe,Mn, Zn) Iron is an important metal in pedogenesis studies [361. In fact in this study Fe behaviour is very different among the sites (cf. fig.7). Profiles of total Fe in soil outline a depletion in the A2 layer of the Podzol and in the 50 top cm of the cambisol. These high depletions are not compatible with the current fluxes (cf. Table 7) because Fe, like Cr, is mainly in particulate form [391. Fe shows an original and well-known behaviour in Podzol [40,41,42] with a depletion in the A2 layer. Accumulation of Fe in amorphous form in the spodic layer leads t o a total content twice as high as that of the bedrock. Lateral movements of solutions in the top layers amplify this accumulation [14,15].V and Cr follow the same accumulation pattern as Fe in this soil type. Mn and Zn are only released in acid soils and particularly at LRY (Acid cambisol), concentrations at this site being 5 to 12 times higher than those a t ARG. Following the same calculation of total depletion time done in previous section, table 7 shows that current Mn and Zn fluxes are high compared t o the historical depletion. Release of Mn and Zn at the LRY site has recently increased tremendously. Moreover, fig.10 shows that concentrations of these elements in a water extract of LRY soil increase when soil pH drops below 5. Actual Mn and Zn release in soil solution seems t o be related to recent soil acidification. This behaviour is similar to that of A1 during soil acidification 1431,and has been forecasted by models of soil acidification [441, by laboratory experiments [3 11 and ecosystem studies [lo].

250

u I

Figure 10. Evolution with depth of H+, Mn and Zn concentrations in water extracts of LRY soil. At ARG, Mn and Zn concentrations are much lower than those at LRY while soil pH is slightly more acidic. The low value of Zn concentrations can be

97

explained by the almost total depletion of Zn in the A2 and Bh layers. Concerning Mn, redox potential role seems t o be insignificant because Mn concentration does not vary with the level of the water table. Moreover, ARG site is more affected by soil saturation than LRY site. Up t o 50% of Mn is extractable with Tamm reactant in LRY soil mineral layers whereas this value is less than 5% in ARG. This could be explained by the existence of Ferro-magnesians minerals a t the LRY site. SOpH is not the only factor governing Mn and Zn release; existing mineralogy plays a significant role. Fe, Mn and Zn exist at very low concentrations in spring water. This is due to the pH of the water. A t such a pH these elements precipitate in hydroxides.

Elements released by the aquifers Sr is the only trace element released from all soils and aquifers to solutions. In almost all solutions, Sr concentrations are correlated to Ca ones. In carbonated soils the Sr/Ca ratio is equal to that of the underlying rock proving that Sr originates from calcareous gravel existing in soils. Sr concentrations in these soils are lower than those of spring because equilibrium with calcite and Sr-CO3 is not reached in soils. This is due to the low contact time between calcareous gravel and soil solutions. Sr can be considered to be a major element in certain springs (concentration of several mg/l), this fact being characteristic of aquifers containing evaporites [451. These high concentrations come from the high solubility of celestine (SrS04). On other sites, Sr mainly comes from SrC03 dissolution, its behaviour being close to that of Ca. Numerous metals (V, Co, Mn, Fe, Zn, Cu, Pb, Ni) exist at low concentrations (app. 1 pgA) in springs compared to soil solutions. They are retained by different reactions: precipitation in carbonate and hydroxides forms, adsorption on clays existing in rock fissures, etc. Considering the p H of the spring water each of these reactions are likely to occur and we have no means to precise the prominent one. The relative importance of each reaction is determined by differences in reactions kinetic. The major exceptions to this scheme are the evaporite aquifers. In the spring water coming from these aquifers, concentrations of these metals are higher than 1 pgA or even than soil solution concentration. This is mainly explained by the association of these metals with highly soluble sulphur minerals lying in the aquifer. In evaporitic aquifers other traces such as Li and Rb are released in solution, these alcaline elements could be associated with halide minerals knowing their chemical structure. Behaviour of B and Ba differs highly among aquifers as they are stopped in numerous aquifers and liberated in others. B is released by evaporite and cristalline aquifer and Baby flyschs. More knowledge about the clay and feldspar content of these aquifers is required t o understand the behaviour of these elements. Cristalline aquifer is particular in the sense that trace elements existing in soils are absent in the spring water whereas totally different elements (U, Mo, W) are released by the aquifers. The conditions of rock forming can explain the high content of these elements, generally existing in reduced conditions. Cr present high concentrations at LRY spring, and considering the water pH, it can only be in Cr2072-form [39]. As in soil, Cr is in Cr3+ form [46] it has

98

undergone an oxidation between soil and spring. This phenomenon is rather surprising but it has been observed by Bartktt 1471 who associates this oxidation to Mn reduction at specific Redox potential. All elements typically released by aquifers have been found i n solutions of rocks powder mixed with CO:!acidified water. The origin of these elements can therefore be related directly to rock weathering in aquifers. Along the discussion we have compared concentrations at different levels of the ecosystem, but we have explained in methodology section that this comparison could be erroneous in mountainous aquifers. In fact cases must be differentiated: First (i), in small aquifers, such as LRY, where flow passes are known, the assumptions concerning water flow are almost verified. Second (ii), in large aquifers of highly soluble rocks, pH retains the metals existing in soil solutions and composition of the springs are quickly determined by the composition of the rocks. This is the case of calcareous and evaporite aquifers. Third (iii), in cristalline aquifers direct infiltration from atmosphere and fracture flow are important, but in that case most of the trace elements contained in spring are specific to aquifers and their origin cannot be attributed to soil. The only elements for which the roles played by soil and aquifers are not clear are Ca, Sr, B and Ba. For Ca and Sr the question is the proportion of Ca and Sr coming from calcareous soil’s gravels. We have determined a proportion of 2/3 in one site, but it vanes according to the microbial activity. Moreover, it is not possible to determine the proportion of Ca and Sr coming from bare parts of the catchment. As B and Ba can be released or stopped in aquifers, the proportion of B and Ba due to soil cannot be estimated if the proportion of the aquifer covered by soil is not known.

CONCLUSION: The first task of this study was to define the respective roles of atmosphere, soils and aquifers in the release of trace elements in springs. In a general thinking we can say that few trace elements (B) pass through the whole system without any influence. Most of the elements released in the soil are stopped in the aquifer due to their pH. New kind of mineral or chemical conditions induces the solubilisation of specific trace elements in aquifers. The liberated elements differ with lithologies of aquifers, therefore they can be used as tracers of specific rocks. This characterisation is the aim of the AQUImP project. If we look at the pollution aspect, studied springs are currently free of trace contamination. This is only valid for dissolved trace elements. Nevertheless, the three types of ecosystems will react differently if the current scenario of atmospheric inputs continues. - On carbonated rocks, all trace metals are fixed in the first meters of the aquifer, the size of the aquifer protects them against trace element’s contamination. This fact does not exclude nitrate or organic pollution - On Podzol, organic matter carries a lot of heavy metals but its fixation in spodic layers is a good trace stabiliser. A problem can arise when organic matter passes directly to spring or rivers as in the thin soil - aquifer system in Scandinavia. In this case, rivers and lakes can be highly polluted.

99

Acid cambisol soil releases high levels of pollutants (Mn, Zn) which mainly originate from soil acidification rather than atmospheric deposition. For the LRY example when the decarbonatation wave will reach the spring level, water composition will be above the law values. In similar sites, existence of preferential flow could lead to the contamination of aquifer even now or in the close hture. Concerning trace elements dynamics the most important and new point concerns B and Ba behaviour related to soil hydraulic status and apparently t o clay swelling. This original process deserves further research, particularly for Ba, which solubilisation mechanisms need to be clarified. The important features t o study trace elements dynamics are: origin of organic matter mobility, role of biological cycling, actual redox conditions in soils and aquifer and precise location of elements in the solid phase. Numerous minerals and chemical conditions found in aquifers induce important variability in the content of these elements in springs.

ACKNOWLEDGEMENTS We thank M. Bensimon for ICP-MS analysis, J.D. Dubois, Y. Mandia and P. Basabe for use of the material gathered in their PhD dissertation, and the Swiss National Research Fondation for the project financial support.

REFERENCES Wedepohl KH. Springer Verlag, Berlin-Heidelbrg-New York, 1968-1979. Drever J I (Ed). The chemistry of weathering D. Reidel publishing company (Holland), 1985. Tardy Y. MQmoires du service g6ol. d'Alsace et de Lorraine, n"31, 3 Strasbourg, 1969. 4 Mosser C. ULP Strasbourg, Institut de gbologie, MBmoire n"63, 1980, 227p. Oden S. Proceedings of the first international symposium on acid 5 precipitation and the forest ecosystem, Dochinger L.S., T.A. Seliga (Eds). USDA Forest service general technical report, 1976, 1-36. 6 Heinrichs H and R Mayer . J. of Envir. Quality, 1980; 9: 111-118. 7 Schmitt H W and Sticher H. In Metals and their componds the environment Merian E. (Ed), VCH Weinheim New York Basel Cambridge, 1990: 311-332. Heinrichs H and R Mayer. J . of Envir. Quality, 1977; 6: 402-407. 8 Turner RS. PHD, Univ. of Pennsylvania, Philadelphia PA. (Diss. Abstr. 839 160981, 1983. 10 Bergkvist B. Water, Air and Soil Poll., 1987; 33: 131-154. 11 Paniaux A, JD Dubois, Y Mandia, P Basabe and M Bensimon. Memoires of the 22nd congress of IAH,Paniaux Ed, Lausanne, 1990; : 254-262. 12 PQdroG. Revue gQogr.phys. gQol.dyn., X, 1968: 157-170. 1 2

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13 Atteia 0,JC Vedy, A Parriaux, E Dambrine. Memoires of the 22nd congress of IAH,Parriaux Ed., Lausanne, 1990. 14 Dambrine E. PhD Diss. Univ. Paris VII, 265 p, 1985. 15 Atteia 0. EPFL Lausanne, PhD Diss. no 1031,1992,253pp. 16 Parriaux A and M Bensimon. Memoires of the 22nd congress of IAH, Parriaux Ed, Lausanne, 1990: 719-727. 17 Atteia 0.Submitted to Journal of Hydrology, 1993. 18 Bensimon M, Gabus JH,Parriaux A. J. Trace and Microprobe techniques, 1991;9: 81-93. 19 Lake DL, PWW Kirk and J N Lester. J. of Envir. Quality, 1984;13: 175-183. 20 Jeanroy E. PhD Diss. UniversitR de Nancy, 1983,157p. 21 Cavallaro N and MB McBride. Soil Sci. SOC. Am.J., 1984;48: 1050-1054. 22 DeCoursey DG. Proceedings of the NATO advanced study institute on recent advances the modeling of hydrologic systems Sintra, Portugal. Luwer Academic, 1988:35-48. 23 Atteia 0.Submitted to Atmospheric Environment, 1993. 24 Atteia 0.Annales ScientXques Forestibres. (in press). 25 Schultz. Vergleichende Betrachtung des schwertemetallhaushalts verschiedener waldokosystems, Waldsterben busgenweg 2,3400 Gottingen, 1987,245~. 26 Keller C. PhD Diss. EPFL Lausanne, 1991,171p 27 Hofken KD. Effects of Accumulation of air pollutants forest ecosystems B Ulrich, J Pankrrath Ed., 1986: 57-64. 28 Godt J , Schmidt M, R Mayer. Atmospheric pollutants forest areas, Georgii H.W. Ed., Reidel, 1986:263-274. 29 PBdro G and AB Delmas. Ann. Agron. 1970;21: 483-518. 30 Bourg ACM and J-C Vedy . Geoderma, 1986;38: 279-292. 31 Tyler G.Water, Air and Soil Poll., 1981;15:353-369. 32 Bergrren D.Intern. J. Environ. Anal.Chem., 1989; 35: 1-24. 33 Buffle J and RS Altmann . Aquatic surface chemistry Stumm W. (Ed). Wiley, New York, 1987:351-383. 34 Keren R and H Tdpaz. Soil Sci. Soc. Am.J., 1984;48:555-559. Am. J., 1981;45:478-482. 35 Keren R and RG Gast. Soil Sci. SOC. 36 Duchaufour P. PBdologie, tome 2: constituants et propriBt6s des sols. Masson Ed. Paris, 1983. 37 Mahan BH. Chimie, InterEditions, Paris, S.A., 1977. 38 Bergback B, S Anderberg and U Lohm. Water, Air and Soil Poll., 1989;48: 391-407. 39 Stumm W and JJ Morgan. John Wiley, 1970. 40 Tonkonogov VD,BP Gradusov, NY Rubilina et al. Pochvovedeniye, 1987;3: 68-81. 41 Sanborn Pand LM Lavkulich. Soil Sci. Soc. Am.J. 1989;53: 517-526. 42 Farmer VG. Soil. Sci. Plant Nutr. 1982;28: 571-578. 43 Driscoll CT,N Van Breemen, J D Mulder. Soil Sci. SOC. Am. J., 1985;49: 437-444.

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44 De Vries W, M Posch and J Kamari. Water, Air and Soil Poll., 1989; 48: 349390. 45 Mandia Y PhD Diss, Lausanne EPFL, 1992,318 p. 46 Bartlett RJ and JM Kimble. J. of Envir. Quality, 1976a; 5: 379-383. 47 Bartlett RJ and B James. J. of Envir. Quality, 1979; 8: 31-34.


103

Leaching behaviour of granulated non-ferrous metal slags F.M.G. Tack, P.H. Masscheleyn & M.G. Verloo Laboratory for Analytical Chemistry and Agrochemistry, University of Gent, Coupure Links 653, B-9000 Gent, Belgium ABSTRACT The leaching behaviour of selected heavy metals from granulated non-ferrous metal slags was investigated. Experimental results and sorptiorddesorption data of selected soils were used to estimate metal migration from a hypothetical slag disposal site into the underlying soil. In a heavy textured soil, metals leaching from the slag site were predicted to accumulate mainly in the upper 10 cm. I n more sandy soils, the metals were dispersed over larger soil depths, resulting in lower accumulated levels. When comparing the modelled migration results with soil and soil water quality standards, only the migration of Zn was of practical importance and may limit reuse of the slags in bulk form. INTRODUCTION Metal slags are a by-product of non-ferrous metal processing and refining. These slags consist of an inert matrix, in which residual elements are included that can no longer be extracted in an economically feasible way. As opposed to slags resulting from iron and steel production, those from non-ferrous metallurgical industries do contain, besides high amounts of Fe and Ca, relatively high levels of several toxic elements (e.g. Cd, Cu, Pb, Zn, As, Ni, Cr) as compared to levels found in the environment. Formerly, those slags were produced as coarse blocks appearing like rocks. They were widely used to reinforce shores and as building material in dike construction. In the early eighties, new extraction technologies became available resulting in the production of slags with lower residual amounts of non-ferrous metals. Those slags, however, are granulated and other ways of reuse had to be considered. Because of their high bulk density, good and quick settling and resistance against frost, the granulated non-ferrous metal slags were found to be an excellent material for use in road construction. They were found suited to replace conventional materials in drains, foundations and capillary screens. Due to increased environmental awareness, investigations of possibilities for reuse also had to consider environmental aspects. Despite of the lower total toxic trace element content as compared to the old slags, the leachability of elements from the granulated slags is expected to be higher due to the finely granulated structure and resulting larger specific surface area. As slags may periodically or permanently be percolated by drain water or ground water, this aspect is critical when using the slags in road construction. Using leaching tests we assessed the leaching behaviour of selected elements including Cd, Zn, Pb, Cu, Cr, Ni and As from the granulated non-ferrous metal slags. The results were modelled to predict migration of metals from disposed slags

104

with time. Based on model estimations, the environmental restrictions of reusing the slags in bulk form were evaluated.

LEACHING BEHAVIOUR

Introduction To assess the leaching behaviour of the granulated non-ferrous metal slags we used the Standard Leaching Test of the Netherland Energy Research Center (SOSUV-test) [l]. This test, originally developed for investigation of combustion residues, allows to evaluate the leaching behaviour at different WS-ratios (Liquid to Solid ratios). The US-ratio is used as a relative time scale. Relating the relative time scale to a real time scale of a particular field situation allows for short, medium and long term leaching behaviour to be estimated. The velocity by which a particular US-ratio is reached depends on the porosity of the material and on the rate of advection of fresh leaching medium. A column percolation test is run to test leaching behaviour at cumulative USratios u p to 10. For higher cumulative L/S-ratios (20-loo), a batch equilibration cascade test is used. The advantage as compared to the column test is that high WSratios can be reached in shorter time spans. Distilled water acidified to pH 4 is proposed as standard leaching agent and simulates the most significant effects of acid rain. If a particular field situation is known, the actual leaching agent can be used instead. Materials and methods

Column percolation test Leaching agent was prepared by acidifying demineralized water with diluted ultrapure nitric acid to pH 4. A 30 cm long, 3.5 cm diameter polyethylene column was filled with 0.4 kg of the metal slags. At the bottom and top of the column a 0.45 pm membrane filter prevented fine particles from entering the leachate. The column was percolated from bottom to top at a rate of 10 mWhr. Fractions were collected up to a cumulative L/Sratio of 10. pH and conductivity were determined using standard equipment. Samples were acidified to pH 2 with nitric acid and kept in polyetylene bottles prior to analysis.

Butch equilibrufion test In 1 L polyethylene flasks, 40 g of the metal slags were shaken for 23 hrs with leaching agent (distilled water acidified to pH 4 with HNO,). After 15 minutes of settling, the liquid was decanted and filtered through a 0.45 pm membrane filter. The membrane filter and collected solids were added to the slags in the flask and, after addition of fresh leaching agent, the extraction was repeated until five fractions were collected. After determination of pH and conductivity, the samples were acidified to pH 2 with nitric acid and stored for analysis.

105

Amlysis Concentrations in the extracts were determined with flame atomic absorption (Varian AA-1475). Concentrations below the detection limit were measured with graphite furnace atomic absorption. Mercury was determined with flameless absorption using a Coleman Mercury Analyser. Calcium, K and Na were determined with flame emission (Eppendorf ELEX 6361) and anions with ion chromatography (Dionex 2000i/SP).

Results and discussion

Total contents Total contents and physical characteristics of the metal slags tested are given in Table 1. Zinc is, amounting to 4 %, a main constituent of the slags. Several elements (Pb, Mn, Ba, Cu, Sn and Cr) are found in the range of 0.1 to 0.5 %. Many other elements are present at the m d k g level. Mercury is not present.

Table 1 Total element contents and some physical characteristics of the metal slags Total element contents (mg/kg) Al As

B Be Cd

co Cr cu Fe Hg

17727 325 713 7 123 455 1175 2876 246320 0.001

Mn

M0 Ni Pb Sb Se Sn Ti TI

v

5840 223 89 5437 220 257 2590 760 26 108

Zn Ba Ca K Mg Na

s

41816 2920.0 74173 2775 15085 11920 11000

Physical characteristics Density Bulk density Granulometry (%) 0 - 0.2 mm 0.2 - 1 mm 1-2mm >2mm

3567 k d m 3 2200 k d m 3 5.7 36.4 46.8

11.1

As compared to levels accepted for soils (Table Z), the total concentration of several heavy metals is very high. However, total contents should not be the only criteriurn to decide on possibilities for reusing the slags. As the metals are bound in an inert matrix, leaching may be so low that, even in the long term, no negative environmental effects can result. In the latter case, beneficial reuse is preferred over

106

Table 2 Reference levels in soil, surface water and ground water used in the Netherlands [2] Soil quality (mg/kg) A B

Element Cr co Ni cu Zn As Cd Sn Ba

Hg Pb

*

50 + 2L 20 10 + L 15 + 0.6(L + H) 50 + 1.5(2L + H) 15 + 0.4(L + H) 0.4 + O.O07(L + 3H) 20 200 0.2 + O.O017(2L H) 50 L + H

+

+

250 50 100 100 500 30 5 50 400 2 150

Water quality (Ilg/L)

C

A

B

C

800 300

1 20 15 15 150 10 1.5 10 50 0.05 15

50 50 50 50 200 30 2.5 30 100 0.5 50

200 200 200 200 800 100 10 150 500 2 200

500 500 000 50 20 300 000 10 600

A background level; some levels are calculated depending on clay content L (fraction c 2 pm) and organic matter content H

B: warning level for (closer) investigation C: warning level for sanitation (investigation) confined disposal. Leachability and mobility of the contained metals should be considered in the evaluation of the environmental effects of reusing of the slags.

leaching behaviour At US-ratios of 0.1 to 1, concentrations of several elements (Table 3) are high, compared to ground water quality standards (Table 2). The levels rapidly decrease in the subsequent fractions. At the highest US-ratios, only Pb and Zn persist in the leachate. Their concentrations exceed the Dutch C-levels for surface and ground water quality, even at the highest US-ratios. Metal slags are a potentially continuous source of Zn and Pb release in the environment. Concentrations of leached Zn are an order of magnitude higher than the concentrations of other elements. This is due to its much higher total content but also to the higher solubility of Zn as compared to other metals [3]. The cumulative leaching behaviour of several elements is depicted in Figure 1. The leached quantities generally amount to 0.1 - 0.5 7% of the total content at the highest US-ratios. Up to 1 % of the total Zn is leached. Even when leached concentrations are higher than accepted concentrations in surface water or ground water, adverse environmental effects resulting from beneficial reuse can be unimportant compared to the environmental and economical drawbacks of confined disposal. Particular field situations have to be modelled in order to assess the extent of metal migration. From the results, potential adverse effects and risks can be identified. Only when those effects are considered acceptable, reuse of the slags in bulk form can be allowed.

I07

Table 3 Concentrations in leaching test fractions f o r the different cumulative WS-ratios (L/S cum.)

ws

cum.

Vol. (mL)

PH

EC S/cm

As

40 160 200 400 400 800 2000 800 800 800 800

20.2 0.9 0.4 0.2 0.1 0.1 0.0 0.1 0.0 0.0 0.0 0.0

630.0 170.0 40.0 20.0 20.0 11.5 9.5 11.7 3.3 1.2 1.4 1.7

880.0 250.0 60.0 37.0 32.5 0.8 12.5 23.1 6.8 3.8 3.5 2.7

Concentration (pg/L) MO Ni

Pb

Se

1540 450 180 120 100 30 27 51 15 9 9 8

1720 660 460 410 320 300 270 315 250 170 225 340

161 27

Concentration (mg/L) Mg Na

c1

so,

140.0 7 7.0 2.0 I .o 1.o 0.0 0.0 0.0 0.0 0.0 0.0 0.0

208.0 93.0 60.0 20.0 1.0 1.3 0.7 1.a 3.3 3.1 3.0 1.1

684.0 173.0 42.0 15.0 8.4 4.8 3.2 6.4 2.6 1.3 2.0 1.o

1200 980 90 62 62 42 37 37

800

5.42 6.09 6.15 6.52 6.22 6.22 6.51 6.44 6.40 6.17 6.47 6.45

cum.

Cr

cu

Mn

0.1 0.5 1 2 3 5 10 20 40 60 80

9 1 6 0 1 0 3 53 1 10 12 3

190.0 45.0 14.0 21.o 170.0 43.0 39.0 18.0 11.0 8.O 12.0 49.0

1010.0 220.0 80.0 70.0 65.0 101.0 58.0 12.6 7.1 11.3 9.9 10.9

cum.

Zn

Ca

K

0.1 0.5 1 2 3 5 70 20 40 60 80 100

3.7 4.5 6.3 3.9 9.2 7.1 7.0 8.5 5.9 4.2 3.6 2.7

73.0 9.0 1.o 1.o 1.o 0.0 1.o 0.0 0.0 0.0 0.0 0.0

12.0 1.0 0.0 0.0 0.0 7 .0 0.0 0.0 0.0 0.0 0.0 0.0

WS

100

US

CO

~-

~

~~

0.1 0.5 1 2 3 5 10 20 40 60 80 100

Concentration (p g/L) Be Cd

23

21 18 17

255.0 37.5 0.0 3.1 0.9 0.0 7.0 7.2 4.6 3.5 2.3 0.0

5 0 0 0 0 0 0 13 7 11 10

11

11.7 2.3 0.2 0.2 0.3 0.2 1.2 0.0 0.0 0.0 0.0 0.0

2 0 1 2 6 2 2 3 3 3

108

CUMULATIVELEACHING (%of total)

I1

/

I I

..a

I

m....m'

II

0.001 ,,,......'.

0.0001

0.1

II

Column Test I Shaking , , , ,,

,

,

, ,,

1

10

I TBSt,

, ,,

J

100

US-RATIO

Figure 1. Cumulative leaching behaviour of selected metals from the metal slags.

ESTIMATING METAL MIGRATION IN THE ENVIRONMENT Introduction Results of a leaching test can only readily be understood in terms of potential environmental effects when they are applied to field situations. Besides the observed leaching behaviour, the particular characteristics of a disposal site also determine the actual migration of metals. These characteristics include geometry and dimensions of the slag mass involved, sources of leachate (precipitation, surface water, ground water), percolate flows and rates, and element fate determining mechanisms involved (adsorption, convection, dilution, plant uptake). Once a field situation is described, the time scale of the leaching process can be related to the US-scale of the leaching test. Next, from the leaching test results, concentrations and flows of metals from the slags can be estimated in function of time. Finally, the resulting risks and potential hazards can be evaluated. When worst case assumptions are made, one can estimate the most unfavourable effects. Whether these effects are acceptable or not will determine the feasibility of reusing the slags in bulk form. It can however remain dificult, especially when migration is low, to decide whether the observed effects are acceptable or outweigh the disadvantages of having to dispose the slags in isolated sites as a waste material. When metal slags are disposed on the land, metals may migrate with run-off water and contaminate surrounding fields where they can be taken up by plants. They also may enter surface waters where they can be taken u p by aquatic

109

organisms, or sorb on suspended material and settle in sediments. When the soil permeability is sufficiently low, the percolating flow can penetrate into the underlying soil where sorption processes retard metal migration. Eventually, contamination of ground water may occur. TO provide a better understanding of the significance of the observed leaching test results, metal migration from a hypothetical slag disposal is estimated. The ecological significance of the resulting contamination is evaluated using Dutch reference levels for quality of soil, ground water and surface water.

The field situation Migration from a hypothetical metal slag disposal is calculated. The slags are accumulated 1 m high on an area of 1 ha. The yearly precipitation extends to 780 mm. The bulk density of the slags is 2200 kg/m?. The underlying soil is assumed to have a bulk density of 1600 kg/m3 and a soil moisture content of 0.2. Relating the L/S-ratio scale to a time scale Since 1 m2 of the slag disposal area or 2200 kg of slag material yearly receives 780 L of precipitation, an US-ratio of 780/2200 or 0.35 is reached after 1 year. Inversely, a US-ratio of 1 is reached after 2.82 years. In reality, evaporation o f precipitated water will occur which will delay the predicted effects. In Table 4,some US-ratios are given with the corresponding time scale (denoted as Time) for the proposed field situation. Metal concentrations in the leachate and in a receiving surface water If all of the leachate of the slag disposal was collected by means of an impermeable layer and discharged in a surface water, a percolate volume of 0.247 L per second and per ha would be produced. The evolution of the concentrations in this leachate with time are directly estimated f r o m the observed concentrations of the leaching test (Table 3). The B-level, the level above which pollution can be suspected, is 200 pg/L for Zn (Table 2). As the concentrations in the first percolate amount to 123700 pg/L (Table 3) a dilution of 618 is needed to lower the concentrations to 200 p@. If the volumetric rate of a receiving surface water is at least 0,247 x 618 = 153 L/s, the Znconcentration will not raise above the B-level. After 28 years (WS-ratio of lo), a volumetric rate of only 8.6 L/s is needed. The pollution thus would not result in observable effects, even at the long term. Resulting environmental effects would thus be negligible. However, upon generalized reuse of the slags in road construction, it would be necessary to prevent leachates from entering directly into surface waters. The presence of many of those weak contamination sources in the environment would cause a slow but continuous increase of natural background concentrations in surface waters and sediments.

110

Table 4 Calculated average penetration depth (x), metal concentration in solution (c) and adsorbed concentration (s) for different L/S ratios and three types of underlying soil (L = light sandy loam, M = light loam and H = heavy clay)

ws cum. 0.1 1 10

20 100

Time (Years)

Volume (Urn31

0.3 2.8 28 56 282

220 2200

22000 44OOo

m o o 0

Adsorbed quantities (s)

L

(mg/kg) M

H

L

Depth (x) (cm)

M

H

Cadmium

0.006

35.5 8.6 1.5 0.9 0.3

316.5 80.1 14.1 8.9 3.2

280.1 69.4 12.1 7.7 2.7

0.2 2.4 24.1 48.2 240.7

0.0 0.3 2.6 5.1 25.5

0.0 0.3 3.0 5.9 29.6

0.190 0.044 0.052 0.029 0.023

146 35 42 24 18

198 48 56 32 25

1377 404 468 276 215

0.0 0.2 1.7 3.4 16.9

0.0 0.1 1.3 2.5 12.6

0.0 0.0

0.1 1 10 20 100

1.720 0.666 0.335 0.322

1047

2464 1167 630 608 510

10593 6016 3542 3427 2927

0.0 0.2 2.0 4.0 20.0

0.0 0.1

0.0

0.1 1 10 20 100

123.7 34.3 10.7 9.2 5.5

2103 7 488 783 709 476

4843 4478 3635 3490 2894

1.4 7.6 59.9 117.8

0.8 3.2 18.7 35.8 158.5

0.1 1 10 20 100

0.630 0.151 0.026

0.017

Copper 0.1 1 10 20 100

0.2 0.3 1.4

Lead

0.267

442 229 220 184

0.7 1.5 7.2

0.0 0.1 0.3 1.3

Zn

1233 617 243 215 133

563.7

0.4 1.1 4.0 7.3

26.1

Ill

Migration in the underlying soil

Estimation of average penetration depth When the underground is permeable, a considerable amount of the leachate can penetrate. Migration of metals however will severely be retarded as compared to the migration velocity of the bulk liquid phase because of sorption processes [4,5]. When a column is percolated with a volume V (L) of a solution, containing a concentration c (m@) of an element, the average penetration depth x (mm) of the element can easily be estimated from [5, 61: x = V/(e

+ r.s/c)

(1)

where 0 is the volumetric soil moisture ratio, is the bulk density of the soil (g/cm’) and s is the adsorbed amount (mglkg), in equilibrium with the concentration c (m@) in solution. The equation may more readily be understood when rewritten as: x(0c +

rS) =

vc

The original amount of metals added (right side of the equation) is distributed over a depth x in the soil, both in the liquid (ec) and the solid phase (rs). The relation between c and s is provided by an adsorption isotherm. For our example, a Langniuir adsorption isotherm [5, 71 was used with experimentally determined constants for three soils (Table 5). s = K.c/(1

+ K.c/s,)

(2)

To apply equations (1) and (2) on the data from the leaching test, weighed average concentrations of the total volumes corresponding with each US-ratio are calculated (denoted as c in Table 4). For Zn f.i., the weighed average concentration of the total volume at a L/S-ratio of 1 is calculated as: 123740 x 0 . 1

+

34530 x 0.4 + 16320 x 0 . 5

(0.1 + 0 . 4 + 0 . 5 )

= 34346 pg/L

Those concentrations are used in equation (2) to yield the corresponding adsorbed amount. After substituting the total volume, the corresponding weighed average concentration and adsorbed concentration in equation (l), the average penetration depth after each period of time is estimated. The results for the three soils are given in Table 4.

Modelling the concentration profile A prediction of the actual shape of the concentration profile in function of depth is obtained by using a one dimensional transport model, which includes adsorption phenomena [8]. Since sorption and desorption mass transfers occur much faster than the transport processes, adsorption equilibrium can be assumed. Sorption-desorption processes in soils have been shown to occur in a few minutes [5]. At any time, the relation between the concentration in solution and the amount sorbed is given by the Langmuir isotherm (2). The transport equation is expressed as [3]:

112

Table 5 Langmuir-adsorption parameters for the light sandy loam (L), light loam (M) and heavy clay soil (H) (S, in nig/kg, K in kg/L) [7] Medium (M)

Light (L)

s,

K

Sm

K ~~

Cd cu Pb Zn

3333 2480 7665 2000

4540 3845 8290 2500

57 818 705 26

Heavy (H) Sm

K

~

6370 5000 20400 5000

540 1096 2039 107

465 10000 12811 1250

where R is termed the retardation factor given by: z

ds

@

dc

R = I+--=

1 +

z

K/@

(1 + Kc/s,)*

with: o = volumetric moisture ratio (L3/L3) c = pollutant concentration (M/M) D = coefficient of dispersion (L2/T), calculated as Do + E V D,, = molecular diffusion coefficient ( L V ) , macroscopic velocity (L/T) s = adsorbed quantity (WM) s, = maximal adsorbed quantity (M/M) z = bulk density of the solid phase (M/L') t = time

c

=

dispersivity (L) en v =

This model was solved numerically, using the finite difference technique. The same conditions were assumed as in the estimation of the average penetration depth. In addition, the following assumptions were made: the volumetric rate being 780 mndyear, the macroscopic pore water velocity is calculated to be 1.068 cm/day. With Do = 1 cm?/day and E = 5 cm [8], D becomes 6.34. The input of metals in function of time was given by the concentrations observed in the leaching test, the L/S-scale being related to the appropriate time scale as illustrated before. The concentration profiles of Cd, Cu, Pb and Zn in function of time are shown in Figures 2 and 3.

I13

SORBED CADMIUM PROFILES HEAW CLAY

LIGHT LOAM

j U / , + 90

1ooo

1

2

3

4

5

looo

, 1

,

, 2

,

, . , 3

4

,I :I,, 5

1ooo

1

. , . , . ,

2

3

4

ADSORBED AMOUNTS (mg/kg) ADSORBED AMOUNTS (mg/kg) ADSORBED AMOUNTS (mg/kg)

SORBED COPPER PROFILES LIGHT LOAM

5 0 ~ 10

20

30

HEAW CLAY

L L 5 0 ~ 20 30 2~ 30

5 0 ~ 10

ADSORBED AMOUNTS (mg/kg) ADSORBED AMOUNTS (mg/kg) ADSORBED AMOUNTS (mg/kg)

2.8 years (US= 1) 28 years (US= 10) 56 years (L/S = 20) 282 years (US= 100)

-----

------_---

Fib'ure 2. Concentration and adsorption fronts of Cd and Cu resulting from leaching of a slag disposal after different time spans in different underlying soils.

114

SORBED LEAD PROFILES

or

. HEAW CLAY r- ---_- (“A) Hermionr Bays*

1064

152

2539

341

7323

2500

140

364

228

299

20

37

160

40

430

33

101

168

Saronikos Gulf163i7 Thermaikos Gulf17 Patraikos Bayi8 Navarino

59

Kalamata Bay2()

40

86

193

352

29

66

151

355

43

56

173

Itha’ci Bay21

106

106

359

28

55

150

Argostoli Bay22

206

43

I4 18

85

20

10

984

131

0 16(Mn0)

1604(Fe2O3)

Method total

250

0 46

05NHCI

107

185

0 5 N HCI

210

06

total

414

39

233

total

418

0 36

6 09

total

383

0 19

4 38

total

005

2.90

total

132

24

46

106

205

Pagassitikos Gulf17

34

74

14

32

114

395

N Euboicos Bay23

27

58

336

28

41000

106

2 N HCI

2.33

2 N HCI

908

510

227

278

2 N HCI

Lesbos Island l 7

39

43

12

247

2 N HCI

Shallow Water24

22

Kavala Bay’’

*This study

1

92

13

56

35

60

145

0.085

6.5

total

Table 5

e

N

4

128

10 m In

a13

c

01.5

II

3

u12

ui 4

X

018 YMlN = -.27

Figure 3.

Sketch map of the area studied showing the location of the sediments analysed

RESULTS The results from the bulk chemical analysis of the Hermioni offshore sediments are showtl in Table 3 Metal Concentration Levels It is seen that significant enrichments in Hg, Ag, Pb, Cd, Cu, Zn and Fe relative t o normal nearshore sediments occur in the region studied. Mercury reaches the value of 2082 ppb, this value being much higher than that of the average shale, mean crust and mean sediment (14). The maximum Hg vatue is also greater than that reported for Navarino Bay sediments ( I 5). Silver reaches the value of 76 ppm being significantly enriched relative to the average shale, mean crust, mean sediment and the Navarino Bay sediments. The maximum value of Pb is 1604 ppm, being 73 times higher than the average concentration of Pb in shallow water sediments. The above value is greater than the maximum Pb values reported for all Greek Bays Studied (Table 4). Cadmium varies between 0.5 ppm and 14 ppm its maximum value being 64 times higher than its average value in shales. The above value is also greater than the maximum Cd values reported for all Greek Bays studied. Copper varies between 40 ppm and 7323 ppm, showing a 130-fold enrichment relative to normal shallow water sediments. Significant, but lower Cu enrichments in shallow water sediments were reported from the Vassilikos Bay, Cyprus (25). A 28-fold enrichment in Zn was found relative to normal shallow water sediments, its maximum value being 2540 ppm. This value is comparable with the maximum concentration of Zn reported for the heavily polluted Saronikos Bay sediments. Generally, cobalt is not among the metals markedly enriched in the metal polluted coastal sediments However, in the sediments

I29

508

F i g u r e 4 . V a r i a t i o n s of Fe2O3, Cu and Zn r e l a t i v e t o t h e i r average v a l u e s

I30

Figure 5. Variations of Hg, Co and P b relative to their average values

131

Figure 6. Variations of Cd and Be relative to their average values

I32

Figure 7. Variations of Mo, Ba and V relative to their average values

133

Figure 8. Variations of M n and Ag relative to their average values

134

E-

Figure 9 . Variations of Cr, Ni and P2O5 relative to their average values

135

studied a 26-fold Co enrichment is observed, this enrichment being greater than that observed in Northern Euboekos Bay metal polluted sediments (16-26) Figures 4 to 9 demonstrate the variations in the element concentration levels relative to their average values in the area studied. On the X axis of the above figures the sample coding numbers are shown in order of increasing number. On the Y axis the 0- level represents the average concentration value which appears on the right margin, while the positive and negative values represent the positive or negative deviations from the average value. The average concentration of Fe203 is 4.8% with one major geochemical anormaly at station 4 and a minor one at station 9 At these two stations similar geochemical anomalies were observed for Cu, Zn, Hg and Co, while Pb and Cd exhibit marked anomalies only at station 4 . By contrast Be shows a major anomaly at station 4 and minor anomalies at stations 9, 16 and 20 Station 4 is characterized by a depletion in Al, Mo, Ni and Ti Manganese exhibits a major anomaly at station 7 and a minor one at station IS Silver shows a distinct peak at station 5 . Cr, Ni and P show no significant variations except for a peak at station 23 Sr exhibits two major anomalies, one at station 3 and another at station 22. Geostatistical Analysis In order to investigate the geochemical behaviour ofthe elements studied and their mode of incorporation in the sediments the correlation coeffrcients among their concentrations were calculated The results are shown i n Table 5 The application of factor analysis on the geochemical data showed that Factor I accounts for 41.1% of the data variance and shows strong loadings of Fe, Zn, Pb, Cd, Hg, Cu, Co and Be. This factor represents the sulphide-bearing fraction of the sediments and demonstrates the strong input of sulphide related metals to the marine environment. Factor 2 accounts for 23. I % of the data variance, it shows strong loadings of Al, Zn, V, Ba, Mo and Ag and represents the clay fraction of the sediments. Elements like Ag which originally were associated with the sulphide minerals may be adsorbed after their oxidation on the clays. Factor 3 accounts for 15.0% of the data variance and shows strong loadings of P. Cr and Ni and may represent weathering products of magmatic rocks Factor 4 accounts for 9 4% of the data variance It shows strong loadings of Ca and Sr representing the carbonate fraction of the sediments. Factor 5 accounts for 5 9% of the data variance and shows strong loading only of Mn, representing the manganese oxide fraction of the sediments The fact that no loadings of trace metals are observed in this factor supports the different geochemical behaviour of Mn.

~

~~

I -0 0 10 -0 200 0 953 -0 287 0 19.3 0 164 0 I20 0 0-39 0 974 0 390 -0 188 0 060 -0 727 0 005 0 078 -0 0 5 I 0 967 0044 0 088 -0 174 -0 1 I 5 TI I

~~~

~

~

2 -~

~

-0.184 0 914 0.086 0 854 0 029 -0 318 -0 177 -0 552 -0 177 0 882 -0 I 9 4 -0 I93 0 236 03 5 -0 1.33 0 788 -0 22.3 -0 159

12; 0 908 0 575 13 I -0

~~~~

~~

-0005

-0 969 -0 075 0 027 -0 1 1 1 0 069 0 859 -0 143 0 746 0 084 0 001 -0 173 0 064 0 109 -0 023 0 088 -0 I 6 6 0 055 -0 006 0 048

-0 01 I 0 152 15 0

-0 I 3 6 04

0 105 -0 234 -0 071 0 248 -0 014 0 066 0 957 -0 174 -0 042 0 004 0 933 -0 0.39 0 91s -0 250 -0 0 1 5 -0 5 1 3 -0024 -0003

~

-0 006 0 088 0 227 -0 167 0 962 0 19s -0 0 10 -0 140 0 069 0 260 -0 060 0 072 0 052 0 027 0 I10 -0 080 -0 012 -0 I07 0 107 -0 052 -0 167 59

-0 038 ~~

~~

5

A

)

~~~~~

~~

Figures I0 and I I denionstrate the dilt'ei-ent geochemical behaviour of Ag and Flg in thc area s1tidied

Ile~iierrtGeographic I)istributiori 111 order t o investigate the eleriient geographic variations, the geochemical niap of each zleirient was tlrawii tising ii specilic coniputing piogi-ainrne lie, ZII,1'11, Cd, Ilg, 4

I

10 4

4.4 Metal Transfer between Inorganic and Organic Substrates

With respect to the modeling metal partitioning between dissolved and particulate phases in a natural system, e.g, for estuarine sediments, the following reqirements have been listed by Luorna and Davis [65]: 6

6

the determination of binding intensities and capacities for important sediment components the determination of relative abundance of these components; the assessment of the effect of particle coatings and of multi-component aggregation on binding capacity of each substrate, the consideration of the effect of major competitors (Ca2’, Mg”, Na’, C1-), the evaluation of kinetics of metal redistribution among sediment components.

280

It seems that thermodynamic models are still restricted because of various reasons: (i) adsorption characteristics are related not only to the system conditions (i.e., solid types, concentrations and adsorbing species), but also to changes in the net system surface properties resulting from particldparticle interactions such as coagulation; (i) influences of organic ligands in the aqueous phase can rarely be predicted as yet; (iii) effects of competition between various sorption sites, and (iv) reaction kinetics of the individual constituents cannot be evaluated in a mixture of sedimentary components. At present, experimental studies on the dissolvedsolid interactions in such complex systems seem to be more promising. One approach uses a six-chambered device, where the individual components are separated by membranes, which still permit phase interactions via solute transport of the elements [66]; in this way, exchange reactions and biological uptake can be studied for individual phases under the influence of pH, redox, ionic strength, solid and solute concentration, and other parameters. The system is made of a central chamber connected with 6 external chambers and separated by membranes of 0.45 pm pore diameter (Figure 1Sa). The volume of the central chamber is litres and each of the external chambers contains 250 ml. Either solution or suspension can be inserted into the central chamber. In each external chamber the single solid components are kept in suspension by magnetic stirring. Redox, pH and other parameters may be controlled and adjusted in each chamber. In an experimental series on the effect of salinity (i.e., disposal of anoxic dredged mud into sea water) quantities of model components were chosen in analogy to an average sediment composition: 0.5 g algal cell walls (=5%), 3 g bentonite (=30%), 0.2 g manganese oxide (=2%), 0.5 g goethite (=5%), and 5 g quartz powder (=SO%). In the central chamber, 100 g of anoxic mud from Hamburg harbour was inserted; salts were added corresponding to the composition of sea water. After 3 weeks solid samples and filtered water samples were collected from each chamber and analysed.

'"

pg g-1

In 100s Sludge: 32200 ug

J.

Treatment wirh Seawater

.1

MobiiizPdCopper: 417 pg

ibrane '

(Dissolved: 156 pg

n

Algal

Cell Walls

Figure 15: Metal Transfer Between Sedimentary Components. 15a: Schematic View of the Multichamber Device. 15b: Transfer of Copper from Anoxic Harbor Mud into Different Model Substrates after Treatment with Artificial Seawater [66]

28 1

The effect of salinity on metal remobiliation from contaminated sediments is different for the individual elements. While approximately 16% and 9% of cadmium and zinc, respectively, in the dredged mud from Hamburg harbour is released, for metals such as copper the factor salinity increase seems to be less important in the transfer both among sediment substrates and to aquatic biota. This is, however, untrue as can be demonstrated from a mass balance for the element copper in Figure 1%: It is indicated that only 1.3 % of the inventory of Cu of the sludge sample is released when treating with seawater. Only one third stays in solution, equivalent to approx. 40 pgl-', and there is no sigmficant difference to the conditions before salt addition. Two thirds of the released copper is readsorbed at different aflinities to the model substrates. Slight enrichment of copper is observed in the iron hydroxide (approx. 80 ppm) and manganese oxide (1 00 ppm), whereas the cell walls - a minor component in the model sediment - has accumulated nearly 300 ppm of copper. The dominant role of organic substrates in the binding of metals such as Cd and Cu is of particular relevance for the transfer of these elements into biological systems. It can be expected that even at relatively small percentages of organic substrates these materials are primarily involved in metabolic processes and thus may constitute the major carriers by which metals are transferred within the food chain.

4.5 Reaction Cell Experiments for Discrimination of Mobilization and Scavenging Processes of Metals at Sediment Resuspension

Resuspension of sediments from the Elbe River can significantly decrease pH values due to high acid producing potential and low neutralizing capacity. Metals can be released into the dissolved phase, but may subsequently be readsorbed or precipitated in part to solid phases. To evaluate a regional and long-term perspective experiments were undertaken at stable neutral pH-values, and both total release and scavenging rates were extrapolated from time series of measured net release values over 630 hours [67]. In laboratory simulation experiments, wet sediment material was transferred to a 180 ml reaction cell; 135 ml of artificial river water (2 mM CaCI, and 5 mh4 NaCI) was added to give a solid/liquid ratio of 1:64. The suspension was sampled at regular intervals and then centrifbged and filtered through a 0.2 pm membrane. The same volume of artificial river water used to wash the centrihgation tube was added into the reaction cell in order to keep the suspension volume constant. The experiments were carried out under exclusion of light. The example in Figure 16 shows that the initial lead release was high but then decreased continuously due to readsorption. After about 100 hours the relative equilibrium between release and readsorption was reached. Curves for copper indicate continuous release until about 350 hour, when scavenging rate became faster than the release rate. With respect to cadmium and zinc, release was found to be higher than the scavenging rate during the whole experiment. The order of total release from the sediment has been found to be Cd (5 %) > Zn (1.5 %) > Cu (1%) > Pb (0.3 YO). The percentage scavenging of released metals was in the order of Pb (85%) > Cu (53 YO)> Zn (35 %) > Cd (30 %). Dominant processes are adsorption on organic substances, adsorptiodcoprecipitation by fresh Fe-Mn oxides and precipitation of metal phosphates originating from the decomposition of organic matter. Based on experimental results and relevant literature data, a four stage interaction model has been developed for metals in anoxic sediments to describe the behaviour of heavy metals in the system subsequent to oxidation.

282

0.0

200.0

100.0

300.0

460.0

500.0

600.0

700.0

Figure 16 Lead Release from Elbe Sediments. Full Quadrangles: Net Release; Full Circles: Total Release; Open Circles: Scavenging (AdsorptiodCoprecipitation) The time at which the highest concentration (peak value) of metal release is found can be called "characteristic value" of metal release and represented by an index Omeghe in hours. When the oxidation time is longer than this value, the dominating process is scavenging or relative equilibrium; when the oxidation time is less than O m e g h , the dominating process is release of metals (Figure 17). It has been suggested that the latter situation in particular will pose long-term problems to the aquatic ecosystem.

I

.b',

Cu

i

t

! QZ*

@Cd

Time Figure 17: Schematic Diagram Showing Changes in Relative Concentrations of Metals During ResuspensiodOxidation [67]

283 5. Remediation Procedures

"Geochemical and biological engineering" emphasise the increasing efforts to use natural resources available at the disposal site for reducing the negative environmental effects of all types of waste material, in particular ofacid mine wastes [3, 68, 691. 5.1 Storage Under Permanent Anoxic Conditions

Regarding the various containment strategies is has been argued that upland containment (e.g., on heap-like deposits) could provide a more controlled management than containment in the marine environment. However, contaminants released either gradually from an imperfect impermeable barrier (also to groundwater) or catastrophically from failure of the barrier could produce substantial damage [70]. On the other hand, near-shore marine containment (e.g., in capped mound deposits, offers several advantages, particularly with respect to the protection of groundwater resources, since the underlying water is saline and inherent chemical processes are favourable for the irnmobilisation or degradation of priority pollutants. In a review of various marine disposal options, Kester et al. [70] suggested that the best strategy for disposing of contaminated sediments is to isolate them in a permanently reducing environment. Disposal in capped mound deposits above the prevailing sea-floor, disposal in subaqueous depressions, and capping deposits in depressions provide procedures for contaminated sediment [71]. In some instances it may be worthwhile to excavate a depression for the disposal site of contaminated sediment which can be capped with clean sediment. This type of waste deposition under stable anoxic conditions, where large masses of polluted materials are covered with inert sediment became known as "subsediment-deposit". The first example was planned for highly contaminated sludges from Stamford Harbour in the Central Long Island Sound following intensive discussions in the U.S. Congress 1721.

Figure 18 Solubility of Metal Sulphides and Metal Oxides (from Ehrenfeld and Bass [76]

284

Under sub-sediment conditions there is a particular low solubility of metal sulphides, compared to the respective carbonate, phosphate, and oxide compounds (Figure 18). One major prerequisite is the microbial reduction of sulphate. Thus, this process is particularly important in the marine environment, whereas in anoxic freshwaters milieu there is a tendency for enhancing metal mobility due to the formation of stable complexes with ligands from decomposing organic matter. Marine sulphidic conditions, in addition, seem to repress the formation of mono-methyl mercury, one of the most toxic substances in the aquatic environment, by a process of disproportionation into volatile dimethyl mercury and insoluble mercury sulphide [73].There are indications that degradation of highly toxic chlorinated hydrocarbons is enhanced in the sulphidic environment relative to oxic conditions [74,751. 5.2 Chemical Stabilisation by Additives

In general, solidificatiodstabilization technology is considered a last approach to the management of hazardous wastes. The aim of these techniques is a stronger fixation of contaminants to reduce the emission rate to the biosphere and to retard exchange processes. Most of the stabilization techniques aimed for the immobilization of metal-containing wastes are based on additions of cement, water glass (alkali silicate), coal fly ash, lime or gypsum [77-791. Laboratory studies on the evaluation and efficiency of stabilization processes were performed by Calmano et al. [80].As an example Figure 19 shows acid titration curves for Hamburg harbour mud without and after addition of limestone and cement/fly ash stabilizers. Best results are attained with calcium carbonate, since the pH-conditions are not changed significantly upon addition of CaCO,. Generally, maintainance of a pH of neutrality or slightly beyond favours adsorption or precipitation of soluble metals [81]. On the other hand it can be expected that both low and high pH-values will have unfavourable effects on the mobility of heavy metals.

\

-,

'.,

\

+ 10%cement

+IO%flyash

'\

+ 20% \

caicium carbonate

\ original sample 2

4

6

11

to

1

It

mmol [H+] / g solids Figure 19: Effect of Calcium Carbonate and CementLFly Ash Additives on Chemical Stabilization of Fine-Grained Sediment from Hamburg harbour (Calmano et al. [SO]) Experimental studies of the processes taking place with mixed residues fiom lignite coal incineration indicate favorable effects of incorporation of both chloride and heavy metals in newly formed minerals (Figure 20).

285

Figure 20: Schematic Representation of Mneral Formations in "Stabilisates" from Coal Residues [82] Ettringite (i.e., calcium-aluminate-monosulphate-hydratephases) in particular, can act as "storage minerals" for chloride and metal ions. The former may be incorporated at up to 4 kg CaC4 per m3of the mineral mixture. Calcium-silicate-hydrate phases may be formed in a subsequent process, and by filling hrther pore space these minerals can significantly reduce permeability of the waste body for percolating solutions. Experimental studies of the leachability of salts and trace elements from samples of "stabilisates", with a pressure-filtration method, indicate relative high rates of release for sulphate ions, but not for zinc and cadmium in the eluate [82]. An overview on various fields of environmental research and management to which mineralogical methods can be successfdly applied has recently been given by Bambauer [83].

5.3 Final Storage Quality of Municipal Solid Waste Materials

From a geochemical point of view the "reactor landfill" is characterized by labile conditions during the intial aerobic and acid anaerobic phases, the former mainly due to uncontrolled interactions with organic solutes (see above). Particular problems occur when leachate collection pipes are plugging during the acidic decomposition period. Sludge-only Iandfills have even lower permeabilities and interparticular porosities than municipal solid waste landfills.

On the other hand, solid residues with final storage quality should have properties very similar to the earth crust (natural sediments, rocks, ores, soil; Table 9). This can be achieved in several ways, e.g., by assortment or thermal, chemical and biological treatment. In most cases this standard is not attained by simple incineration of municipal waste, i.e., by reduction of organic fractions only. There is, in particular, the problem of easily soluble minerals such as sodium chloride.

286

Table 9: Comparison of Inventories of Chemical Components in the Two Landfill Alternatives and in the Earth Crust [84] Reaktor Landfill

I Final Storage

I Earth's Crust

Major solid constituents Solid "inert" waste

Silicates, oxides

Quartz, Fe-oxide, clay, carbonates

Putrefactive waste

[Gypsum, NaCI]'

(Gypsum, NaCI)

Grease trap waste

(Char)b

Keragenic compounds

Organic micropollutants

Organic micropollutants

-

Metals in reactive chemical forms

Metal-bearing minerals, mainly oxides

Metals mainly in inert forms

Protons, electrons

(Protons)

@H,acid rain)

Organic compounds

(Organic residues)

(Humic acids)

Dissolved salts

pissolved salts]'

(Dissolved salts)

Measures before incineration include the separate collection of (organic) kitchen and garden wastes (containing chlorine and sulhr), which can be transferred into compost; a major decrease of chlorine content, however, would require a significant reduction of PVC in municipal solid waste. After incineration washing of the residues can be performed either with neutral or acidified water. Future efforts should be aimed for optimizing the incineration process in a sense, that: critical components are concentrated in the filter ash and in the washing sludge, whereas the quality of the bottom ash is improved - by application of high-temperature procedures [85] in a way, that deposition is facilitated and even a reuse of this material is possible due to either the low concentrations or chemically inert bonding forms of metals [86]. Recent advances mainly concern the application of high temperature techniques for re-melting precipitator ashes and flue gas purification product. An example is the RedMelt-process, designed by Faulstich and colleagues [87], which - in a reductive milieu - afiects differentiation into three phases (Table 10): (1) metal-rich (Fe, Cu) bottom phase; (2) "condensate" phase with enrichment of zinc, lead and cadmium in the flue gas (from where metals can be recovered and transferred to a base metal smelter); and (3) relatively metal-poor silicate phase, which can easily be used for construction purposes, Table I 1 summarized available leaching data on different materials and trePtments municipal solid waste incineration products. All values relate to the pH-4 conditions as prescribed by the Swiss Guidelines in the Technical Advise on Solid Waste Materials (1991). While untreated and even washed electrostatic precipitator dusts do not meet the requirements given by the TVA-supplement, cement-solidified ashes are near the limit. Satisfylng results can only reached by using high-temperature processes, and best results are shown by the R e m e l t procedure. With regard to energy consumption, initial data suggest, that the average consumption is in the range of 1-2 kWh/kg residue, and this is nearly equivalent to the energy content of the original municipal solid waste material.

287

Element

Content (Mass-%) Input

Products Silicate

Metal

Condensate

3,4 02 0,6

64 03 03 1,4 0,3 0,03

Silicium Aluminium Calcium

22,o 9,o

26,2 6,8 10,6

Iron Copper Chromium

10,l 0,3 0,04

4s 0,03 0,04

85,O

Zinc Lead Cadmium

0,6

0,09 0,99700 t

FECAL COLIFORMS [CFU/l OOml]

v,

2 W 0,99600 0

2oool t

0

1 {

0,99900

-

0,99800mE

+

SAPROPHYTE COUNTS [CFUlmI]

DENSITY [g/cm3]

s

120rn(B=12rn)DEPTH]

-

9

0,99700 $

1000

v,

- 0,99600 0

a

0,99500

B C

A

z

x

c_l

0,99500

A

B C

D

SAMPLING STATIONS

Figure 6 . Water density, saprophyte counts and fecal coliforms at the sampling sites in the gulf of Agno (August 19, 1991).

333

In Figure 7 we have presented in diagrammatic form the situation in summer, when the density of the surface water of the gulf is lower than that of the affluent. In this situation the contamination with allochthonous bacteria is located at a depth of 7 m, where the density of the lake water is the same as that of the water of the affluent with its high bacterial content. In this case the gradient of fecal colifoms measured in the water coumn between 0 and 20 meters varies from 250 CFU/lOOml at 0.5 m, with a maximum of 2400 CFU/100ml at 7 m, and decreasing again at 20 m (50 CFu/100mI).

Figure 7 . Density and distribution of fecal colifoms carried into the lake by the contaminated affluent. Diagrammatic presentation of a summer situation.

334

CONCLUSION

By means of determination of the presence of bacteria on the water column of a meromictic lake, with a permanent stratification, we were able to ascertain how the maximum bacterial concentrations are connected to the trophogenous zones. In fact it was possible to observe an abundance of heterotrophic bacteria at the level of the supplementary primary production due to bacterial photosynthesis. In the chemocline, the metabolites released from the trophogenous bacterial zone (Chromatium okenii) at a depth of 12 m and the sedimenting biomass allow a proliferation of heterotrophic bacteria, which we already located at 11 m. The original supply of oxygen in this layer adjacent to the anoxic zone allows a marked bacterial catabolism to take place, whch is demonstrated by the saprophyte counts, by the increased turbidity and, as a consequence of the bacterial metabolism, by the subsequent disappearance of the oxygen. The aerobic mineralisation activity is thus situated 1 m above the layer of water with a high hydrogen sulphide content. In the second lake ecosystem investigated it was possible to determine the vertical distribution of allochthonous bacteria of faecal origin canied into the lake by a contaminated affluent. These germs are distributed in the water column following the flow of a passive transport due to the stratification of the water of the affluent, which reaches the zone of the lake where the water is of equal density. These results demonstrate that the gradient formed by the amount of nutrient elements and organic material present is not the only factor which governs the distribution of bacteria in a body of water. Consequently, in the study of the impact of biological contaminants of faecal origin carried into lakes it seems essential to take into account the differences between the density of the water of the affluent and that of the body of water into whch it flows. The distribution of bacteria of faecal origin and of pathogens shows seasonal variations which are reflected in the public-health and sanitary assessments of the water destined for bathing or dmking, which are the most important uses for which water is treated and protected.

335

REFERENCES [ l ] Rheinheimer G. In:Rheinheimer G, eds. Wiley & Sons: New York, 1991; 18524 1 [2] Fenchel TM, Jorgensen BB. In: Alexander M,eds. Plenum Press. New York, 1977; VOI 1; 1-41. [3] Pointdexter JS. In: Fletcher M, Gray TRG and Jones JG, eds Cambridge University Press, 1987;283-3 17. [4] Cole JJ., Findlay S, Pace M.L. Mar Ecol Prog Ser 1988; 43: 1-10. [5] Peduzzi R, Demarta A, Tonolla M. Aquatic Sciences 1992; 54: 3/4: 33 1-337. [6] Kunetsov SI. In: Droop MR, Jannasch H W , eds. Academic Press. London, 1977; 1-45, [7] Leclerc H, Mossel DAA, In: Leclerc H, Mossel DAA,eds. Doin, Paris, 1989; 273-278. [8] Demarta A, Ph Thesis No 2290, University of Geneva, 1989. [9] Martin G, In: Martin G, eds. Lavoisier: Paris, 1982; 279-292. [lo] Geldreich E. In: Mitchell R,eds. Wiley & Sons. New York, 1972; 235-235. [ 111 Peduzzi R,Tonolla M, Demarta A, Del Don C, Hanselmann K, Bachofen R. In: Vemet J-P, eds. Troisitme Conference Internationale des Limnologue &Expression Franqaise. Morges, 199 1; 183-189. [ 121 Tonolla M, Diplom Work University of Zurich. [ 131 Fried1 C, unpublished data. [I41 Buhrer H, Ambuhl H. Schweiz. Z. Hydrol. 1975; 37: 347-369.


331

THE ROLE OF THE BACTERIAL COMMUNITY IN THE RADIONUCLIDE TRANSFERS IN FRESHWATER ECOSYSTEMS.

F.Hambuckers-Berhin, A.Hambuckers, J.Remacle

Microbial Ecology, Department of Botany B22, University ot Liege, B 4000 LIEGE, BELGIUM.

INTRODUCTION

During the last forty years extensive industrial use of radioactive and fissile materials has taken place in Europe and elsewhere. Large scale nuclear facilities, for civil and defence purposes, were developed some of which giving rise to significant radioactive contamination of the biosphere. Therefore several questions are to be asked concerning human health hazards following releases of radioactive substances in the environment either continuously or accidentally. Four nuclear plants are settled in two sites along the river Meuse, Chooz (1) and Tihange (3). Moreover two new plants are being build in the site of Chooz (France). The river Meuse flows through France, Belgium and Netherlands and play a great role for the human welfare of the populations living in these countries. The river Meuse supplies five millions people with drinking water. Eight hundred km2 (agricultural lands, meadows) are flooded with water pumped from the river. It must also be mentioned that seventy tons of fishes were caught in 1982 in the river and its tributaries. Our laboratory is involved in a multidisciplinary research programme dealing with the study of the impact of nuclear power plants on the aquatic ecosystem more especially on rivers [l-2-31. Two neutron activation products contained in the waste effluents of the above nuclear facilities are mainly studied 6oCo and 134Cs. For example, the nuclear plant, Tihange 2, released lo00 MBq of 334Cs in 1982 [3]. It can be supposed that the transfer and the immobilization of these radionuclides could occur along the aquatic food webs since: (i) Co ion is a trace metal required for normal growth of all living organisms, (ii) Cs competes with K and ammonium ions [4] which are important metabolites. Moreover they could effectively radiocontaminate the living components of the food web owing to their periods (60Co: 5.2 years; 134Cs: 2.05 years). The final aim of the programme is to elaborate a general model of radionuclide tranfers through the food webs. Unfortunately the role of bacteria was neglected up to the recent years which impeded the finalization of the model. It is indeed obvious that bacteria could take a significant part in the radionuclide cycle at three levels. Firstly, they are vectors of radionuclide transport because they are able to take up, store or immobilize radionuclides generally in organic molecules or to precipitate them as insoluble salts e.g. carbonates,

338

sulphides as observed for Cd and Zn ions 15-61, Besides under some conditions bacteria release the radionuclides. Secondly bacteria can be an important step in further intake of radionuclide by the other living components of the food web. Thirdly bacteria are responsible, through the biodegradation, of the release of radionuclides sequestered in the organic matter.

w I I

I

I

T

Ll -

I

BOM

A T E R

CRAZING 2 /' I

CATABOLISM

3

I

Figure 1. Conceptual scheme of the radionuclide fluxes by considering the interactions between bacteria and the organic matter. POM: Particulate Organic Matter ROM: Refractory Organic Matter UDOM: Utilizable Dissolved Organic Matter BOM: Bacterial Organic Matter

Radionuclides Organic matter transfers Radionuclide transfers

The radionuclide fluxes (Figure 1) can be conceived by considering the interactions between bacteria and the organic matter described by Billen and Fontigny 171. These fluxes occur in the water column and in the interstitial water of the sediments. The particulate organic matter or polymeric organic matter (POM) is hydrolyzed by bacterial

339

ectoenzymes and transformed in utilizable dissolved organic matter (UDOM) i. e. small molecules such as amino acids, organic acids, mono- and oligosaccharides which can be easily taken up by hcterotrophic bacteria. A part of POM, not or slowly metabolizable by bacteria, constitutes the refractory organic matter (ROM). A part of this durable organic matter arises from cell envelopes and walls of degraded bacteria. These components can remain in individual form or can be adsorbed on clays (specially smectite or kaolinite). In this case, the radionuclide-binding capacity of ROM would he reduced because the sites normally available to radionuclides-binding would be masked or chemically neutralized as shown for metal-binding capacity of walls- and cell envelopes-clays composites 18-91. During organic matter hydrolysis, radionuclides sequestered in POM are either transfered to the water phase, or complexed with UDOM. On the other hand, bacterial yields were evaluated in the river Meuse [lo] and averaged 30%. It means that 70% of consumed substrate are catabolyzed while the remaining 30% are transformed into bacterial biomass (BOM). It results that a part of sequestered radionuclides is soluhilized through bacterial activities. Finally bacterial biomass can be destroyed by cell lysis or grazed by aquatic predators. It leads to organic matter recycling and transfers of immobilized radionuclides either to the water column or to POM or to higher trophic levels of the food web. The density of the bacterial community is difficult to assess in river ecosystems. However it can be assumed that 30% of the primary production is consumed by planctonic bacteria [ 1 I]. In steady state conditions the bacterial concentration remains relatively unchanged due to grazing which is of the same order of magnitude as bacterial growth rate 0.1 h-l 110-121. Moreover, the bacterial density is controlled by environmental conditions. So, a pulse of UDOM provokes bacterial blooms [13], and the initial bacterial level will be reached again through predation. All the above processes exemplify the significant bacterial contribution to the radionuclide cycling. The first part of this paper deals with the comparison of the aerobic bacterial communities growing in the water column and the sediments in order to evaluate the homogeneity of bacterial colonization of the river and to know whether it is allowed to extrapolate the radionuclide flux kinetics to the two main compartments of the river. The second part is devoted to the study of 6oCo and 134Cs fluxes between bacteria and water in the river Meuse.

COMPARISON OF THE AEROBIC BACTERIAL COMMUNITIES COLONIZING THE WATER COLUMN AND THE SEDIMENTS Material and methods Water and sediments were collected at Hastiere along the river Meuse. The samples were aseptically taken, stored in sterile bottles until analysis. The bacterial strains were collected from the water column following the scheme depicted in Figure 2 .

340

100 ml Meuse water (5 aliquots)

100 ml Meuse water (5 aliquots)

/

\

filtration (0.22pm pore size)

/

\

filter on PCA

bacterial film dispersed in 15 ml

10 fold'dilution solution spread on PCA (3 replicates)

I

100 pl of dilution spread on PCA (3 replicates)

Figure 2. Analysis of the bacterial communities of the river Meuse. Isolation of bacterial strains from the water column (PCA = Plate Count Agar).

The bacterial strains were harvested from the sediments following the scheme of figure 3. The absence of contamination was checked by following the same schemes with sterile Meuse water and sterile distilled water. The Petri dishes were incubated at 30°C during maximum 5 days. The strains were then isolated and purified. Fifty strains were selected and characterized by their morphology, gram-stain and biochemical profile ( API strip, Bio-Merieux, Oxy-Ferm tube, Roche). The characters were coded 1 for positive or present, 0 for negative or absent and 1 for missing test. Strain similarity was estimated by the simple matching coefficient of Sokal and Mitchener [14]. Cluster analysis was carried out by using the average linkage method (procedure CLUSTER, UPGMA [15]). The strains showing the highest similarity coefficient are proximally gathered in a hierarchical structure of more and more large groups. The results of the cluster analysis are graphically represented by dendrograms using the TREE procedure [ 151.

+

34 1

sediments (250 g)in suspension in 400 ml of sterile Meuse water (5 aliquots) shaking 48 h, room temperature

/ / 100 PI of siispension

\

\

decantation

I

spread on PCA (3 replicates)

filtration of supernatant (20 ml, 0.22 pm pore size)

I

( 3 replicates)

/

100 pI on PCA

bacterial film dispersed in 15 ml of sterile Meuse water

I

10 fold dilution

(3 replicates) 100 pl bn PCA (3 replicates)

Figure 3. Analysis of the bacterial communities of the river Meuse. Isolation of the bacterial strains from the sediments (PCA = Plate Count Agar).

Results The dendrogram (Figure 4) is composed of all the strains (50) which are gathered by hierarchical level. The minimal distance between the strains is 0 which means full similarity between the strains, on the ground on the chosen tests. So, the dendrogram gives an evaluation of the similarity between the strains and each strain is characterized by its level of aggregation I 161. All the 50 strains were recovered in two main clusters. Cluster 1 which joined 70% of all the strains, is composed of 63 % of water column strains and 37 % of sediment strains. Cluster 2 contains the remaining 30 % with 1 strain from the water column and 14 strains from the sediments. The strains clustered in this group were different from the strains of cluster 1 by the ability to use carbohydrates such as saccharose, mannitol, sorbitol and rhamnose (Table 1). The cluster analysis shows that the aerobic bacterial community isolated from the sediment was constituted of two groups : one group presenting the same biochemical features as the

342

aerobic bacterial community isolated from the water column, and an other group composed of strains only present in the sediments.

Average d i s t a n c e between c u s t e r s 0.0 w U Y U

S

2

18 6 15 8

W

7

S

w

11

5

I

Q) 4-J

5

5

w

7

UY 3

W W

20 21

Y

2

-

w

8

0

Y W

3 4

w

5

w

la

S

21 23 6

S

S S U Y S

w S 9 S Y S S

c\l L

0

+ m 3 -

0

1.2

1.4

14 19 3 13 19

ao 23 22

15 17 26 24 16

S S

7 14

S S

25

S S

22 9

S

13

S

27 18

S

1 .o

12 1

S S

3

0.8

11

x Y

0.6

16

S

L

0.4

1 10 9 17

w

w

0.2

10

Figure 4. Dendrogram of the cluster analysis of 50 strains isolated from the water column (w) and the sediments (s). I t appears that a great proportion of the bacteria colonizing the sediment and the bacterial community of the water column share similar biochemical features. It could be due to the fact that a part of the bacteria of the water column settle down and colonize the interstitial water and the sediments.

343

'Table 1 Main characteritics of strains of clusters 1 and 2 (see Figure 4) CLUSTER 1 number of strains total from water from sediment..

CLUSTER 2

35 22

15

13

14

1

% of positive responses in each cluster

Gram + Gram-

34 66

20 80

oxidase catalase arginine dihydrolase indole ONPG~@)

63 86 29 0 14

20 80 53 20 100

degradation of gelatin

60

13

acetate malate citrate

0 0

0 0 60

glucose saccharose arabinose mannitol sorbitol melibiose rhamnose

43

utilisation of

?dl

51

0 6 0 0 6 0

ONPG, o-Nitrophenyl---galactopyranoside

100 67 93 87 80 80 87

344

Consequently the role of bacteria in the radionuclide transfers in aquatic ecosystem will be investigated by analyzing the aerobic bacterial community isolated from the sediments. It will be considered in a first step that the responses could be representative of the aerobic bacterial communities of the river.

KINETICS OF 6oC0 AND 134CS TRANSFERS MEDIATED BY BACTERIA. It was decided to study the radionuclide transfers in a bacterial community instead of in a monospecific bacterial culture in order to obtain more reliable results from an ecosystemic point of view. The inoculum of the bacterial community was isolated from sediments sampled in the river Meuse at Hastikre (Belgium). The sediment! were stored in sterile bottles at 4'C until the experiments. A new preculture was carried out before each experiment in order to obtain a bacterial community as close as possible to the natural community. It is well known indeed that several strains cannot develop in laboratory conditions so that subcultured bacterial communities lose a lot of strains with uncontroled changes of the specific composition.

a. Radianuclide immobilisation by the bacte rial community, Material and methods

Preculture and cultures conditions. Precultures are carried out by inoculating sterile Meuse water with sediments. The Meuse water is supplemented with starch (2 g.1-l) and bacto-peptone (10 g.1-l) in order to improve the bacterial growth. The preculture vessels are shaked and incubated at 20'C until the mid exponential phase of bacterial growth i.e. during 48 h. At this stage, the bacterial community of the sediment is harvested by centrifugation (10 min, 2,OOOxg). The pellet is discarded and the supernatant is again centrifugated 15 min at 15,200xg. The last pellet is used to inoculate the experimental cultures which are carried out in the same Meuse water medium as preculture. The initial O.D. of cultures are f0.06. The cultures are shaked 24h at 20'C. The bulk growth rate of the bacterial community rankes between 0.13 and 0.15 h'l which is of the same order of magnitude as mentioned in earlier observations [lo].

Radionuclide uptake by bacteria. The culture medium is contaminated with 6oCo and 134Cs (up to 2,000 Bq.1-l) before the inoculation. At regular time intervals, an aliquot of the culture is centrifugated

345

(3 min, 12,500xg). The dry weight of the bacterial biomass is determined and after ashing in nitric acid, the radioactivity level is detected by gamma spectrometry.

Results

The uptake of 6oCo and 334Cs by the bacterial communities collected from the Meuse river was investigated in presence of increasing radiocontamination (up to 2000 Bq ml-l) in a Meuse water medium. When the data were expressed in the form of a double-reciprocal Lineweaver-Burk plot, two distinct phases were observed for 6oCo. One phase characterized the low contamination levels betwen 24 and 90 Bq.ml-' whereas another phase occured in the high contamination levels between 90 and 2,000 Bq.rn1-l (Figure 5). Each phase followed the Michaelis-Menten kinetics: 1/V = (l/Vmax) + (Km/Vmax). (l/[S]) where, V was the rate of uptake (Bq.g-I.hl), [S] was the activity of the medium (Bq.ml-'), Vmax was the maximal rate and K, was the Michaelis constant. Tdbk 2 (Bq.g-I.h-I) values obtained from Lineweaver-Burk plots of the Km (Bq,ml-l) and V uptake of 6oCo and ly%s.

134cs

Radiocontamination range of w ter (Bq.ml-B)

Vmax

24 - 90 90 - 2,000

2,500 18,692

22 - 295

1,286

Km

63.70 1,145.79 240

Vmax and Km values were calculated from the Lineweaver-Burk plot (Table 2). Km value was significantly greater in the high contamination levels. The two different phases of 6oCo uptake were therefore assumed to reflect high and low affinity uptake systems. The level of radiocontamination respectively explained 58% and 98% of the variation of the uptake rates. The 134Cs uptake kinetic only showed one phase (Figure 6). 73% of the variation of the contamination rate were explained by the level of radiocontamination of the water column. The Km values of Table 2 show that the highest affinity transport system is observed for %3in the low radiocontamination levels. During the radiocontamination,

346

33% and 24% of the variations of the concentration factor could be explained by the contact time between the biomass and the radionuclides for 6oCo (from 22 to 2,000 Bq 60Co.m1-1) and 134Cs (from 16 to 300 Bq 134Cs.ml-1) respectively. At the end of the uptake experiment (24 h; bacterial biomass: 500 mg dw.l-I), the activity remaining in the water column ranked between 25% and 40% of the initial activity in the case of 6oCo (initial activity: 0 to 2000 Bq.ml-I) and between 45 and 95% in the case of 134Cs (initial activity: 0 to 300 Bq.ml-I). For the same radiocontamination levels e.g.fl50 Bq.ml-l, the activity level of the bacterial community was +2.5 times higher for 6oCo (+50,000 Bq.g-' d.w.) than for 334Cs(+20 OOO Bq.g-l d.w.). Thus, an important part of &Co and 134Cs can be stored by the aerobic bacterial community which constitutes therefore a pool of radionuclides in the river ecosystem.

Figure 5. Lineweaver-Burk plot of 6oCo uptake by the bacterial community. [S] : radiocontamination eve s of the medium (Bq.mi-l) V : rate of uptake (Bq.g' .h- ) : maximal uptake rate (fjq.g-l.h-l) ::?affinity constant (Bq.ml- ) Equations: (1) low radiocon 'nation lev sy = 0.4 loy+ 23.8 lo9 x (r = 0.7595) (2) high radioconta ination leve : y = 53.5 61.3 10-?fx (r = 0.9924) (r : correlation coefficient)

\ I

347

l/CSI .lo-3

Figure 6. Lineweaver-Burk plot of 134Cs uptake by the bacterial community. IS] : radiocontamination eve s of the medium-based Meuse water (Bq.ml-l) V : rate of uptake (Bq.g' .h ) Equation: y = 0.8 + 186.7 x (r = 0.8570) (r : correlation coefficient)

I -\

b. Radionuclide release by the bacterial community.

Material and methods The bacterial community is contaminated with 6oCo and 134Cs during the exponential growth phase. Bacteria are harversted by centrifugation (3 min, 12,5OOxg), washed and suspended in a dialysis tube with sterile Meuse water (Spectrapor 5 ; thickness 0.09 mm; molecular sieve 14,000 D). The dialysis tubes are incubated in running sterile Meuse water at different temperatures (13'C, 20'C) and pH (6.5;7.0;7.5;8.0;8.5;9.0). At time intervals, the radioactivity levels detected either in the bacterial biomass or in the Meuse water outside the dialyse tubes.

Results The decontamination of the bacterial community loaded with 6oCo and 134Cs was investigated in relation to two environmental parameters, the temperature and the pH. The

348

activity levels of bacterial biomass at time zero ranked between 200 Bq.g-' d.w.and 900 MBq.g-l d.w. for 6oCo and between 1 and 50 P4Bq.g-l d.w. for 134Cs. When the temperature of the water column wdS maintained at 20°C, the kinetic of the decontamination of the bacterial community was described by a double negative exponential model. It was fitted using the procedure NLIN of SAS 1151 following the DUD method of Raltson and Jennrich [17]: y = m e-ax + n e-bx, where y : 6oCo or 134Cs levels in bacteria (Bq.g-I d.w.) x : time (h) m,n : ordinate values for each exponential (Bq.g-l .d.w.) a : slope of the line with equation In y = In m + ax b : slope of the line with equation In y = In n + bx a and b parameters depended on the desorption rate and were considered as an estimation of the biological half-lives (Tbl and Tb2) calculated as follow: Tbl = In 2 I a Tb2 = In 21 b Examples of adjusted curves were given for 6oCo (Figure 7) and 134Cs (Figure 8).

Figure 7. Desorption of 6oCo by the aerobic bacterial community. Observed values: square; Predicted values: line; y = 69 e-350.347x + 165 e-o.046x (r = 0.9121) y : 6oCo levels in bacteria (Bq.gl d.w.) x : time (h)

349

A

1500

0,

\

D

m

x

1000

+.> .4

u

500

0

0

200

100

300

400

500

Time (h)

Figure 8. Desorption of 134Cs by the aerobic bacterial community. Observed values: square; Predicted values: line; y = 916 e-7.82x 879 e-o.018x r = 0.9504) y : 134Cs levels in bacteria (Bq.g' d.w.) x : time (h).

+

I

For both radionuclides, the biological half-lives Tbl were found to be extremely short, of the order of a few seconds or minutes whereas the biological half-lives Tb2 were longer (1Sh to 461h for 6oCo and 39h to 8,976h for 134Cs). Except for a high value of Tb2 observed with 134Cs (374 d) and a very small value observed with 6oCo (0.6 d), both corresponding with the lowest radiocontamination level of the medium during the radiocontamination of the bacteria communities, the values of biological half-lives were of the same magnitude as observed for a cyanobacterium Scenedesmus sp . [ 18- 191. When the temperature of the water column was maintained at 13OC (average temperature of the Meuse river), the results showed that radionuclides fixed by bacteria were not released. At 20°C, the decontamination of the bacterial community was followed as the pH increased from 6.5 to 9. The chosen criteria was the increase of radioactivity in the water column which was a consequence of the decontamination of bacterial biomass. In any cases, the data were fitted by the following mathematical relation: y = m(l-e-ax) + n(1-ebx), where y : 6oCo or the 134Cs levels in the water column (Bq.ml-l) x : time (h) m,n : ordinate values (Bq.g-l .d.w.) a : slope of the line with equation In y = In m + ax

350

b : slope ofthe line with equation In y = In n + bx This mathematical adjustment showed that the evolution of radiocontamination level of the water column was adequately described by a biphasique process with a first phase characterized by a rapid desorption of radioactivity from the bacterial biomass into the water column and a second phase during which the radioactivity level of the water column increased slowly. The parameters a and b depended on the contamination rate of the water column and allow to compare the biological half-lives of the bacterial biomass, Tbl and Th2 respectively. They were calculated as follow: Tbl = I n 2 l a Tb2 = In 21 b. In the case of 60Co,the longest Tbl and Tb2 were found at pH 8, the average pH of the Meuse river (Figures 9 and 10). For the other pH values of the water column, Tbl and Tb2 were shorter which means a more rapid desorption of the 6oCo by the bacterial biomass and consequently an increasin radioactivity level in the water column. The same observations were made in the case of f34Cs. Given that the mean temperature in the Meuse river is 13'C and pH averages 8 , bacteria are likely to act as radionuclide sinks for much of the time but, when the temperature increases in summer they could act as sources.

6.5

7.0

7.5

8.0

t

8.5

a v e r a g e Meuse value

9.0 PH

Figure 9. Biological half-lives Tbl (h) of the 6oCo desorption of the bacterial community vs the water column pH.

35 I

Figure 10. Biological half-lives Tb2 (h) of the "Co desorption of the bacterial community vs the water column pH values.

CONCLUSlON AND SUMMARY.

The radionuclide fluxes between the bacteria and the water in aquatic ecosystem were studied by examining the bulk transfers mediated by a bacterial community isolated From the river sediments. The comparison of the aerobic bacterial communities colonizing the sediments and the water column shows that the bacterial community of the sediments is composed of two sub-communities. The fist one is similar to the water column community by its biochemical feahires; the other one displays quite different characteristics and appears to be more representative of the sediments.

352

An important part of 6oCo and 134Cs can be immobilized by the bacterial biomass which therefore constitutes a pool of radionuclides, their transfers to the water column being controlled by tern erature and pH. The uptake of 68Co and 134Cs by bacteria can be described by the Michaelis Menten model. The uptake kinetics depend on the type of radionuclide and the level of radiocontarnination in the water column. The highest affinity uptake system is observed for 6oCo at low radiocontarnination levels. The decontamination of bacterial biomass develops in two phases. The first phase is characterized by a very short biological half-life, a few seconds or minutes while the second phase is longer, the biological half-lives reach between 15 h to 461 h for 6oCo and between 39 h to 8,976 h for 134Cs. It could be inferred from the study of the influence of temperature and pH that in the river Meuse where temperature averages 13’C and H is around 8, the rates of radionuclide releases are very low so that 6oCo and lY4Cs are effectively trapped by bacterial biomass. When the conditions change i.e. the temperature raises, the radionuclides immobilized by the bacterial biomass are released and radioactivity consequently increases in the water column. It appears thus that the environmental parameters play an important role in the radionuclide transfers mediated by bacteria and are to be more deeply investigated in the further.

REFERENCES

De Clercq-Versele H, Kirchmann R. L’impact des rejets de la centrale nuclkaire de Tihange (Belgique) sur l’kcosysteme Meuse: cinq annCes d’ktude in situ et d’approche expdrimentale (1976-1980). BLG 555, 1982. KirchmaM R. L’irnpact des rejets de la central nuclCaire de Tihange (Belgique) sur I’Ccosystkme Meuse: dtudes in situ et recherches expkrimentales durant la pCriode 1981-1984. BLG 573, 1985. Kirchrnann R, Vandecasteele CM, Foulquier L, Lambinon J., Sombre L. La radiokcologie des grands fleuves: des donnCes de sites et de l’expkrirnentation B la moddlisation (application la Meuse et au RhOne). BLG 635, 1992. Crerners A. Modelling the transport of radionuclides through the freshwater environment. CCE Project Bi7-008, Final Report, 1992. Rernacle J, Vercheval C. Can J Microbiol 1991: 37: 875-877. Remacle J, Muguruza I, Fransolet M. Wat Res 1992: 26: 923-926. Billen G, Servais P. in Bianchi M, Marty D, Bemand JC, Caumette P, Gauthier M eds Micro-organismes dans les Ccosystkmes ocdaniques. Paris: Masson, 1989; 2 19-245. Walker SG, Flemming CA, Ferris FG, Beveridge TJ, Bailey GW. Appl Environ Microb 1989: 55: 2976-2984.

353

9 10 11 12

13 14 15 16

17 18 19

Flemming CA, Ferris FG, Bevendge TJ, Bailey GW. Appl Environ Microb 1990: 56: 3191-3203, Servais P, Billen G, Hascoet MC. Wat Res 1987: 21 : 445-450. Azarn F, Fenchel T, Field JG, Gray JS, Meyer-Reil LA, Thingstad F. Mar Ecol Progr Ser 1983: 10: 257-263. Servais P. Etude de la degradation de la matikre organique par les bacteries hiterotrophes en rivkre. DCveloppement d’une demarche methodologique et application a la Meuse belge. Thkse de Doctorat. Universitk Libre de Bruxelles, 1986. Daurnas R, Bianchi M. Arch Hydrobiol Beih Ergebn Limnol 1984: 19: 289-294. Sokal RR, Michener CD. Univ Kansas Sci Bull 1958: 38: 1409-1458. SAS. User’s Guide: Statistics. Cary NC: Sas Institute Inc., 1985. Le Minor L, Viron M. Bacteriologie medicale. Paris: Flarnarion MedecineSciences, 1982. Ralston ML, Jennrich RI. Technometrics 1979: 1: 7-14 Nucho R, Baudin JP. Sci Eau 1986: 5: 361-376. Sombre’ L. Contribution h 1’Ctude du transfert du radiocesium (Cs- 134 et Cs- 137) dans une chaine alimentaire d’eau douce simplifiee: eau - algue verte (Scenedesmus obtiquw) - mollusque filtreur (Dreissena polymorpha) - Poisson (Barbus harhus). These de Doctorat. Universitk de Provence (Aix-Marseille I), 1987.


355

Effects of plants on the accumulation of Zn, Pb, Cu and Cd in sediments of the Tagus estuary salt marshes, Portugal. Isabel Caqador", Carlos Valeb and Fernando Catarino" Departamento de Biologia Vegetal, Universidade de Lisboa; Campo Grande, Bloco C2, 1700 Lisboa, Portugal

a

Instituto Nacional de InvestigaqSo das Pescas, Av. Brasilia 1900 Lisboa

ABSTRACT Vertical distribution of Zn, Pb, Cu and Cd concentrations, organic matter content, pH and Eh in sediment between roots and in intertidal non-vegetated sediments from two sites on the Tagus salt marshes are compared. A sub-surface metal enrichment was observed a t depths where roots are abundant. Increase of the non-available metals to plants in the salt marsh sediments indicates different sink capacities for metals in the Tagus salt marshes: Pb >> Zn > Cu > Cd.

INTRODUCTION Several studies have been carried out in wetlands suggesting these areas may act as natural sinks for metals (1,2). The recognition of the importance of this action led recently to the construction of artificial wetlands in order t o concentrate metals from industrial waste water and mine seepage (3). In fact, many salt marshes lie close to heavily industrialized areas in estuaries, facing important discharges of metals and other toxic wastes (4).Transported by the tidal currents, both in dissolved and particulate forms, metals are eventually incorporated on the surface sediment in salt marshes (5). Vegetation may act as sediment traps (6), playing thus an important role in the settling of suspended estuarine material arid their associated mctals. Available metals may be uptaken through the roots and subsequently incorporated in the plant tissues (7). Litter plant material enriched in heavy metals may return to the sediments. Incorporated in several organic and inorganic forms in sediments, metals may not be readily available for uptake by plant roots (8). The complexing reactions involving natural organic matter in the sediments may play a key role in controlling the metal bioavailability (9). Beyond organic matter input, emergent vegetation appears to have an important role in the metal cycling as root-induced changes in the rhizosphere may modify substantially potential redox, pH and root-associated microorganisms (7). For example, reddish-brown deposits on plant roots (10) and rhizoconcretion formation around the roots (11)on several salt mnrshtls and have t w n att,iibuted to the oxidative capacity of the roots. In this paper we report the concentrations of Zn, Pb, Cu and Cd in intertidal sediments and in salt marsh sediments of the Tagus estuary and assess the role of vascular plants on the metal enrichment in the sediment sub-surface.

356

THE TAGUS ESTUARY The Tagus estuary is one of the largest estuary on the Atlantic coast of Europe, covering an area of 300 km2a t low tide and 340 km2a t extreme high tide. The estuary is composed of a deep, straight and narrow inlet channel and a broad, shallow inner bay (Figure 1).The southern and eastern parts of the bay contain extensive inter-tidal mud-flat areas. A striking feature of these areas is the presence of halophytic vegetation. The Tagus river is the principal source of freshwater to the estuary, the average annual discharge being 400 m3.s1.The tides are semi-diurnal, with amplitudes that range from 1m at neap tide to 4 m a t spring tide and its effect reaches 80 km landward of Lisbon. Contrary to many european rivers crossing industrial regions, most pollutants are discharged directly into the estuary. The Tagus estuary receives the inflow of effluents from about 2.5 million inhabitants living in the Great Lisbon area, together with the effluents from its industries (chemicals, steelmaking and shipbuilding).

I

. 5 Krn

Figure 1. The Tagus estuary; location of the sampling sites.

357

MATERIALS AND METHODS Sediments were sampled in May 1991 from two sites on the Tagus estuary salt marsh with different degrees of metal contamination (Figure 1):Corroios (site 1)and Rosario (site 2). At both sites, three 60-cm long sediment cores were collected from intertidal non-vegets,cd zones and one from each of the areas on the salt marsh dominated by Halimione portulacoides, Spartina maritima and Arthrocnemum fruticosum. Sediment cores were sliced in situ a t the following depth layers: 0-5; 5-15; 15-25; 25-35; 35-45 and 45-55 cm. Redox potential (Eh) and pH of the sediment layers were measured using a Crison pWmV meter. The pH was measured after a short period of stabilization and the E h electrode was allowed to reach the equilibrium for 15 minutes. Both pH a n d E h measurements a t each layer were repeated four times. I n the laboratory, sediment samples were dried, roots removed and homogenized. The organic matter content (loss of ignition a t 600 'C for 2 hours, LOI), grain size distribution (by wet sieving) and concentrations of Zn, Cu, Cd and P b were determined. After homogenizing, sediments were extracted with the following solutions: 2.5 % acetic acid, 0.005 M DTPA i n CaC1, and 0.1 M triethanolamine a t pH=7.3, 1 M HNO, (9) and HNO,/HCl (12). For the acetic acid, DTPA and HNO, extractants, 10 nil of solution were added to 2 g dry sediment in polypropylene bottles and shaken on a rotating shaker for two hours a t room temperature. The suspension was then filtered. For the HNO,/HCl extraction 10 mL of HNO,/HC1(3:1 v/v) was added to 2 g of dry soil twice a t 130 degrees Centigrade (12). Heavy metal contents were determined by atomic absorption spectrometry. The strong HNO,/HCl digestion was considered to remove the total amount of Zn, Pb, Cu and Cd in the sediment samples. At both sites on the salt marsh, eight plants of each species were collected and transported to the laboratory in plastic bags. The plant roots were washed with demineralized water, diied and homogenized. For Zn, Cd, Pb a n d Cu analysis 10 ml HNOJHC10, were added to 100 mg of plant material according to the methodology described in (12). Standard additions and sludge plus vegetal reference materials were used for sediment and plant analysis, respectively.

RESULTS AND DISCUSSION All sediment samples were made up of silt and clay, containing less than 3 Redox potential and pH measurements were quite 1% sand (fraction ~ 6 um). variable. E h values recorded in sediment layers from non-vegetated sites and below roots were always negative, while in vegetated sediment they varied within a broader interval (from -290 to +210 mV). Positive values were recorded in upper sediment layers where larger amounts of roots are present. Values of pH were always below 7.0 and changed from 5.7 to 6.4 i n the upper 25 cm. E h a n d pH varied inversely (Figure 2) indicating that the rhizosphere is a more oxidative and acid environment than sediments where plants are absent or in layers below the root depth.

358

+++

0

T

:* + ** *** *#f

++

-100-

+

-200

-

t

-900

-

*t

-400 6.2

6.8

f* :#

*

8.0

8.4

0.8

1 2

PH

Figure 2 - Redox potential (mV) versus pH in sediment between roots (+) and in non-vegetated sediment (including sediment below roots) (*) a t sites 1 and 2 on the Tagus salt marsh.

Lot (%) 30 26

+

-

20 -

f

+ + + 4wt+

*

1610 -

6.2

6.6

6.0

6.4

6.8

7.2

PH

Figure 3 - Organic matter content (LOI, %) versus pH in sediment between roots (+) and in non-vegetated sediment (including sediment below roots) (*) a t sites 1 and 2 on the Tagus salt marshes.

359

Organic matter content (LOU was higher in Sediments colonised by plants: in non-vegetated sediment LO1 values remained below 15%, while in vegetated areas they varied from 12% to 30%. However, in the upper 25 cm where roots are present, displayed the highest organic matter content and the lowest pH, as shown by the LOI=pH relationship presented in Figure 3. This indicate8 the organic matter enrichment of' the salt marsh sediments as compared to the intertidal sediments. Vertical distributions of mean concentrations of total Zn, Pb, Cu and Cd, with standard deviations in sediments from sites 1 and 2 are shown in Figures 4. Concentration profiles in the vegetated sediment (x> were plotted separately from the non-vegetated sediment (0).In sediments from the non-vegetated areas a progressive decrease of metal concentrations with depth was observed in the upper 20 cm. Below that depth, metal levels remained rather constant. Metal enrichment has been observed in the surface sediments of the Tagus estuary and attributed to environmental contamination (13). Concentrations of Zn, Pb, Cu and Cd in the surface sediments a t site 2 were higher than at site 1, possibly due t o the proximity of metal pollution sources. Standard deviations obtained for the non-vegetated sediment samples were low, indicating that metal concentrations were comparatively uniform at each site. Profiles of total metal concentrations in the salt marsh sediments were quite different: sub-surface peak values were found and, in general, salt marsh sediments contained more metals than intertidal sediments (compare 0 and x profiles, in Figure 4). In particular, Zn concentration went up to 1150 ug.g-1 and Pb to 500 ug.g-1 a t site 2, while levels in intertidal surface sediments reached only 730 ug.g-1 and 200 ug.g-1, respectively. The increase of metal concentrations, and principally the sub-surface peak values, appear to be related to the presence and the activities of plant roots. Previous work on the Tagus salt marshes (14) have proved that roots are present in higher quantities in the sediments a t these depths. Standard deviation was found to be relatively high between 5 and 15 cm, suggesting that sub-surface enrichment does not have the same intensity all over the salt marsh. The percentages of metals removed by weak extractions (acetic acid, nitric acid and DTPA) in relation to the total concentrations vaiied with depth and commonly explained by the post-depositional processes. Ratios for Zn, Pb, Cu and Cd recorded in the 5-15 cm depth layer of non-vegetated sediments are compared values found a t the same depth with roots (Table 1). In sediments colonised by plants the percentages of metals removed by those solutions were smaller than in non-vegetated sediments. In spite of the high total concentrations of metals found at sediment between roots, the amount of metals removed is small, which means that the larger fraction is non-available for plants.

360

80

1

0

Po0

400

800

800

1000

I

1200

Zn (use-1)

v,

90

Pb

40

60

80 0 80

100

200

300

400

600rite 1600

0

100

200

300

400

600

I

Pb (ug.g-1)

I

600

so I 0

dtu 2 100

200

300

400

600

I

800

Pb (ug.g-1)

Figure 4 - Vegetal profiles of total Zn, Pb, Cu and Cd concentrations (ugg') in sediment between roots (x) and in intertidal sediment ( 0 )from sites 1 and 2; mean concentrations (full line) and standard deviation (dashed line).

36 1 0

fT

w

60-9

00

alb 2 80

Cd

Figure 4 - (cont)

-Dk

site 1

Cd

362

Table 1 - Percentages of Zn, Pb, Cu and Cd recovered with 2.5 % acetic acid, 1 M HNO,, and DTPA in relation to total concentrations in the 5-15 cm sediment depth layers of vegetated and non-vegetated areas of site 1. ~

~

Samples

Zn

Pb

cu

Cd

63 43 41

12 70 73

14 37 42

36 40 60

18 4 3

1 4 3

2 21 11

1 26 8

non-vegetated sediment HAdtotal HN03/total DTPNto tal vegetated sediment HAdtotal HN03/total DTPMtotal

'

Extraction with the strong chelating agent DTPA estimates metals bound to the organic matter, and the use of 1 M HNO, measures metal cations incorporated into carbonates, phosphates and partially organic matter. These forms may all be available to plant roots (9). The reduction of the percentages of those forms at sediment depths where plant roots are abundant and the observed increase of the forms easily exchangeable (acetic acid extraction) may result from several interactions between the plant and the sediment environment (7,9). Furthermore, metals are uptaken by the plants (4). Metal concentrations in the roots of the three most abundant vascular plants are presented in Table 2. Residues of Cd and Cu were lower than those of Zn and Pb. However, the amount of metals incorporated in the roots vaiied substantially from plant to plant. Table 2 - Concentrations of Pb, Cd, Cu and Zn (ug.g-1) in roots of Spartina maritima, Arthrocnemum fruticosum and Halimioneportulacoidesfrom the Tagus salt marshes. Areas

Plants

Zn

cu

Pb

Cd

Site 1

Spartina Arthrocnemum Halimione Spa rtina Arth rocnemu m Halimione

478 420 799 432 346 644

53 10 101 37 31 171

282 60 422 180 90 840

3 > Zn > Cu> Cd.

REFERENCES 1 2

3

4

5 6

7 8

9 10 11

12 13 14

Oenema 0, Steneker R, Reynders J . Hydrobiological Bull. 1988; 22: 21-30. Orson RA, SimpsonR L, Good RE. Estuar Coast Shelf Sci 1992; 34: 171186. Dunbabin JS, Bowmer KH. The Science of the Total Environment, 1992; 111; 151-168. Otte ML. In: Rozema J, Verkleij AC, eds. Ecological Responses t o Environmental Stresses. the Netherlands: Muwer Academic Publishers, 1991; 126-133. Bourg ACM. In: Salomons W,Forstner H, eds. Chemistry and Biology of Solid Waste, Dredge material and Mine Tailings. the Netherlands: Springer Verlag, 1987;3-32. Chenhall BE, Yassini I, Jones BG. The Science of the Total Environment 1992; 125: 203-225. Rozema J, Otte ML, Broekman R, Kamber G, Punte H. In Grank AJ, Benham PEM, eds. Spartina anglica - a research rewiew. London: HMSO, 1990; 63-68. Weimin Y, Batley GE, Ahsanullah M. The Science of the Total Environment 1992; 125: 67-84. Piccolo A. The Science of the Total Environment 1989; 81/82: 607-614. Otte ML, Rozema J, Koster L, Haarsma MS, Broekman RA. New Phytol. 1989; 111: 309-317. Vale C, Catarino FM, Cortesgo C, Caqador MI. The Science of the Total Environment 1990; 97/98: 617-626. Otte, ML. Amsterdam: Free Univ. 1991; 188p. Vale C. The Science of the Total Environment 1990; 97/98: 137-154. Broteriana 1981; 54: 387-403. Catarino FM, Caqador MI. Boletim da SOC.

365

Studies on hea metals of eriph ton and its host plant /Phragmites australis (Cav.) )iyrien ex Steu&l/ in &allow lakes G. Lakatos Department of Ecology, Lajos Kossuth University, H-4010 Debrecen, Hungary

INTRODUCTION

In Europe the most important plant communities in the littoral zone of lakes are the reed communities (Scirpo- Phragmitetuni) which play multifold role in point of view of nature conservation and environmental protection. They have significant role especially in the management of water quality. The effectivity of reed belt in improvement of water quality is enhanced by the periphyton (epiphytic biotecton) formed on the under-water parts of the reed. Through the habitat pecularities reed community offers appropriate biotop for many organisms. Although the significance of reed belt and its periphyton has been recognized, only some papers [l-61 deal with it in detail and with adequate deepness, especially in Hungary and in Austria (7-121. The phenomenon of reed decline was first observed more than 20 years ago in Hungary, the serious damages have been reported since the 80ies [ 13-15] and mostly from our large lakes (Lake Velencei and Lake Balaton). As the accelerated and large-scale reed decline appears as a great problem of environmental protection and nature conservation all over the Europe many research institutions and scientists [18-191 make efforts on revealing its main causes. The main aim of this paper are the followings: To corn are the heavy metals content of periphyton and its host plant in different shallow akes. To demonstrate the spatial changes in the heavy metals contents of periphyton in the reed stands of the same lakes. To estimate the possible reasons of common reed decline in eutrophic shallow lakes. To reveal the correlation among the contents of heavy metals in the periphyton and young reed, applying regression and path analysis.

P

MATERIAL AND METHODS

In the littoral zone of Lake Balaton the periphyton-reed complex has been studied in four transects in the bays at Paloznak, Bozsa, Szigliget and Keszthely for four years (Fig.1.).

366

Hungary

SZ

-

bay Szigliget

Figure 1. Sampling sites on the northern part of Lake Balaton. The reed stand in Paloznak and Bozsa bays can be considered healthy and undisturbed stands of the lake. The reed stands in the Bays at Keszthely and Szigliget are much richer in nutrient elements and heavy metals than in the other areas of the lake. In Bay Keszthely there has occured extensive eutrophication which offerred the reed stands negatively and the accelerated die-back of reed took place /lS/. The reed-periphyton samples were taken near the open water in every transect. In case of Lake Velencei 14 sampling points were selected at the open water side of reed stands in regions differing in water quality (Fig.2.). Unfortunately the reed-decline phenomenon can also be observed in this shoreline of this lake /20/.

367

Figure 2. The study area and sampling sites of Lake Velencei.

Lake Fertij (Neusiedlersee) is the second largest eutrophic shallow lake in Central Europe. The larger part of the lake belongs to Austria and Hungary has the smaller part of it. During the study of this lake reed-periphyton samples were taken in both Hungarian and Austrian parts of the lake. 34 sampling plots were selected on the eastern, northern, western and southern parts of the lake altogether (Fig.3.).Figure 4 shows the sketch of reed /Phragrnites/ and the samples which were analyzed.

8 sediment

Figure 3. Sampling sites of Lake Fertij /Neusiedlersee/

Figure 4. Sketch of reed and the samples which were analyzed.

368

Simultaneously with the reed and periphyton sampling, water samples were also taken for investigation. The wet mass of periphyton was estimated and after drying it at 105 Co, the dry mass was measured. The ash contents of reed and its periphyton samples were also determined. The concentration of 27 cations was measured from the periphyton, reed and water samples by ICP method. Concentration factors were also calculated. Linear regression analysis was used to reveal the relationships between the measured parameters. The direct and the indirect effects of the individual effective factors (independent variables) on the variance of the investigated parameters were estimated by means of path analysis.

RESULTS AND DISCUSSION

O n the basis of the ash content of periphyton samples from Lake Velencei and Ferto are grouped into the same category and can be described by the dominance of inorganic components. At the same time samples collected in Lake Balaton belong to the other category [21]. The ash content of green reed samples are smaller, than that of the old reed ones, only the green reed samples of Lake Ferto exceed the 10 percent. The lowest iron concentration was measured in the periphyton and green reed samples of Lake Balaton, and the highest one in case of Lake Velencei /Table 1./. In contrast to it the manganese content of the samples collected in Lake Balaton was about threefold higher than in the samples of the other lakes. Consequently the Fe/Mn ratio differed considerably in the periphyton of three lakes /Lake Balaton 4; Lake Ferto 20; Lake Velencei 391. Table 1 Heavy metal content of periphyton formed on green reeds in three shallow lakes (%) Ash

Fe

Mn

Zn

cu

42.6 (6.5) 0.71 (0.22)

0.178 (0.138) 0.066 (0.048) 0.024 (0.022)

Lake Velencei 69.0 (8.3) 1.85 (0.42)

0.047 (0.022) 0.028 (0.019) 0.023 (0.019)

Lake Ferto

0.049 (0.034) 0.01 1 (0.010) 0.005 (0.003)

Lake Balaton

71.2 (8.4) 1.01 (0.29)

Similarly to the manganese the highest concentration of zinc also occurred in case of Lake Balaton, which indicated twofold accumulation in comparison with the green reed samples and it is sixtimes higher than in the periphyton samples collected in Lake Ferto. Differences in the heavy metal content of periphyton appeared not only among the three lakes but among the different water regions of a given lake too. T h e heavy metal content of the periphyton was especially high in the water areas that were impacted by a large anthropogenic load or were situated in the neighborhood of sewage effluent. Considering Lake Velencei as an example Table 2 illustrates results to compare the heavy metal concentration of periphyton in the brown water region of the lake /southern part/ which is still in natural state and in the green water area /northern part/ where considerable planktonic eutrophication takes place due to the outer loading of plant nutrient and other inorganic and organic pollutants [22-231.

369

Table 2 Heavy metal content of periphyton formed on green reeds in two water regions of Lake Velencei (%) Mn

Zn

cu

Brown water region 1.66 (0.16)

0.039 (0.019)

0.023 (0.012)

0.018 (0.014)

Green water region 2.24 (0.26)

0.056 (0.003)

0.036 (0.010)

0.032 (0.028)

Fe

In case of Austrian part of Lake Ferto the periphyton samples collected in the southern part of the lake contain Mg in high concentration. The concentration of Mn, Cr and Pb was high in the western part of this lake which is more intensively used for recreation. These cations indicate the human impacts or anthropogenic loads for the water quality of lakes /Fig 51. The values of concentration factors which express the ratio give good evidence for the filtering, accumulating and eliminating role of the reed periphyton.

water

( IWI I ) Pb (pjm)

li (pPrn)

r

03

02

r----

-

0 , -

.... ~.

Figure 5. Concentration of Cr and Pb in water and periphyton. To reveal the ossible reasons of reed decline healthy reed and diseased reed stems were analyzed. d e samples were taken from the same time and site of the studied lakes. Consequent differences can not be stated for the concentration of macro cations, nitrogen, phosphorus and some heavy metals /Fe, Mn, Cd, Pb, e t d , but there are some

3 70

elements which show significant differences between the healthy and diseased reed stems. The date are in percentage of the healthy reed stems /Table 3/. Table 3 The percentage changes of element content of reed stems due to the disease

ash

Zn

cu

Mo

S

%

Lake Balaton

117.6

189.2

202.0

27.3

81.2

Lake Velencei

131.4

217.0

303.5

26.3

68.4

Lake Ferto

105.2

151.6

225.8

28.2

71.1

The results of stud of element content of reed did not help us to much in recognizing the direct causes ofYthe decline of reed community, but the higher concentration of ash, Zn and Cu and lower content of Mo and sulphur of the dieseased reed stems are interesting and warning. In case of iron and zinc positive correlation can be found between of the green reed and its periphyton /Fig.6./. For the periphyton the analysis showed positive correlation between the individual heavy metals and the magnesium which underlines the importance of the absorption processes.

BOZSA I bay p i p h y t on

Fe% 2,o

i -

J

o o

iron

contents of reed

I

Figure 6. Relationship between iron content of young reed and its periphyton

37 I

In case of zinc the variance of the concentration in the periphyton was directly determined by the concentrations in the green reed beside the significant influence of other factors lFig.7.1

Diagram 04 'Pa+h analyris ( Z L ) I

I

' ""

\young

J

+ I reed

"I'

' 0

1

Figure 7. Diagram of path analysis /Zn/ CONCLUSION

The most important plant communities in the littoral zone of shallow lakes are the reed communities. The effectivity of reed belt in improvement of water quality is enhanced by the periphyton. Differences in the heavy metal content of periphyton and its host plant appeared among the shallow lakes. The heavy metal content of periphyton was especially high where impacted by large anthropogenic load. On the basis of our results the biofilter role of periphyton-reed complex can be confirmed. Since the main reasons of reed decline have not been yet revealed and therefore the future research has to focus on studying the role of heavy metals and sulphur. ACKNOWLEDGEMENTS

This study was made within the framework of the Project No 1881 financed by OTKA (National Scientific Research Fund) and partly with the support from CEU (Central European University) which are greatly acknowledged. We are thankful to the Balaton Limnological Research Institute in Tihany and Kossuth University in Debrecen for providing us with good working facilities at their laboratories.

372

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

13 14 15 16 17 18 19 20 21 22 23

Kornhrkovh J, Kornhrek J. Symp. Biol. Hung. 1975; 15: 77-95. Kowalczewski A. Ekol. Pol. 1975; 23: 509-543. Lakatos G, Voros L, Entz B. BFB-Bericht 1982; 43: 40-61. Meulemans JT. Arch. Hydrobiol. 1988; 112: 21-42. Szczepanska W. Pol. Arch. Hydrobiol. 1970; 17: 397-418. Wetzel RG. Oikos 1960; 11: 223-236. Buczk6 K. Studia Bot. Hung. 1986; 19: 63-71. Dokulil M. Verh. Internat. Verein. Limnol. 1975; 19:1295-1304. Lakatos G. BFB-Bericht 1989; 71: 125-134. Padishk J. BFB-Bericht 1982; 43: 95-105. Schiemer F, Prosser M. Aquat. Bot. 1976; 2: 289-307. Sommer U. Osterr. Akad. Wiss. Math.-naturw. Abt. I. 1977; 186: 219-246. Dinka M. Folia Geobot. Phytotaxon. 1986; 21: 65-84. Lakatos G. Proc. Int. Symp. Aquat. Macrophytes Nijmegen, 1983; 117-122. K o v h M. et al. Symp. Biol. Hung. 1989; 38: 461-471. Haslem SM. Ann. Bot. 1970; 34: 573-591. Dykyjova D. et al. Photosynthetica 1970; 4: 280-287. Den Hartog C . Aquat. Bot. 1989; 35: 1-4. Ostendorp W. Aquat. Bot. 1989; 35: 5-26. Lakatos G. BFB-Bericht 1989; 71: 157-164. Lakatos G, Bir6 P. BFB-Bericht 1991; 77: 157-164. Bartha Zs. Acta Bot. Acad. Sci. Hung. 1977; 23: 1-11. Lakatos G. Acta Biol. Debrecina 1978; 15: 147-168.

373

Origin and pathways of Cadmium Contamination in the Gironde estuary, Garonne river and tributaries. J.M. Jouanneau, Y. Lapaquellerie and C. Latouche Departement de Geologie et Oceanographie Universite de Bordeaux I 351, Cours de la Liberation 33405 Talence, France

INTRODUCTION The Gironde estuary, SW of France, (fig.1) is being to become a model for metallic pollution, due to industrial activities located far away, 300 km upstream of the estuary itself. Nevertheless, for a long time, the Gironde estuary had been considered as one of the rare example of an undisturbed estuary in the highly industrialized temperate climatic zone of the northern hemisphere. But, in 1979 "R.N.0."- the "Reseau National d'observation de la Qualite du Mlilieu Marin" (national network in charge of monitoring the quality of the marine environment) made known that there were high metallic element contents, particulary Zinc and Cadmium, in oysters of the lower part of the Estuary. These oysters, which developped mostly from natural beds, are not commercialized but only used for the spawn production which is later transplanted in the oyster farming areas. Given the very high rate of toxicity of cadmium - responsible for the "ltai - Itai" disease (in Nriagu, 1980) a plan was immediatly set out to study the distribution of this element in molluscs, water and sediments of the Gironde Estuary and invarious embayments of the southern part of Biscay coast. Following the different works carried out within the framework of the University of Bordeaux, (1-2-3-4-5), or research contracts (IFREMER, Ministry of the Environment 1984, 1985), IGBA studied the environmental distribution of cadmium in suspended matter and sediments within the estuary as well as within the adjacent marine zones : Ouest-Gironde and Marennes-Oleron Bay. The surveys were carried out either individually or in collaboration with national organizations in charge of the natural environments such as IFREMER (estuary and coastal zone) or the "Agence de I'Eau Adour-Garonne" (rivers). Hydrological and sedimentological characteristics of the Gironde estuary have been already discussed at great length (6). However, a brief review is given only of essential data for the interpretation of the results presented here. Below the junction between the Garonne and Dordogne rivers at the Bec d'Amb&s (fig.1) this estuary covers at high tide an area of 625 Km2. Total average river discharge of both rivers is about 766 m3s-1. The inlet tidal prism (marine water volume introduced into the estuary by current flow) is estimated at 2.109m3 for spring tide and at I .I .109m3 for neap tide. It decreases exponentially towards the upper estuary. The lower part of the 2 rivers which is subject to the dynamic tide but not to the saline one is considered as a fluvio-estuarine zone (Garonne upstream limit : La Reole ; Dordogne upstream limit : Pessac). In this zone, the limit of saline intrusion (salinity range 2 0.5 %o) reaches Bordeaux at low water level periods.

314

As in most estuaries there exists in the Gironde Estuary a residual circulation system, which is a major cause of the turbidity maximum (water masses loaded with suspended sediment : 0.1 to 10 9.1-1) and fluid mud (very turbid water lenses 100 to 300 g.1-l). The total suspended sediment reaches approximately 4.2 to 5.3.1 0%

'ertuis breton

Bellevue srtuis de Maurnusson

F'

France

Spain 0

f 20 km

Gironde estuary

Arcachon Bassin

Figure 1. Map of Gironde estuary and neighbouring nearshore Atlantic area.

375

Both turbidity maximum (1 .I to 2.0.106 t) and fluid mud (2.2 to 2 106 t) are subject to fluctuations, cyclical displacements (associated with tides) and especially seasonal effects (associated with river discharge). During flooding, turbidity maximum and fluid mud may go further than 80 km downstream Bordeaux, whereas at low water level they go as far as Cadillac (approximately 50 km upstream from Bordeaux). Further upstream, suspended matter becomes less important (about 10 mg.1-1) except during flood events (more than 100 mg.1-I). Sedimentological studies have shown that for suspended sediments ( 7 ) ,the downstream estuary supply by marine mineral material is minimal. Evidence of marine sand bed load transport has been reported, but only close to the estuary inlet.

SETTLEMENT AND CHARACTERISTICS GIRONDE ESTUARY AND TRIBUTARIES.

OF

CD

POLLUTION

IN

Estuarine fauna contamination The comparative studies made throughout specimen of oysters fields of the French coast between 1979 and 1988 (RNO, 1988) showed the exceptionally high level of Cd content in the Gironde oysters (20 to 1OOpg.g -1). Comparisons, for instance, with the Britanny and Arcachon oysters revealed that the Gironde oysters contained 10 times more Cd. Results obtained also showed that Cd content was a bit higher in the oysters of the Marennes-Oleron Bay than in the Arcachon oysters (fig.2).

Figure 2. Cd contents in oysters from the French coastal zone (from Boutier, 1987) Cd content in various components of the specific estuary fauna (shrimps, fishes) are compared (tabl. 1 ) with the mean value measured in great number of oyster specimens.

316

Table. 1 Cadmium in Gironde estuary animals

n

Mean contents dry weight pg.g-1

Crassostrea gigas

40

48.9

Crangon crangon

7

0.33

Pleuronectes platessa (muscle)

11

0,lO

Solea vulgaris (m uscle)

11

0,02

Platichthys flesus (muscle)

2

0.02

Limanda limanda (foie)

1

4,OO

SPECIES

Engraulis encrassicolus (total animal) 1 0,Ol R.N.O., 1979 B 1982 - n = number of measurements - data from Dumas, 1' 85. Mean results obtained from a great number of samples are given tabl. 1. It clearly appears that the oysters are, from far, the more affected by Cd pollution. A comparison (3) with several mean values measured in the same oyster species (Crassostrea gigas) all over the world (tabl. 2), shows, that the Cd contents of Gironde estuary oysters are the highest ones after those of the oysters of the Derwent Estuary (Australia) a well known very polluted area. Even in highly industrialized U.K. estuaries and harbours (Bristol Channel, Pool Harbour) oyster Cd contents are lower.

377

Table 2 Teneurs en Cadmium relevees chez Crassostrea gigas.

LOCATION

Cd conten

(pg.9-1dry weight)

Mean value

Extreme values

2,Oto 2,6

English Channe coast

REFERENCES

R.N.O. bull.

1979 to 1982 1,0a 116,5

Atlantic coast Gironde Estuary

48,9

until 116.5

Bristol Channel (R.U.)

23

4 to 43

Poole Harbour (R.U.)

4,6

BOYDEN et ROMERIL

I974 BOYDEN. 1975

4 to 6

THORTON, 1976

Saldanha Bay (South Africa)

9,o

WATLING et WATLING, (1 976 b.)

Knysna Estuary (South Africa)

3,7

WATLING et WATLING, (1976a.)

ramar Estuary (Australia)

33

22 to 45

AYLING, 1974

56,7

0 to 204

RATKOWSKY et al.,

Derwent Estuary and South coast of

1974 35. (a) : including Gironde estuary.

378

Cd in the estuarine environment

Regarding dissolved cadmium (fig. 3) the following distribution was observed : in the upper part of the fluvio-estuarine system, during low river discharge periods, either in Garonne or in the Dordogne branches of the estuary, the concentration are low (-20 ng.1-1) and similar to those usually found in fresh waters (8-9). Conversely, during high river discharge periods, the concentrations could reach 250 ng.1'' in the Garonne fluvio-estuarine part. Downstream, a large increase of dissolved cadmium concentration is always observed in relation with the turbidity maximum and for salinity range between 15 and 20 %o. In this part of the estuary the dissolved cadmium concentrations reach 400 ng.1-1, five times higher than normal (10). From the inlet of the estuary to the coastal waters, values decrease again from 50 ng.1-1 to 15 ng.1-1 that are normal values for coastal waters (1 1). Estuarine particulate Cd Compared to the mean world and to the natural regional background values (3), the Cd content in the surficial sediments of the lower part of Gironde Estuary (fig. 4) proved to be slighty higher than normal. By taking into consideration a reference value of 0,6 pg.g -1, it appears that the enrichment coefficient characterising the deposits of the lower part of the estuary reached a value slightly higher than 2. In contrast, Cd contents are quite normal both in the shelf sediments and in the deposits of Marennes-Oleron Bay.

Ia

C( 4 0 nq 4 0 (C

(

1-1

IOOnq I-'

1 3 c > loono

l

1-1

Figure 3. Dissolved Cd.

Figure 4. Cd contents in sediments.

319

The study of the sediment cores recovered in the oyster fields (4) allowed us to reconstruct the evolution in time of the Cd content in this part of the Gironde Estuary. For Le Verdon, La Coubre and above all Talmont sites, an important rise of Cd was reported in the deposits corresponding to about the last 30 years (fig. 5). Le V e r d o n

Ta Im on 1

Lo C o u b r e

Bellevue

Figure 5. Cd in cored sediments.

In the suspended matter S (J,l from the whole estuary (1) abnormal values were recorded in suspended matter of the lower part of the Garonne (La Reole) at the riverlestuary junction, especially during high water level periods when values attained 14 pg.g -1. Comparison of the Cd content in the suspended matter from La Reole and from dowstream to Bordeaux, showed that the wastes of the city, often thought to be the cause of pollution, cannot be considered as the main source of Cd for the Gironde estuary. Cd contents in sediments, SM and water of the drainage basin In order to precize the origin of Cd contamination of SM in the upper part of the estuary, surveys were carried out over the entire drainage basins of the two rivers flowing into the Gironde estuary, Garonne river and Dordogne river. Sampling of sediments in 70 selected stations allowed to carefully recognize the distribution of heavy metals (Cd, Zn, Cu, Ni) in bottom sediments of the various tributaries. Cd contents are presented fig. 6 . Bottom sediments from the Lot river -one of the main tributary- appear heavily contaminated by Cd. High level of several other elements (Zn, As, Ba, Cu) were also measured in the Lot fluvial deposits. The source of these high metallic concentrations was attributed to the drainage of the waste area of an

380

old industrial plant (Zn) in Decazeville, a town overflown by the Riou-Mort river, a small tributary of the Lot river located 250 km from the junction of Lot with Garonne rivers. Accidental metallic pollutions affecting Riou-Mort river have been often reported since 1979 (12). Sampling and analysis of water and SM confirmed the very high level of dissolved and particulate Cd and various other metals, in the Riou-Mort water. Conversely, lower -and often normal- dissolved and particulate Cd contents were measured in the other Garonne and Dordogne bottom sediment tributaries.

Figure 6. Particulate Cd content from the Lot drainage basin

PATHWAYS OF OYSTERS CD CONTAMINATION

In order to understand the series of processes responsible of the final oyster contamination, several phenomenons have been considered : - the Cd mobilization in the Riou-Mort and Lot river, - the Cd transfer into the estuary, - the kinetics for Cd accumulation in oysters.

38 I

The source of Cd and the riverine processes. As soon as the source of pollution was recognized, the waste area of the old plant has been isolated by building a dam. Since this construction, this waste is no longer a major source of pollution. But in spite of this work, the Lot river remains a source of Cd pollution. Indeed bottom deposits all along the Lot river contain very high contents of heavy metals, particularly Cd. These deposits are accumulated behind a series of dams which have been built all along the Lot river in order to regularize the water fluxes. The accumulation of sediment behind these major dams (fig. 7a) has been estimated by using seismic profiles (fig. 7b) and coring. Then, geochemical studies on cored sediments allowed to recognized the thickness of contaminated bottom sediments (fig. 7c). From these results, the global stock of Cd, in the bottom deposits of the whole Lot valley has been estimated to 200 T. The more recent studies (13) have for objectives to predict the risks of mobilization of this contaminated stock either by reworking of sediment or by diffusion of pore water at the water/sediment interface (fig. 8). Preliminary results show that resuspension of fine polluted sediments can occur during the river floods, when currents reach 30 cm.s-l (measurements at 1 rn above the bottom). In addition, major changes of bottom physico-chemical conditions (e.g. acidification andlor Eh increase) could have for consequence an increase of dissolved fluxes at the watedsediment interface. Preliminary calculations (13) applying Fick law to the pore water profiles (14) indicate that Cd molecular diffusion at the water sediment interface varies from 17.10 -6ng.cm2.s-1 (reducing environment) to 41.1O-6ng.cme.s-1 (oxydizing conditions) such fluxes have nevertheless to be precised through the study of a greater number of pore water profiles realised in the framework on the on going studies. The transfer of Cd, within the estuary Fluvial inputs of Cd within the estuary have been monitored since 1982. These inputs widely vary, according to the hydrological conditions of the rivers (7). For average riverine discharge conditions (Garonne 1 13,6109 m3.year-1 ; Dordogne : lO,2.lO9 m3.year-1) the fluvial contribution of Cd has been estimated to 23 t.year-1 for particulate Cd and 2 t.year-l for dissolved Cd. As for many metals (7) an abrupt upstream/downstream decrease of Cd content of SM (fig.9 [a]) has been observed along the estuary (2). Conversely, maximum dissolved Cd (fig. 9 [b]) concentrations have been observed for salinity of about 20 %o, i.e. in sites very close of the inlet. Far away towards the sea, at higher salinities, the Cd content decreases through dilution with the marine water. This repartition of particulate and dissolved Cd has been explained (15) by a solubilisation along the estuary of the Cd initially associated with fluvial SM.

382

a

I

) Molndornt

b)

C)

Selsmlc profile

Cojarc

d)

Core2 Sandy mud

I0 Muddy sand

20 24 Sand

32

30 40

Muddy sand

50 cm

Fig. 7. Cd in bottom sediments of Lot river (Cajarc).

crn

Colorc

Cd

content

383

o 10

,

'

"

"

3

1

20 30 40

50. cm

j_

Figure 8. Cd in pore water (Cajarc).

L

0 - $0 La Reole

10

b

'

30VeO S o l i n i t y

Oyster IbOKrn from Bordeoux area LeVerdon

Bordeaux

-GAAONNE RIVER-+ (estuorine port)

20

+--ESTUARY-PF

SHELF-

Figure 9. Upstream downstream evolution of particulate and dissolved Cd contents (low and high river dischage periods).

384

IGBA studies, by monitoring the behaviour of particulate and dissolved metallic elements along the estuary and by using radioactive zinc tracers previously showed that when metals such as Zn enter the Gironde estuary, they mainly occur in a particulate form but become soluble when they reach the estuary (16). As cadmium has similar chemical characteristics as zinc, it is logical to assume that it reacts in the same manner within the estuary and this would be the reason why, downstream, metal mainly occurs in a dissolved form. On the basis of various surveys devoted to the study of the behaviour of metals within the estuary and its outlet (7-l), it was possible to estimate the flow budget for various elements particularly for Cd. Estimated on the basis of hydrological conditions average year, the Cd budget (tabl. 3) showed that the lower part of the estuary receives an input of dissolved Cd exceeding 20 tons per year. Table 3 Annual budget of Cd in the Gironde estuary Ocean

Rivers

Estuary

IM concentrations, but significant differences in performance between different batches of tubes were noted [63] 3.2.3.5. Tube material and tube design For wall atomization, pyrolytically coated graphite tubed are superior for the determination of Se. In the use of platforms, however, optimum conditions seem to depend on the kind of sample and the kind of apparatus used.

430

Soaking with W- solution lead to a change of the surface structure because of carbide formation, but to no improvement for the determination of Se from dilute nitric acid solutions[24 I]. Coating of graphite tubes with Nb or Ta resulted in better reproducibility and accuracy, at equal sensitivity for Se in presence of 50 pg Ni [233]. TaC - coated graphite tubes lead to additional thermal stabilization of Se in presence of excess Pd, but also caused some memory effects [224]. In boron-nitride coated tubes, retention of S e in HNO, was possible up to 500", but losses of Se(IV) from 0.1M- HCI occurred to 20-30% within the range of 200-500". The boronnitride -coating, however, was destroyed after 30-35 firings [216]. Coating with Mo was achieved by injecting 4 times 5 5 pI of 5% Mo solution and heating to 2500". The Mo- coating significantly suppressed the interference of Fe resulting from overcompensation of the D,- signal, due to delay of the Se in the peak height mode, in presence of P t N matrix modifier [214]. 3.2.3.6.Use of auxiliary gases Addition of 10% H, to the inert gas (Ar) reduced the peaks of organic molecules during graphite- furnace atomization (like pentane, hexane, methanol, ethanol, and chloroform; except aromates) and doubled the peaks of Se- containing organicals. Thus, D,compensation was sufficient to use a graphite furnace as a detector in gas-chromatographic separation of volatile selenium- containing species collected from ambient air [246]. 5% H? in Ar was used to reduce the Pt- modifier to its elemental state, which made i t far more effective [225]. Ashing of biological samples with O? inside the graphite tube may be risky, because Se forms volatile oxides [247]. Charring of dithiocarbamate- extracts at 300" in air doubled the signal of Se in TaCcoated tubes, because of improved destruction of the reagent [224]. CO was used as an auxiliary gas during charring up to 1000" at a flow rate of 0. I llmin to remove O? from the furnace. This prevented the formation and subsequent atomization of iron oxide, which interfered in the background correction of the Se- signal with D2 [225]. 3.2.3.7. Solid sampling and slurry atomization For Se, solid sample atomization is only reasonable since the introduction of powerful Zeeman-background compensation systems, and availability of sufficient homogenous solid standard material. As best and universal modifier, a mixture with graphite powder was recommended. For Se, however, few reference materials were available, and the detection limit was not sufficient in any case [248]. Introduction of micro-amounts of solid samples cannot only done rized about 60pg sample from a homogenized tablet of geological material, which was introduced into the tube together with the carrier gas [249]. For the analysis of Se in metal chips, about I mg of the alloy was inserted into the furnace and atomized. For Se, an atomization temperature from Ni-base alloys of 2600" was needed, and a detection limit of 0.2 pg/g for 1 mg sample could be reached only Standardization had to be made against doped alloys of similar composition [2SO], because the sample was not completely vapourized.

43 I In solid sampling of liver homogenate, the Ni/Ag- matrix- modifier was added to the sample before freeze drying and homogenization, to ensure sufficient mixing For Img sample weight of liver, a detection limit of 10 ppb could be achieved [247] In slurry atomization, the solid sample is mixed with a fluid for reasons of dilution, and pipetted into the graphite tube. For analysis of coal and coal fly ash samples, a solid slurry was prepared by ultrasonic mixing. Calibration was done against similar NBS- standard materials. Atomization was done from a platform inserted in a standards grooved tube. The solid was diluted with slurry was 5% H N 0 3 / 0.04% Triton X-100. A Zeeman compensation device was essential. Se was more difficult to determine than As, Pb, and TI. At 1900°, severe matrix absorbance appeared, which necessitated to atomize as low as 1 8 5 0 O . Pd as modifier sharpened the peak, but did not improve the accuracy [25 I]. For the analysis of milk powder, 0.5 g sample was mixed with 3 ml H,O, and 200 pl of this slurry added to lml RhiMg- matrix modifier (0.4% Rh(NO,), + 0.25% Mg(N0,)2.6Hz0 + 0.4% HNO, + 3% Triton x-100). Charring was done from a platform, Mg- nitrate acted as ashing agent. Evaluation was done via peak area and standard addition [252].

3.3 ATOMLC EMISSION SPECTROSCOPIC METHODS

3.3.1 INDUCTIVELY COUPLED PLASMA

- DIRECT ASPIRATION

The main atomic line of Se at 196.026 nm has a rather high excitation and ionization energy and therefore tends to peak rather high i n the plasma [253], nearly like an ionic line. The increase of observation height in the pure Ar plasma, however, lowers the signalibackground ratio, because of strongly increasing background Similarly, the signal/ background ratio is worse in the high power-range and upon addition of 5% N, to the plasma gas [254]. For Se, no improvement by means of switching the plasma excitation frequency from 50 MHz to 100 MHz has been achieved [255]. As far as spectral overlaps are concerned, the main emission line of Se at I96 026nm may be interfered by an Fe-line at 196.059nm, if the resolution of the spectrometer is insufficient [2 561. Detection limits of Se emission in the ICP using direct aspiration, are within the range 10-70 ngiml [255,257] and thus insufficient for the analysis of natural waters and digests of biological material [258]. Direct coupling of the ICP to a liquid chromatographic system with a flow rate matched to the usual nebulizer uptake, lead to a detection limit of about 0.1 pg Se at the insensitive line at 203.985 nm [89]. 3.3.2 DIRECT CURRENT PLASMA

- DIRECT ASPIRATION

In the DCP, matrix effects upon the emission signal of Se at 196 026nm by ionization enhancement are negligible The signal was regained in presence of up to 1 5 g/l Na, and Ig/l AI, Ca, Fe, Mg [259] The precision, however, decreases significantly at lower concentration levels [259] The detection limit of about 0 25 pg/ml [260] is insufficient for environmental samples

432 3.3.3 HYDFUDE METHODS The hydrides of interest, together with a constant amount of hydrogen, are produced in a continuous- flow system, and introduced into an analytical plasma without further nebulization. Evolution of a smooth stream of hydrogen without bumping and splashing is essential. Low spectral background and lack of nebulization losses improve the detection capability about two orders of magnitude, which is sufficient for many environmental samples. The multielement capability of the instrument is limited to the hydride-forming elements, only Se(IV) is monitored, and all interferences of hydride evolvation (see chapter 3.2, 3.8) have to be considered. The AES signal stabilized about 1 min after introduction of the sample to the hydride system [261]. The optimum range of acidity of the sample depends on the exact geometry of the device, and the carrier gas stream used. Starting from HNO, gave a slightly lower sensitivity compared to HC1 [262]. For the stabilization of the gas introduction into the plasma, in the simplest approach, the spray chamber of the conventional ICP- nebulizer was replaced by the mixing chamber of the hydride system [257]. Alternatively, the hydrides and hydrogen have been separated from the reaction mixture via a 2m or 5m long silicon rubber tube placed in the stream of Ar-plasma gas [263]. A long and narrow gas- liquid separator was filled with Pyrex beads, which provided smoother separation of liquid and gas [264]. If the hydrides are introduced into a DCP, half of the Ar flow is directed through the hydride generator to carry the evolved hydrides into the plasma, while the other half passes its normal way through the nebulizer system. The reaction gases (mainly hydrogen) are dried by passing CaCI,, and delayed in a suitable delay tube before nebulization, to provide a smooth gas stream. To obtain lower detectable concentrations, hydride evolution from a lOml sample batch is also possible, leading to a detection limit of 15 ng Se. Matrix Ni was successfully masked with phenanthroline [260,265]. 3.3.4 OTHER EXCITATION TECHNIQUES Excitation of electrically conducting solid samples in a DC-arc between crbon electrodes at 30 A suffer from the short wavelength of the Se- main emission line at 196.026 nm. A stream of argon shielding gas (7.5 l/min), long exposure time (65 sec) and optical grating of 1200 grooves/mm were necessary to obtain reasonable signals on the spectrographic plate. This method was used to determine Se in various sulfides after mixing with equal amounts of ultrapure graphite [266]. Analytical lines which can be measured without a vacuum are found in the visible region, but the energy required for their excitation is so high that they do not appear in arc and spark spectra. If the sample is dried upon an Al-cup, which is subsequently used as the target of a hollow cathode discharge of 240W, ionic lines at 444.62nm and 444.95nm appear, which can be easily detected spectrographically, with a detection limit of Ing/ml, achievable in human serum [267]. Contrary to other excitation sources, in a microwave induced plasma (MIP) within a flow of 1.1 I/min He, the Se- lines at 206.279nm and at 203.985nm emit signals as well, but applications for real matrices are not given [268]. Vaporization from a Ta- strip into an Ar-fed MIP allows to evaporate the solvent after solvent extraction, and subsequent pulsed vaporization of the analyte [ 161. After pressure

433 decomposition, the detection limit of the MIP, however, was insufficient for the determination of Se in fish and vegetables Molecular emission cavity analysis (MECA) uses the broad emission of SeO, Se2, and S e 0 2 in the range 330-530 nm, with a maximum at 413 nm Molecular eniission is achieved by injecting the sample into a steel cavity placed in an air-H1-N, flame, and cooled on the backside As many metal ions interfere, and H2Se is much more sensitively detected than Se- oxyanions, the sample is introduced after hydride formation For reasons of enrichment, the H2Se can be caught in a cool trap, and evolved by immersing the trap into hot water [269,270] 3.4 DETERMINATION OF SELENIUM BY MASS- SPECTROMETRIC METHODS

3.4.1 GENERAL Five selenium isotopes are naturally abundant: Isotope Natural for more details see 3.6.1 abundance 74-Se 0.87 YO 76-Se 9.02 YO 78-Se 23.52 % 80-Se 49.82 YO 82-Se 9.19 YO Mass spectrometric techniques have the advantage of large multielement capabilities, relatively simple spectra, and offer the possibility of isotope ratio measurements and isotope dilution analysis. 3.4.2 ICP

- MS

A horizontally mounted ICP - torch serves as sample introduction device for a quadrupol mass spectrometer [271]. Since this technique is rather new (first papers published about 1980), only few applications for the analysis of Se in real matrices can be found in the present literature. In case of Se, isobaric interferences of neighbouring elements hardly occur. Arsenic is monoisotopic at 75 m/z, where there is no S e isotope, and bromine is hardly found as a cation [272]. In matrices with high chloride content, or in case chloride containing acids have been used for sample dissolution, polyatomic ions containing CI are encountered in the mass spectrum [273]. Additionally, Se coincides with background peaks of the Ar-plasma itself, emanating from various Ar and C1 isotopes. 76 Se 36Ar - 40Ar 77 Se 40Ar - 37CI 78 Se 38Ar - 40Ar 80 Se 40Ar - 40Ar Only 82 Se is free from such interferences 12721. Spiking the sample solution with 10% propan-2-01 or introduction of 3% N2 into the nebulizer gas flow reduces these polyatomic signals; maybe there is competitive formation of Arc', ArO' and ArN' [273].

434 In ICP-MS of positively charged ions, without any matrix, a detection limit of only 0.8 pg/l could be reached within 10 sec single ion monitoring, according to its high ionization potential of 9.75 eV [274,275]. In multielement analysis of natural lake waters, the detection limit of only 3pg/1 Se was insufficient ( 2 sec integration) [275]. In acetate buffer leaches of soils, monitoring of 78-Se and 82-Se resulted in a detection limit of 5 pgil in a 20 sec total measurement time. Noteably, the sampling for the mass spectrometer was done at 22 mm above coil, which is unusually high, but reasons were not given. However, the precision of the recovery of spikes to the acetate buffer was only 128% [276]. Direct use of non-digested urine and serum samples into the ICP-MS yields polyatomic fragments. At mass 82, where there is no Ar-Ar or Ar-CI background like for the other Seisotopes, the results were anomalously high, in comparison with ICP-AES and AAS determinations. This may be due to the CCI" ion, which also appears in the spectrum of pure trichloro-acetic acid [277]. In the direct analysis of red blood cells, 74-Se coincided with FeO' [278]. In the negative ion scan, there are considerably fewer background peaks. The background peak at 78-Se disappears, but at 80-Se it is still present [279]. Introduction of hydrides into the ICP-torch reduces the number of interfering molecular ions in the mass spectrum, e.g. oxides and chlorides, but the start from Se(1V) and interferences in hydride generation have to be considered. A debubbler to remove excess air prior to the mixing zone greatly improves the results [278,280]. ICP-MS via the hydride was applicated in hunian metabolic studies employing stable isotope tracers [278]. 3.4.3 OTHER MASS SPECTROMETRIC METHODS

In spark source mass spectrometry, the electrical conductivity of selenium in most of its compounds as well as in most of the matrices of interest is insufficient. It needs to be electrolytically preconcentrated at gold- electrodes after suitable sample decomposition by wet chemical methods, or reduced by means of hypophosphoric acid after addition of gold chloride spike. The resulting gold containing Se, Te and some other trace metals, is directly sparked. For quantitation, isotope dilution with 78-Se [45] or with 82-Se [281] has been applied. Thus, analysis of Se in coal, in steel and Ni- based alloys down to 0 1 pgig is reported. Both selenite and selenate evaporate as negatively charged ions from a hot metallic ribbon, which has been used to determine Se- traces in natural waters ("negative thermionic MS"). Selenite and selenate could be discriminated after separation by anion exchange. Within the range pH 1-12, no isotope exchange between selenite and selenate could be detected [282]. 3.5 ELECTROCHEMICAL TECHNIQUES 3.5.1 PREFACE As selenium can occur in various oxidational states, For pure aqueous acids or alkalis, the normal potential of the respective redox reactions are [283]: 1M acid: H,Se Se H,SeO, Se0,'I M alkali: Se'. Se ~ 0 . 3 V> 7 SeO,'. SeO,'-

435

Common to all techniques for real samples (except for detection of compounds i n HPLC) is to start from Se (IV) The formation of compounds with the electrode material on its surface can lead to shifts in the peaks obtained, as well result i n electrochemically inactive compounds Direct applications of electrochemical methods have been given only for saline waters, and for soil extracts obtained with neutral salts 3.5.2 DETERMINATION OF SELENITE IN AQUEOUS SOLUTIONS 3.5.2.1 Mercury electrodes 3.5.2.I . 1 Dropping mercury electrodes In 0.2M - HCI, the polarographic wave at -0.54V vs.SCE, deriving from the reduction of HgSe to Hg + H,Se, yielded a non- linear calibration graph Pb, Cu, and Fe influence both peak position and current, because of formation of selenides at the electrode surface [284] In 0.1M- HCIO, as well as in O.1M- KNO,, excess of Pb, Cd, and Cu decrease the Sepeak at -0.61 V vs. SCE, without change of its shape [285]. Even at pH=I, Pb can be masked with EDTA [286]. In weak acid ammonium sulfate, nitrate interferes, and has to be removed, eg by short warming the sample with ethanol [ 1291. In 1M NH,CI or NH,- acetate soil extracts, in'the range pH 5 4 - 9.0, an analytically useful peak could be obtained within the range - 1 . I V to - I .6V vs SCE. The best separation of the selenium current peak from that of the supporting electrolyte was obtained at pH 8. with peak potential of -1.34V versus SCE, and a detection limit of 5ng/ml. In this buffer solution, Fe, Pb, and Cu do not affect in 1000-fold excess, and Co, V, Te, Cr(lIl), Mo do not affect in 100 fold excess over Se. The interference from Zn is removed by addition of EDTA, which simultaneously causes e peak from the reduction of Pb-EDTA at - I .2V, but in real samples the Pb- concentration is usually too low to interfere Among the organics, at the potential and pH conditions used for the reduction, the most likely interferents are expected to be simple organic compounds containing carbonyl and carboxyl groups Only maleic acid was found to interfere, and could easily be removed by mild acid hydrolysis [284]. In acetate/borate/phosphate buffer pH 4.0, the second peak of the Se- reduction at - I .3 I V vs. SCE could be used in presence of large excess of Cu, Pb, and Cd [285]. In order to avoid precipitation of selenides, formation of selenosulfate from selenite was achieved by adding Na-sulfite to the acidified sample solution, with subsequent adjusting to pH 7-8 [ 129,2851. At pH 9-1 1 in ammonium sulfite solution, Se can be measured by reduction of selenosulfate at -0.95 V down to 5 ngiinl [129]. 3 5 2 1 2 Stationary Hg

- drop electrode

In a first step, Se is deposited and enriched by electrolytical reduction at the working electrode Cyclic voltammograms show, that a solid HgSe - film develops on the drop surface, which dissolves at a more cathodic polarization [287] In the presence of halide ions, the cyclic curve is modified, and the wave shifted to more negative values From selenous solutions containing halide ions, a selenium compound is deposited on the surface of the Hg drop, which transfers no electricity, but is cathodically reducible, termed HgSe

436

The observed shift in the peak potential is due to complex formation [287]. At increasing plating time, an additional peak at more negative potential appears in cathodic stripping, which is less at lower pH, and does not occur at pHc4.2. The sum of currents of the two peaks is constant, and equal to the intensity of the peak when only one is present. This is interpreted by the formation of a mixed deposit of Hg-Se", and the formation of Se" by reaction between H,SeO, and H,Se [288]. In presence of Cu, a selenide presumably containing Cu(1) is formed during the electroplating step, instead of the Hg-Se compound [289]. Anodic stripping voltammetry The position of the anodic stripping peak strongly depends on the pH of the solution. The determination is interfered from many elements deposited together with Se, and also from compound formation at the electrode surface [290]. In the analysis of milk powder, a matrix very low with respect to trace metals, the direct anodic stripping voltammogram of a decomposition solution with HCIO, yielded high background currents, which could be reduced by fuming with HCIO, [291]. Increased sensitivity could be reached by scanning in the differential pulse mode, with an amplitude o f f 20 to 50 mV. Cathodic stripping voltammetry Besides As, Se, and Te, also V, Cr(VI), Mo and W have been determined by this technique [292]. Cyclic voltammetric studies at a stationary hanging mercury drop electrode in acetate buffer in the range pH 3.85 - 5.55 shows three electrode reactions. At -0.3 V vs SCE, a broad and completely irreversible peak appears, which is due to the reduction of selenite to elemental selenium, and which is pH dependent and temperature - dependent. In cathodic stripping voltammetry, the two other peaks, at -0.75 V and at -0.90 V are utilizable. The third peak at -0.90 V is not observed at pH 4.2/30 sec plating time, but increases with increasing plating time. This is explained by both the formation of a mixed deposit og Hg-Se", and the formation of Se" by reaction of selenite with hydrogen selenide [288]. The selenium peak in the cathodic stripping voltammogram is shifted with pH, beause of participation of protons on the reduction to H,Se. The dissolution current is proportional to the electrode surface, which indicates, that the reduction takes only place at the electrode surface [293]. Increase of pH reduces the peak current, and shifts the dissolution peak to more negative values. At pH '8, no cathodic dissolution peak was observed [294]. The cathodic stripping voltammetry of Se in 0. IM HCI yielded a non- linear calibration graph for to 10"M Se. Variation of the deposition potential in the range of +0.05 to -0.30 V VS. SCE was of low influence [295]. The presence of metal ions may shift the stripping peak to a more negative potential. The determination of Se is interfered by Pb and Cd, which strongly suppress the peak at 1 mg/l already. The suppression is less in presence of some Cu, when deposition at -0.3V is used [36]. Zn and Cd make no peak themselves, but they severely suppress the Se- signal [294]. In dilute HNO,, 20 - 100 fold excess of selenide forming cations can be tolerated, but Pb and As interfere [296]. Arsenic forms a new peak at -0.3V vs. Ag/AgCl along with Se. In ammonium sulfate at pH 4.5, addition of Cu (up to 1 pg/ml) enhances the deposition of Se at Hg, and shifts the peak potential to more negative values. During cathodic

437

stripping, presumably a Cu-Se compound is dissolved within the range -0.8 1 to -0.84 V. Like Cu, also Bi, Ag, and Au shift the stripping peak potential to more negative values, with respect to the Hg-Se peak, which leads to peak enhancement and to peak splitting at deposition potentials more positive than -0.5 V. From the position of the peak potentials it can be concluded, that Cu-Se is the most stable compound investigated [290]. Whereas the cathodic dissolution peak of selenium in dilute H,SO, and H,PO, is found at -0.51 V vs. SCE, it moves to -0.704V in IM- ammonium sulfate/0.4M- EDTA/pH 4, and to -0.81 V in 0.3M KNa- tartrate/0.2M-EDTA/pH 6 [294]. Interference from Cd, Zn, As, Cr, Pb and W was minimized by utilizing ammonium sulfate/EDTA/pH 4 as supporting electrolyte. Cu, Sb, Ti and TI in excess still interfere. The surfactant Triton X-100 at > 1 O-'% completely suppresses the peak [294,295]. Cathodic stripping voltammetry in ammonium sulfate /EDTA at pH 2.3 in presence of 2 pg/ml Cu, enables the specific determination of selenium and tellurium in presence of large excess of each other [293]. Cu concentrations of 3 pg/l or more enhance the height of the cathodic stripping peak. In presence of Cu, peak height and peak potential largely depend on the deposition potential At deposition potentials more positive than the formation of Cu?Se, the stripping peak shifts to more positive potentials, and the peak height decreases [289,297]. Cathodic stripping of Cu-Se yields a narrower and higher peak than for Hg-Se. Optimum sensitivity has been achieved at pH 1.6 and 40 pM Cu [289]. In dilute HCI, Fe, Pb, Zn, Cd, and Te decrease the cathodic stripping peak; Zn and Pb can be masked with EDTA. After deposition at -0.35V vs.Ag/AgCI, the cathodic stripping peak is obtained by sweeping the potential of the working electrode to -0.9 Vwith a speed of 10 mV/sec [297]. Addition of Cd caused a decrease of the stripping peak of Cu,Se, while 2 new peaks were formed, due to Cd reduction, and the cathodic dissolution of CdSe. Cd can be partially masked with EDTA [297]. Cathodic stripping voltammetry at the hanging mercury drop electrode can be directly applicated to the analysis of drinking water. Ammonium sulfate as the supporting electrolyte, as well as EDTA to mask interfering cations, are added to the sample, the pH adjusted with sulfuric acid to pH 2.2, and 0.1 mg/l Cu to improve sensitivity. After 5 min plating at -0.254 vs Ag/AgCI, a detection limit of 1 pg/l at the stripping peak at -0.68 V could be reached. Sulphide strongly interferes [298]. For the determination of Se(IV) in sea water, only 1/1 HCI and Cu have to be added to adjust to pH 1.6. After 15 min of deposition at -0.4 V, down to 0 7 ng/l Se could be detected [289]. After volatilization of Se from most of the matrix in a stream of oxyge, and dissolution from the cool finger with dilute HCI, Se was determined by cathodic stripping voltammetry with ammonium sulfate/EDTA pH 4.5 as supporting electrolyte, and addition of some Cu [621. Direct use of the digestion solution from biological materials (liver, rapeseeds) resulted in no peaks in cathodic stripping voltammetry because of high background. Separation via extraction of a piazselenol with subsequent wet ashing of the extract lead to suitable sample solutions ready for the determination. Without the evaporation of the organic solvent, however, resulted in sensitivities less than 30 % with respect to the aqueous sample [26]. For the determination of selenium in biological materials, Se was separated and enriched from the acid digest by sorption upon an anion- exchange resin in the acetate form at pH 3, prior to either cathodic stripping or anodic stripping voltammetry [36].

438 3.5.2.2 Au- electrodes

In anodic stripping voltammetry of selenium from a Au-disk electrode in 1M- H,SO,, stripping peaks appear at +0.64V, +0.86V, and at -1.03 V. Hg interferes [299]. Similarly, in 0. IM-HCIO,, three anodic stripping peaks are observed for large quantities of deposited Se. At the beginning, and at low concentrations, approximately a monolayer is deposited. Thus, the anodic stripping of very small quantities of Se following deposition at a low flux yields a single anodic peak at 0.8V. Further electrolytic reduction of Se leads to irreversible diffusional transport of Se into the electrode, forming a Au-Se alloy of unknown stoichiometry [300]. In the anodic stripping voltammogram following deposition at high fluxes of Se on the Au- surface, two additional peaks, at +0.63V and at + l . 15V are obtained [300]. These peaks can be interpreted as due to bulk Se, adsorbed Se, and as intermetallic Au-Se compound of unknown stoichiometry. Only the adsorbed amount of Se is analytically usefull. To enable only formation of adsorbed Se upon the electrode surface, a deposition potential of 0.15V vs. Ag/AgCI is preferable over the deposition at more negative values. The optimum deposition potential at the Au-electrode is thus far more positive than the -0.35 V for the mercury drop electrode for the same sample solutions. This also reduces the extent of codeposition of interferents, such as Cu and Pb [36]. The Au- electrode was pretreated prior to each experiment by polishing the surface, and preconditioned in 0.2M- HCIO, by applying alternate cycles. However, difficulties in obtaining reproducible Au electrode surface area from one experiment to the other arose. The anodic stripping procedure at the rotating gold electrode equalled the sensitivity of the cathodic stripping procedure at the mercury electrode, but the reproducibility was worse [36]. To make the Au- surface continuously renewable, the Au can be plated upon glassy carbon prior to each run. Detection limits were 4 times less than for pure Au electrodes, but only I@'M Au was needed in solution, which made it rather cheap [301]. 3.5.2.3 Graphite and carbon electrodes At a graphite-pin-electrode, made a mixture of graphite and polyethylene- powder, the reduction of selenite gives two polarographic waves, within the range 0 to -0.2 V VS. SCE, and at -0.6 to -0.8V , which correspond to the reduction of selenite to elemental selenium, and of elemental selenium to selenide. At electrolysis in the range of the more negative maximum, the electrode surface is partially covered with red selenium, another part precipitates as colloid near the electrode in the solution. This amorphous Se is electrochemically incative. Its relative amount incrreases with pH, and is at maximum at a plating potential of -0.3 to -0.4V [302]. Increasing acid concentration increases both the anodic and the cathodic stripping peak of elemental hexagonal metallic selenium. During electrolysis in concentrated HCI, however, elemental chlorine formed at the counter electrode interferes by oxidation of selenite to electrochemically inactive selenate [302].

Anodic stripping voltammetry At a graphite disk, impregnated with wax polyethylene = 3 I , after electrolysis at -0 6 V vs SCE, anodic stripping voltametry can be performed in dilute HCI, which results in a peak at +O 22V for selenium Cu is added to increase the electroactivity of Se deposits because of formation of an intermetallic compound, but produces a second peak at -0 15V To obtain maximum sensitivity, Cu Se must exceed 30 1 [303] Acidic electrolytes give much better sensitivity for Se than neutral or basic ones After 1 min depositlon time a

439 detection limit of 0 1 pg/l was achieved for pure solutions [304] Upon the addition of some Cu, the peak of selenite ion increases and moves towards a more positive potential, which suggests the formation of an intermetallic compound At large excess of Cu, the anodic stripping peak for Cu appears at about 0 (vs AgIAgCI), which overlaps the Sepeak [304] Matrix silver delivers a peak at +O IZV, and has to be removed beforehand, whereas Pb increases the Cu-Se signal, and Fe and Mn are of no influence [303] Cathodic stripping voltammetry: At a graphitelwax electrode plated with Hg, contrary to the Hg drop-electrode, it is possible to determine Se by cathodic stripping in acid as well as also in alkaline sample solutions (> IM- NaOH), which masks Cu and other metals The sensitivity of the cathodic stripping peak is effected by the deposition potential and the time, the scan rate, and the thickness of the Hg- film. As optimum conditions, plating at -0.10 to -0.30 V, and potential sweep of 60 mV/s has been found [305]. For cathodic stripping voltammetry upon a rotating graphite disk electrode, Se is plated for 1-5 min in 0.4M H,SO,, in presence of Cu and chromate, and subsequently stripped in the pulse mode, to obtain a peak at -1.07V The method was applicated to the determination of Se in matrix Ga, after separation from the matrix and from nitrate, deriving from the dissolution procedure [306]. 3.5.3 DETERMINATION OF PIAZSELENOLS Isolation of selenium as a piazselenol, and the subsequent polarographic determination of the piazselenol itself enables separation and enrichment from interfering matrices (see 211,234,311) In 0 IM NH,CIO, i n formiate buffer at pH 2 5 , polarographic waves of the piazselenol from 3,3'diaminobenzidine at -0 1 1 V and at -0 63 V vs SCE are obtained at the dropping Hg- electrode The reagent itself gives reduction peaks at -0 4 I V and -0 97V A detection limit of 0 4 ngIml could be obtained [307] Similarly, at the hanging mercury drop electrode, adsorptive stripping voltammetry of the complex of Se with 3,3'dianiinobenzidiile in the differential pulse mode enabled the detection down to 0 2 pg/l [295] For the analysis of effluents from the mining industry, Se is extracted with o-phenylenediamine from ammonium perchlorate buffer pH=9 with toluene After addition of acetone and NH4CI/HCI - buffer pH1 3 , a homogenous solution with sufficient electrolytical conductivity is achieved to enable the polarography of the extract with a dropping Hg electrode Under conditions of anodic polarization, only one polarographic wave of Se is achieved, which yields niaximuin current at pH I 3 , leading to a detection limit of lygll in the original solution Prioi to the extraction, C u was removed by extraction with dithizone, and large excess of Fe was masked with EDTA [308]

3.5.4 ELECTROCHEMICAL DETECTORS IN LIQUID CHROMATOGRAPHY Liquid chromatographical separation on a cation exchange resin with on-line detection of Se at a tubular Au- electrode helps to cope with interferences of codeposited metals in anodic stripping voltammetry When Se is eluted from the column, the potential I S set to -0 30V, and the reduction current is monitored The electrode surface IS cleaned by rapid

440 cycling between -0.30 and +1.20 V, and kept inert during regeneration of the ion-exchange column at +1.30 V [92]. Selenols, diselenides, and selenenyl sulfides could be detected by group- selective electrochemical reactions after reversed phase liquid chromatography at two dual Hg/Au amalgam electrodes in series in the eluent stream. The upstream electrode is set at - 1.1OV VS. Ag/AgCI to reduce disulfides, diselenides and seleny sulfides, or at -0.5SV to reduce only the Se- compounds. The selenols are detected directly by facilitation of the oxidation of Hg from the downstream electrode, when it is set to +O. 15 V vs.Ag/AgCI. Diselenides and selenenyl sulfides are determined by first reducing them to the selenol and/or thio form at the upstream electrode followed by detection of the selenol and/or thiol at the downstream electrode [90].

3.6 RADIOCHEMICAL AND NUCLEAR METHODS

3.6.1 GENERAL As can be seen from table 1 , activation with neutrons leads to isotopes with overall soft b-lines. For their detection, the NaJ(TI)- detector is up to 10 times more sensitive, but lack

of selectivity requires chemical separation procedures. With the Ge(Li)- detector, direct counting after activation is possible in some cases [9]. --Tables 1 and 2

3.6.2 ACTIVATION WITH THERMAL NEUTRONS AND PURELY INSTRUMENTAL DETECTION OF 75-SE After sealing in polyethylene or quartz vials, 30h irradiation with thermal neutrons and 2-3 weeks storage, 75-Se could be determined by y- counting at 265 keV with a Ge(Li) detector. In particulate matter from riverine and marine waters [3 171, in atmospheric particulates [ 145,3181, coal, fuel oil and fly-ash [319], detection limits of 25 ng abs. [313] resp. 10 ng/g [145,3 181 were achieved. For biological materials, like tissues, blood plasma, serum and erythrocytes, even 5 days of irradiation, 6 weeks of cooling, and half an hour counting time with a Ge(Li)- detector were necessary [3 19,320,3211. At 264.6 kEv, 182-Ta at 264.6 keV cannot be discriminated from selenium because of its quite similar half- life [3 131, but this is not crucial for biological matrices. 3.6.3 ACTIVATION WITH EPITHERMAL NEUTRONS During activation with epithermal neutrons only, contrary to interfering concomitant elements, the cross section of 74-Se to yield the y- radiating isotope 75-Se, is not entirely lowered, which improves selectivity [32 I]. The production of 24-Na with epithermal neutrons is much smaller than with thermal ones, and the overall activity is about 20-fold less after Cd shielding [323]. Similarly, in epithermal neutron activation of silicate rocks and sediments, the interference of 181-Hf and 131-Ba at the 75-Se line at 136 keV is

44 I lowered, when epithermal neutrons are used [312]. However, at the 136 keV- line, the radiation from 99m-Tc at 140 keV, which is a daughter of Mo, is of equal sensitivity to the radiation of 75-Se at 136 keV [315]. The line at 265 keV is not observed in any sample because of insufficient detection limit [3 121. For the analysis of selenium in coal and fly ash, epithermal irradiation was found preferable, because of better precision, and equal sensitivity with respect to conventional y- irradiation method. Ash samples had to be irradiated for 1 day, and coal samples for 2 days. After 20 days of decay time, the 75-Se could be counted at 265 keV, achieving a detection limit of 0.1 pg/g [314]. In case of biological matrices, the improvement of using epithermal neutrons with respect to thermal ones was found to be negligible [321]. 3.6.4 METHODS OF ACTIVATION AND SUBSEQUENT CHEMICAL SEPARATION For counting the low-energy y- emission of 75-Se, chemical separation from interfering matrix elements allows to use the less selective, but more sensitive scintillation detector Addition of up to 1 OOmg inactive Se- carrier allows to do the clean-up on a macro- scale, and to control the regain of the analyte After coprecipitation or adsorption, active Se can easily be counted on the solid phase 3.6.4.1 Geological materials From acid digests of geological materials, Se was separated and counted on the solid after precipitation with sulfite [51,54,3 16,3241, with thioacetamide [325], with MnO, [IS], or sorbed on A&O, [326]. Interfering activated Ta was removed by coprecipitation with inactive Ta carrier from acid solution [326]. Alternatively, distillation as the bromide [54,57] as well as various extraction methods were utilized with activated samples [ 54,325,3271. Decomposition with alkaline fluxes after irradiation and addition of Se- carrier quantitatively yields soluble selenate, which is not precipitated along with hydroxides (Fe) or sulfides in alkaline sodium sulfide solution [325]. For large sample weights, Se and Te were extracted together with A u and Ag by fire assay in a flux consisting of lead oxide, soda, quartz, sugar, and borax, into newly formed metallic lead. Only this lead button was activated. The active material in matrix Pb is less hazardous to handle. However, 203-Pb emitted at 279 keV close to Se, and had to separated prior to counting. After dissolution in nitric acid, Se was reduced to the elemental state with hydroxylamine, and counted on filter [328,329]. 3.6.4.2 Biological samples After irradiation with thermal neutrons and 1-2 weeks of cooling, inactive carrier I S added If the samples are burnt in a stream of oxygen, Hg and Br come along with Se, which may interfere in radioactive counting This has to be also considered after distillation as bromide [309,333] Se was separated from Hg and Br by solvent extraction [327,330], adsorption on charcoal [64], or 82-Br was just allowed to decay within 20 days [I21 From acid digests of activated biological samples, Se was separated by liquid-liquid extraction as piazselenol [9,105], as carbamate [327], or as iodide into CC14 [53] Direct counting was possible after sorption as the dithizonate on carbon powder at pH=8, or after coprecipitation with Fe- hydroxide in presence of ascorbic acid [332]

442 3.6.4.3 Water samples Water samples were reduced with ascorbic acid at pH=2, and Se was adsorbed on active charcoal, which was finally activated and counted [ 1361. 3.6.5 USE OF THE SHORT-LIVED 77M-SE 76-Se can be activated to yield the short-lived 77m-Se with a half-life of only 17.5 seconds. No separations are possible, but nearly no handling is necessary. 28-AI, together with Ti, Mg, V, Mn and U, is also accessible to short time activation, decaying with a half-life of 2.3 min at 1779 keV [334]. In geological and coal matrices, short-lived Se can be interfered with 28-AI, causing high dead times and high background [313]; therefore, 77m-Se activation has been mainly used for biological and related matrices. Short activation of 77m-Se minimizes matrix activation. The main sources of error are the variation of the neutron flux [335], and exact timing. Biological samples were simultaneously treated with aqueous standards in cycles of 20 sec irradiation, 3 sec waiting, and 18 sec counting at 162 keV with a Ge(Li) detector The delay of 3 sec between irradiation and counting periods was necessary to decay 38m-CI with its half-life of 0.7 sec [336]. Food samples were irradiated for 20 sec, decayed for 20 sec, and counted for 20 sec at I62 keV for 77m-Se. The precision significantly improved by recycling the samples up to 4 times [337,338]. 200 mg of biological material was irradiated for only 2-4 seconds in a high flux of 10'' n/cm2.s, and measured I 5 sec after the end of the bombardment at 162 keV, thus achieving a detection limit of 5ng/g [335]. Similarly, for determination of Se in liver tissue, samples and liver reference standards were irradiated separately and in fixed sequence for 90 sec. The integrated value of the current neutron flux was determined by use of gold monitor samples [339]. In atmospheric particulates, a detection limit of I8 ng Se and a precision of 13% could be achieved after 60 sec irradiation with 5.10'' n/cm'.s, 5 sec delay and 30 sec counting [3 131. Alternatively to thermal neutrons, the 77m-Se could be also produced by 2 min photon activation from a 5-10 kCi co-60 source, but the detection limit of 1 mg Se in a pellet of 20g of animal food stuff was rather poor [340]. 3.6.6 PHOTON ACTIVATION With high energy photons of about 35 MeV , two nuclear reactions with Se- isotopes can occur [341,342,343,344]: 76-Se (y,n) 75-Se 120 d 265 keV/ 136 keV 82-Se (y,n) 8lm-Se 57.3 m 103 keV High energy photon activation has been applicated to determine Se in the multi-element analysis of river sediments, atmospheric particulates, and related materials. The samples were pelletized with Li,SO,, encapsulated in Al for neutron capture, and simultaneously irradiated with standards.After irradiation, the Al-foil had to be discarded because of to reduce background from 27-Al(n,y)24-Na [344]. Similarly, atmospheric particulates on polystyrene filters were pelletized along with elemental and flux monitors, and submitted to multielement photon activation with Ge(Li) detection [345]. Soil and fly-ash samples were irradiated, cooled 1 day, pelletized with cellulose powder, and finally measured with a planar intrinsic Ge-diode [343]. At 136 keV,

443 the limit of practical determination was estimated to be 0 2 pg/g, there was some interference by 57-co

3.7 X-RAY - SPECTROMETRIC METHODS FOR T H E DETERMINATION O F SELENIUM 3.7.1 GENERAL For analytical purposes of Se, the K a - emission lines are exclusively used. They hardly coincide with neighbouring elements, or with L - lines of heavy elements (see table 3). Table 3 Wavelength and energy of some X-ray lines [346] Element line wavelength energy keV As Ka2KL2 1 1799 10 508 KalKL3 11759 10 544 Pb LalL3M5 1 1750 10 552 Bi La2L3M4 I 1554 10 731 LalM3L5 1 1439 10 839 Ge KR3KM2 1 1294 10 978 KBlKM3 11289 10 982 Se Ka2KL2 1 1088 I I 181 KaIKL3 1 1048 I I222 As Kfi3KM2 I0578 I 1 720 KDlKM3 I0573 I I 726 Br Ka2KL2 1 1 878 10438 KaIKL3 10397 I1 924 In the wavelength dispersive mode, an overlap between the Se KOL2,3 line, the second order of the Hg L l M 2 and the Hg L2M4 line [347] is reported. 3.7.2 WAVELENGTH

- DISPERSIVE XRF

The best efficiency of Se-Ka line excitation is obtained with a Mo-target tube, which I S used in most cited cases, but continuous background of bremsstrahlung may be high, thus decreasing the accuracy Excitation with a Ag- target tube yielded nearly the same power of detection, but with much less background radiation [348] As a compromise for the determination of 12 elements in pharmaceuticals after coprecipitation enrichment, a Cr-tube has been used [147], which allows to detect S e by a factor of 2-3 less with respect to a Mo-tube The excitation spectrum is usually discriminated with a 100-LiF- crystal, at an angle 26 = 31 87" After excitation with a Mo-target tube, backgiound correction on both sides of the Se-peak was done at 3 I 10/32 60°[ 1441, or at 3 I 0132 74"[349] In case less than 10 mg of solid per cm' are available for final determination, a correction for non-infinite sample depth has to be applied [350] For water samples, Se can be collected by coprecipitation and sorption methods, and the resulting uniform solid finally counted on the membrane filter As collectors, ZnS or CdS freshly prepared on a membrane filter, were successfull at pH 3-4 Coprecipitation with

444 subsequent XRF- measurement has been done with polyvinylpyrrolidon/ thionalide at pH 4 [152], with diethyldithiocarbamate at pH 4.7-5.0 [150], or with combined dibenzylammonium and Na- dibenzyldithiocarbamates at pH 2.5 - 8.0 [ 1461. Similarly, from acid digests of biological materials and pharmazeutical preparations, Se was coprecipitated with the dibenzylaminosalt of dibenzyldithiocarbamic acid at pH 4 *I [147], or with Te [348,349]. Nitric acid interferes with the Te- reduction, and has to be reduced beforehand. Diethyldithio-carbamate as precipitating agent is not suitable in presence of large excess of Fe, which is also precipitated and can dominate the spectrum [349]. For the selective XRF- measurement of Se together with As and Sb, the corresponding hydrides were evolved from hydrochloric acid samples, and adsorbed on filter paper impregnated with AgNO,, which could be directly submitted to XRF measurement [347]. 3.7.3 ENERGY - DISPERSIVE XRF For Se, the line at 11.21 keV is used throughout, after Mo-K- excitation. Most authors prefer peak area evaluation, and background correction. Sample preparation procedures are not entirely different to wavelength dispersive methods. In soils and atmospheric particulates without chemical pretreatment, however, the detection limit of about 3 pg/g is not sufficient [350]. To reach the common geochemical level, wetchemical separation from the matrix and reductive coprecipitation with Te has been proposed [ 1281. Se in serum and whole blood could be directly counted after drying a 0.75 pI sample on a filter spot. The peak intensities were corrected according to the intensity of the back-scattered Mo K X-rays in the range 14-19 keV. As detection limit, 60 ng/ml in 100 sec was reached (normal value: 80-120 ng/m1)[351]. For the determination of Se in water samples, elementary Se was adsorbed on active carbon after reduction of selenite with ascorbic acid. Both Selenate and selenite were reduced by refluxing with thiourea in sulfuric acid. For excitation of Se in the active carbon, a W-target tube was used, and the detection limit could be extended to SO ng/l by an extraordinary long counting time [134]. 3.7.4 X-RAY SPECTROMETRY IN TOTAL-REFLECTION GEOMETRY Total reflection of the incident beam yields excitation only at the sample surface [353,358]. Aerosol samples could be excited directly in the total reflection geometry, but digestion or low-temperature plasma ashing was preferred to obtain a more representativ sample [359]. For serum, 30 pl samples were pipetted onto a Si X-ray mirror, spiked with Ge as internal standard, and air dried. A W- target tube was used to excite a Ni- secondary target, yielding a narrow band-pass of primary X-rays (K-edge at 12.66 KeV), to excite selective Se at 13.0 and 13.3 keV, while at the same time preventing the excitation of a relatively high concentration of Br (K edge at 13.475 keV) in the serum [358]. 3.7.5 PROTON - INDUCED X-RAY EMISSION (PIXE) The determination of Se by PIXE has only been reported for biological matrices yet. If biological specimen directly interact with the proton beam, they should be placed on pure backing thin targets backed with nuclear graphite, to avoid problems with the stability of

445

the shape of the sample. When the proton current exceeds 200 nA, evaporation of Se, Ca, As and K occurs [352]. The depth of penetration of exciting protons into the sample decreases with increasing average atom number of the sample, and ranges from 0.5 - 50 pm [353]. To reduce intense low-energy radiation, an Al- absorber of 78 pm thickness was placed in front of the detector [354]. By direct excitation with protons of 1.7 MeV, 0.08 pg/g Se could be detected [352] in a biological matrix Blood serum was dry ashed at 60°, ground, mixed with 20% graphite powder and Pd as internal standard, and formed to a thermally stable pellet. As the detection limit, 10 ng/g was found for 100 min counting after excitation with 1.8 MeV protons, or after 30 min counts after excitation with 4-MeV protons [ 3 5 5 ] . Alternatively, freeze- dried tissue and plant samples were converted into a fine powder, doped with Ag as internal standard, and fixed at the target frame by means of 1% solution of polystyrene in benzene [356]. From acid digests of biological matrices, Se can be selectively separated by coprecipitation with Te, and excited with 1.8 MeV protons, which leads to a detection limit of about 3 ng Se [42,354,357].

3.8 CONVERSION TO SE(IV)

In potable waters as well as after oxidational digestion, a substantial part of Se is present as selenate, which is advantageous because it is far less volatile and adsorbable on solids. Prior to most methods involving chemical reactions, Se has to be converted to the quadrivalent form. Besides the conversion reaction, excess nitrous oxides and chlorine have to be driven off, which interfere with subsequent hydride formation, colour reactions, reductive precipitation, electrochemical reactions etc.. In addition to interferents present in the sample, incomplete recovery is possible due to incomplete reduction of Se(VI), further reduction to Se", or volatilization. Losses due to volatilization or precipitation at vessel walls can be traced with 75- Se labelled compounds Most authors use reduction reaction of selenate with hydrochloric acid: H,SeO, + 2 HCI = H,SeO, + H,O + CI, Different optimum conditions found in the literature depend on the acid mixture used, the amount of sample to be oxidized, as well as on residual organics and cations (Fe) in the sample (see table 4). Remaining nitrite, which may interfere further, can be destroyed by addition of hydroxylamine, sulfanilamide,or amido-sulphuric acid [ 18 1,360,3611. Dry-ashing with Mg(NO,), and dissolution with 6M-HCI on the boiling water bath yields selenite and complete oxidation of organics, which renders a conversion step unnecessary. Besides heating with HCI, conversion was also achieved by UV- irradiation at pH> 7.5

W I

446

Table 4. Quantitative conversion to selenite in HCI min "C [HCII HCI only 180 20 3M [362] 30 105 4M [363] , 30 85 5M [363] I, 30 65 6M [363] potable water 5 boiling 4M (1811 , 6M [210] dil. KMn04 8 acid digests 30 95 5M [364] 30 SM [365] 90 15 95 SM [6,366] 10 80 6M [I871 15 80 6M 1181 4 boiling 4M * [205] 10 boiling 4M [I881 20 boiling 6M [368] 30 boiling 6M [I991 120 boiling water bath 6M [471 20 boiling water bath 6M [411 Cont.flow S heated coil 7.SM [I981 $9

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Table 1) NUCLEAR REACTIONS OF SELENIUM ISOTOPES [54,309,310] Isotope 74-Se 76-Se 78-Se

Natural abundance 0.87 YO 9.02 YO 23.52 YO

80-Se

49.82 %

82-Se

9.19 YO

thermal neutron cross section 30 barns 21 0.33 0.05 0.5 0.08 0.004 0.04

Some reactions with epithermal neutrons: 77-Se 7.65 % (n,p) 78-Se 23.52 YO (n,2n) 80-Se 49.82 YO 0.04 (n,cr) 82-Se 9.19 % 1.5 (n,2n)

product halfprincipal isotope life y-lines (keV) 75-Se 120.4 d 12 11136/265/279/401 77m-Se 17.5 s 161 79m-Se 3.91 m 96 weak 79-Se 65000 a no y 8 1-Se 18.6 m 280 weak 8 1m-Se 56.8 m 103 weak 25 m 360/520/830/13 10 83-Se 83m-Se 70 s 650/10 10/2020

77-As 77m-Se 77-Ge 81m-Se

39 h 17.5 s 11.3 h 56.8 m

Table 2) NUCLIDES INTERFERING WITH T H E y-RADIATION EMISSION SPECTRUM OF 75-SE 75-Se line keV 121.1

135 9

264 5

275 5 400 7

interferent line keV Eu 121 8 Ba 1 2 7 7 Hf 133 I

Yb Mo Ta Cd Hg

1307 1404 264 1 2609 2792

A u 411

resulting nuclide 152-Eu 131-Ba 181-Hf

half-I ife

169-Yb 99-TC 182-Ta 115-Cd 203-Hg

318d 66 7 h

l98-A~

27d

12.7 y 11 5 d 42 5 d

115 d

53 5 h 46 6 d

g Se/g interferent

Ref thermal n epith n 97 22 [3111 02 02 [3111 7 23 [3111 [3 121 [3131 24 17 [311] [3 I4,3 151 15 3 7 [311,313] 06 [3111 04 5 I 131 11 [3161 [91


459

SELENIUM OCCURRENCE AND ECOLOGY

1. SELENIUM IN T H E ATMOSPHERE For many urban areas in the Northern hemisphere, total selenium contents in the range of 0.3-20 ng/m' have been reported. In remote areas, Se goes down to values lower than 0.1ng/m3, the minimum has been measured on the South Pole (0 007 ng/m')[l-6]. The ChacaltaydBolivia has been taken as representative for remote locations in the southern hemisphere, the sampling station in the Canadian Northwest territories for the background of the northern hemisphere, and the Jungfraujoch/Switzerland for the background of Western Europe [7]. Notably, the arctic aerosol at Alert Bay /Northem Canada is somewhat higher in selenium than the aerosol at the South Pole [2]. An unusual high value has been reported for Ankara in winter 1976 [ l ] With respect to the average abundance in the earth crust of 50ng/g [8], Se in aerosols is generally enriched, even at its lowest values at the South Pole Enrichment of selenium is often estimated via the proportion over Fe in the atmosphere, with respect to SeRe of the average crustal abundance ("enrichment factor"). Enrichment factors of the aerosol in the range of 600- 10000 with respect to the surrounding soils are reported, which are quite high. In Ankara, in spite of the top values for the aerosols, the enrichment factor is lower, because Se in the soil is rather high [ I ] . Similarities of occurrence and chemical properties in coals and other fuels lead to the conclusion, that the output of SO,, which is responsible further for the acidification of rain, is accompanied by a parallel output of Se [6] (compare chapter 5. "Selenium in rocks"). Major sources of selenium input into the atmosphere are thus the combustion of fossil fuels (coal, oil), but also the processing of non ferrous metals (roasting of sulfide ores), as well as municipal incinerators [S]. In urban areas, distinct seasonal cycles can be observed, with maxima in December and minima in June, because of coal or fuel oil combustion in winter time. Rooted plants as well as soil microorganisms volatilize Se from soil (see chapter 7 ) In the background aerosols, however, there are no seasonal cycles [ 4 ] . Maximum to minimum concentration ratios are only at 3-4, whereas they are 30-40 for typically crustal elements [4] In Glasgow, however, a seasonal maximum in summer has been found, contrary to most other trace elements [3]. Selenium in western atlantic precipitation strongly correlates with acid concentration (0.959/n=15), and also with non sea-salt sulfate. Lack of correlation between Se and Na means, that it does not derive from marine sources [9] In the aerosol of St LouisiUSA, 75% of Se- contents can be statistically explained by a factor marked by sulfate as main component [lo]. Contrary to Pb, Br etc., traffic is no significant source of Se input into the atmosphere [ I ] . The residence time of aerosol particles depends on the aerodynamic diameter, which results from source, age and history of the particles [ 7 ] .High enrichment of Se in the atmosphere can be explained by the fact, that Se enters the atmosphere mainly as condensation aerosol and is thus present in very fine particles, or even gaseous During sampling on dry filter media, 19-35% of the selenium got lost, irrespective of the filter material, either because of volatility, or of small size of the particles [ 1 I ] This delays precipitation in comparison with "conservative" elements. like Fe or Al, which emanate mainly from soil abrasion. Enrichment of Se in the atmosphere is thus a result of increased volatility and slower precipitation. Se is removed from the atmosphere by processes such as wet deposition, through 11s association with sub-micrometre aerosols, and the efficient scavenging of these small

460 particles during precipitation events [9]. Removal by wet deposition predominates [ 9 ] , though the Se contents of rain water is lower than many surface waters Atmospheric deposition is an important input to oceanic surface waters [9]. Similarly, increasing precipitation correlates strongly positive with the selenium contents in humus layers of Norwegian forest humus soil or farmland soil [ 121. In Glasgow, however, the daily average weather conditions do not significantly influence the Se contents of the aerosol [3]. Only few papers deal with the speciation of selenium in aerosols. In rainwater samples from Japan and coastal California, selenite is the major Se- species. Rain and snow in Belgium contain variable quantities of selenite and selenate. In the precipitation of the Western Atlantic, however, the ratio of Se(IV)/Se(VI) lies within the narrow range of 1.26*0.95 [9]. Selenide and elemental selenium have been found in a laboratoy atmosphere [6], but they are negligible for outdoor sampling sites [9]. SeOz and Se" are released during fuel combustion, but also natural emissions of organic selenium compounds, mostly dimethylselenide, are known [9] (see also chapter 4. "Selenium in soils"). During the combustion of coal, the generated SO2 may partially reduce SeO, to elemental Se, which preferentially deposits on aerosol particles. On the other hand, oxidants on the troposphere, e.g. 0, or H,O,, may cause the production of selenate, which may serve as an indication reaction for strong oxidants [ 6 ] .

2. SELENIUM INNATURAL WATERS 2.1 TOTAL CONTENTS The selenium contents of natural waters can vary within a broad range, from 200mV (important for potable water resources), or at

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