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ACIDIFICATION RESEARCH: EVALUATION AND POLICY APPLICATIONS

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Studies in Environmental Science 50

ACIDIFICATION RESEARCH: ETVALUATION AND POLICY APP LICAT10NS Proceedings of an International Conference, Maastricht, The Netherlands, 14-1 8 October 1991

Edited by

T. Schneider Rijksinstituut voor Volksgezondheid en Milieuh ygiene (RIVM), 3720 BA Bilthoven, The Netherlands

ELSEVIER Amsterdam - London - New York - Tokyo 1992

ELSEVIER SCIENCE PUBLISHERS B.V. Molenwerf 1 P.O. Box 21 1,1000 AE Amsterdam, The Netherlands

@ 1992 Elsevier Science Publishers B.V. All rights reserved.

No part of this publication may be reproduced, stored in a retrieval system or transmitted in any form or by any means, electronic, mechanical, photocopying, recording or otherwise without the prior written permission of the publisher, Elsevier Science Publishers B.V., Copyright and 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 from the CCC about conditions under which photocopies of parts of this publication may be made in the USA. All other copyright questions, including photocopying outside of the USA, should be referred to the publisher.

No responsibility is assumed by the Publisher for any injury and/or damage to persons or property as a matter of products liability, negligence or otherwise, or from any use or operation of any methods, products, instructions or ideas contained in the material herein. This book is printed on acid-free paper. Printed in 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 34 35 36 37

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. Jergensen Trade and Environment: A Theoretical Enquiry by 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. MBszbros 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. Jergensen and I.Johnsen Disposal of Radioactive Wastes by Z. Dlouhq 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 by Nitrogen Oxides edited by T. Schneider and L. Grant Environmental Radioanalysis by H.A. Das, A. Faanhof and H.A. van der Sloot Chemistry for 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. VeziroQlu Chemical Events in the Atmosphere and their Impact on the Environment edited by G.B. Marini-Bettblo 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. Nhdasy and Y. Andriska Principles of Environmental Science and Technology (second revised edition) by S.E. Jorgensen and I.Johnsen 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 the Natural Environment by H. Schreier

38 39 40 41 42 43 44 45 46 47 48

49

H o w 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 i n 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. Privarovh 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. BBr Waste Materials in Construction edited by J.J.J.M. Goumans, H.A. van der Sloot and Th.G. Aalbers Statistical Methods i n Water Resources bv D.R. Helsel and R.M. Hirsch

vii

Fomword SESSION A

xii Opening session

1

Acidification: a n international problem J.G.M .Alders

3

Acidification as an example of the link between science and policy G.J.R.Wolters and H.Marseille

7

Acidification research and policy in the province Limburg H.W.Riem, B.R.Pasma and D.van Nierop

SESSIONB

State-of-the-artof acidificationresearch

17

25

Forest vegetation and acidification: a critical review RSchlaepfer

27

Global environmental change: implications for acid deposition research D.J.Waters and P.G.Whitehead

45

The role of ammonia in acidification: perspective for the prevention and reduction of emissions from livestock operations A.A.Jongebreur and J.H.Voorburg

55

Emissions of acidifying components M.Amann

85

Acidification of forests and forest soils: current status E.Matzner

77

Stress combinations in forests J.L.Innes

87

Effects of increasing nitrogen deposition and acidification on heathlands J.A.Lee, S.J.M.Caporn and D.J.Read

97

The interaction of forest vegetation and soils with the aquatic environment; effects of catchment liming on lakes T.R.K.Dalzie1, G.Howells and R.A.Skeffington

107

...

vlll

Higher order effects L.Reijnders

127

Acidifying effects on groundwater JSoveri

135

Monitoring for the future: integrated biogeochemical cycles in representative catchments T.Paces

145

The critical loads concept for the control of acidification J.-P.Hettelingh, R.J.Downing and P.A.M.de Smet

161

SESSIONC

175

AcidXcation policy

Canadian acid rain policy S.Milburn-Hopwood and K.J.Puckett

177

Acidification policy in Finland E. Lumme

185

Acidification policy - control of acidifying emissions in Germany B.Schtirer

191

Acidification policy in Hungary E .KovBcs

203

Acidification abatement policy - The Netherlands experience G.J.A.Al and V.G.Keizer

211

The convention on long-range transboundary air pollution: its achievements and its potential H.Wuster

221

Acidification policy - Sweden K.LGvgren, G.Persson and E.ThornelSf

241

Air pollution control policy in Switzerland B.C.Galli Purghart

247

Acidification research: evaluation and policy applications; a United Kingdom policy response R.G.Derwent and R.B.Wilson

253

Acidification policy in the United States D.Leaf

257

ix

SESSIOND

New mstxmh d t s on the acidBcation problem

263

Setting priorities for the measurement of acid aerosols and gases: 3 examples for Switzerland P.A.Alean-Kirkpatric and J.Hertz

266

High resolution assessment of acid deposition fluxes W.A.J.van Pul, J.W.Erisman, J. A.van Jaarsveld and F.A.A.M.de Leeuw

277

Measuring and modelling atmospheric dry deposition in complex forest terrain G.P.J.Draaijers, R.van Ek, W.Bleuten and R.Meijers

285

The transplantation of four species of Lobaria lichens to demonstrate a field acid rain effect A.M.Farmer, J.W.Bates and J.N.B.Bel1 Acidification research activities in Poland W.A. Mill

301

Critical loads for Dutch forest soils W.de Vries, J.Kros, R.M.Hootsmans, J.G.van Uffelen and J.C.H .Voogd

307

Scenario analysis with the Dutch Acidification Systems (DAS) model; an example for forests and forest soils A.Tiktak, A.H.Bakema, K.F.de Boer, J.W.Erisman, J.J.M.van Grinsven, C.van Heerden, G.J.Heij, J.Kros, F.A.A.M.de Leeuw, J.G.van Minnen, C.van der Salm, J.C.H.Voogd and W.de Vries

319

Acid rain abatement in Belgium: lessons of cost-effectiveness studies C.Cuijpers and S.Proost

341

Base content in soil and problems arising in connection with acidification L.Werner

349

Measurements of tree growth and health in the Liphook Forest Fumigation project: an evaluation of large scale open air fumigation experiments M. R. Holland and P.W. Mueller

357

SESSIONE

Results &om national research programmes

The United States national acid precipitation assessment program P.M.Irving

363

366

X

The United Kingdom research programme and its implications for policy, now and in the future R.B.Wilson

375

Research into forest decline and air pollution in France; major findings and relevance for policy applications G.Landmann

383

Background, results and conclusions of the Dutch Priority Programme on Acidification G.J.Heij and T.Schneider

397

Acidification research in Sweden H.Staaf and U.Bertills

415

Finnish research programma on acidification (HAPRO) 1985 - 1990 P.E .Kauppi

431

Status of acidification research in Czechoslovakia and its relationship to politics and economics in Europe T.Paces

443

The Swiss national research program “Forest Damage and Air Pollution” F.Haemmerli, N.Kriiuchi and MStark

449

SESSIONF

461

Closing seasion

A comparison of some national assessments J.Nilsson and E.Cowling

463

po6TFIRSESSION

519

The Dutch Acidification Systems (DAS) model: the emissions and air transport modules K.F.de Boer, J.W.Erisman, F.A.A.M.de Leeuw, T.N.Olsthoorn and R.Thomas

521

The relationship between research and policy on acidification impacts in the nature conservation agencies of Great Britain A.M.Farmer

523

The Dutch Acidification Systems (DAS)model: the forest module SOILVEG J.J.M.van Grinsven, C.van Heerden and J.G.van Minnen

525

A modelled assessment of critical load exceedences for sulphur over the United Kingdom G.W.Campbel1 and J.G.Irwin

527

xi Dry deposition over grassland: seasonal influences, chemical equilibria and surface wetness M.A.H.G.Plantaz, A.T.Vermeulen, P.J.de Wild, G.P.Wyers and J.Slanina

529

The Dutch Acidification Systems (DAS) model A.Tiktak, A.H.Bakema, K.F.de Boer, R.M.Kok and T.N.Olsthoorn

531

Long-term impact of three deposition scenarios on Dutch forest soils W.de Vries, J.Kros, C.van der Salm and J.C.H.Voogd

533

Soil and soil solution composition of 150 forest stands i n The Netherlands i n 1990 W.de Vries, E.E.J.M.Leeters, C.M.Hendriks, W.Balkema, M.M.T.Meulenbrugge, R.Zwijnen and J.C.H.Voogd

535

Automated denuder systems for dry deposition studies of acidifying compounds G.P.Wyers, A.T.Vermeulen, R.P.Otjes, A.Wayers, J.J.Mo1 and J S l a n i n a

537

AQUACID: modelling the acidification of shallow heathland lakes in The Netherlands; the aquatic systems module of DAS F.G.Wortelboer

539

An international research program on acid rain and emissions i n Asia W.K.Foel1

541

Deposition of acidifying compounds D.Fowler, J.N.Cape, M.A.Sutton, R.Mourne, K.J.Hargreaves, J.H.Duyzer and M.W.Gallagher

553

List of plu-ticipanb

573

xii

The International Conference on ACIDIFICATION RESEARCH, EVALUATION AND POLICY APPLICATIONS, organized by the Ministry of Housing, Physical Planning and Environment, was held in Maastricht, The Netherlands, from 14 - 17 October 1991. A first Conference of this kind was held in Amsterdam, from 5 - 9 May 1986. During that Conference 24 official delegations from ECE member countries discussed the available results of Acidification Research Programmes and Projects and evaluated these results with regard to the implications for Policy Actions. In 1986 already a number of results indicated the need to study and describe the (mostly negative) effects of acidification. More substantial evidence for the relative importance of acid deposition in the whole field of environmental stresses on the endangered ecosystems or environmental compartments however, surfaced during the execution of nationwide coordinated research programmes. Examples of such a programme could or still can be found in the USA,Canada, Finland, France and The Netherlands. A number of individual research results have been reported and discussed already in the Glasgow Conference held in 1990. The relation between scientific results of integrated research programmes and policy actions to prevent, reduce and limit the widespread damage caused by acidification was, however, not presented as yet. In cooperation with the members of the MARC group (Meeting of Acidification Research Coordinators) the Organizing Committee therefore suggested a programme for the Conference containing several types of presentations: - thematic reviews on specific topics of acidification research by invited key speakers; - summaries of national research programmes by programme coordinators; - overviews of acidification policy plans and actual abatement programmes by national or supra-national representatives; - selected poster presentations or short papers on recent research results. The Proceedings of the Conference contain the opening statements by J.C.M.Alders, Minister of Housing, Physical Planning and Environment; G.J.R.Wolters, Conference Chairman and W.H.Riem, Deputy Commissioner of the Queen in the Province of Limburg. After the Opening Session three half day sessions were devoted to the stateof-the-art presentations of key elements within the Acidification Research Programmes. In parallel sessions these were followed by presentations on: 1. Acidification Policy by national representatives 2. Results from National Research Programmes and 3. Recent Research Results on the Acidification Problem. During several days of the Conference also a Poster Session was held. A t this Session detailed information was given on results from specific (sub) items out of the research programmes.

...

Xlll

The final Session consisted of the presentation and discussion of a paper on “A comparison of some national assessments”. This comparison comprised an “all out effort” by two senior experts (Jan Nilsson and Ellis Cowling) in the field of Acidification Research. They succeeded in evaluating and comparing research results from a number of different national research programmes. A work of quality, carried out in a (relative) brief period before, during and shortly after the Conference. The editor would like to express here his admiration for this outstanding effort. The successful conduct of a meeting between research scientists and policy makers from a large number of countries, depends on the dedication of numerous individuals. As editor and Chairman of the Organizing Committee I would like to acknowledge here all those who have contributed to the organization of the Conference and the associated events. The preparation of the final programme, and the selection of the invited speakers, was carried out with the support of the members of the National Advisory Committee, listed in this volume. Members of this Committee also acted as chairmen of the individual conference sessions. I also would like to recognise the excellent work performed by my Organization Committee: Joop van Ham of SCMO-TNO who served as scientific secretary together with Bert Jan Heij who also organized the Poster Session. Ottelien van Steenis not only performed the secretarial work of the Committee but also formed with the excellent help of Nel Venis-Pols and Marianne Vonk the Registration and Information Centre, and last but not least she also took care of all the preparations for these Proceedings. I am grateful for the, as usual, excellent organization of the programme for accompanying guests by Mini Schneider. I would also like to thank the Burgomaster and Aldermen of the City of Maastricht for their welcome reception in the Town Hall of this most bourgondic town of The Netherlands. Finally thanks are due to the members of the MECC Conference and Hotel Services for their expert performance and good standard of Dutch hospitality. I hope that these proceedings, published by Elsevier Science Publishers in their usually rapid and professional way, will be used as a work of reference, both by research scientists as well as by policy makers.

T.Schneider Chairman Organizing Committee

xiv

CONFERENCE CHAIRMAN Ministry of Housing, Physical Planning and G.J.R.Wo1ters Environment

ADVISORY COMMITTEE G.J.R.Wolters, chairman J.van Ham, secretary G.J.Heij B.A.Kleinbloesem L.Koster L.Reijnders T.Schneider K.Verhoeff G.H.Vonkeman

Ministry of Housing, Physical Planning and Environment TNO Study and Information Centre for Environmental Research National Institute of Public Health and Environmental Protection Dutch Electricity Generating Board Shell Nederland B.V. Netherlands Society for Nature and Environment National Institute of Public Health and Environmental Protection Ministry of Agriculture, Nature Management and Fisheries Committee for Long Term Environmental Policy

ORGANIZING COMMITTEE TSchneider, chairman Mrs.O.van Steenis, secretary J.van Ham G.J.Heij PARTN'JZRSPROGRAMME Mrs.M.Schneider - Ferrageau de St.Amand scientxcsecxt2tariat

J.van Ham

Registration and Information Centre Mrs.O.van Steenis

SESSION A OPENING SESSION

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T. Schneider (Editor). Acidification Research. Evaluationand Policy Applications 1992 Elsevier Science Publishers 0 . V .

3

ACIDIFICATION: AN INTERNATIONAL PROBLEM J.G.M. Alders Minister of Housing, Physical Planning and Environment, P.O. Box 20951, 2500 EZ Is-Gravenhage, The Netherlands

Mr. Chairman, Ladies and Gentlemen, I would like to start by welcoming you here at this conference in Maastricht and I wish you a useful and pleasant stay. Undoubtedly, you are all familiar with the problems caused by acidification. These problems are emerging nowadays in all industrialized countries and have a wide variety of effects on our ecosystems. Therefore, acidification is an example of the type of problem that can be dealt with effectively only by international measures. Of course, national efforts are indispensible, but because of the transboundary aspects of the problem, only a co-ordinated international abatement strategy can effectively address this problem. Acidification is a complex scientific problem. A large number of specialist fields are involved and, hence, international cooperation is also needed in scientific research. A conference, such as this one, can play an important role in facilitating the exchange of the latest information on acidification: not only with respect to research results but also with respect to the development and implementation of abatement policies. It is very useful for all countries to be informed about each other's results: this prevents overlap in research and has a steering and stimulating effect. In the Netherlands a national research programme on acidification was set up to study the effects of the emissions of sulphur oxides, nitrogen oxides and ammonia. The most important results of the second phase of this programme are presented during this conference. The international aspects of acidification are obvious. In most countries the share of emissions from neighbouring countries in the total deposition is significant. Not only in small countries like the Netherlands, but also in large countries with relatively low emissions, a high percentage of the acid deposition can be traced to foreign origin.

4

Foreign contribution is for example very high in Norway, Sweden and the Netherlands. Spain and Great Britain however, are to a large extent responsible themselves for the SO,-deposition in their countries. The first step in an international abatement strategy was the United Nations Convention on Long Range Transboundary Air Pollution, signed in Geneva in 1979, in which countries declared their firm intention to reduce emissions that contribute to depositions in other countries. This convention, once considered a paper tiger, already provides a basis for coordinated SO,- and NO,-abatement. A VOC Protocol will be signed next month and there is still more to come. For use in international negotiations, the UN-ECE is developing the so called "critical loads concept". In this concept the maximum deposition of sulphur and nitrogen on a specific area is determined, below which no negative effects will occur: the critical load. To stay below this critical load, emission reductions of 80 to 90 percent are needed. In a step-wise approach towards this critical load, target loads will be determined. This concept creates an unique opportunity for an abatement strategy as effective as possible and will lead towards international agreements with differentiated obligations. From our experience with modelling, we know that dramatic reductions are necessary. Some people experienced this as a shock after years of belief that severe abatement measures could be avoided. Once more this is an example that a good judgement comes from experience, but that experience comes from bad judgement

.

It is not likely that the required reductions will be realized in the short term. A comparison of the emission reductions, needed to reach the target loads, with the reductions foreseen in the current policies of European countries demonstrate that considerable supplementary efforts are needed. At this moment we are confronted with a related problem that should get all of our attention. From the point of view of cost effectiveness, measures taken in countries in Central and Eastern Europe should have priority. However, most of these countries do not have the financial means to actually take these measures. The need for creative international solutions becomes more and more urgent and maybe we should adjust the "polluter pays principle" to the new situation. The countries in Western Europe should assist countries in Central- and Eastern Europe that lack both technical and financial means. The European Energy Charter offers a good possibility to turn this idea into action, if it is used to improve the energy infrastructure in the East-European countries.

5

Negotiations on reduction percentages will play an important role during the coming years: an international abatement scheme has to be developed. The considerable costs involved in reduction measures are increasing the hesitation to really take these measures. Hopefully this will be avoided by a step-wise approach. In view of the seriousness of the effects, this development of an international abatement policy cannot await still more scientific certainty. Yes, there are still gaps in our knowledge, but knowledge alone will not solve our problems. To conclude, I can state that the Netherlands is supporting the critical loads concept and has accepted the resulting necessary emission reductions. A first step to reach our critical loads has already been taken. I would like to invite other countries to do the same. Furthermore, it is clear that the West-European countries have to assist East-European countries in order to reach our common targets as soon and as effectively as possible. I thank you for your attention.

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T Schnetder (Editor), Acidification Research Evaluationand Policy Applications

0 1992 Elsevier Science Publishers E V All rights resewed

I

ACIDIFICATION AS AN EXAMPLE OF THE LINK BETWEEN SCIENCE AND POLICY G.J.R. Wolters (Conference Chairman) and H. Marseille Ministry of Directorate Directorate, Netherlands.

Housing, Physical Planning and Environment, General for Environmental Protection, Air P.O. Box 450, 2260 MB Leidschendam, The

Ladies and gentlemen, I am very pleased to welcome you all at this international conference about acidification research, evaluation and policy application. Modern environmental issues, like acidification, depletion of the ozone layer and climate change, ask for an approach in which a close link is made between science and policy. In the past, when the first environmental problems became evident, we only had the possibility to treat the problem by using a technology based approach. Emission reduction by applying best available technology was the most reasonable thing we could do. Nowadays our knowledge about the whole chain from emissions to environmental effects has increased in such a way that scientific information about the seriousness and urgency of the environmental problem can give us accurate guidance in policy development. This means that, if the problem is very urgent and application of best practicable technology is insufficient to avoid irreversible changes in ecosystems, science provides a sound basis for the additional structural measures that have to be taken eventually. For example, measures to change production processes or reduce energy use or car use could be applied. "No effect levels", or "critical levels and loads", can be used to establish a policy objective. The time scale in which the policy objective should be reached, depends on the sensitivity of the systems. If the ecosystems concerned have a large buffer capacity so that they can stand a higher level of pollution than the "no effect level" for a certain period, it could be possible to prolong the time scale. But with this we should be careful because scientific information always deals with uncertainties. Even if science tells us that an ecosystem is less sensitive than we thought, and even if we are not 100% sure of serious ecological impacts, this could never be an excuse to do less than use the Best Available Technology and to produce more pollution than necessary. Mr. Joris Al, who will speak on the acidification policy of the Netherlands (tomorrow

8 or Thursday), will explain this two-track approach, based on effect-oriented environmental quality objectives on the one hand and source-oriented Best Available Technology on the other hand. International agreements tend to reflect the close relationship between science and policy more and more. I only have to refer to the speech of Minister Hans Alders, who told you about the critical loads approach, being used for international negotiations about further reductions of nitrogen and sulphur emissions. In my opinion, the acidification policy in the Netherlands is a good example of the link between science and policy, and I will try to tell you in short how acidification policy has developed on a dynamic scientific and technical basis. The purpose of presenting these developments is to give you food for thought on the various ways science and policy may be linked with respect to acidification. Possibly our experience could be of help for policy development in other countries. I will especially try to clarify how we translated scientific information into policy goals and how we dealt with uncertainties and changing insights. Approach The process of setting policy goals for the acidification theme is given in fig. 1. The theme "acidification" was recognized in Dutch environmental policy from 1976 onwards. Swedish data about acidification of lakes had caused concern about the situation in the Netherlands, but in fact very little was known about the damaging effects of acid on ecosystems and the first document handling this theme did not set air quality objectives as a policy goal for reducing the problem. A big change however occured when photographs of dying forests in Germany appeared in our newspapers. Questions in parliament and pressure from NGOs led to the first deposition objectives in the Indicative Multi-year Programme for Air '84-'881, based on Swedish and Canadian studies of the acceptable levels of sulphur and sulphate deposition. The Swedish and Canadian critical loads were translated into a preliminary critical load of 1800 acid equivalents per hectare per year for combined deposition of SO,, NO, an NH,. Complete conversion of these compounds into acid was assumed and a rough estimate of soil sensitivity in the Netherlands was made. An emission reduction factor of 3 to 4 for total acidifying emissions on a European scale was calculated to be necessary to reach the critical load for the Netherlands. Based on the approach of equal emissions per capita, the necessary reductions of SO,, being rather mild, were assumed to be possible, but the necessary technology for reducing NO, and NH, emissions was very inadequate. Interim emission ceilings for SO, and NO, were established on the same level as the current emissions, and the intention of lowering the emission ceilings on a European scale was mentioned.

9 Start of the research programme The questions from parliament, asking for abatement measures and a research programme, also led to an overview of current scientific knowledge in the Netherlands and to the memorandum "the problem of acidification' ", after which a large national research programme, the "Dutch Priority Programme on Acidification", was started. It was financed by several ministries as well as by the Electricity Generating Board and the Oil Companies. The first phase of this programme lasted until 1988.

SOCIETY

POLICY GOALS

SCIENCE

Questions in Parliament1 Pressure from N G O s

--

International Data

Survey of Current Knowledge Acidification Problem ( ' 8 4 )

AIR '85-'89 ( ' 8 4 )

,d."

9"'

AcidificatiOn Research Programme I ( ' 8 5 - ' 8 8 )

~

'

--

Environmental Policy Plan ('89)/

Acidification Abatement Plan ('89)

4'

Fall o f

*

*l/

Acid Res Prog II

('88-'90)

,

Environmental Pollcy Plan PLUS ('90)

Acid Res

; ; ; : : ;p

\

Environmental Policy Plan # 2 ('93)

fig.1: flow chart with the interactions policy goals and society.

between

science,

10

In 1984, in the Indicative Multi-year Programme for Air '85'89,, the deposition objective was changed. On the basis of an extensive inventory of available scientific knowledge, nitrogen deposition below 1600 equivalents per hectare per year was no longer considered to be acidifying. With unaltered assumptions of sensitivity of the Dutch soils, compared to the Swedish, a deposition of 3000 equivalents, of which 1600 in the form of nitrogen, was now considered to be the critical load for our ecosystems. Emission reductions of 70% for SO,, 33% for NO, and 50% for NH, were calculated to be necessary for the Netherlands to reach this deposition level, under the condition of comparable SO, and NO, -reductions in the neighbouring countries. The first comprehensive abatement programme was drawn up. The abatement technology for ammonia still had to be developed. Little was known about ammonia emissions and a decision about additional reduction measures was planned for 1988, when final deposition objectives should be established. In 1987, the first results of the research formed the basis of the "Interim Evaluation of Acidification Policy in the Netherlands'". Ecosystems appeared to be more sensitive for acidification than formerly was assumed and the effect of the measures appeared to be overestimated. A number of additional measures was described to attain at least the original 1984emission reduction targets. Final deposition objectives After the final report of the Dutch Priority Programme on Acidification was completed in 19885, final deposition objectives were established in 1989 in the National Environmental Policy Plan6 and in the Acidification Abatement Plan'. Because of the high sensitivity of ecosystems, the available abatement technology was inadequate to attain the necessary emission reductions in a short period. For that reason a set of deposition objectives was developed, based on critical loads, with a different time scale (table 1). year

acid load (eq./ha,yr)

N-load (eq./ha,yr)

ozone (1 hr./gr.s) ( w / m 31

400

-

120/50

I2010

1400

1000

240/100

2000

2400

1600

240' /lo0

middle of next cent

I

max. number of exceedances per year: 2 table 1:

policy goals for the acidifying components and ozone in the Acidification Abatement Plan and the National Environmental Policy Plan, 1989.

11

For the acidifying components, a true critical load or no effect level of 400 acid equivalents was set as a target level, to be reached around the middle of the next century. This will require more than 95% reduction of emissions in all European countries, attainable only on the basis of completely new technology and fundamental changes in industrial production, agriculture and transport. An objective of 1400 acid equivalents was set for 2010, to be reached as a mean deposition on Dutch forests, in order to safeguard the forests against the most serious adverse effects. From this 1400, a maximum value of 1000 equivalents of nitrogen was established, to prevent nitrogen accumulation in soil and vegetation and to prevent exceedance of the target value for nitrate in groundwater used as drinking water. These objectives will require an 80-90% emission reduction in Europe, which will require extensive energy conservation, decrease of car use and structural changes in agriculture. Abatement measures will be decided upon in the second National Environmental Policy Plan, which will be drawn up in 1993. An interim target load of 2400 acid equivalents, including 1600 equivalents of nitrogen at most, was established for the year 2000. This value would protect, according to the scientific knowledge at that moment, the less sensitive parts of the ecosystems in the Netherlands, for example the continuous forest areas on richer soils. It would also prevent heathland from changing into grassland (with some management measures) and groundwater under forests and nature reserves would satisfy the standards set for the preparation of drinking water. Emission reduction objectives To reach the interim deposition objective for the year 2000, emission reduction percentages (relative to 1980 emission levels) of 80% for SO,, 50% for NO, and 70% for NH,, have been established for the Netherlands, based on Best Available Technology and additional structural measures such as energy conservation and reduction of car use (table 2). The measures are described in detail in the Acidification Abatement Plan. For the neighbouring countries comparable reductions for SO, and NO, are required, and a 25% reduction of NH,. EMMISSION REDUCTION OBJECTIVES (2000):

so*: NO,: NH,: VOC:

80% 50%

70% 60%

table 2: Emission reduction objectives for the acidifying components and VOC in the Acidification Abatement Plan and the National Environmental Policy Plan, 1989.

12

Though the uncertainty regarding current deposition and the formulated critical loads was rather great, the emission reduction objectives were rather easily accepted by all the emitter-categories. The gap between both was so big that for the period until the year 2000, the uncertainty was overshadowed by the necessary reductions to reach the critical loads. The present acid load and the development in the last ten years are presented in next table. 1980

1981

1982

1983

1984

1985

1986

1987

1988 1989

6800

6900

6500

6300

6400

6300

6200

5900

5000

table 3:

4800

Deposition of total acidifying components (acid equivalents per hectare per year) from 1980 to 1989. (source: RIVM)

The decline in deposition results mainly from the decrease in SO,-emissions in the Netherlands and neighbouring countries. However, meteorological conditions are responsible for fluctuations. Ozone

(see table 1 and 2)

In order to reduce the role of ozone in the effects of acidification on nature to such an extent that it becomes negligible, the ozone level, averaged over the growing season, should not exceed 50 pg/m3. The harmful effects on humans would be negligible when ozone levels would not exceed 120 pg/m3 as an hourly average. These values can be approached only during the course of the next century, requiring more than 9095% reduction of NO, and VOC emissions in Europe and a large reduction in carbon monoxide and methane emissions over the whole world. For that reason they are referred to as final goals. Most serious effects of ozone on nature and humans can be prevented at the values which are chosen as objectives for the year 2010: 100 pg/m3 averaged over the growing season and 240 pg/m3 as a maximum hourly average. Drastic structural measures f o r NO, and VOC are needed in an international context to reach those objectives. A s an interim objective for the year 2000, the same values are established, but the hourly average of 240 pg/m3 is alowed to be exceeded twice a year. Additional to a NO, reduction of 50%, VOC emissions will be reduced with 60% in order to reach the interim objectives f o r the year 2000.

13

Further developments

(see fig 1)

A political discussion about the structural measures in traffic caused the fall of Dutch cabinet just after the National Environmental Policy Plan was decided upon. The new government decided that the present plan was not strong enough and announced the National Environmental Policy Plan - "Plus' ' I , in which, one year later, the emission reduction objectives for SO, and NO, for the year 2000 were accelerated with several years. For NH, , extra attention to local peaks of deposition was paid by "object oriented policy", protecting sensitive ecosystems from extremely high ammonia loads. In April this year the second phase of the Dutch Priority Programme on Acidification was finished9. Important conclusions are that the deposition of NH, appears to have been underestimated and the NOI deposition overestimated. The contribution of ammonia and ammonium compounds to total acid deposition in the Netherlands is almost 50%. By the enlarged contribution of ammonia, the Dutch contribution to the deposition in our own country is also enlarged. It is now more than 50%, while we formerly thought it to be about 40%. This means that our policy of emission reductions will be more effective than we thought to reduce the deposition in the Netherlands, especially for ammonia. This is one of the reasons that, in spite of a higher present load than presumed, the prognosis for the year 2000 is that the interim deposition goal will be attained if the proposed abatement measures would have their expected effects on national emissions and if neighbouring countries would reduce according to expectations. However, a recently published evaluation study by RIVM concludes that the abatement programme is less effective. This was based on an analysis of the proposed measures in so far as effectuated hitherto or considered suitable to be effectuated well before 2000. Also the level of reduction in neigbouring countries in 2000 was concluded to be insufficient. Although part of the proposed national measures was not taken into account and further international abatement is still being negotiated, the conclusions of the study are alarming. Its policy implications will be thoroughly analysed. The new deposition figures also have implications for regional differences within the Netherlands. In regions which used to be relatively heavily polluted, acid deposition has increased more (compared to former insights) than in the less heavily polluted regions. So the problems in the southern half of the Netherlands, where deposition is relatively high, are greater than we presumed. In fig. 2, the deposition in the year 1989, according to the latest insights, is shown. The deposition varies from 3000 acid equivalents per hectare per year in the north to more than 8000 in the south. Another change in scientific insights concerns the values of the critical loads. A level of 2400 acid equivalents per hectare per year is no longer considered as a critical load for certain types of forests. For all types of forests on sensitive soils the critical load is now established to be about 1400 equivalents.

14

Model calculations show us that an abatement policy resulting in a deposition of 1400 equivalents as an average on forests in the year 2010 will in the long term give complete protection to the forests as far as the critical Al/Ca-ratio in the soil solution is concerned, and that irreversible effects like depletion of the aluminum hydroxide buffer will be prevented. Evaluation of the present acidification policy will take place in 1992 and will be partly based on the results of the research programme. But we can already conclude that the deposition objective of 1400 equivalents has a stronger basis now. A second National Environmental Policy Plan will be drawn up in 1993. This autumn the third and last phase of the Research Programme will start. The final results will be used after 1994.

fig. 2: Total potential acid deposition in 1989 (mol H+ ha- yr-’ ) (source: RIVM9 )

15

Conclusion Concluding I can say that, during the last eight years acidification policy has developed in a consistent way. Present measures are derived from the most recent scientific insights with respect to critical loads. It was a challenge for scientists to help the policy makers in setting environmental objectives. The interaction between science and policy has been very fruitful and stimulating to both sides. Though it takes some time to give answers to urgent political questions, finally it is very satisfactory to have a policy which is based on the best available knowledge. Changing scientific insights and uncertainties have not kept us from adopting far reaching emission reduction measures. Nowadays we are at the eve of deciding upon even more drastic measures, concerning structural changes in production, extra measures for regions which are heavily polluted and reduction of volume of various activities. Those measures will only be acceptable when grounded on a firm scientific basis. I hope that this international conference will promote scientific understanding and will contribute to further agreements about the reduction of emissions, not only in this country but also in the rest of Europe.

References 1. Parliamentary Documents 11, 1983/84, 18 100, nr. 7. 2. Parliamentary Documents 11, 1983/84, 18 225, nrs 1-2. 3. Parliamentary Documents 11, 1984/85, 18 605, nrs 1-2. 4. Parliamentary Documents 11, 1987/88, 18 225, nr. 22. 5. Dutch Priority Programme on Acidification: Summary Report Acidification Research 1984-1988, Publication 00-06, November 1988, RIVM. 6. Parliamentary Documents 11, 1988/89, 21 137, nrs 1-2, issued to Parliament May 25th, 1989. 7. Parliamentary Documents 11, 1988/89, 18 225, nr. 31, issued to Parliament July 20th, 1989. 8. Parliamentary Documents 11, 1989/90, 21 137, nr. 20, issued to Parliament June 14 th, 1990. 9. Dutch Priority Programme on Acidification: Final Report second phase Dutch Priority Programme on Acidification, report no. 200-09, G. J. Heij and T. Schneider (eds), April 1991, RIVM.

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17

ACIDIFICATION RESEARCH AND POLICY IN THE PROVINCE LIMBURG

H.W. Riema, B.R. Pasmab and D. van NieropC a Deputy of Environment in the Province of Limburg, P.O. Box 5700, NL 6202 MA Maastricht, The Netherlands and Department of Environment in the Province of Limburg, P.O. Box 5700, NL 6202 MA Maastricht, The Netherlands

Abseact Acidification is an important environmental problem in The Netherlands. Both national and provincial government have tasks in the prevention of further damage due to acidification and the reduction of emissions. In this paper it is shown how a policy on acidification is formulated in th e province of Limburg. The national deposition targets have been adopted. Feasibility studies were carried out to establish the emission reduction necessary to reach th e deposition targets. These studies also show what emission reduction is feasible in a cost-effective way and how cost-effectiveness decreases as reduction targets a r e set higher. Thus emission reduction targets were set and a provincial abatement strategy, including measures to be taken by various target groups, was formulated. 1.

NATIONAL AND PROVINCIAL ACIDIFICATION POLICY Provincial governments play an important role in Dutch environmental policy. Provinces a re administrative bodies, controlled by elected administrators, and take in an intermediate position between national and local government. Provinces have several tasks in maintaining air quality and emission control. One of t h e most important tasks is the issuing of licenses t o companies tha t cause air pollution. Furthermore provinces have to report to national authorities on (trends in) air quality. In this paper i t is shown how a higher scale (national) policy is used to formulate a lower scale (provincial, regional) policy. In figure 1 i t is shown how the province of Limburg is situated in The Netherlands.

Figure 1. Limburg is the southernmost Dutch province.

18 Acidification is an important environmental problem in the province of Limburg. Due t o high Ht and N-deposition levels severe damage is already visible in different parts of Limburg. The damage known a t present is listed in table 1. Table 1. Damage due to acidification in Limburg [ l ] Crops and wood (harvest losses) Buildings, materials, monuments Forest Heathlands, nature Heathland pools

DFL 30 million (13 million ECU) DFL 3,5 - 5 million (1,5 - 2 million ECU) 40 % of the forests is damaged or dying; in North-Limburg 75 - 90 % damaged > 60 % severely damaged 90 % severely damaged

The Limburg policy on acidification is of course based on the national acidification policy. The main strategic goal of this policy is the prevention of the most severe damage t o forests and nature in 2010. The national policy sets well defined deposition and emission reduction targets and has identified so called target groups. Target groups for acidification policy in t he Netherlands are: power plants, industry, agriculture and traffic. This approach is translated t o the provincial scale. The first s t e p in translating national policy to the provincial scale consists of determining the most important sources of acid emissions and the actual level of deposition throughout the province. The second step consists of a study of possible measures and their cost and effectiveness for each of t h e target groups. These studies give insight in the decrease of cost-effectiveness as t h e goals for emission reduction a re set higher. An economic evaluation of the policy showing costs and effects can thus be made. This paper describes t h e results of t h e application of this method in Limburg. 2.

MO!3 IMPORTANT SOURCES

The most important sources of acid emissions in Limburg a re shown in figure 2. I t is clear that power plants and industry a r e the most important contributors to SO2 emissions. For NO, th e largest sources a r e power plants, industry and traffic. NHY is mainly emitted by agricultural activities, while the most important VOC-source is traffic. Insight in the contribution of sources to t h e emission of the various components is needed for an effective abatement strategy. The large differences in contribution of various types of sources to emissions of acid compounds is obvious but the regional differences in contribution t o the acid emissions is also very important. In Limburg this regional differentiation is large. In the south of the province t he contribution of NO, is 62 %. In the north NH3 is the most important component: 46 %. In all of Limburg S02-emissions play a minor role. These figures a r e illustrated in figure 3.

19 Traffic 5%

SO,

;

28.9ktonly

Various 5%

NO, : 61.8kton/y Industry 7%

Various 21%

NH, : 19.4 ktonly

Traffic 49%

VOC : 2 4 . 7 ktonly

Figure 2. Contribution of different sources to emissions of SOP, NO,, NHy and VOC in Limburg (1988)

\ Region I

I Total

NO,

0

NHy

1

1

382 5 (62%) 119.6(19%)

I 1

360.2 (26%) 169 9 (22%)

552.4 (40%)

360.3(46%)

479.9 (34%)

Figure 3. Regional differentiation in contribution of various components to acid emissions in Limburg (kg H+/ha/y, 1988)

20 In the north of Limburg ammonia plays a major role in acidification. Emissions of over 200 kg/ha/y are common in this region [2, 31. Most ammonia evaporates directly from stables and manure storages, particularly from poultry sheds and storages. (figure 4) The second important source is the application of manure on fields, while relatively little emission is caused by animals in pastures. The contribution of different agricultural sources is shown in figure 4. Cattle

Pigs

0 manure storage 53%

39%

/

Figure 4. Contribution of different agricultural sources to ammonia emission in northern Limburg (1988) In the southern part of Limburg the most important sources of acidification are power plants and industry. Considerable transboundary transport from sources in Belgium and Germany occurs, but sources in Limburg have an even greater influence on foreign air quality and acidification. Limburg appears to be a net exporter of acid components (see figure 5).

.....

Figure 5. Acidification balance in Limburg (1988) 3.

DEPOSITION

In the National Environmental Policy Plan [4] t h e deposition goals for the Netherlands are set. These targets are based on critical loads for sensitive forest and heather vegetations. The Limburg targets a r e set according to this national approach.

21 The main goal of acidification policy is to prevent occurance of the most severe damage due t o acidification in the year 2010. Therefore the mean deposition levels of H+ and N have t o be reduced as quickly as possible t o the following levels (table 2). Table 2 Deposition targets in Limburg (mean values in eq.H+/ha/y and eq.N/ha/y) year

2000

H+ deposition 2.400 N deposition 1.600

2010 1.400 1.000

To reach these targets a reduction in deposition of 70 - 75 % (relative t o t h e level in 1988) is needed in the next 20 years. The Dutch strategy for the reduction of acidifying components is published in several documents 14, 5, 61. These national documents give a set of measures that should lead to the mentioned targets. This national acidification policy will reduce the mean Dutch depositionlevel, which was 4.900 equivalents H+/ha/y in 1988, to an average of 2.200 equivalents H+/ha/y in the year 2000 (71. Thus i t seems that the national deposition goal of 2.400 equivalents H+/ha/y will be realized on a national scale. If one takes a closer look a t the calculations however, the distribution of deposition over the country appears to be inhomogeneous. In the northern provinces the depositionlevel is lower than 2.400 eq.H+/ha/y and in the south i t is considerably higher. Table 3 shows deposition levels in the Netherlands and in different parts of Limburg.

Table 3 Mean deposition levels in the Netherlands and in Limburg in 1988 and 2000 (equivalents H+/ha/y). Scale

1988

2000

The Netherlands Limburg North-Limburg South-Limburg

4.900 5.360 5.590 4.860

2.200 2.960 3.060 2.750

Implementing t h e national emission policy in the province of Limburg therefore is insufficient t o reach th e deposition goals. Calculations show that in all of Limburg acid deposition remains about 25 % above the target levels. In the north of Limburg this is even more: 30 %. These figures show that important natural values will remain threatened if the national abatement strategy is implemented in Limburg. For the protection of these values a more stringent povincial policy is needed. Therefore the province of Limburg has decided on a complementary policy [ I ] . This policy includes limitation of emissions per component and per targetgroup.

22

EMISSION TARGETS IN LIMBURG

4.

Emission targets in Limburg are primarily based on the need to protect natural values and on technical and economical feasibility of measures. In order t o establish the latter several feasibility studies were carried out 12, 8, 9, lo]. The studies identified technically possible measures for each target group and determined the attainable reductions and their cost. Thus cost-effectiveness of the measures could be determined. From these studies conclusions were drawn regarding the targets for emission reductions by each target group and a set of cost-effective measures was determined. The emission reductions aimed a t a r e illustrated in figure 6. Emissions for traffic are not included in this figure, because the provincial administration has very little influence on them. For reduction of traffic emissions the national policy is carried out.

t

L

k tonly

.--. 0-0 A-A

*-*

SO, : Industry

+ power plants

+ industry + power plants

NH, : Agriculture NO, : Industry VOC : Industry

Figure 6. Targets for t he reduction of emissions of S02, NO,, Limburg.

NHy and VOC in

Table 4 shows for the different target groups and components the relation between national and provincial emission targets and the technically maximum reduction. From this table i t is obvious that the provincial government of Limburg aims a t a further reduction of ammonia emission. I t was found tha t because of the enormous acid load in t h e north of Limburg, which is primarily caused by agricultural ammonia emissions, a further reduction of ammonia emissions is needed and feasible for both industry and agriculture [2, 81. They both have to bring down emissions with an extra 5 % in 1994 above national policy. In t he north of the province agricultural emissions of ammonia have t o be reduced by 80 % relative to the emissions in 1980. Though the technical means ar e not fully available a t present i t is thought that their development can be completed in time. Although an extra effort in reducing ammonia emissions is found feasible not all farmers will be able to finance measures. I t is estimated tha t maybe 40 % of the farmers have insufficient financial means to invest in environmental measures [2]. Therefore a regional ammonia abatement program is required. SO2 emissions of industrial sources can be reduced in a cost-effective way by an extra 15 YO above national policy in 1994 [8]. On the long te rm the national target of 85 % reduction has t o be reached.

23 SO2 emissions in power plants will be reduced by 99 % in 2000 through the closure of several older units. This is more than national policy requires and therefore i t is allowed that in 1994 the emission will still be well above national targets. NO, and VOC emissions will be reduced according to the national policy.

Table 4 Goals for the reduction of emissions in Limburg per target group and for various components (% and kton/y; emission 1980 = 100 %) Target group and component

National targets

Technical maximum

Targets in Limburg

Emission (kton/y)

1994

2000

1980

1994

2000

1994

2000

voe

50 40 50 35

85 60 60 65

90 60 80 50

65 40 55 35

85 60 60 65

16.7 29.1 2.4 7.3

5.6 17.2 1.0 4.7

2.5 11.6 0.8 2.5

Power plants SO2 NO,

85 30

285 > 50

95 40

70 30

99 60

29.1 10.2

7.9 7.2

0.3 4.0

30

70 - 90

50 - 70

40 45

75 80

17.4 13.0

10.0 7.2

4.4 2.6

Industry SO2

:OH.

Agriculture

NH3 North-limburg

5.

IMPLEMENTATION OF THE LIMBURG ACIDIFICATION POLICY

As stated before Dutch provinces have important tasks in air quality control. Several instruments a r e available for the implementation of acidification policy (permit, subsidy, information). Measures that a r e found to be most cost-effective a re effectuated first. The priority ranking is based on feasibility studies 12, 8, 9, lo]. When carrying out t h e measures however sometimes further and more detailed study appears to be necessary and shall be carried out. Emission reducing measures have already been agreed upon with the power production board. These measures will result in a decrease of NO, emission of 80 - 90 % to a remaining 1.3 - 3.1 kton/y and of 99 % for SO2 in 2000. The remaining emission of SO2 by power plants will be 250 kton/y. An additional stimulation program for energy saving will be formulated in 1992. For the industrial sources a program for revision of permits will b e carried out. Furthermore there is a program for stimulating the development and use of new abatement techniques in industry and agriculture. A communication program will be developed in which the importance of action and possibilities to solve the acidification problem will be shown. With respect t o the transboundary aspects international cooperation is sought, primarily with adjoining provinces in Germany and Belgium.

24 6.

POLICY AND RESEARCH

When formulating a regional policy which is more stringent then national and supra national policies the necessity of a solid scientific basis is strongly felt. Target groups have t o make a tremendous effort to reduce their emissions and always ask for motivation for further measures. Although i t is not always possible to present realistic figures, t h e feasibility studies have provided a basis for this motivation. Policy makers in Limburg need however more study yet on two subjects in particular: i.e. on methods for monitoring t he effects of measures taken and on abatement strategies and techniques on a more detailed level. Particularly proces-integrated techniques or - if yo prefer - clean technology needs further development. With respect to monitoring the effect of measures on a regional scale methods are needed, t h a t a r e far more precise then those using general models and those following trends on a national scale. For monitoring the effectiveness of a provincial cornplementory policy more exact information is needed. Policy on a provincial scale regards saving regional natural values. Therefore i t is needed tha t critical levels a re reached a t the right places, which is primarily on acid sensitive natural values themselves. A monitoring method has to be exact enough for this scale. Regarding abatement strategies and techniques information is needed on technical options for emission reduction in various processes. This includes effects and side effects of measures in terms of environmental effectiveness and in terms of cost effectiveness. Being the licensing authority the provincial government needs t o know what emission reduction is possible in specific situations. When working out concrete measures i t appears in some cases th a t the recommended techniques, studied in feasibility studies, supply no solution in a specific process or a re far more expensive than estimated. More detailed study on various processes has to supply t h e information needed in t h e governmental licensing activities. Research and policy will have to work together t o find solutions for the problem of acidification. Policymakers have t o learn from science and vice versa. Cooperation between research and policymakers on an international scale will enhance the solution of the acidification problem on an international, national and regional scale. References 1 Province of Limburg, Environmental Policy Plan 1991 - 1994 (1991). 2 Heidemij, Haalbaarheidsstudie reductie NH3-emissie landbouw, Eindrapport; Province of Limburg (1989). 3 Heidemij, Haalbaarheidsstudie reductie NH3-emissie landbouw, Aanvulling 1989 t o t en m e t 1991; Province of Limburg (1990). 4 Ministry of Housing, Physical Planning and Environment, National Environmental Policy Plan, Second Chamber, session 1988 - 1989, 21137 nos 1-2. 5 Ministry of Housing, Physical Planning and Environment, Bestrijdingsplan Verzuring, Second Chamber, session 1988 - 1989, 18225 no 31. 6 Ministry of Agriculture, Nature and Fishery and Ministry of Housing, Physical Planning and Environment, Plan van aanpak beperking ammoniak emissies van d e landbouw, Second Chamber, session 1990 - 1991, 18225 no 43. 7 G.J. Heij and T. Schneider (eds), Final report second phase Dutch priority program on acidification (1991). 8 Badger, Onderzoek emissiebeperkende maatregelen in de industrie, Province of Limburg (1989). 9 Badger, Verkennend onderzoek naar de mogelijkheden voor het opwekken van schone energie, Province of Limburg (1989). 10 Grontmij, Haalbaarheidsstudie beperking emissies verkeer, Province of Limburg

SESSIONB STATE-OF-THE-ARTOF ACIDIFICATIONRESEARCH

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T Schneider (Editor). Acidification Research Evaluation and Policy Applications 0 1992 Elsevier Science Publishers B V All rights reserved

27

ION AND ACIDI FICATION: A CRITICAL REVIEW

Rod01 p h e S c h l aepfer

Paper presented at the International Conference ACIDIFICATION RESEARCH, EVALUATION AND POLICY APPLICATIONS, Maastricht (NL), October 14-17, 1992

Abstract The objective of this paper is to present a review of the most important scientific knowledge and uncertainties concerning forest vegetation and acidification, and to draw some conclusions for research policy. Forest vegetation and acidification are considered as parts of a global environmental system. The phenomenon of "Forest Decline" is introduced. The methods used for causal research and monitoring are presented, as well as their related difficulties. The scientific knowledge gained from epidemiological approaches (forest damage inventories), experiments and mechanistic studies are evaluated. The paper ends with some conclusions for research policy and emphasizes the need to intensify the use of integrated approaches in environmental sciences.

Contents 1. Introduction 2.

Forest vegetation and acidification as parts of a global environmental system

3.

Methodological considerations

4.

What do we know about forest damage?

5.

What do we know, at an epidemiological level, about the association between defoliation and its hypothetical causes?

6. Can we reproduce experimentally the observed symptoms?

7. Can we explain the mechanisms leading to forest decline? 8.

Summary and conclusions

9. Bibliography

28 1. INTRODUCTION

A large number of contributions concerning forest vegetation and acidification have been published. Many papers and books are available which give good reviews about the known facts as well as about of what is known, as well as about the existing gaps and uncertainties. Examples of these are:

-

The annual Reports on the Forest Damage Survey in Europe [Anonymous, 1991 a]

-

Air Pollution and Forest Health in the European Community, an Assessment of the Current Scientific Evidence [Ashmore, M.R., Bell, J.N.B. and Brown, I.J., 19901

-

Interim Report on Cause-Effect Relationships in Forest Decline [Anonymous, 1990 a]

-

Proceedings of the 5th Meeting of Acidification Research Coordinators [Anonymous, 1990 b]

-

Different reports on national research programmes about forest decline and pollution in France, Germany, Netherlands, Finland, Switzerland, US etc. [see Bibliography Nos 3, 5, 7, 9, 10, 12, 13, 15-17, 20-22, 241

The term Ifforestdeclineffis usually used to describe a decrease of the vitality in the forest ecosystem which can lead in some cases to the death of the stands. The most important criterion used to describe the phenomenon is crown defoliation. There are some generally accepted ideas about forest damage in Europe and in North America. It is clear that a number of different types of decline are present, each being spatially delimited, characterised by a specific set of symptoms and resulting from a certain combination of climate, soil and pollution. Every country has accepted the idea that forest decline is due to a complex set of factors. However, some countries differ over the interpretation of the role of air pollution. This paper is based on the assumption that our society is interested in understanding the phenomenon of ffforestdeclinef1. My objective is to draw conclusions for future research concerning forest vegetation and acidification. For this purpose, I will try to analyse the available information and to discuss how far they allow us a) to evaluate the normality of the situation, b) to describe the evolution of the damages and c) to explain the causes. The problems of global change and critical loads will not be discussed.

29 2.

FOREST VEGETATION AND ACIDIFICATION AS PARTS OF A GLOBAL ENVIRONMENTAL SYSTEM

Difficulties in discussions and decision-making in environmental sciences and policy sometimes come from a biased and narrow-minded view due to ideological thinking or a lack of understanding of the need for a global approach to the problem. We scientists can contribute towards minimizing these difficulties by presenting our results as impartially as possible and by integrating our research topics within higher order systems. In this sense it should be remembered that the forest vegetation is only a part of the forest ecosystem, that acidification is only a part of the pollution problem and that both forest ecosystems and pollution are parts of a global environmental system (Figure 1).

Forest ecosystem .............._.. ..............._. ____________._._. ~

Vegetation Fauna Mlcroorganlsms Soil Water Air

Figure 1: Global environmental system

30

The forest ecosystem is composed of the interacting elements of vegetation, fauna, microorganisms, soil, air and water. It itself interacts with other systems like llClimate*f, "Management practices" , ltPollutionlland l1Pathogensl1. The word Ilacidification" can be divided into two parts: tfacidll and "ficationll In order to understand the relationship between forest ecosystems and acidification we have to consider both parts. The vtacidlt part is the input into the ecosystem of compounds which release protons. The second part, flficationll, is the effects of acid deposition. The term "acid depositionll is often used as a synonym for "acid rain" or "acid precipitationll. Historically, these expressions have been repeatedly used as "forerunnersttin discussions and research on atmospheric pollution. They are thus frequently applied in a non-specific manner to various kinds of emissions, especially in political discussions and as titles for research programmes and conferences. In a more narrow sense, the term "acid deposition" can today be defined as the input of all components into an ecosystem which determine the net proton flux into the system. Acid rain in this context is only a part of acid deposition. For the determination of total acidity input, the interception of sulfur- and nitrogen-containing particles as well as the dry deposition of nitrogen dioxide, nitric acid, sulfur dioxide and ammonia have to be included.

.

3.

THOD DO LOGICAL CONSIDERATIONS

Ecosystems usually vary in time and in space. The changes can be normal or abnormal, compared to a defined standard. Abnormal changes are closely related to diseases, damage, decline or a decrease of vitality. We are therefore interested in the importance, the dynamics, the spatial variation and the causes of changes in the forest ecosystem. This information can serve as basis for decision making. The certainty of existing information about changes depends on the magnitude of the changes, the complexity of the processes involved, the natural variability of the system studi.edand the methods used to obtain the information. Causal research is relatively easy when the effects are due to a single factor. In ecological studies, the situation is generally more complex. Rothman (1986) presents a model of causation which helps the conceptualization of epidemiological problems like forest decline. In his view, a disease is produced by a constellation of components that act in concert. He defines the "sufficient causeu1as a set of minimal conditions and events that inevitably produce disease. ltMinimalll implies that none of the conditions (components) are superfluous. Two component causes in a "single sufficient cause" are considered to have mutual biological interaction. A component is considered as a "necessary cause" if without this component there is no

31

disease. Research indicates that it is reasonable to assume that forest decline is also produced by a constellation of components that act in concert. These components may be certain constellations of climate, soil, pollution, species and management practice. One objective of research could be the identification of the set of minimal conditions that produce forest decline. Other interesting tasks are to find out if there is a "necessary cause" without which there is no forest decline and to understand the mechanisms leading to changes. Methodological aspects of causal inference in epidemiology and ecology have caught the attention of many authors (for example: Koch, R., 1884; Mosteller, F. and Tukey, J.W., 1977; Rothman, K.J., 1986; Schlaepfer, R., 1988; Oren, R., Werk, K.S., Meyer, J. and Schulze, E.-D., 1989 or Ashmore, M.R., Bell, J.N.B. and Brown, I.J., 1990, Innes, J., 1991). Based on their contributions we can describe an integrated and iterative approach for forest decline research which should give us information about the magnitude, dynamics and variability, as well as about causes of the phenomenon: 1. Detection and definition of the problem. We have first to

detect if there are abnormal deviations and to find out their symptoms (y) and hypothetical causes (x). 2. Description of magnitude, dynamics and variability of the

phenomenon. In this step, the spatial distribution and the evolution in time of the symptoms and the hypothetical causes are described. Surveys are used to do this, the best being an integrated forest damage survey. 3.

Detection of associations in space and in time between the symptoms and the hypothetical causes (epidemiological approach). When causality exists and when other variables are equal, the relationship between the symptoms and the causes is consistent across comparable populations in direction and perhaps in amount. We can verify the consistency of relationships by using surveys. The data are analysed using multivariate techniques such as multiple regression. The results of the analysis are used to formulate hypotheses about the form and the intensity of the cause-effect relationships. The multivariate technique has the advantage of giving information about effects as well as about interactions between causal agents.

4. Experimental reproduction of the observed symptoms. Experi-

mentation is the only scientific method which allows an unequivocal assessment of the impact of a pollutant. An experiment allows us to study, under given conditions, the response of the research object when submitted to a controlled change of a parameter of interest. If the relationship between the symptoms and the hypothetical causes is indeed causal, we can, by intervening and changing the level of the hypothetical cause, reproduce experimentally the observed

32

symptoms. This step allows us to test hypotheses about the cause-effect relationships which are suggested by research or by other sources of information (i.e. surveys and case studies) and to quantify the relationships. A limiting factor in forest decline research is that experiments are difficu1.t to perform with mature trees. 5. Explanation of the mechanisms. In this step, models are pro-

posed to explain the processes by which the causes produce the symptoms of the different types of forest decline, in particular defoliation and discoloration. The hypotheses about the mechanisms are based on an integrated analysis of all the information available about the phenomena (basic research in physiology and biochemistry, surveys, experiments, field studies). 6. Validation of the models. Are the proposed models about the

mechanisms leading to the observed symptoms reasonable? The most important method used to answer this question is to verify if the models match available data derived from field studies, surveys or experiments. It is important to realise that confirmation of the importance and the causes of a particular type of forest decline should be the product of a global evaluation of the results of all the steps mentioned above. Such an integrated approach is a guideline for ecological research and helps us to recognise the difficulties of Ifprovingvv the causal nature of an association, as well as the limitations of surveys and experiments.

4.

WHAT W WE KNOW ABOUT FOREST DAMAGE?

There are numerous symptoms associated with forest decline. They usually vary in importance from region to region and from species to species. Examples of symptoms related to forest vegetation are:

-

crown defoliation, i.e. changes in the density and the size of leaves and needles and their premature senescence and abscission,

-

discoloration, for example the yellowing of spruce needles,

- changing in the branching structure of trees,

-

increment change,

-

root damage,

- increased sensitivity to stress,

33

-

nutrient deficiencies,

- degradation of surface wax structure. The International Cooperative Programme on Assessment and Monitoring of Air Pollution Effects on Forests (ICP Forests) annually publishes a report on the forest damage survey in Europe. These surveys started in 1986, and are based on guidelines laid down in a Manual on 94ethodologies and Criteria for Harmonized Sampling, Assessment, Monitoring and Analysis of the Effects of Air Pollution on Forestsv1.31 countries participate in the programme. The most important criterion used in the surveys is the defoliation of the crown.

Conifers

Broadleaves

% of trees with needle loss >25%

% of trees with leave loss >25%

60

50

50

40

30

-----------I--

86

87

-------.-I ........

............

88

89

Austria Czechoslovakia Finland Germany

90

86

1

1

I

87

88

89

Netherlands ....-.. .- Sweden -- - - - - Switzerland United Kingdom " . . " I . "

Figure 2: Forest Damage in Some European Countries

90

34 conifarm 86

87

88

89

Broadlaavam 90

n

86

87

88

89

90

1990

Aumtria Czachoelovakia

UK

4,5

3,5

4,0

n 1990

5,5

4,l

3,3

3533

16,4

15,6

27,O

32,O

50,3

10505

-

-

23,O

27,O

34,O

43,O

1036

-

7,s

3,5

6,7

6,9

79'1

-

29,l

37.0

33,9

3056

10,O

20.0

21,O

28,s

635

Source: International Cooperative Programme on Assessment and Monitoring of Air Pollution Effects on Forests: Report on the 1990 Forest Damage Survey in Europe (Draft) Table 1: Forest damage in some European countries from 1986 to 1990, defoliation classes 2-4 ( % trees with more than 25 % defoliation)

Table 1 and Figure 2 present for some countries, from 1986 to 1990, the percentage of trees, with more than 25 % defoliation.

The results show that the level and the course of defoliation can be very different both between countries and between conifers and broadleaves. One question is whether the observed differences are due to different assessment methods or to possible causes like climate, soil, pollution and management practices. We see from this example that an important problem related to the forest damage surveys is the interpretation of the results. It is now accepted that despite the progress made in recent of years, many uncertainties and gaps remain which make it difficult to draw conclusions from the data:

-

The nature and extent of the differences in the manner in which defoliation is being assessed by different countries have not been quantified. Comparisons between the results obtained are therefore invalid (Anonymous, 1990 a).

-

Little is known about defoliation in stands under different natural conditions. The estimates of defoliation alone provide only a little information about the severity of the damage.

- Average estimates of defoliation for countries provide no information on the different$ypes of forest decline in Europe.

35

-

In view of the difficulty of applying a standard assessment of defoliation across a wide range of growth habits (Ashmore, Bell, Brown, 1990), comparisons of the health of individual species must be made with extreme caution.

-

The observed trends in defoliation are difficult to interpret because we know little about the evolution of defoliation over the last 50 years.

-

Defoliation is not a cause-specific symptom. Estimates of defoliation alone give nearly no information on its causes.

-

There is no evidence of a widespread decline in the growth of European forests associated with the widespread reports of defoliation, although reductions have been demonstrated at a local or regional scale (Anonymous, 1990 a).

These gaps suggest some recommendations for the execution of future surveys:

-

-

Quantify the uncertainties of forest damage surveys and improve the harmonization of assessment methods in different countries. Develop criteria for the evaluation of the severity of damage. Present results according to the type of forest decline rather than by individual countries. Establish whether an observed form of damage is unprecedented or whether it has existed for some time. Incorporate other indices of tree condition into large-scale surveys (for example increment, better use of tree mortality, branching, nutrient content in needles and leaves).

5. WHAT DO WE KNOW, AT AN EPIDEMIOLOGICAL LEVEL, ABOUT THE

ASSOCIATION BETWEEN DEFOLIATION AND THE HYPOTHETICAL CAUSES?

Ideally, the data necessary for a multivariate approach to the study of associations in forests between the observed damage and hypothetical causes are given by a sampling survey, in which not only the damage, but also the levels of the hypothetical causes like climate, soil, pollution, pathogens and management practices are observed at each point. Such an integrated survey does not exist in Europe. Separate long-term monitoring programmes with no interconnection have been run for air, forest or soil. These separate programmes enable conclusions to be known concerning the variations in time of individual variables. If a high enough number of observation

36

points is used, it is possible to obtain information about the spatial distribution of the variable. There are also many integrated descriptive case-studies in Europe. Examples of these are given by the three field studies Lageren, Alptal and Davos in Switzerland (Stark, 1991). A case-study enables a detailed description of variables and relationships between variables in time to be made, but gives no information about the spatial association between defoliation and causes. Theoretically, it is possible to use the data from other monitoring programmes to estimate the value of variables of interest (for example pollution or meteorological data) which are not measured on the forest survey plots. The success of this will depend on the density and the reliability of the networks involved. In relation to pollution, it is difficult to obtain reliable monitoring data of a consistent quality from different European countries (Ashmore et al., 1990). Because of this, the European Monitoring and Evaluation Programme (EMEP) was created. The EMEP is a coordination centre which is linked through the UNECE Convention of Long-Range Transboundary Air Pollution to international agreements on pollution reduction. The EMEP network is based on a 150 km grid and therefore cannot identify particular concentrations in localised areas. The spatial distribution cannot be mapped with great certainty. However, the EMEP data provide a broad picture of the main features of the distribution of sulphur and nitrogen pollutants across Europe. Because of the above problems, it is difficult to study the relationship in the forest between damage and its hypothetical causes, especially pollution, in Europe. The available information does not show a general association between defoliation and pollution in Central Europe (Rehfuess, K . E . 1991). There are even contradictory results at the regional scale. We do not know if this situation is due to an absence of a real association or to the inadequacy monitoring methods. To fill this gap we have to support any effort which is made in Europe to:

-

create an integrated large-scale survey, including observation of forest damage criteria as well as the hypothetical causes like climate, soil, pollution, pathogens and management practices:

-

improve and apply simple field methods for air pollution monitoring:

-

improve the existing long-term monitoring programmes (more reliable data, higher density of the network);

-

improve the coordination between the existing long-term programmes for monitoring soil, air and vegetation.

37 6. CAN WE REPRODUCE EXPERIWENTALLY THE OBSERVED SYMPTOMS?

Different experimental techniques are used in forest decline research. The most important are open- and closed-top chambers, controlled environmental chambers, laboratory based experiments and field experiments. Much of the work done is based not on experiments, but on observations of forest stands, or analysis of collected material, usually making comparisons between healthy and damaged stands. Ashmore et al. (1990) give a useful picture of the current emphasis of research on air pollution and forest decline in Europe. They point out that of all work reported, approximately 75 % is on coniferous species and only 25 % is on the broadleaves. Nearly 70 % of work using coniferous species is devoted to only one species, Norway spruce, and 60 % of work using broadleaves is devoted to beech (Fagus sylvatica). They also found that the majority of experimental work has been carried out using either ozone ( - 30 % ) or simulated acid deposition (-25 % ) . The majority of studies only use single pollutants. There is little work on nitrogen pollutants. A considerable number of experiments have used pollutant concentrations in excess of those expected in ambient air across Europe. Ashmore et al. identified nine main categories of processes that have been studied:

- leaf metabolism

-

-

-

(-

30 % )

nutrient balances (-15 % ) leaf surfaces ( * 10 % ) growth/biomass ( - 10 % ) root/mycorrhizal responses (5-10 % ) water relations (5-10 % ) reproductive processes and genetics (0-5 % ) response to secondary stresses (0-5 % ) soil chemical changes (0-5 % )

It is impossible to summarize in this paper all the important experimental results obtained over the past few years. I will simply mention some conclusions reached by Ashmore et al. (1990):

- "It is clear that disruption of the processes of photosynthesis, carbon metabolism and transport is observed frequently in damaged stands of Norway spruce and other species. Similar disruptions have been demonstrated in some experiments with air pollutants, but this does not prove that direct impacts of air pollutants are responsible for the effects observed in the field. It seems more probable that nutrient deficiencies are the prime cause of loss of chlorophyl, reduced rates of photosynthesis and disrupted translocation patterns on many sites.

38

- "There is evidence that exposure of conifers to air pollutants or wet deposition can lead to changes in epicuticular waxes similar to those observed in declining stands. .. However, some experiments at realistic, or above ambient, levels of exposure have not reported significant effects on leaf surface properties."

-

IIPollutant deposition may affect soils in three major ways: by increasing acidity, by decreasing nutrient availability (either as a result of leaching or competitive inhibition of uptake), and by increasing the solubility of toxic ions such as aluminium. It

- "There is evidence that fine root and mycorrhizal vitality may be affected by both above- and below-ground pollutant impacts. It is not clear which of these is of greatest significance in different field situations, or, indeed, whether damage to the root system is necessarily associated with forest decline.

- "It seems likely that foliar leaching of mineral nutrients due to acid deposition or other pollutants is not a key factor in the development of the mineral deficiencies associated with forest decline. Rather, the supply of mineral nutrients from the soil is the critical factor."

-

"There is clear evidence that atmospheric pollutants may affect the water relations of a tree. However, the lack of consistency in the experimental data makes it difficult to generalise about the significance of this factor in forest decline.

-

"It is impossible to assess whether current levels of pollutant deposition are increasing, decreasing, or having no effect on tree growth: it is likely that all three possibilities exist on different species in different regions of the CEC .

-

"It is clear from experimental studies that air pollutants at realistic concentrations may have quite subtle effects on the pattern of seedling growth and development."

-

"Evidence suggests that pollution may enhance tree sensitivity to other environmental stresses - the most studied example being winter damage. There is a general concensus that exposure to pollutants increases the susceptibility of trees to low temperatures. This has been demonstrated for summer pollutants, such as ozone, as well as those more prevalent during the winter.

In the conclusions from Ashmore et al. the words llrnayll or Ilcan" appear frequently. The reason for these cautious formulations certainly lies in the limitations of the available experimental works, for example:

39

-

Most experiments are performed with young trees and fail to simulate field conditions.

- Often, higher concentrations of pollutants are used than are normally found in areas with forest decline.

-

Most experiments consider the effect of a single pollutant and therefore do not give information about combinations of pollutants or interaction between pollutants and other factors of influence.

- Few experiments are performed with the aim of studying directly the main symptom observed in the forest damage survey, i.e. defoliation.

- The main effort put into experimental work has been for Norway spruce and beech. It follows from these conclusions and the limitations listed above that the main symptoms of forest decline can only be partially reproduced experimentally. This situation could be improved if we encourage the following steps:

-

more emphasis on field experiments, especially those involving field manipulation,

-

more emphasis on experiments with more than one pollutant, and on experiments designed for the study of interactions between pollutants and other hypothetical causes,

-

better coordination of the objectives of experiments with the identified types of forest decline.

7. CAN WE EXPLAIN THE MECHANISMS LEADING TO FOREST DECLINE?

Ten years ago, many people attributed forest decline mainly to air pollution. The research efforts of the past decade give us a more complicated picture of the phenomenon. Different models are now proposed to explain specific processes or the whole mechanism leading to forest damage. For example, canopy photosynthesis is a well explained process. However, much has to be learnt about the processes which control the water and nutrient balances of trees, and the partitioning of carbon and other essential elements. Unfortunately, those processes for which the fundamental understanding necessary for successful modelling is poorest are precisely those which appear to be the most important in terms of pollutant impacts (Ashmore et al., 1990).

40

We also know that at the cellular or leaf level in seedlings and trees, the same basic processes occur (photosynthetic pathways, respiration, conversion to metabolites, translocation to other organs). However, due to different heterogeneity and proportions of leaf, wood and root tissue, the relationship and balance between these processes differs. It is therefore likely that differences between the responses of seedlings and mature trees to the factors of influence, particularly pollutant stress, are important. This means that experiments with seedlings may be useful for our understanding of individual processes, but are unreliable as an aid to understanding the forest ecosystem as a whole. An interesting example of an explanation of a particular type of forest decline is given by the Fichtelgebirge study (Schulze et al., 1989). In this study, a decline was recognized only if needle-yellowing was accompanied by reduced stand growth rate per ground area. Schulze et al. deduce that direct effects of air pollutant are not of major importance, that stemwood growth is most closely related to the supply of magnesium from the soil and that the symptoms of needle yellowing reflect a complex balance between the uptake of base cations from the soil and nitrogen compounds from both the soil and the atmosphere. The authors also mention that the conclusions of the study must be considered only as a set of plausible hypotheses which can best explain the patterns in the data without major contradictions. Large-scale weather stress is also considered as a inciting factor (Landmann, G., 1991) or as the possible cause of the synchronization of different disease types in coniferous stands during the period 1980-1985 (Rehfuess, K.E., 1991). Most of the proposed explanations of the mechanisms remain hypotheses and models which have still to be tested and validated in the field. Many of them were elaborated for very special situations and cannot be easily generalized. The ways in which all the key factors interact to produce the symptoms of forest decline (defoliation, discoloration) and to produce different suites of symptoms on different species, different soils and different pollution situations are unclear. This conclusion shows that a great research effort is required if we wish to have a better understanding of the mechanisms. For example, it is necessary to intensify:

-

basic research, especially in tree physiology and biochemistry,

-

the use of modelling as an approach allowing us to integrate the information coming from the different methods used in forest decline research.

41 8.

SUMMARY AND CONCLUSIONS

I will try to summarize and to draw the conclusions by answering three questions: What do we know? What are the gaps? What should we do? What do we know? The most important agreements among scientists about forest decline in Europe are:

-

There are different types of forest decline characterised by defoliation. Each of these types is associated with different sets of additional symptoms with different sets of plausible stress factors.

-

Most of the types of forest decline are multiple-stress phenomena in which climatic factors, soil conditions, pollution, pathogens and management practices may be components which act together.

-

No general spatial or temporal association between defoliation and pollution has been observed in Europe.

- There is no evidence of a widespread decline in the growth of European forests.

-

There is much experimental evidence about the form of the cause-effect relationships between pollutants and their impacts on specific processes. Unfortunately, these results cannot be easily extrapolated to field conditions.

-

There are only a few cases of forest decline for which a plausible explanation of the mechanisms exists. One of them is the yellowing of Norway spruce at higher altitudes on acidic soils in Germany. Nutrient imbalances are associated with the decline. Changes in soil chemistry seem to be a major cause of nutrient disharmony.

-

Although the direct effects of pollution on forest vegetation are not so important as was thought some years ago, they, together with the long-term impacts of acid deposition (including nitrogen) on forest soils, have to be considered as a potential threat to the forest ecosystem.

What are the gaps in our knowledge?

-

Because of our lack of knowledge about the ffnormalfl situation, we do not know how severe the observed levels of defoliation are.

42

-

The results of the forest damage surveys in different countries are not comparable.

- We do not know if the non-observed association between defoliation and pollution is due to an absence of such an association or due to the inability of existing methods to detect it.

- We do not know how far the results of chamber experiments can be extrapolated to field conditions.

-

We have only an incomplete picture of the importance and the dynamics of different types of forest decline.

- For most of the types of forest decline the mechanisms with which the influencing factors affect the ecosystem are unknown.

- We do not know the interactions between the different possible causes, in particular between pollution and climate factors. What should we do?

An important duty of research is to improve scientific knowledge concerning forest decline. As scientists, we see that the methods used have their limits. We see also that, despite of the efforts that have been made, the lack of international coordination remains an important problem. We should therefore encourage:

-

the elaboration of large-scale and regional integrated multifactorial monitoring;

-

a better coordination of experimental work with the observations in the field,

-

mechanistic studies to improve our knowledge about processes in the ecosystem, and

-

the intensification of the use of multi-disciplinary modelling techniques for explaining the phenomenon.

I believe that better progress can be made if more of the national ambitions and efforts in research are put into international projects. Only a European approach to the problem will allow monitoring and causal research in forest decline which is based on ecological realities rather than artificial spatial units like countries. In this sense, the participation of non member countries to research programmes of the European Community should be facilitated.

43 Forest decline has scientific, psychological, economic and political implications. Decision-makers in many fields have to take account of the problems related to forest vegetation and acidification. It is therefore important that scientists provide impartial information about the existing facts, gaps and uncertainties. This information should be intensified and presented in languages which can be understood by the audience. We should also make it clear that the existing knowledge about the risks facing forests is enough to justify any effort to reduce air pollution in Europe. We also have to point out that the available information about the cause-effect relationships in forest decline does not allow us to give useful advice for fixing the priorities of a European pollution control policy.

9.

BIBLIOGRAPHY

Anonymous: Interim Report on Cause-Effect Relationships in Forest Decline. Convention on Long-Range Transboundary Air Pollution, International Cooperative Programme on Assessment and Monitoring of Air Pollution Effects on Forests (1990 a). Anonymous: MARC V. 5th Meeting of Acidification Research Coordinators, Nancy, September 25-27 (1990 b). Anonymous: NAPAP Annual Report 1989 and Findings Update. National Acid Precipitation Assessment Program, Washington DC (1990 c). Anonymous: Report on the 1990 Forest Damage Survey in Europe. Convention on Long-Range Transboundary Air Pollution, International Cooperative Programme on Assessment and Monitoring of Air Pollution Effects on Forests (1991 a). Anonymous: 7. Statuskolloquium des PEF (Projekt Europaisches Forschungszentrum fiir Massnahmen zur Luftreinhaltung) vom 5. bis 7. Mlrz 1991 im Kernforschungszentrum Karlsruhe. Kernforschungszentrum Karlsruhe (1991 b). Ashmore, M.R.; Bell, J.N.B.; Brown, I.J.: Air Pollution and Forest Health in the European Community: An Assessment of the Current Scientific Evidence. Commission of the European Communities, Air Pollution Research Report 29 (1990). Haemmerli, F.; Schlaepfer, R.: Forest Decline in Switzerland. In: Huettl, R.F.: Forest Decline in Atlantic and Pacific Region, Springer Verlag, Berlin, New York. (Paper submitted: June 1991). Hauhs, M.; Wright, R.F.: Regional Pattern of Acid Precipitation and Forest Decline Along a Cross Section Through Europe. Water, Air and Soil Pollution 31 (1986). 463-474, D. Reidel Publishing Company. Innes, J.L.; Boswell, R.C.: Monitoring of Forest Condition in Great Britain 1989. Forestry Commission. Bulletin 94 (1990).

44 10

11

Innes, J.L.: The Application of Cause-Effect Criteria to the Relationship between Air Pollution and Forest Decline in Europe. In: James W.S. Longhurst (ed.): Acid Deposition, Springer Verlag, Berlin (1991). Kauppi, P.; Anttila, P.; Kenttamies, K. (eds.): Acidification in Finland: Finnish Acidification Research Programme HAPRO 1985-1990. Springer Verlag, Berlin, Heidelberg (1990).

12 13

14 15 16

17 18 19 20 21

Koch, R.: Die Aetiologie der Tuberkulose. Mitt. Kaiserl. Gesundheitsamt, 2: 1-88 (1884). Krause, G.H.M.; Prinz, B.: Experimentelle Untersuchungen der LIS zur Aufklarung mdglicher Ursachen der neuartigen Waldschaden. Landesanstalt fur Immissionsschutz, NordrheinWestfalen, Bericht Nr. 80 (1989). Landmann, G.: Les recherches en France sur le deperissement des for&ts. Programme DEFORPA, 28me rapport. Ecole nationale du genie rural, des eaux et des forkts, Nancy (1991). Mosteller, F.; Tukey, J.W.: Data Analysis and Regression. Addison-Wesley, Reading, Massachusetts (1977). Oren, R.; Werk, K.S.; Meyer, J.; Schulze, E.-D.: Potentials and Limitations of Field Studies on Forest Decline Associated with Anthropogenic Pollution. In: Schulze, E.-D.; Lange, O.L.; Oren, R. (eds.): Ecological Studies, Vol. 77, 23-36, Springer Verlag Berlin, Heidelberg (1989). Orthofer, R.: Effects of Air Pollutants on Ecosystems: Research Experiences and Future Strategies in Austria. Oesterreichisches Forschungszentrum Seibersdorf (1990). Rehfuess, K.E.: Review of Forest Decline Research Activities and Results in the Federal Republic of Germany. J. Environ. Sci. Health. A26 (3), 415-445 (1991). Rothman, K.J.: Modern Epidemiology. Little, Brown and Company, Boston, Toronto (1986). Schlaepfer, R.: Waldsterben: Eine Analyse der Kenntnisse aus der Forschung. Eidg. Anstalt fur das forstliche Versuchswesen, CH-8903 Birmensdorf, Bericht Nr. 36, (1988). Schlaepfer, R.; Haemmerli, F.: Das "Waldsterben" in der Schweiz aus heutiger Sicht. Schweiz. Z. Forstwes. 141 (3); 163-188, (1990).

22

23

24

25

Schneider, T.; Heij, G.J.: Dutch Priority Programme on Acidification: Thematic Reports. National Institute of Public Health and Environmental Protection, Bilthoven, The Netherlands (1990). Schulze, E.-D.; Lange, O.L.; Oren, R.: Forest Decline and Air Pollution: A Study of Spruce (Picea abies) on Acid Soils. Ecological Studies, Vol. 77, Springer Verlag, Berlin, Heidelberg (1989). Schulze, E.-D.; Oren, R.; Lange, O.L.: Processes Leading to Forest Decline: A Synthesis. In: Schulze, E.-D.; Lange, O.L.; Oren R. (eds.): Ecological Studies, Vol. 77, 459-467, Springer Verlag Berlin, Heidelberg (1989). Stark, M. (ed.): Luftschadstoffe und Wald: Ergebnisse aus dem Nationalen Forschungsprogramm 14. Lufthaushalt, Luftverschmutzung und Waldschaden in der Schweiz, Programmleitung NFP14. Verlag der Fachvereine, Zurich (1991).

T Schneider (Editor), Acidification Research Evaluation and Policy Applications @ 1992 Elsevier Science Publishers B V All rights resewed

45

Global E n v i r o n m e n t a l Change: Implications for Acid Deposition Research D. J. Waters and P.G. Whitehead

Institute of Hydrology, Wallingford, Oxon, OX10 SBB, UK Abstract Global environmental change has implications for acid deposition research. Many of the complex processes underlying acidification are also influenced by climate change and land use. It will be very difficult to attribute any perturbation in the ecosystem t o any specific factor, i.e acid deposition, climate change or land use. The following paper attempts to summarise the impact and interactions between climate change, land use and acidification processes and suggests key areas for research.

1. I n t r o d u c t i o n Research into the effects of acid deposition has received a relatively high profile over the last twenty years, However, increasing media awareness i n recent years has diverted some of o u r attention to affects on materials, e.g. historic and cultural monuments, health effects (i.e. increasing concern of high photochemical oxidant concentrations) and quality of life related parameters, e.g. visibility. Knowledge concerning primary pollutant emissions, atmospheric transport and chemical transformation and deposition has improved enormously. The resultant effects o n ecosystems particularly soils, surface waters and flora and fauna has progressed such that the ability for models to predict the environmental impacts has improved. Spatial variability often complicates model prediction and long term monitoring must proceed so that complex processes can be refined and improve our chances of model extrapolation. Temporal variations i n atmospheric pollutant concentrations mean that many processes are dynamic and not in steady state, e.g. sulphate adsorption. Relative contributions of acidic pollutants and their impacts o n ecosystems have changed, systems which were dominated by sulphurous compounds are feeling an increased impact of nitrogen. Our understanding of the effects of nitrogen are less well understood. However, it is clear that once concentrations exceed the biological requirements, this highly mobile anion has an acidifying effect. Other atmospheric gas concentrations have

46

changed, particularly ozone and ammonia and present research needs to incorporate their synergistic, additive or antagonistic effects. Increased emission of greenhouse gases and the resultant implications for climatic change has ramifications for acidification research. Changing temperatures will affect evapotranspiration and the rate of biochemical reactions in soil and water. Changing seasonality and volume of rainfall will effect the flushing and leaching of soils as well as the frequency of acid episodes and perhaps dilution effects. The implications for soil moisture and perhaps water pathways and, therefore, mixing will control inputs t o the surface water system. It is quite clear many of these climatic changes will have resulting implications for acidification studies. Land use change has other obvious implications for acidification research. Natural vegetation succession and colonisation might arise from climate changes. However, it is the land management practice that will have the greatest effect. Afforestation or deforestation have well documented effects as well as the implications of fertiliser and pesticide usage under more intensively managed systems. It is quite clear that environmental change has implications for many of the processes important in acidification studies. This paper attempts to summarise the implications of environmental change and perhaps suggest future approaches to continue acidification research per se and incorporate other modifications to the ecosystem.

2. Global Environmental Change

2.1 Atmospheric Pollution Trends in atmospheric pollution will have direct effects on ecosystems. However, assessing the immediate influence of changes in deposition will be complicated due to the ecosystem response time. National policies for pollution abatement differ and the resultant local deposition pattern will still be influenced by long range transport. Much time and expense has been dedicated to sulphur emission reductions and more recently abatement strategies for nitrogen emissions have been set. This reflects the increasing importance of nitrogen and its increasing contribution to the total acidic input to catchments. Policies for ammonia emission reductions is still largely a national rather than international problem. Photochemical oxidants such as ozone concentrations have also been rising, and its implicated impact on human health, materials and vegetation might see future protocols for emission control. To date most protocols for the reductions of sulphur emissions have been based on linear changes, i.e signatory to the 60% club, which requires EEC countries to reduce emissions by 30% of the 1980 levels by 1993 and 60% by 2005. Recent research into critical loads of sulphur to ecosystems has provided an alternative strategy for sulphur emission abatement. With the use of critical and target loads, various emission reductions and their total environmental resource impact can be assessed. Using this technique national policies can be equated to reflect direct ecosystem benefits. Of course reductions in emissions will not yield

47

a linear decrease in deposition but the critical loads approach is much more scientifically valid. It is, however, more complex requiring a n understanding of processes so that critical loads can be evaluated. Reductions in sulphur loadings to ecosystems will be beneficial but the magnitude and speed of response will depend on the timescales to reach new equilibrium concentrations as well as the total reduction in sulphur deposition. For example, predictions of future acidification trends using the MAGIC model (Figure I), suggests very large emission cuts are needed in some catchments if the acidification status is t o improve (Cosby et al., 1986; Whitehead et al., 1988a; Whitehead et al., 198813). I t is clear however, that as sulphur emissions and, therefore, atmospheric concentrations are reduced nitrogen will assume the role as the dominant acidic pollutant. 6.0

5.5

5.0

4.5

4.0

1840 1880 1920 1960 2000 2040 2080 2120

-Historical levels to 1984 and constant 1984 levels thereafter

..... Historical levels to 1984 and 1984 levels reduced by 50% by the year 2000 and constant thereafter - - - Historical levels to 1970 and constant 1970 levels thereafter Figure 1

Simulated historical and future pH at Dargall Lane (SW Scotland)

Abatement strategies for nitrogen are much less stringent. To date protocols concerning the control of emissions of nitrogen oxides or their transboundary fluxes commit national emissions a t 1987 levels by 1994. As our knowledge of the effects of nitrogen increases, no doubt greater control measures will be required. In some of our research programmes the role of nitrogen has attained a higher profile as rising surface water concentrations have been observed. Nitrate is a highly mobile anion and is as efficient as sulphate in leaching base cations from the soil. It is, therefore, a n acidifying agent. It is also extremely important as a nutrient source for algae and increased levels of nitrogen will lead t o more eutrophic conditions in upland streams and lakes. The

48 increasing importance of nitrogen has resulted in greater modelling effort but understanding the processes controlling nitrogen dynamics is still rather limited. Critical loads of nitrogen have not been attempted in many countries and, therefore, critical loads of acidity are preliminary. Of course increasing nitrogen deposition will not be detrimental for all areas and for all parts of the ecosystem, for example, in nitrogen deficient areas increased deposition might increase forest growth. As nitrogen is a very important nutrient it is difficult to predict when saturation will commence and, therefore, when it will act as an acidifying agent. Ammonia emissions have also been rising, particularly in regions of high livestock density. The importance of this acidifying and fertiliser compound varies nationally, for example, it requires a high priority in the Netherlands where its contribution to total acidity is greater than 45%. The highest ammonia sources are from intensively managed livestock systems, particularly from manure, land spreading and housing systems. Trends in these activities will have implications for the increasing importance of ammonia emissions. The Dutch government have already imposed an ammonia emission reduction of 30% by 1994 and a further reduction of 70% by the year 2000. In a future environment where nitrogen species are liltely to increase in importance other national abatement strategies must adopt more stringent protocols for nitrogen species. There has been increasing interest in levels of secondary pollutants, i.e. photochemical oxidants over the last two decades. In particular tropospheric ozone concentrations have been monitored and it has been implicated that elevated concentrations have a direct affect on human health, materials and vegetation. Although there are a number of factors which control the sensitivity of vegetation to ozone it has been shown that if concentrations are maintained at 50 ppb for several weeks adverse affects on plant growth have been observed. The forest decline experienced in the Federal Republic of Germany has been attributed to elevated ozone concentrations. Mean ozone concentrations recorded at these high elevation sites are comparable with areas of the USA where ozone is known to damage the forest. The main effect is the impact on cellular permeability which allows occult deposition to leach key nutrients at an accelerated rate. Changes in the concentrations of acidic pollutants will have obvious ramifications for acidification research. However, there is increasing evidence that the scientific community is targeting nitrogen species, as their relative importance is increasing. In addition, although not a primary or secondary pollutant the deposition of base cations provides an important neutralisation source and should not be ignored. Estimates of critical loads would change if there was a n increasing trend in base cation deposition to catchments. The rest of the paper deals with other global environmental changes and considers the indirect implications for acidification studies. 2.2 L a n d Use Change

Although natural changes in vegetation succession and colonisation might arise from other global environmental changes this section will consider the politically driven changes in land management, particularly, afforestation, deforestation and agricultural practice and the implications for acidification research. Land use change has already formed and integral part of acidification

49

studies. Many comparative studies have implicated the exacerbating effect of coniferous a f k e s t a t i o n on stream water quality. Similarly modelling approaches have been employed, to predict future changes in the acidification status of the water under a number of land use change scenarios. MAGIC has been used at a number of sites to predict the impact of afforestation, deforestation and replanting under a number of deposition scenarios (Figure 2) and has recently been used to investigate the influence of land use an critical load estimation (Whitehead et al., 198813, Jenkins et al., (in Press)).

7 6

5

5 4

3

1840

1900

1960 2000

21 00

0.8

0.6

-

a

0.4

0.2

18.40

=1900- <960

2000

21 00

- Conifers planted 1958:

. 0

deposition unchanged from 1984 levels - Conifers planted 1958:deposition reduced by 50% - No planting:deposition unchanged from 1984 levels - No planting: deposition reduced by 50%

Figure 2

Simulated pH and aluminium levels for different deposition and land use regimes (moorland C15)

The possible ef'fects for conifer afforestation exacerbating surface water

50

acidification are listed below; a) b) c) d) e) f) g)

Increased evapotranspiration (concentration effect) Increased scavenging (dry and occult deposition) Uptake of base cations Increased depth of litter layer Improved drainage (drying of the soil) Modification of water pathways Removal of surface vegetation (neutralisation source)

Increased water usage and interception loss leads to higher evapotranspiration and therefore, the concentration of the pollutant i n the remaining water. The forest vegetation is a much more effective filter of occult and dry deposition over grassland. These pollutants accumulate on the needle surface and enhance the acidity of throughfall. Growth of the forest naturally entails net base cation retention and as the crop is harvested this represents a direct removal from the catchment. Ultimately this denudes the soil base cation status and, therefore, the ability t o neutralise the acidic inputs. The accumulation of organic matter results in a very acidic surface soil horizon which increases the solubility of aluminium. However, if the leached aluminium is in a non-labile organic form this would decrease its bioavailability in surface waters. Modification of the water pathways can change the quality of the throughflow. Improving the drainage can result in more rapid surface runoff and, therefore, reduce the potential to neutralise the acidic inputs. It can also increase the hydrological response time of the catchment, promoting acidic episodes. The presence of a ground cover can provide a n extra source of acid neutralisation. With the development of the coniferous forest canopy, less light is able to penetrate to the forest floor, which results in a barren surface. The only ground vegetation remaining is at the forest edge where sufficient light is available. Although it is difficult to apportion any of the enhanced acidification effect t o any one process it is fair to say that conifer forests i n areas of high acid deposition enhances surface water acidification (Battarbee, 1989). Deforestation results in a decrease in acid loading to the system. However, following tree felling, acidification pulses have been associated with elevated nitrogen concentrations (Neal et al., in press). Whilst the presence of the forest retains and cycles the nitrogen in the soil its removal creates a nitrogen saturation effect. Replanting the system has a n enhanced effect unless a source of base cations are added. The already low base saturation of the soil will be further diminished by I'orest growth and will cause further surface water acidification. Future agricultural practice will have significant implications for acidification research. Intensive livestock farming has already been mentioned in connection with ammonia emissions. I t is quite clear the Netherlands have significant problems and if a similar situation develops in other countries, national abatement policies will need to be invoked. Fertiliser application significantly outweighs any acidic deposition particularly in managed areas. Nitrate leaching to surface and groundwaters is already a problem i n some localities although it is very much restricted to lowland areas. However, if the buffering capacity of the higher order trihutaries are reduced acidification might penetrate further down the river system. The application of persistent organics for increasing biological

51

control might deteriorate surface water quality and inflict biological stress. The influx of acidic waters might have synergistic effects on biota. It is quite clear that as land becomes more intensively managed an increase in fertiliser use and pesticide control will be required for increased food production this might create added problems in lowland catchments. The need for optimal land management strategies including timing and form of application will be necessary to reduce pesticide and fertiliser leaching. The only widespread farming practice in the uplands is improved pasture for sheep grazing. Limestone is applied to most of these improved sites and, therefore, provides a possible acid neutralisation source. However, as the improved pastures tend to be on the upper slopes which are not hydrological source areas their influence on surface water quality is minimal (Waters et al., 1991). Trends in agricultural practice and location are more likely to be politically driven, however, climate change might impose some areal restrictions. The future scenario for the south east of the United Kingdom is increased temperatures and decreased rainfall. Hence, farmers may wish t o extend into the uplands t o receive more rainfall unless irrigation is to be utilised more widely.

2.3 Climate Change Climate change is certainly the global environmental problem of the nineties. Many future scenarios have been suggested and the implications on a global scale are variable. The absolute magnitude of the changes are also tentative but it is quite clear that they will interact with acidification research. Many of the predicted changes will have direct affects on many processes, others will influence fluxes in an indirect way. The following paragraphs incorporate the main climate change scenarios envisaged for Northern Europe and summarises the impact on soil and water. Of course the implications of these changes can be extrapolated to a global scale. The following list includes the main changes which will have direct implications for acidification research; a ) Weather patterns (frequency of air masses) b) Temperature (mineralisation of N, rate of biochemical reactions) c) Precipitation (snow vs rainfall, volume, seasonality) d) Evapotranspiration (concentration effects) e ) Soil moisture (waterlogging, drying, water pathways) fl Vegetation periods (i.e. growing season, timing of fertiliser application)

As climate changes o n a global scale it is likely this will affect local weather conditions. The frequency of air masses might change and a shift from less pollutant loaded to more pollutant loaded air masses will increase local deposition, of course the reverse is also applicable. The implications for temperature change are rather more obvious. The rate of biochemical reactions will increase with an increase in temperature, e.g. nitrification and denitrification. An increase in temperature will increase organic nitrogen mineralisation which with a n increase in nitrification will lead to nitrate and H' production, acidification and possibly eutrophication. Perhaps more important are the predictions for precipitation change, both

52

in annual volume and seasonality. An increase in rainfall as forecast for much of northern Britain which will increase the amount of wet and occult (mist) deposition. In turn an increased volume of water will pass through the soil and, therefore, leach larger amounts of cations. Increased rainfall will also lead to a decrease in the soil moisture deficit and in some cases to waterlogging. These reducing conditions might restrict the mobility of some elements and, therefore, However, i t might increase heavy metal transport t o surface waters. concentrations such as manganese and iron. This increase in water source area within the catchment will increase the proportion of land area having a direct influence on surface water quality. With changes in the height of the water table, hydrological pathways will change and influence the quality of leachate reaching surface waters. It might follow that an increase in precipitation concurrent with a rise in temperature will increase the fraction falling as rain, thereby, reducing the winter storage of accumulated snowfall and perhaps reducing the i d u e n c e of spring meltwater. It should also follow that an increase in temperature would increase the evapotranspiration rate and, therefore, the concentration of pollutants. However, the seasonality of precipitation will also play an important role, perhaps increasing the frequency of acid episodes through more frequent and greater storm runoff and altering the distribution of deposition with wetter winters and drier summers. The converse of course might be true for the south of Britain. The climate change scenario equates to a small reduction in the amount of rain with again an increase in temperature. This will further reduce the effective rainfall and no doubt lead to larger soil moisture deficits. The drying of the surface soil horizons will increase the solubility and, therefore, mobility of some elements, particularly, sulphate, nitrate, organic acids and aluminium. Increasing the mobile anion source will lead to enhanced leaching of cations, both base metals, H' and aluminium. Increased mobilisation of organics, however, will provide a suitable ligand to reduce the bioavailability of aluminium and, therefore, the toxicity of the surface waters. It is quite clear that the possible ramifications of climate change are complex and do not necessarily exacerbate acidification in all cases. However, i t is obvious that climate plays a vital role in controlling many of the major processes and, therefore, needs to be incorporated in future acidification studies.

3. Conclusions a n d F u t u r e Research Any future change or trend in primary and secondary pollutants will have a direct affect o n acidification research. As protocols for emission abatement are modified for certain pollutnnts and invoked in the future for other pollutants, these changes must be monitored and affects recorded. Many of the major processes studied within the acidification field are influenced by a number of factors. The affects of land use change are already being addressed in many research programmes and i t is quite clear that future land management strategies need to be investigated. The influence of various land use types still needs clarification and the processes operating in coniferous forest catchments to enhance surface water acidification need to be quantified. The implications of

53 climate change, including weather modifications, regional variations in precipitation and temperature obviously effect a lot of major processes both chemical and physical and require study. Future research programmes must address these interactions. Integrated monitoring programmes are a necessity and must run in parallel with modelling studies. Further refinement of processes will improve model extrapolation and, therefore, ease in predicting the environmental impacts. Models such as MAGIC already incorporate changes in land use and runoff i n conjunction with deposition scenarios and this must continue, for alternative land use types in other localities, to estimate the resource a t risk. New models on nitrates in upland soil systems are required. New proposals such as the CLIMEX programme should receive support such that climate change interactions can be studied. This experiment plans to investigate the effect of CO, enrichment on a forested headwater catchment and record its effect on many processes. Development of links between hydrochemical and biological models should be encouraged so that inputs to the ecosystem can be related to effects. In this way future abatement strategies for emission control will have greater ecological significance. A new research programme has just commenced a t the Institute of Hydrology to investigate the impacts of hydrology and climate change on biochemical and ecological systems (Eatherall et al., 1991). The project 'ICE' plans not only to link models but also couple the models to databases and a GIs. It is this type of approach that will help us to understand the impacts of environmental change on acidification.

4. References

Cosby, B.J., Whitehead, P.G., and Neal, C. (1986) A preliminary model of long term changes in stream acidity in South West Scotland. J. Hydrol, 84, 381-401. Eatherall, A , , .Jcnltins, A. and Finch, J. (1991) Modelling the climate change impacts 011 biocheniical and ecological systems: Core model project. Progress report to DOE, November 1991, Institute of Hydrology, Wallingford, Uk. Neal, C., Fisher, R., Smith, C.J., Hill, S., Neal, M., Conway, T., Ryland, G.P. and Jeffrey, H.A. (In Press) The effects of tree harvesting on stream water quality at an acidic and acid sensitive spruce forested area: Plynlimon mid-Wales. J. of Hydrol. Waters, D.J., Jcnltins, A , , Staples, T.B. and Donald, A. (1991) The importance of hydrological source areas in terrestrial liming. ZWEM, 5, 3, 336-341. Whitehead, P.G., Reynolds, B., Hornung, M., Neal, C., Cosby, J. and Paricos, P. (1988a) Modelling long term stream acidification trends in upland Wales at Plynlinion. Flydid. Proc., 2, 357-368. Whitehead, P.G., Bird, S., Hornung, M., Neal, C., Cosby, J. and Paricos, P. (1988b3 A modelling study of lhe Llyn Brianne catchments. J. of Hydrol., 101,191-212.

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T. Schneider (Editor), Acidification Research. Evaluation and Policy Applications 0 1992 Elsevier Science Publishers B.V. All rights reserved

55

The role of ammonia in acidification. Perspectives for the prevention and reduction of emissions from livestock operations A.A.

Jongebreur and J.H. Voorburg

Institute of Agricultural Engineering Wageningen, The Netherlands

(IMAG-DLO),

P.O.

Box

43, 6700 AA

Abstract Ammonia emissions from livestock operations have a considerable contribution to acid deposition. Almost 50% of the potential acid deposition is in the form of ammonia. As ammonia is a nitrogen fertilizer it moreover contributes t o the eutrophication of forests and other natural vegetations. Prevention and reduction of ammonia emission is connected with the composition, the handling and storage and the spreading technology of the liquid manure. Adaptation of the feed rations for dairy cows, fattening pigs, laying hens and broilers to the need of the animals can result in a reduced excretion of nitrogen and thus influence the ammonia emission. More exact figures of this type of reduction must be delivered by further research. Landspreading of liquid manure can be carried out with a reduced emission of ammonia in comparison with above-ground application with the help of tine injection and trenching techniques. However, attention must be paid to special soil types (a.0. peat soil) and the protection of hatching birds. Research and development in the field of livestock housing gives rise t o the perspective that within the coming years the ammonia emissions from this source can be reduced with about 50%. In practice the application of livestock housing systems with reduced ammonia emission will also depend on the extra costs and the standards for animal welfare.

INTRODUCTION The results of the experiments by Van Breemen et al. [ l ] indicating that ammonia gave an important contribution to the problems of acidification was an unpleasant surprise f o r the Dutch livestock production. In that period the Government and the farmers' organizations were just considering the possibilities for a limitation of the loading rate of animal manure per ha of cultivated land. In that approach the attention was concentrated on phosphates. The fact that nitrogen, when evaporated as ammonia could have an impact on the environment complicated considerably the problems regarding livestock production on the one side and environment at the other side. This paper describes shortly the effects of ammonia deposition on forests and heathlands, as became apparent from the research carried out in the framework of the Dutch Priority Programme on Acidification. The main part however of this paper is connected to the research aiming at a reduction of the ammonia emission and the application of the relevant technology in livestock production enterpriees.

56 1. EFFECTS OF W O N I A DEPOSITION

Old farming handbooks learn that ammonium sulphate is a nitrogen fertilizer with as a side-effect acidification of the soil. This also holds more or less for the effects of ammonia deposition. In the final report second phase Dutch Priority Programme on Acidification 121 it is concluded that with regard to nitrogen the acidification problem is closely linked with the problem of N eutrophication. As ammonia is a basic compound deposition of ammonia is a potential acid load. The difference between potential acid load and actual acidification of the soil is mainly the result of uptake by plants and denitrification. The current contribution of N to the actual soil acidification is estimated to be about 35%, the contribution of S being about 65 %.It should be taken into account that only a part of the N deposition is in the form of ammonia. Almost one third is deposition of NO, (Table 1).

Table 1 Estimated deposition of acidifying compounds in the Netherlands (mol ha-' yr-') in 1989 670 1160 2190 100

Total N

3350

Total acid

4800

Source: Erisman and Heij, [3].

The means tions times

'fertilizing' effects of ammonia are more complex. Acidification that ammonia is converted into nitrates by nitrification. In situawith little nitrification in the top soil the NH4/K ratio is someabove the critical value of 5 which implicates a low availability of

K.

In forest ecosystems in the Netherlands N is usually not a growth limiting factor. An increased availability of N strongly stimulates the growth of nitrogen loving grasses and herbs. The increased N content leads in combination with soil acidification to disturbed nutrients ratios, including nutrient deficiency. Thus the forest ie becoming more vulnerable to damage resulting from diseases, pests, frost and drought. A distinct relationship has been found between increased N contents and the occurrence of Sphaeropsis. The effect of a high N deposition on heathland ecosystems should be considered in combination with stress factors such as frost, drought, heather beetle plagues and natural aging. If as a result of these factors the Calluna canopy is opened, the heather will be crowded out by grasses. The critical nitrogen load for the crowding out an open Calluna canopy by grasses is about 10-15 kg N ha-' yr-'. Finally the high nitrate content in forest soil leads to a high nitrate content in the groundwater [ 2 ] . Research on the composition of shallow

57 groundwater at 150 different locations (forests and heathland) has shown that the nitrate content was higher than the drinking water standard in almost 30% of the coniferous forest sites investigated. A survey of the average critical nitrogen load on sandy soils in the Netherlands is given in Table 2. Heij and Schneider [ 2 ] draw the attention to a correct interpretation of the definition critical load. If 'critical loads' are exceeded temporarily this will certainly not mean the death of the forests. It does, however, involve a certain risk to the vitality of forests, and this risk will increase as critical loads are exceeded further and for a longer period. Evaluation of critical loads shows that most effects of nitrogen deposition will not occur below loads of 400-600 mol ha" yr-'.

Table 2 Average critical nitrogen loads (mol ha'' yr-') for terrestrial ecosystems on well drained sandy soils in the Netherlands Coniferous forests Vegetation changes Elimination by grasses Frost damage/fungal diseases Nutrient imbalance Nutrient leaching to groundwater

400-1400

Deciduous forests

Heathlands

600-1400 700-1100

1500-3000 800-1500 900-1500

1700-2900

2000-3600

Source: Heij and Schneider, (21.

2. VOLUME OT AEMONIA EMISSIONS The contribution of NH, to potential acid deposition in 1989 was 46%. Different from SO, and NOy 81% of deposition of ammonia is from national origin. This in comparison with the fact that 55% of the total potential acid deposition was caused by emissions in the Netherlands itself (21. This means that national measures to reduce ammonia emission are relatively more effective. Initial estimations of the ammonia emissions from different sources suffered from lack of information. Increasing research efforts resulted in more reliable emission factors though the information of some sources is still insufficient or disputed. Within the Dutch Priority Programme it has been decided t o estimate the total ammonia emission by using the consistent set of emission factors. The calculated emission factors are based on assumptions regarding the evaporation rates of NH3 from the manure under the different conditions. The comparison with the measurements clearly demonstrate that the uncertainty in the calculated emission factors is very large. Therefore it is necessary that the measurements of ammonia emissions will be continued in order to obtain more accurate figures [2]. The most recent estimations are at a level of about 250 k ton NH3 per year mainly caused by livestock production (Table 3). Asman and Jaarsveld [ 4 ] mentioned that the ammonia emission in 1980 was a factor 5 higher than 1870 and the largest increase has occurred since 1950 [4]. Table 3 shows that livestock production is by far the most important

58 source of ammonia emission. Prom the heading "other sources" it is worthwhile to mention fertilizers (10 k ton yr-') and industry (7.6 k ton yr-l). The awareness of the important contribution of livestock production to the problems related with acidification has lead to the extension of the existing research program devoted t o the utilization of animal manure with a programme aiming at the measurement, prevention and reduction of ammonia emission. Table 3 Total NU3 emissions in the Netherlands in 1988 in k tons yr'l tages %

CATTLE stable and storage manure spreading pastures total cattle

and percen-

of total emiseioq

35.5 58.5 24.5

118.5

47.8

67.1

27.1

PIGS

stable and storage manure spreading total pigs POULTRY stable and storage manure spreading total poultry Sub total Other sources Total

26.0

41.1 17.1

12.3 11.9 13.2

Source: Heij et al., [5]

3. PREVENTION AND REDUCTION OF AMMONIA EMISSION

Due t o the total quality of the ammonia emission we cannot state: The best solution for pollution is dilution. Prevention and reduction are the key-words for action. The prominent role of ammonia in acidification and eutrophication of Dutch forests has lead t o special attention to the reduction of the ammonia emission in the National Acidification Abatement Plan. It is the aim of this plan to realize a reduction of the ammonia emission of 30% in 1994 and of 70% in the year 2000 which brings back the ammonia emission level t o the average of about 1950. As a first step more information had to be collected on the contribution of different sources of ammonia emission and o n the factors influencing this process. During the first phase of the Dutch Priority Programme on acidification measurement techniques have been developed t o determined emissions during and after landspreading of manure and from manure stores. Housing systems for pigs and poultry often have a forced ventilation. An automatic system was developed for continuous registration of the emission from these houses.

59 Housing systems for cattle usually have a natural ventilation. This complicates the measurement of the ammonia emission considerably. A model is being developed to determine the emission under conditions of natural ventilation. Meanwhile some cattle houses at the experimental farm of IMAGDLO at Duiven are equipped with a forced ventilation and an automatic monitoring system t o registrate the ammonia emission. Also in the research on and in the development of systems with a reduced ammonia emission it is necessary to realise that ammonia-N is a crop nutrient. The evaporation of ammonia should be considered in the framework of the total N balance of agricultural production (Table 4).

Table 4 Rough N balance of Dutch agriculture in 1985-1986 (1000 ton N-yr') N invut Inorganic fertilizers Animal feeding stuffs Different sources Total

498 483 139

1.120

N outuut Crops Animal products Difference

95 170 855

1.120

Source: Van der Meer, [ 6 ] .

The difference between input and output in Table 4 is an indication of the yearly losses of nitrogen in Dutch agriculture. A considerable but still unknown part of these losses is caused by denitrification. From the point of view of protection of the environment attention has to be paid t o the part of the difference which is supposed t o be caused by the evaporation of ammonia and leaching and run-off of nitrogen both in the order of roughly 200.000 ton N per year. 3.1. Input of nitrogen From the nitrogen balance (table 4 ) it is evident that too much nitrogen is lost in Dutch agriculture. A research programme aiming at the prevention of losses anyway should try to reduce the input. Most efforts in this research programme are concentrated on animal nutrition since half of the nitrogen excreted by the animals is in the form of ammonia. Research carried out in cooperation with the feed stuff industry has proved that a more exact adaptation of the composition of the ration to live-weight and productivity of the animals results in a saving of the nutrients P and N. For pigs and poultry in this way a reduction of P input of 10-15% and of the N input up to 5% has been realized. A stronger reduction of the N excretion by pigs and poultry is possible when the composition of the proteins in feeding rations can be optimized by the addition of amino-acids. In this way a reduction of the N excretion with 20-25% is possible. The utilization of synthetic amino-acids in the feed rations for pigs and poultry is mainly influenced by economic factors (prices of the synthetic amino acids) [7]. The N excretion by cattle depends on a number of factors connected with the composition of the roughage and the more complicated digestion system of these ruminants. An important factor is the application rate of fertilizer N. Too high application rates result in a surplus of proteins in the

60 grass in comparison with the needs of the cows. During grazing of the cows 3-13% of the excreted N is evaporated as ammonia [ a ] . 3.2 Landaprmading

The main source of ammonia evaporation is landspreading of manure. Table 3 shows that landspreading is responsible for roughly 50% of the ammonia emission. It is Sometimes assumed that the conventional spreading technique is responsible for most of these losses. This impression is possibly influenced by the strong smell produced by these activities. Experiments by Pain and Klarenbeek (91 have shown that emissions during spreading are less than 1% of the total ammonia emission of spreading and the liquid manure while laying on the land. The NH3 emission is mainly caused by the manure laying on the surface of the cultivated land or on the crops. The best remedy for the prevention of these emissions is fast incorporation into the soil by ploughing or by injection techniques. In this way the emission can be reduced on average with at least 80%. Other techniques are being developed for situations where injection or ploughing immediately after application of the slurry is not possibly due to soil type and composition or the necessity to protect hatching birds in the meadows. Tine injection of liquid manure is not so good applicable on peat and heavy clay soils. Without incorporation of the liquid manure in the soil 20-100% of the ammonia from animal manures evaporates after landspreading. Next to the positive effect of reduction of ammonia emission by direct incorporation on acidification there are two other important effects namely: - The farmer can save on the fertilizer cost; As the natural evaporation of ammonia is strongly depending on weather conditions, it is a very uncertain factor. Prevention of the evaporation enables the farmer to the application of a more exact amount of the manure. In this way, the risks for leaching and crop damage can be reduced. The further development of appropriate systems is stimulated by an extra facility for measuring the ammonia emission of different types of spreading techniques. The introduction of the mentioned techniques at farm scale is stimulated by demonstration projects and other activities of the Advisory Service. An increasing number of farmers is applying injection, sod injection and trenching (disc injection into the sward layer) techniques for at least a part of the produced animal manure. Since the reduction of acidification is a high priority the incorporation of manure is obliged in a part of the Netherlands namely on sandy soils. This obligation will be extended to other soils in the course of the coming years (up to 1994), when all manure must be incorporated into the soil.

-

3 . 3 . Live8tOCk houaing Sy8t.l.

Housing systems are the second source of ammonia emission. From the point of view of emission prevention and reduction housing is much more complicated than landspreading. Except for grazing animals the manure is excreted in the house. The housing systems contain mostly a capacity for manure storage but moreover, the floor is polluted with freshly excreted manure and urine. The emiseion from a livestock house is not only caused by the manure storage but also by the surface polluted with fresh excreta. The pollution

61 of the floor is strongly influenced by animal behaviour. Most livestock housing systems for cattle and fattening pigs have a slatted floor with a manure storage under the slats. In cattle houses with a cubicle system and slatted floors roughly half of the NH3 emission is caused by the manure storage and the other half by the liquid manure on the stable floor. The emission from a closed floor with manure removal by mechanical scrapers is of the same magnitude [lo]. If the animals are kept on litter, the stable floor and the manure storage are combined. Thus far there is no information available about the quantity of the emission of this husbrandy system. In a short state of the art recent developments in the livestock housing systems the aspects with relevance for the ammonia emission are discussed. CATTLE Cattle is usually housed in loose housing systems with cubicles. In theory it would be possible to have a strong reduction of the NHg-emission by returning to the old-fashioned housing with the cows tied up. From the point of view of labour organization and animal welfare such a development is not favourable. The research carried out is aiming at a reduction of the emission from the stable floor by flushing systems and from the manure storage under the floor by acidification realized through adding of nitric acid. Flushing has the disadvantage that the volume of the manure is increased. Acidification with nitric acid can be attractive from an economic point of view as the costs of nitric acid are compensated by savings on the bill for fertilizer N. This depends, however, on the livestock density and the grassland management system on the dairy farm. The system is complicated by the risk of denitrification which results in a breakdown of the nitric acid and volatilization of the nitrogen. PIGS Like cattle the surplus N in the ration of pigs is excreted in the form of urea. Thus far no practical tool is available to prevent the mineralization of urea. The surface of the floor polluted by the excretion of manure and urine is in a pig house rather limited and influenced by climate in the house and the feeding system. This has as an effect that only one third of the emission is coming from the stable floor. Nevertheless, further efforts to influence the animal behaviour to keep the solid floor clean are worthwhile. Most efforts should be concentrated on the liquid manure stored under the slatted floor. Flushing systems with the aim of a fast removal of the manure with restricted ammonia evaporation are being developed and studied in experiments and research. Hoeksma et al. [12] report reductions of ammonia emissions in houses for fattening pigs by flushing up to 709. POULTRY Layers are mainly kept in cages, so the droppings fall directly into a storage pit or on a belt under the cage. Fast removal with this transport belt to a closed storage tank outside the house can reduce the NH3 emission with 60% or more [ll). To improve animal welfare the development of houses for layers to be kept on a floor with litter, perches and nests is stimulated. A floor with litter (and droppings) will increase the ammonia emission in comparison with the cage systems with drying of manure and fast removal of manure. A key for a further emission reduction is the fact that birds excrete the N surplus in the form of uric acid. Fast drying of the manure prevents the mineralization of uric acid. Alternative housing systems for broilers on the basis of fast drying of the manure have been developed and experiments

62 are carried out with these type of housing. According to more recent results these systems seems to reduce the ammonia emission by more than 70%. OUTSIDE STORAGE

Improved utilization of animal manure with prevention of nitrate leaching is only possible if landspreading is limited to a period during and just before the growing season of the crops. This necessitates to the extension of storage capacity on the farms. Namely on dairy cattle farms, this is frequently realized in the form of a storage facility outside the house. In fact this is a new source of ammonia emission. This emission from outside storage basins can be reduced by 70-90% in covering by a sufficiently closed roof or a floating layer. On cattle slurry mixed with some straw such a floating layer usually is formed spontaneously in the form of a crust. Such a crust is able to reduce the emission with SO-lS%, however, the formation of crust depends on the climate and the composition of the manure [13]. THE INTRODUCTION OF NEW HOUSING SYSTEUS

The knowledge on emission prevention and reduction from systems generated by research institutes is just one step in the development of a housing system. For the development of a fully integrated system an active participation of the farm building industry and practical farmers is necessary. To stimulate the realization of these practical developments the different commodities in Feed Stuff Industry, Meat Industry and farmers' organizations have provided a fund to subsidize practical experiments. Moreover, promising housing systems are demonstrated in a demonstration project called 'Pro-Pro' (County Programme on the prevention and reduction of ammonia emissions) in Noord-Brabant, which is supported by the Ministry of Housing, Physical Planning and Environment and the Ministry of Agriculture, Nature Management and Fisheries. To support practical experiments with housing systems the Ministry of Agriculture, Nature Management and Fisheries supports the facilities for measuring the ammonia emissions. The measurement techniques are based on the experiences of research institutes measuring ammonia emissions. 3.4 Manure processing The high number of livestock in the Netherlands results in a total production of liquid manure which can not be completely utilized on cultivated land situated within a reasonable distance. Since there is a legal limitation of the application rate of manure, some regions have a manure surplus. The total volume of this surplus is estimated to be 6 million tons on short term (1994) and 20 million tons on the long run (2000).

This surplus has to be processed in special manure processing plants. The actual state of the manure processing technology learns that the main endproduct of manure processing will be a dry manure with a high and standardized quality. The surplus manure will not contribute t o the ammonia emission caused by landspreading. The processing plant itself has hardly any emission of ammonia. The end-product on its turn will also cause little evaporation of ammonia. The technology of the manure processing plants developed thus far has proved to meet these conditions. Livestock enterprises which deliver manure to a processing plant don't need a large storage capacity. Depending on the type of storage these farms have a somewhat smaller emission. The emission of the housing system itself is not influenced by regular delivering of the manure to the processing plant.

63 4 . DISCUSSION

A relevant question to be put forward in discussions on the actual 'state of the art' is in how far a reduction of the ammonia emission originating from livestock farms with 70% can be realised. Assuming that the volume of livestock production does not change Table 5 gives an estimate of the possible emission reduction by different options a.0. adaptation of the feed ration, techniques for incorporation of liquid manure into the soil and adaptation of the livestock houses.

Table 5 Effect of different measures on the total NH3 emission (in Reduction percentage Actual emission Adaptation of the feed ration Techniques for incorporation liquid manure into the soil Adaptation of livestock houses

Housingstorage

Grazing

%)

Landspreading

Total

37

11

52

100

20

30

9

42

80

80

30

9

8

47

50

15

9

8

32

For the evaluation of this table the following points are of importance. A reduction of the N excretion of 20% by pigs and poultry on the short term is realistic [13]. More complicated is the role of grazing and utilization of proteins by cattle. The relation between the 19-content of the faeces and urine and the emissions demands further research. A reduction of the emission from landspreading with more than 80% is possible on soils where injection and trenching machines or ploughing immediately after application of the slurry are applicable. - In housing systems for broilers a reduction of more than 70% seems to be possible. For cattle housing a 50% reduction seems to be realistic. Housing systems for pigs and layers are not yet sufficient developed to make an estimate. The state of the art in the technology of livestock housing, storage tanks for liquid manure and spreading techniques shows that for the livestock branches a reduction of ammonia emission of about 70% can be achieved. However, attention must be paid to such a design of the housing systems that the conditions for the standards of animal welfare can be fulfilled. In this phase it is not possible to calculate exact costs for the adaptation of livestock housing [14].

-

5 . REFERENCES

1. N. van Breemen, P.A.

Burrough, E.J. Velthorst, H.F. van Dobben, T. de Wit, T.B. de Ridder and H.F.R. Reijnders, Soil acidification from atmospheric ammonium sulphate in forest canopy throughfall. Nature Vol. 299 (1982) 548-550.

64 2. G.J. Heij and T. Schneider (eds.), Final report second phase Dutch priority programme on acidification, RIVM, Bilthoven, No. 200-09 (1991) 234. 3. J.W. Erisman and G.J. Heij, Concentration and deposition of acidifying compounds. In: Heij, G.J. and T. Schneider (eds.), Final report s e c o n d phase Dutch priority programme on acidification, RIVM, Bilthoven, No. 200-09 (1991) 51-96. 4. W. Asman en J.A. Jaarsveld, Gedrag van atmosferisch ammoniak. In: P. Del Castilho, W.H. Rulkena en W. Salomons (eds.), Dierlijke meet problemen en oplossingen, KNCV Symposia (1990), KNCV, 's-Gravenhage, 127-165. 5. G.J. Heij, J.W. Erisman and J.H. Voorburg, Ammonia emissions and abatement. In: Heij, G.J. and T. Schneider (eds.), Final report second phase Dutch priority programme on acidification, RIVM, Bilthoven, No. 200-09 (1991) 37-50. van der Weer, Stikstofbenutting en verliezen van gras- en 6. H.G. maisland, DLO, Wageningen, (1991) 134. Jongbloed and J. Coppoolse, Mestproblematiek: Aanpak via de 7. A.W. voeding van varkens en pluimvee, IWO-DLO, Lelystad, (1990) 131. 8. J.H. Voorburg and G.J. Monteny, Review of research into ammonia emission from liveetock. In: Schneider, T. and G.J. Heij (eds.), Thematic reports Dutch priority programme acidification, RIVM, Bilthoven, No. 200-07 (1990) 23. 9. B.F. Pain and J.V. Klarenbeek, Anglo-Dutch experiments on odour and ammonia emissions from landspreading livestock wastes, IMAG-DLO, Wageningen, Research Report 80-2 (1988) 36. 10. J. Oosthoek and W. Kroodsma, Ammonia emissions from cowsheds, Agricultural and Food Processing waste, Proc. of the Sixth International Symposium on Agricultural and Food. Proc. Wastes. Am. SOC. of Agr. Eng., St. Josehp, Michigan, (1990) 442-449. 11. J. Oosthoek, W. Kroodsma and P. Hoeksma, Betriebliche Massnahmen zur Minderung von Ammoniakemissions aus Stallen. In: Ammoniak in der Umwelt, KTBL, Darmstadt, VDI, Dusseldorf, (1990) 29.1-29.23. 12. P. Hoeksma, M. Daanen and J.A.M. Voermans, Reduction of ammonia emission from pig houses. In: Environmental challenges and solutions in agricultural engineering, University of Norway, Dept. of Agr.Eng., (1991) 81-91. 13. M.J.C. de Bode, Vermindering van ammoniak-emissie door korstvorming op rundveemengmest, IMAG-DLO, Wageningen, No.226 (1990). 14. A.W. Jongbloed and N.P. Lenis, Nutrition as a mean to reduce environmental pollution by pigs, Paper 4nd Ann. Meeting EAAP, Berlin, 8-12 Sept. 1991.

T Schneider (Editor), Acidification Research Evaluation and Policy Applications @ 1992 Elsevier Science Publishers El V All rights resewed

65

Emissions of acidifying components Markus Amann International Institute for Applied Systems Analysis (IIASA), A-2361 Laxenburg, Austria

Abstract Emissions of SOZ, NO, and ammonia are major contributors to acidification of natural ecosystems. The paper reviews purpose and methods of emission inventories and compares the major international inventories in Europe. For each of the pollutants the major gaps in current knowledge and discrepancies in existing data are briefly discussed. Finally, national differences in per-capita emissions are analyzed for the year 1985. 1. INTRODUCTION Emissions of sulfur dioxide, nitrogen oxides and ammonia are important contributors to acidification of ecosystems that has been of growing concern in many countries in Europe and North America. Due to the complexity of the interaction of emission generation, atmospheric dispersion of pollutants and environmental impacts improved understanding of the major processes seems necessary to design efficient emission reduction strategies. An important prerequisite for such a better understanding is the accurate knowledge of emission rates. This paper summarizes the status of some important aspects of emission accounting in Europe and tries to identify the major gaps in current knowledge.

2. THE PURPOSE OF EMISSION INVENTORIES Attempts to estimate anthropogenic emissions have been made since harmful effects of emissions have been recognized. Early efforts had exploratory character and tried to estimate magnitudes of emission releases. Over time techniques have been refined and emission estimating changed from a scientific adventure to bureaucratic routine, often performed with passion. In order to put resources spent for emission inventorying into relation to the final use of the results the ultimate purpose of emission accounting should be clarified. In general two different objectives can be identified: Emission inventories are considered as an important element of national environmental statistics. They are used to compare present emission rates against historical records and thereby to demonstrate the success (or the failure) of environmental policies. Recent economic research tries to incorporate data on the release of harmful substances to the environment into the general system of national accounts to

66 quantify also negative effects of economic activities. 0

Emission data are an essential input to atmospheric dispersion models, which aim at the understanding of chemical processes and the behavior of air pollutants in the atmosphere.

With the first objective no natural limitation does exist for the ultimate resolution and accuracy of emission estimates. Currently a general trend exists towards a continuous improvement of inventories, e.g., by increasing the temporal, sectoral and spatial resolution of inventories. Resources required to achieve results grow simultaneously. If emission data should provide input to model calculations, the appropriate level of resolution and accuracy can be derived from the specific model structure. It has to be stated that up to now only few cases have been reported in which ’officially’ compiled inventories have actually been used for model calculations (the EMEP model for longrange transport of sulfur and nitrogen compounds [ll]is one example). The majority of modelling studies could not rely on inventories covering their actual data demand and had therefore to estimate the necessary data themselves. Although emission inventories have improved substantially over the last years (not at least with the aim to provide reliable input data for model calculations) it seems questionable if an ’all-purpose’ inventory satisfying all modeller’s demands could ever be developed. Scientific models do by their nature aim at exploring new and unknown aspects of air pollution and will therefore in many cases be one step ahead of institutionalized emission reporting systems. 3. METHODS OF EMISSION INVENTORIES Emission estimates can be established by two different methods: 0

The release of pollutants can be directly monitored at each emission source. In order to obtain reliable overall data continuous measurements over time of all emission sources have to be performed. Such measurement programs are expensive and therefore only applicable to important point sources. Monitoring of small and distributed sources is difficult and resource intensive.

0

More often emissions rates are calculated based on surrogate data. Most commonly, the simple formula

EMISSION = A C T I V I T Y . EMISSION FACTOR

(1)

is applied to estimate total emission from a specific source or source category. Activity rates are usually available from statistics. Proper emission factors can be extrapolated from measurements, extracted from literature, derived from the plant’s commissioning license or calculated based on mass balance approaches. Recently, comprehensive information on emission factors has been compiled in handbooks [7]. The estimation of emissions according to this formula contains a number of inherent uncertainties and potential sources of inaccuracies. Often activity rates are considered as more accurate than emission factors. However, experience shows that also the quality of

67 activity data rapidly declines with finer temporal and spatial resolution. In many eastern European countries statistical data have been classified for a long time and statistics published now are still contradictory in many cases. 4. INTERNATIONAL EMISSION INVENTORIES

The long residence time of many air pollutants in the atmosphere results in strong international interdependencies of air pollution in Europe. Strategies to improve air quality should therefore address domestic and transboundary sources of emissions. Consequently, also emission accounting as one basis for decision making should provide international consistency. In recognizing this demand several efforts were made in Europe to establish international emission inventories for the major air pollutants. The following two tables provide a survey of the most important (completed or envisaged) inventories that are based on nationally submitted data. Although each of these inventories establishes comparability among countries, commonalities among the inventories are limited. Differences occur in their time horizons, the pollutants covered, the sectoral aggregation of emission sources, the spatial resolution, etc. Table 3 compares emission estimates for SOz and NO, for the year 1985 from different sources. Data listed under UN/ECE have been submitted officially to UN/ECE or have been estimated by EMEP [21], [ll]. CORINAIR data are retrieved from the final document of the CORINAIR project, which was carried out for the Commission of the European Communities [6]. For comparison two estimates of independent sources are

Table 1: Spatial resolution of international emission inventories. (Table based on Joerss [13].) Collecting Bodies Area Resolution 150 x 150 km according to EMEP system UN-ECE Convention ECE countries

UN Statistical

ECE countries

National totals

CEC-CORINAIR 85

EC countries

Territorial Units according to administrative regions

CEC-CORINAIR 90

ECE countries

Territorial Units according to administrative regions

OECD/EUROSTAT

OECD countries National totals

OECD/MAP

OECD countries 50 x 50 km according to EMEP system

IPCC (assumed)

global

Commission ECE

National totals

68 Table 2: Comparison of international emission inventories: Types of pollutants covered by the 9.

Pollutant

UN-ECE/ EMEP

UN’

X

x

so2

Data T

X

NO.

S T S

X

T

X

co

X

x

Inventory CORINAIR OECD/ OECD/ 1985 1990 EURO- MAP STAT 1980 x

X

X

X

x

X

X

X

X

X S X X T NH3 X S x x X T NM-VOC X X S T X CHr T NzO X T co2 X PM T X Pb T X T Hg T X Cd Table based on Joerss [13] T: National Totals Inventory S: Spatial Inventory

X

X

X

IPCC (assumed) X

X

X

X

X

X X

X

X

X

X X

X X X

X

X

X

X

X

X

X

X

NM-VOC Non-Methane Hydrocarbons PM Particulate Matters 1 UN/ECE Statistical Commission provided which are calculated with more aggregated information: The inventory of the IIASA-RAINS model for SO1 and NO, emissions is based national energy statistics and fuel characteristics [l].Estimates of European NO, emissions have been compiled by the Norwegian Institute for Air Research (NILU) [18].

69 Table 3: Estimates of emissions of SO2 (in k t SO2) and NO, (in kt NOz) for the year 1985. SO1 (kt SO1) NO, (kt NOz) UN/ECE CORIN- IIASA UN/ECE CORIN- IIASA NILU AIR RAINS EMEP AIR RAINS EMEP Country 1111 161 PI 1111 [GI 111 1181 123 92 Albania 50’ 34 30 Austria 178 250 240 230 199 Belgium 392 452 281 416 342 317 478 Bulgaria 1069 1034 367 284 150 3150 3100 1127 CSFR 769 544 Denmark 333 270 340 268 307 330 258 Finland 382 251 230 287 353 France 1481 1605 1470 1806 1615 1796 1761 Germany, West 2315 2450 1715 2830 2617 2407 2930 Germany, East 5340 5080 7013 a75 876 Greece 288 500 308 500 276 526 746 Hungary 1404 1546 273 265 262 Ireland 141 85 140 136 91 89 97 1573 Italy 2089 2504 1563 1486 2509 1595 Luxembourg 22 16 16 16 19 33 27 Netherlands 200 471 276 588 462 280 544 Norway 98 166 212 100 203 Poland 4300 1248 1374 3638 1500 198 Portugal 198 157 147 263 96 Romania 1738 390’ 1900 604 680 2190 Spain 2190 1065 2220 950 991 Sweden 270 394 339 329 271 Switzerland 214 96 203 210 90 3767 2322 3562 2278 UK 3676 2311 11110 19190 USSR’ 7111 5532 3369 Yugoslavia 1500 1364 400 494 420 Notes: European part of USSR within the EMEP area EMEP estimate 1987

5. CURRENT GAPS AND UNCERTAINTIES OF EMISSION ESTIMATES 5.1 Inventories of SO2 emissions

Since SO2 emission inventories have a relatively long tradition and emission factors are mainly determined by fuel quality and emission control equipment, accuracy of estimates is generally high. Uncertainties have been estimated to be in a range of f 10 percent 18). Aircraft measurements and air monitoring data seem to confirm these evaluation, at least as long as aggregated figures are considered. However, factors discussed above cause

70

important discrepancies and discontinuities in submitted data for some eastern European countries. According to earlier official submissions to UN-ECE/EMEP Romania’s sulfur emissions in the year 1980 amounted to 200 kt of S 0 2 . In 1990, this number was officially increased by a factor of nine to 1800 kt [21]. Major interannual variations in emissions are often caused by economic instabilities. As an example, the development of Romania’s SO2 emissions between 1980 and 1990 are displayed in Table 4. According to these data emissions declined within only two years (between 1988 and 1990) by 40 percent.

...

...

1985 1912 1986 1802 1987 1762 1988 2397 1989 1646 1990 1430 Source: Romanian Ministry for Environment. Uncertainties are not only associated with quantitative estimates of emissions, but also with their spatial distribution. As an example Figures 1 and 2 compare the spatial distribution of Ukrainian emissions for the year 1985 (as estimated in 1988) and for the year 1988, as it has been officially submitted in 1990 when the critical loads approach came into discussion. Whereas in 1985 an emission peak occurred in the EMEP grid 32/27, in 1988 these emissions appear distributed over four neighboring grids cells. The overall accuracy of emission inventories does not only depend on the reliability of data on known emission sources. Maybe more important is the completeness of the inventories. Over the last years increasing attention has been attributed to emissions from ships [4]which turned out to be an important source in north-west Europe, but which have not been considered by earlier inventories. Marine emissions on the North-Sea overrule e.g. the Norwegian land based SO1 emissions by more than a factor of two (Table 5). First estimates of biogenic sulfur emissions from the North Atlantic Ocean indicate also non-negligible amounts of Dimethyl sulphide (DMS) from the marine troposphere (Table 5). 5.2 Inventories of NO, emissions

Estimates of NO, emissions are still associated with major uncertainties. Several methods to estimate emissions from different sectors are currently in use, and even measurements on plant level suggest a wide range of emission factors for the same type of technical equipment. Therefore, also in cases with comparably good statistical material available (e.g. for the Netherlands) the accuracy of overall estimates is considered to be in the range between 10 to 20 percent [3].

71

2 zn o

40

114

5

I 10

27 0

26 0

2

150

~

25.0

Figure 1: Sulfur emissions in the Ukraine in 1985 (in kt S), Source: Eliassen et al., 1988

28 o

27 0

2cC

-.

117

c ~

517

2s c

Figure 2: Sulfur emissions in the Ukraine in 1988 (in kt S). Source: Iversen et al., 1990

72

Table 5: Comparison of anthropogenic and biogenic emissions in the North-West of Europe. Sulfur emissions (kt S’, Yearly Monthly 11988) emissions Emission area Anthropogenic: 121 10 Denmark Ireland 76 6 Norway 33 3 United Kingdom 1832 152 Int. trade, North Sea

87

Biogenic: North Sea (DMS) 21 North Atlantic (DMS) 1035 Source: Tarrason, 1991

7

0.1 - 5 10 - 188

Similar to SO1 substantial discrepancies occur among different estimates for eastern European countries. One striking example concerns the total amount of NO, emissions from the CSFR. Western experts consider official data on NO, emissions from Czechoslovakia (1127 kt NO, in 1985 [21]) as much too high and suggest significantly lower figures (e.g., Pacyna et al. [18] estimate a national total of 544 kt). Large disagreement exists also for emissions from the Soviet Union. Official data report e.g for 1985 3369 kt of NO, for the European part of the Soviet Union [21]. Analysis recently undertaken at IIASA, which is based on new statistical information on regional energy consumption of the Soviet Union computes 60 percent higher emissions (5530 kt NO,), if average emission factors are being applied [19]. Major uncertainties of NO, estimates are associated with emissions from mobile sources. At the moment two different approaches are applied to compute emissions from the transport sector: The more simple method applies a single surrogate emission factor to the amount of fuel consumption. This surrogate emission factor reflects the average of all driving modes and is assumed as representative for the entire car fleet and all driving conditions. Consequently, considerable differences occur in such emission factors used by individual countries. The other approach calculates emissions as composite of numerous driving modes with specific emission factors for individual car categories, driving cycles, etc. This approach has been used by some countries to estimate transport emissions for the CORINAIR inventory. However, this method puts high demand on the availability of accurate statistical data on traffic conditions (e.g., mileage, fleet composition, etc.) which can not always be satisfied. The low accuracy of the underlying statistical material caused serious reservations against this ‘bottom-up’ approach and no general consensus on which approach to prefer does exist at the moment.

73 5.3 Inventories of NHs emissions

Whereas emission estimatesfor SO2 and NO, have some history and are now performed mostly on a routine basis the estimation of ammonia emissions is a relatively new field for most countries. Uncertainties are not only caused by the lack of experience, but also, as recent research indicates, by the general stochasticity of emission factors: NH3 emissions are depending inter alia on the nitrogen content of fodder, on the weather conditions at the time of manure application and are influenced by particular managements practices (e.g., rotational or continuous grazing, etc.). Table 1: Estimates of emissions of NH3 for the year 1987. Inventory Buijsman IIASA Asman EMEP National [51 ~ 4 1 ~ 4 1 121 1101 estimates2 1980/83 1980 1987 1987 1988 Year Country 21 25 27 32 24 Albania Austria 107 85 72 79 79 94 Belgium 102 105 123 82 Bulgaria 147 122 120 123 126 219 170 CSFR 200 128-222 1981 200 197 144 Denmark 129 111 116 103 155-196 1980186 Finland 44 61 43 56 49 52 1984186 841 974 France 709 782 1985 679 650 Germany, West 371 718 380 348-641 1986/88 529 533 274 242 Germany, East 207 90-157 1980/85 228 239 111 Greece 95 112 88 100 179 151 130 156 155 Hungary 90-157 1976187 188 117 130 Ireland 128 128 435 426 Italy 36 1 422 1987 359 366 Luxembourg 7 6 5 5 5 276 Netherlands 154-258 1982187 150 224 239 218 Norway 38 41 64 1987 36 37 47 561 Poland 405 570 528 478 Portugal 76 55 47 66 65 Romania 387 297 340 350 301 365 232 Spain 251 317 273 74 Sweden 52 66 59 62 74 Switzerland 53 64 59 62 64 1987 548 UK 482 492 405 451-560 1983/87 478 1543 1256 USSR' 2288 2446 3182 Yugoslavia 198 214 217 235 235 Notes: European part of USSR within the EMEP area According to [14] I

'

14

Since at the moment national estimates do not exist for all countries in Europe international inventories have been compiled based on centrally available information ([2], [5], [14]). In all these studies emission factors are primarily based on experience gathered in the Netherlands; only in few cases emission factors have been slightly modified to consider national differences. As Menzi has pointed out for the Alpine situation a simple extrapolation of Dutch emission factors to other countries does ignore country specific conditions and agricultural practices and consequently may result in significant inaccuracies of the estimates [15]. Not surprisingly, the similar emission factors applied in these studies limit differences in national estimates to f 15 percent, which is significantly lower than the general uncertainties of such estimates are thought to be. 6. PER-CAPITA EMISSIONS IN EUROPE

Figure 3 displays the potential contribution of S 0 2 , NO, and ammonia emissions in 1985 to acidification on a per-capita basis. Over all of Europe, SO2 emissions contributed roughly 60 percent to potential acidity; the remaining fraction originated from NO, and NH3 emissions at almost equal shares. Particularly high per-capita emissions occur in eastern European countries, mainly caused by the utilization of brown coal leading to substantial ,502 emissions. All six countries with highest per-capita SO2 emissions are located in eastern Europe. This distribution gives some indication on which potential inaccuracies in emission estimates have highest influence on the estimation of total European emissions. 7. CONCLUSIONS

Inventories of acidifying emissions exist for all of Europe, though with variable accuracies. Whereas the sources of the largest contribution to acidification through SO2 emissions are already relatively well documented, still substantial uncertainties exist for estimates of NO, and ammonia emissions. Further refinement of emission inventory techniques seems necessary for these two pollutants.

References [l] Amann M., Sorensen L., The RAINS Energy and Sulfur Emission Database, Status

1991. WP-91-xx, International Institute for Applied Systems Analysis, Laxenburg, Austria, 1991 (forthcoming). [2] Asman W., Ammonia emissions in Europe: updated emission and seasonal variation. Report DMU-Luft-A132, National Research Institute DMU, Roskilde, Denmark, 1990. [3] Baars H.P., Accuracy of emission inventories. Methodology and preliminary results of the Dutch NO, inventory. TNO Report P90/031, TNO Delft, Netherlands, 1990.

75 QDR CSFR HUN POL BUL USR DK LUX IRE FIN ROM UK 9 PA QRE BEL YU F RQ FRA ITA SW E AUS NOR NET AL TK SWI POR

0

2

4

6

I

I

I

8

10

12

10-9 moles =SO2

O N O x

Figure 3: Potential acidity of emissions in Europe on a per-capita basis, 1985 (4) Bremnes P.K., Calculations of exhaust gas emissions from sea transport. Methodology and results. Proceedings of the EMEP Workshop on Emissions from Ships. Oslo, 7-8 June 1990, Norway. (51 Buijsman E., Maas H., Asman W., Anthropogenic NH3 emissions in Europe. Atmos.Env. 32: 1009-1022, 1987.

[6] Commission of the European Communities, Resultats du programme CORINE. SEC(91)958, Commission of the European Communities, Brussels, 1991.

[7] CORINAIR Inventory: Default Emission Factors Handbook. Prepared by CITEPA, Paris, 1991. [8] Egglestone H.S., Accuracy of national air pollutant emission inventories. IIASA/NILU Task Force Meeting on accuracy of emission inventories, International Institute for Applied Systems Analysis (IIASA), Laxenburg, Austria, 1991. [9] Eliassen A., Hov O., Iversen T., Saltbones J., Simpson D., Estimates of Airborne Transboundary Transport of Sulphur and Nitrogen over Europe. EMEP/MSC-W Report 1/88, Norwegian Meteorological Institute, Oslo, 1988.

76 [lo] Iversen T., Halvorsen N., Saltbones J., Sandnes H., Calculated Budgets for Airborne

Sulphur and Nitrogen in Europe. EMEP/MSC-W Report 2/90, Norwegian Meteorological Institute, Oslo, 1990.

T.,Halvorsen N., Mylona S., Sandnes H., Calculated Budgets for Airborne Acidifying Components in Europe, 1985, 1987, 1988, 1989, 1990. EMEP/MSC-W Report 1/91, Norwegian Meteorological Institute, Oslo, 1991.

[ll] Iversen

[I21 Jarvis S.C., Pain B.F., Ammonia Volatilisation from Agricultural Land. Proceedings No. 298, The Fertilizer Society, Peterborough, UK, 1990.

[13] Joerss K.E., Proceedings of the EMEP Workshop on Emission Inventory Techniques, Regensburg, July 1991. [I41 Klaassen G., Past and future emissions of ammonia in Europe. SR-91-01, International Institute for Applied Systems Analysis, Laxenburg, Austria, 1991. [I51 Menzi H.,Neftel A., J.-M. Besson, Stadelmann F.X., Special conditions influencing ammonia emission factors in Switzerland. In: G. Klaassen (ed.): Ammonia emissions in Europe: emission factors and abatement costs. CP-gI-xx, International Institute for Applied Systems Analysis, Laxenburg, Austria, 1991 (forthcoming). [I61 OECD, Emission Inventory of Major Air Pollutants in OECD European Countries. Environment Monographs No. 21, OECD Paris, 1990. [17] Pacyna J., NO, emission from stationary sources in Eastern Europe in 1985. NILU Report 78/88, Norwegian Institute for Air Research, Lillestrom, Norway, 1988. [18] Pacyna J., Larssen S., Semb A., European survey for NO, emissions, 1985. NILU Report 26/89, Norwegian Institute for Air Research, Lillestrom, Norway, 1989. [19] Popov S., The energy- and emission data base of the RAINS model for the European part of the Soviet Union. WP-gl-xx, International Institute for Applied Systems Analysis, Laxenburg, Austria, 1991 (forthcoming). [20] Tarrason L., Biogenic Sulphur Emission from the North-Atlantic Ocean. EMEP/MSC-W Note 3/91, Norwegian Meteorological Institute, Oslo, 1991. [21] UN/ECE, Major review of Strategies and Policies. ECEIEB.AIRI27, United Nations Economic Commission for Europe, Geneva, Switzerland, 1990.

T Schneider (Editor). Acidlflcatlon Research Evaluationand Policy Applications @ 1992 Elsewer Science Publishers B V All rights resewed

I1

Acidification of Forests and forest Soils: Current status E.Matzner Lehrstuhl fiir Bodenkunde, BITOK, University of Bayreuth, Postfach 101251, D-8580 Bayreuth Abstract Soil acidification is defined as a decrease of the acid neutralization capacity of the soil solids. By this definition, forest soils are generally acidifyin under humidic climatic conditions. Some questions are yet to be answered: What is t e rate of acidification? What is the cause and to what extend does acidification influence intensity parameters, like soil solution characteristics? The latter is of greater importance when evaluating the effects of forest soil acidification on trees, deeper soil layers and groundwaters. Evaluating the causes of soil acidification by establishing roton budgets reveals that atmospheric deposition of acidi ing compounds, like sulkric and nitric acids and ammonia, is the dominant cause or the acidification of forest soils in Central Europe. The relative importance of other processes (e.g. tree uptake of excess cations) depends mainly on the rate of deposition and the management regime. In northern Scandinavia and arts of North America, tree uptake may be the dominant cause of forest soil acidiication. Effects of forest soil acidification on the exchangeable cation pools as well as on the composition of the soil solution may be differentiated. While processes like tree uptake and acidic deposition can reduce the base saturation of the cation exchange capacity of the soil in a similar way, the input of acidity by atmospheric deposition In connection with the input of mobile anions has specific effects on the chemical com osition of the soil solution of acid forest soils, esp. on the content of inorganic Al.h e release of cation acids like Al ions leads to the acidification of deeper soil layers and groundwaters and poses a significant stress on tree root functioning in terms of nutrient uptake and root growth. Strong evidence exist that base cation depletion, high Al-levels and unfavourable Mg/Al and Ca/Al ratios of the soil solutions are involved in the forest decline phenomena in Central Europe. Clearly the emission of acidi ing compounds in Europe needs further siGnificant relations are not to be expected, since the reduction. However, direct dose desorption of previously stored sulfate in the soil may prevent a quick recovery of soil chemical status following reduction of the deposition load. In addition to emission control, liming and fertilization of forest soils are appropriate and necessary tools for the stabilization of forest ecosystems.

f

7

1. INTRODUCTION

Numerous papers have been published during the last decade on the symptoms, the extent, the development, and the possible causes of the socalled "novel forest decline". An overview by Ulrich [ 11 is available; several aspects will also be addressed during this conference.

78

Among the different symptoms of forest decline occurring on tree species in Central Europe, the decline of Norway spruce has received the most attention from scientific community. That is why most of the followin statements on the implications of soil acidification on forest decline relate to the I fecline of Norway spruce (Picea abies). Little is known about the possible causes of the decline of other species, especially European beech (Fagus silvatica) and oak (Quercus robur, Quercus petraea). Several hypotheses are still discussed which might explain the wides read nutritional disturbances (esp. Mg-deficiency) of Norway spruce on the one han and the loss of needles on the other. Air pollutants likely lay a major role in most of these problems, the focus rangin from direct effects of pol utants on needles to soil mediated effects of the deposition of acidic compounds, nitrogen and heavy metals. This contribution will concentrate on the possible role of soil acidification in forest decline. It will be shown, that a) detrimental changes of forest soils have occured over the last decades and b) that these changes have altered the stability of forest ecosystems.

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2. DEFINITION OF SOIL ACIDIFICATION According to v. Breemen [2] "soil acidification" is defined as a reduction in the acid neutralization capacity (ANC) of the soil solids. A reduction of ANC is indicated by a net loss of cations from the soil that stem from bindings to weak acid anions like silicate or carbonate. Since losses of cations from soils with seepage water occur naturally in a humidic climate, one can conclude from this definition, that soils are generally acidi 'ng under these conditions. While the question "Do forest soils acidify?" is there ore easily answered, the evaluation of the rate of acidification, its effects on intensity parameters like soil solution com osition, availability of exchangeable cations, and the quantitative contribution of acid eposition to soil acidification need further consideration. Factors influencing the chemistry of soil solutions will be of special importance, since the soil solution represents the environment of tree roots and soil organisms, as well as the link to deeper soil layers and aquatic ecosystems.

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3. CAUSES OF SOIL ACIDIFICATION A decrease of soil ANC may be caused by the deposition of acids (H2SO4, HN03) or by soil internal generation of protons. Several reviews on the theory and the processes responsible for the internal production of protons in soils is available with v. Breemen [3], Bredemeier [4] and Matzner [5]. Major consideration in aerobic forest soils must be given to: - net dissociation of organic acids - net oxidation of organic N and NH4 to HN03 - net dissociation of carbonic acid - excess uptake of cations by the trees The term "net" indicates that the formation rate is lar er than the H + consuming reverse process like mineralization of organic acids, upta e of nitrate, protonation of bicarbonate and mineralization of litter. Quantification of these processes becomes possible by establishing proton budgets based on element budgets of the soil. T r es of soil acidification on non-calcareous soils can be in the range of 1-8kmolc*ha'*ay In summarizing the present status on the reiative im ortance of the above mentioned processes one can conclude, that in Middle Europe e!lt input of acidifying substances (incl. NH4) by atmospheric deposition is the most dominant source of soil acidification, the only exception being calcareous soils where the dissociation of carbonic acid is the

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driving force for soil acidification [3-61. The NH4-de osition followed by nitrification or NHq-uptake is a substantial contribution to the &-load of forest soils in Middle Europe. Its importance may even increase in the future due to "N-saturation" of the ecosystem [7] and to the reduction in S02-emissions relative to NH4-emissions. As was evaluated by Johnson et al. [6], the relative imports ance of the various processes contributing to the reduction of ANC depends on the deposition regime, tree species and forest harvesting practices. In areas of lower deposition rates like Skandinavia and parts of North America, uptake of cations by trees and subsequent storage in the biomass can be the dominant source of ANC reduction in soils. 4. CONSEQUENCES OF SOIL ACIDIFICATION FOR SOIL CHEMISTRY

While the concept of ANC is a usefull tool to describe the rinciples and causes of soil acidification, it is of little help in evaluating the ecophysio ogical relevance of soil acidification. This requires additional consideration of the intensity of soil acidification. As mentioned earlier, the soil solution represents the immediate environment of the tree roots and soil organisms. Its composition is considered an important ecophysiological parameter since it regulates nutrient uptake and potential toxicity of protons and metals. Assessment of the risk of soil acidification to trees should be based on soil solution studies. The chemical composition of the soil solution depends on the input of substances into soils, the physico-chemical interactions between soil solids and the solution and on the influence of or anisms, which results from the uptake of elements as well as from the mobilization of dissolved organic substances (DOC) and minerals by roots and microorganism during decomposition. No direct relationship between intensity parameters of the soil solution and the change in ANC can be expected [2] because of the buffer ability of soils with respect to acid inputs. The physico-chemical reactions of soil solids with the soil solution are of special interest in re lating soil solution chemistry. They were grouped according to typical reactions by Elrich [8] into socalled "buffer ranges". Protons are either buffered by reaction with carbonates, by weathering of silicates, by cation exchange, or by release of ionic Al into the soil solution. The dominant buffer reaction is related to the pH of the soil solution. Acid in uts to soils exceeding the rate of silicate weathering in noncalcareous soils will !ad to a reduction in base saturation of the cation-exchangecapacity (CEC) and thereby reduce the availability of macro-nutrients like Ca, Mg and K to the tree. If the base saturation falls below a level of about 10-1596, the mobilization and release of Al ions into the soil solution represents the dominant buffer system [9111. As was shown in numerous re ional soil inventory studies, these conditions prevail in most of the forest soils in Middfe Europe (for overview see [7]). Furthermore it has been demonstrated that changes of the base saturation and reduction of the H of the soil has occurred in the past decades at a greater rate than Formally assumed In humidic climates, a decrease of soil ANC is a natural process, and acid forest soils with low base saturation are thus also found in areas not subjected to acidic deposition. However, the influence of acidic deposition on such soils results in substantial and s ecific changes of the soil chemistry. While processes like tree uptake and acidic &position can reduce the base saturation of the cation exchange capacity of the soil in a similar way, the input of acidity by atmospheric deposition in connection with the input of mobile anions has specific effects on the chemical composition of the soil solution of acid forest soils, esp. on the content of inorganic Al. Since the mobilization of A1 from Al containing minerals in the soil is strongly pH dependent [ 121, inorganic Al is found in soil solutions at pH < 4.5 in increasing amounts. Due to the law of electroneutrality, the soil solutions must contain equivalent amounts of cations and anions. If the pH becomes

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less than 4.5, the im ortance of the HCO anion goes to zero, and strong mineral acid anions like SO4, N 8 and C1 dominate &e anionic composition of the solutions, thus allowing higher levels of Al, that had not been reached under previous natural conditions. Al ions can be considered as cationic acids [ll]; their transport with seepa e water thus represents an acid input to deeper soil layers and aquatic systems. The acdfication of soil solutions in deeper soil layers (underneath the main rooting zone and Podsol B horizons) must be attributed to acidic deposition in most cases. Soil solution chemistry within a given soil profile exhibits substantial seasonal and spatial patterns. Spatial patterns result from the gradients of roots, or anic matter, organisms, and mineral distribution, as well as from input of substances to t e top soil by throughfall and litter. Seasonal patterns may result from climatic variations, namely from exceptionally high soil temperatures and from drought and rewetting cycles as shown by Matzner [13]. In the latter case, disruption of the ion cycle of the ecosystem by excess nitrification (nitrification exceeds nitrate uptake and immobilization) and subsequent seasonal HNO production is the cause of fluctuations of pH and Al concentrations in the soil sohtion.

a

5. RELEVANCE OF SOIL ACIDIFICATION OF TREES AND FOREST DECLINE Theoretical considerations of forest ecosystem stability [ 141 indicate that the steady state of the elemental budget of the ecosystem is a major supposition for the long lasting persistance of the ecosystem. Steady state conditions are reached if input equals output and thus no change in the state variables of the ecos stems occurs. Steady state conditions are normally not reached in reality because o the permanent variation of climate and inputs. Steady state is considered to be the ideal endpoint of ecosystem development. A long lasting systematic chan e of the input, as represented by the deposition of air pollutants, will ultimately lea to a new steady state of the system, the degree of change in the state variables (like soil solution chemistry, nutrient availability, competition etc.), and the tolerance of the species then determining the new structure and species composition of the system. Acidic deposition in general is thus a major destabilizing factor in terrestrial ecosystems. In the case of forest decline, the changes induced by the alteration of the chemical climate of forest ecosystems seem to overrun the tolerance mechanisms of tree species in many areas. In relation to soil acidity, the mechanisms which induce stress to trees may include the lack of nutrients (esp. Ca, K, Mg) and the occurence of potentially toxic concentrations of H’, Al, Fe and heavy metals. As will be shown later, the lack of nutrients and the occurance of Al-ions in soil solutions have additive and detrimental character.

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5.1. Soil acidification and the occurence of Mg-deficiency symptoms on Norway spruce A hypothesis to explain these symptoms in relation to soil acidification is: ’Soil acidification caused by acid deposition reduces the storage of plant available (exchangeable) Mg and decreases Mg uptake by roots due to the antagonistic effect of A1 ions in soil solutions during Mg uptake. Increased rates of Mg-leachingfrom needles as a result of acidic deposition and excess N availability caused by N deposition will enhance the detrimental effects of soil acidification.” Clorotic yellowing of older needles of Norway spruce is a rather specific symptom related to Mg-deficency [15,16]. The symptoms occur only on acid soils with low content of exchan eable Mg. Liu and Truby ([17] found a good correlation between the M content o f needles and the exchangeable Mg content of soils in the German Blacf Forest area.

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The defiency symptoms are reversible by using a propriate fertilizers [18]. The availability of Mg in the soil is thus a major factor contriguting to this symptom. In addition, Jorns and Hecht-Buchholz [ 191 induced Mg-deficiency and yellowing of needles by ex osing trees to nutrient solutions with different A1 concentrations. The critical molar b g / A l ratio to induce deficiency was 0.2 irrespective of the Mg content. The anta onistic effect of Al ions on the uptake of Mg and Ca by tree roots has been confirmefi in a number of investigations [20-221. The critical Mg/Al ratios observed under controlled conditions in the laboratory are found in the soil solutions of declining spruce stands [23] and coincide with the spatial pattern of decline. Furthermore, Matzner et al. [23] showed a decrease of the Mg/Al ratio with time in a lon -term study in the German Solling area, and Zottl and Hiittl [16] as well as Reemtsma [ 41 reported a decrease in the needle contents of nutrient cations during the last decades. The coincidence of the temporal and s atial pattern of soil changes with the pattern of this symptom is an important criteria For the validity of the hyposthesis. The good correlation between needle and root content of Mg [25] furthermore em hasizes the role of soil-mediated effects. reduction of the soil storage of exchangeable Mg as well as the occurrence of unfavourable Mg/Al ratios of the soil solution is a consequence of buffer processes and must be most1 attributed to soil acidification caused b acidic deposition as stated above. The linl between the symptom of decline and soiYacidification is obvious. It is confirmed by results from experimental acidifications of forest sites. Abrahamsen et al. [26] reported the occurrence of Mg deficiency, growth disturbance, and dieback of Pinus silvestris stands in Norway 7 years after the artificial acid irrigation of the sites had stopped. These findings show the long lasting effects of acidic deposition and emphazise the need for appropriate time scaling of field experiments. However, besides soil acidification additional factors may be involved in triggering M deficiency, for example the leachin of cations from needles and leaves of trees is enfanced b acid irrigation [27-291 anfi by the uptake of NHq by needles [30]. If the uptake of g b roots is inhibited by anta onistic effects and low soil storage of Mg, the re lacement ofYformally leached Mg wilkcease and Mg deficiency will be enhanced. Sc\ulze and coworkers [31] put forward the hypothesis that hi h N inputs to forest ecosystems including the uptake of N by above round parts o the trees induce Mg deficien on acid soils by tri ering the growth o the tree which then faces insufficient Mg supp y. This phenomenon as been also observed in N-fertilization experiments. The widemng of N Mg ratios of spruce needles by the uptake of N throu h above parts of trees rom wet deposition was also shown by Eilers et al. [32fthus con irming this hypothesis. These otential effects of N deposition do not contradict the conclusions drawn from soil acidilcation studies but will have additive character. N-deposition is involved in two different ways: enhancing soil acidification and eutrophying the system. Beside the atmospheric deposition of acids and N, the rate of Mg deposition should have a strong effect on the regional pattern of Mg deficiency by modifying the Mg/A ratio of the soil solution. Matzner [33] showed that the de osition of Mg contributes significantly to the total Mg budget of forest ecos stems wit acid soils. Conversely, M deposition forms a strong gradient from coasta areas with high Me input to Souti? Germany with low rates of input. Thus, the rates of deposition of acidity found to be generally lower in South Germany as compared to Middle and East Germany may induce as low Mg/Al ratios in the soil solution as under high input conditions in other areas. To summarize: the h othesis that soil acidification and acid de osition are related to Mg deficiency, has not een disproved; in contrary, it has reac ed a high degree of quantification.

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82 5.2 Soil acidification and the occurrence of needle loss of Norway spruce

A h othesis relafing soil acidification to this symptom of decline is: 'The chemical to acid (H ,Al) toxicity in the soil solution has increased over the last decades. stress Highest risk of toxicity to tree roots ex& in the deeper rooting zone. H' and A1 toxicity finally causes damage to the root system, a withdrawal o fine roots from deeper layers and disturbance of water uptake. Water stress in periods o drought is the final cause of the needle loss. " The change in soil chemistry induced by acid precipitation and other processes contributing to soil acidification has already been discussed. Especially high levels of Al in soil solutions are characteristic for acid soils under the influence of sulfur and nitrogen de osition. In acid orest soils, t e risk of Al toxicit to tree roots is increasing with soil de th, since in A-horizons A19+ activities in soirsolutions are low compared to B- an Chorizons because of high concentrations of organic substances in the solid phase as well as in the solution phase acting as complexing agents. Matzner and Prenzel [34] and Murach and Matzner [35] reported a decrease of the Ca/Al ratio of about 2 in the top soil to 0.2 in the 40 cm layer of two forest sites in the German Solling area. Under controlled conditions Al ratios of less than 1 induced symptoms of Al toxicity to Murach and V t z n e r [35] concluded that Al-stress is most Norway spruce likely found in the deeper horizons, while H -stress is most probable in the top soil. Effects of high Al concentrations on tree roots may result in a reduction of cation uptake (see above) and in a decrease of root growth and the dieback of root tips. In the last decade, the effects of H t - and Al-ions on tree roots have been studied intensively (for overview see [36]). The results from controlled experiments indicate substantial differences in the tolerance of various tree s ecies [20-22,37,38]. Furthermore, pH and cation concentration of the soil solution in uence the effects of Al. Thus, the concentration of Al must be interpreted in relation to these parameters and, e.g., the Ca/Al ratio of the soil solution proved to be a better indicator for Al-stress than the Al concentration alone. In respect to the most dominant tree species in Europe, Norwa spruce and European beech, beech was found to be more susce table to low pH vaLes while being more tolerant to Al-ions. The opposite was founrffor spruce. These findings from laboratory studies were confirmed by root morphological and root chemical studies in the field [25,38,39]. Murach [25] and Murach and Matzner [35] attributed the shallow rooting of Norway spruce on acid forest soils to the low Al tolerance of this species, causing strong vertical gradients in the distribution of fine roots with highest values of fine root densities in the organic top layers of the soil profile and a sharp decrease in the mineral soil. Gruber [40] investigated the morphology of the needle abcission of Norway s ruce. He concluded that the abcission is a mechanical process induced by severe and cironic water stress. According to Ulrich [36] chronic water stress may be caused by : - a change in climate (e.g. a series of warm/dry seasons - a n increase in transpiration rate (e.g. by the ef ects of pollutants on stomata1 regulation) - a decrease of water uptake - a decrease in the area of conductive xylem tissue connecting water uptake by roots to water loss by transpiration. A decrease of water uptake may be related to soil acidification in different ways: The in respect to water may be negatively influenced by Al ions. found in hydroculture experiments the reduction of the (Al, H t , heavy metals) stress.

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The amount of fine roots capable of water uptake might also be reduced by soil acidification. Stickan et al. [43] reported decreased transpiration rates from mature beech trees that were subjected to artificial acidification, with the consequence being a reduction in fine root biomass [44]. Furthermore, the soil volume used by tree roots for water u take mi ht be reduced because of shallow rooting (see above). This will cause a re&ction o f water uptake during eriods of drought because the top soil will rapidly reach a low water content with sugsequent reduction in transpiration while the water supply from deeper layers is inhibited as compared to a normal situation. All the factors mentioned will contribute to water stress during periods of drought and will increase the susceptibility of the trees to this natural stress factor. The water stress hypothesis also explains the increase of damage followin exceptionally warm/dry climatic conditions with recovery in cool/wet seasons [4$. Soil chemical stress increasing in warm/dry seasons may however contribute to this development of damage

~31. To summarize: the hypothesis proposed to explain needle losses is based on a large number of different observations rather than on controlled experimental evidence. Further assessment of this hypothesis is the subject of ongoing research. 6. CONSEQUENCES FOR POLICY MAKING The detrimental effects of the de osition of acids, nitrogen and other pollutants are not restricted to forest soils and to orest ecosystems. Impacts on other terrestrial and aquatic ecosystems will be evaluated in several presentations during this conference. In general, the problem of acidification points to the risks of environmental pollution which may be considered in general as the uncontrolled distribution of materials in the ecos here. Having no or little control over the input of substances to various ecosystems of ifferent tolerance and stability is equivalent to not having control over the development of these systems. The occurrance of forest decline is one of the most striking examples of this lack of control. Since society relies on the use of terrestrial ecosystems for food and timber production, water supply, erosion-control, climate regulation, recreation etc., three needs must be addressed immediately: - To better understand the functioning of the ecos stems in order to predict the effects of changin environmental factors, includin the eposition of substances, management regime and the possible effects of Global C ange. - To develop critical loads for different types of pollutants and ecosystems and to reduce the emission and deposition of potential destabilizing substances to a level less than the critical load in order to prevent detrimental changes in ecosystems. The latter must be a continuous process based on progress in research and needs better transfer and implementation of scientific results into policy than before. - To mitigate the effects of soil acidification by appropriate liming and fertilization of forest soils. In case of acidifying agents and nitrogen, critical loads have already been proposed ([46], Hettelingh, this conference). Comparing the critical loads with the present de osition rates reveals that a drastic reduction of emissions of NOx, NH3 and SO2 is stii urgently needed. Whith respect to S02, some progress has been made in Western Europe. Reduction of NH3 and NO, emissions and general reduction of pollution in Eastern Europe must be the task of the future. The extent and kinetics of the reversibility of soil acidification by reduction of the deposition load is still the subject of scientific work. While an increase of base saturation of the cation exchange capacity is not likely to occur, improvement of the soil solution chemistry and a mitigation of acid stress can be expected. However, the time scale in

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which significant changes will occur is hard to evaluate. The desorption of formally stored sulfates in the soil may prevent the recovery of soil solution chemistry for a long time, even if the input of sulfate will stop immediately [47]. This holds especially for areas of hi h pollution, and for soils and soil horizons with large sulfate adsorption capacity. d u s , no direct relationship between a reduced input and soil solution chemistry and the vitality of trees can be expected. The present state of knowledge on liming and fertilization of forests under stress is summarized by Zottl and Hiittl [48]. As already stated, the recovery of forest soils under conditions of less deposition is very slow due to the slow rates of mineral weathering. This holds especially for the pool of exchangeable cations which acts simultaneously as a nutrient reservoir and buffer mechanism and thus is a sumosition for the elasticity of forest ecosystems in respect to .. acid stress [ 141. Increasine the base saturation of the exchanee caDacitv will need the addition of lime and/or salt Fertilizers. Both measures are subj&ted i o riiks: While in the case of liming increased rates of nitrification and nitrate leaching is often reported in the first years following the ap lication, the addition of easily soluble salts may lead to mobilization of with subsequent transfer to groundwater [49]. The risk of increased exchangeable nitrate leaching following limng may furthermore be enlarged in the future due to long lasting N-inputs. These restrictions already indicate that liming and fertilization alone cannot be considered sufficient to counteract forest decline. However, liming and fertilizations, used in an appropriate way based on evaluation of soil chemistry and tree nutrition, have positive effects on tree growth, root development, tree nutrition, the degree of decline and seepage water acidity, and thus must be considered as silvicultural tools to support the positive effects of future reduction of emissions and to counteract the present threat from soil acidification.

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Acknowledgement

I like to thank Prof. Harold L. Drake for language editing of the manuscript. 7. REFERENCES

B. Ulrich (ed.), In: International Congress on Forest Decline Research: State of Knowledge and Perspecives, Proc. Kernforschungszentrum Karlsruhe, (1989) 21. N. v. Breemen, In: "Soil Acidity", B. Ulrich and M.E. Sumner (eds.), Springer Verla , (1991) 1. 3 N. v. ireemen, C.T. Driscoll and J. Mulder, Nature 307 (1983) 599. 4 M. Bredemeier, Plant and Soil 101 (1987) 273. 5 E. Matzner, In: "Advances in Environmental Science, Acidic Precipitation", Vol. 1: Case Studies. Adriano, D.C. and M. Havas (eds.) (1989) 39. 6 D.W. Johnson, M.S. Cresser, I.S. Nilsson, J. Turner, B. Ulrich, D. Binkley and D.W. Cole, In: "Acidic deposition: Its nature and impacts", Proc. Royal SOC.of Edinburgh, (1991) in press. 7 J.D. Aber, K.J. Nadelhoffer, P. Steuder and J.M. Melillo, BioScience 39 (1989) 378. 8 B. Ulrich, Zeitschrift fiir Pflanzenernahrung und Bodenkunde 144 (1981) 289. 8 B. Ulrich, In: "Soil Acidity", B. Ulrich and M.B. Sumner (eds.), Springer Verlag, (1991) 28. 10 J.O. Reuss, J. Environ. Qual. 12 (1983) 591. 11 J.O. Reuss and D.W. Johnson, Ecological Studies 59, Springer Verlag, New York, Berlin, Heidelberg, Tokyo, 1986. 12 H.M. May and D.K. Nordstrom, In: "Soil Acidity" B. Ulrich and M.W. Sumner (eds.), Springer Verlag, (1991) 125. 1

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13 E. Matzner, In: "International Congress on Forest Decline Research: State of Knowledge and Perspectives", B. Ulrich (ed.), Proc. Kernforschungszentrum Karlsruhe, (1989) 303. 14 B. Ulrich, In: "Potentials and limitations of exosystems analysis", E.D. Schulze and H. Zwolfer (eds.), Ecol. Studies 61 (1987) 11. 15 C. Bosch, E. Pfannkuch, U. Baum and K.E. Rehfuess, Forstwissenschaftliches Centralblatt 102 (1983) 167. 16 H.W. Zottl and R.F. Hiittl, Allgemeine Forschzeitschrift 9/10 (1985). 17 J.C. Liu and P. Triiby, Zeitschrift fiir Pflanzenernahrung und Bodenkunde 152 (1989) 307. 18 R.F. Hiittl, S. Fink, H.J. Lutz, M. Poth and J. Wisniewski, Forest Ecol. and Management 30 (1990) 341 19 A. Jorns and C. Hecht-Buchholz, Allgemeine Forstzeitschrift 40 (1985) 1248. 20 K. Rost-Siebert, Berichte des Forschungszentrums Waldokosysteme/Waldsterben 12 (1985). 21 D.L. Godbold, K. Dictus and A. Hiittermann, Can.J.For.Res. 18 (1988) 1167. 22 M. Neitzke, Z. Pflanzenernahr. Bodenk. 153 (1990) 229, 23 E. Matzner, U. Blanck, G. Hartmann and R. Stock, In: "IUFRO conference: Air Pollution and Forest Decline", J.B. Bucher and 1. Bucher Wallin (eds.), Interlaken (CH) Proc. Birmendorf, (1989) 195. 24 J.B. Reemtsma, Allgemeine Forst- und Jagdzeitung 157 (1986) 196. 25 D. Murach, In: "International Congress on Forest Decline Research: State of Knowledge and Perspective", B. Ulrich (ed.), Proc. Kernforschungszentrurn Karlsruhe, (1989) 583. 26 G. Abrahamsen, H.M. Sei and A. Sernb, In: "Advances in Environmental Sciences, Acidic Precipitation", Vof 1, Case Studies, D.C. Adriano and M. Havas (eds.), (1989) 137. 27 K. Mengel, H.-J. Lutzer and M.T. Breininger, In: Statusseminar "Wirkungen und Luftverunreiniguen auf Waldbaume und Baldboden", 2.-4. Dez. 1985, KFA Julich, (1986) 292. 28 K. Kreutzer, R. Schierl, A. Gottlein, and P. Probstle, In: "Internationl Congress on Forest Decline Research: State of Knowledge (1989) 29 D.A. Schaefer, and W.A. Reiners, In: "Acidic Precipitation", Vol. 3: Sources, Depostion and Canopy Interactions (Advanced in Environmental Science), S.E. Lindber A.L. Page and S.A. Norton (eds.), Springer-Verlag, (1990) 241. 30 J.G.M. koelofs, A.J. Kempers, A.L. Houdijk and J. Jansen, Plant and Soil 84 (1985) 45. 31 E.D. Schulze, R. Oren and O.L. Lange, In: "Forest Decline and Air Polution, Ecol. Studies 77", E.D. Schulze, O.L. Lange and R. Oren (eds.), Springer Verlag, (1989) 459. 32 G. Eilers, R. Brumme and E. Matzner, Forest Ecol. & Management, (1991) in press. 33 E. Matzner, Berichte des Forschungszentrums Waldokosysteme/Waldsterben der Univ. Gottingen, Reihe A, Bd.40 (1988) 1. 34 E. Matzner and J. Prenzel, Water, Air and Soil Pollution (1991) in press. 35 D. Murach and E. Matzner, In: IUFRO Congress "Woody plant growth in changing physical and chemical environment", Vancouver (CDN) July 1987, D.P. Lavendar (ed.), Univ. of British Columbia, (1989) 171. 36 B. Ulrich, In: "Advances in Environmental Science, Acidic Precipitaion", Vol. 2, A.C. Adriano and M. Havas (eds.), Springer Verlag, New York, (1989) 169. 37 C.S. Cronan, R. April, R.J. Bartlett, P.R. Bloom, C.T. Driscoll, S.A. Gherini, G.S. Henderson, J.D. J o s h , J.M. Kelly, R.M. Newton, R.A. Parnell, H.H. Patterson, D.J. Raynal, M. Schaedle, C.L. Schofield, E.I. Sucoff, H.B. Tepper, F.C. Thornton, Water, Air and Soil Pollution 48 (1989) 181.

86 38 U. Ebben, Berichte des Forschungszentrums Waldokosysteme, Reihe A, Bd. 64 (1991) 1. 39 J.H. Bauch, H. Stienen, B. Ulrich and E. Matzner, Allgemeine Forstzeitschrift 43 (1985) 1148. 40 F. Gruber, In: "International Congress on Forest Decline Research: State of Knowledge and Perspectives", B. Ulrich (ed.), Proc. Kernforschungszentrum Karlsruhe, (1989) 109. 41 B. Klein, In: "Acid Depostion, Environmental and Economic Impacts", D. Adams and W. Page (eds.), Plenum Pub. Co., New York, 1985. 42 H. Stienen, Forstwirtschaftliches Centralblatt 105 (1986) 321. 43 W. Stickan, M. Schulte, Y. Kakubari, F. Niederstadt, J. Schenk and M. Runge, Berichte des Forschungszentrums Waldokosysteme, Reihe B, Bd. 18 (1991) 1. 44 C. Rapp, Berichte des Forschungszentrums Waldokosysteme 49 (1989) Reihe A, 123. 45 FBW - Forschungsbeirat Waldschaden/Luftverunreinigungen der Bundesregierung und der Linder, 3. Bericht, Kernforschungszentrum Karlsruhe (1989). 46 J. Nilsson and P. Grennfelt (eds.), Critical loads for sulfur and nitrogen. Nordic Ministers Rap. 15, Solna, Sweden, 1988. 47 C. Alewell, A. Liikewille and E. Matzner, In: P. Math (ed.), Proc. First European conference on terrestial ecosystems: Forests and wet ands, Firenze May 1991, in ress. 48 R.W. Zottl and R.F. Hiittl (eds.), Management of nutrition of forests under stress. Kluwer Acad. Press, 1991. 49 E. Matzner and K.J. Meiwes, Water, Air and Soil Pollution 54 (1991) 377.

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@ 1992 Elsevier Science Publishers B.V. All rights reserved

Stress combinations in forests J.L. Innes Forestry Commission, Alice Holt Lodge, Wrecclesham, Farnham, Surrey GUlO 4LH, United Kingdom

Abstract Many instances of forest decline, both natural and anthropogenic, can be attributed to combinations of stresses. In most cases, these combinations have proved to be much initially thought. In explaining the decline of trees or forests at any is important to consider all the stresses that may be operating. A divide stresses into predisposinp, triy g and accelerating types, but these may not always be applicable. Generalisations a out the role of pollution in forest decline cannot be made as there are many different types of decline and the importance of specific forms of pollution varies in each.

1. INTRODUCTION

In the 1980s, it became apparent that a substantial decline in the health of some forests was occurring. Such declines have been noted in the past, but the apparent severity and the reportedly widespread nature of the decline was believed to indicated that a new problem had emerged. The areas that was considered to be the most severely affected were located in the then Federal Republic of Germany (FRG). Althou h similar declines had been noted in neighbouring countries to the east, these were we 1 known and had existed for a considerable length of time. The declines in the FRG generated a great deal of concern, coinciding as they did with a rise in environmental interests generally and, in a particular, with a growing awareness of the possible ecolo ical problems associated with environmental pollution. Although air pollution was initial y blamed as the sole cause of the roblem, it has become increasingly apparent that a number of stresses act in com&nation to bring about forest decline.

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2. THE CONCEPT OF STRESS At any given time in a tree's life, it will be under some level of stress. This does not normally represent a problem, as all living organisms are, to greater or lesser extent, adapted to certain levels of stress. For example, all trees harbour insects, providing them with shelter and a source of food. Some defoliation occurs, but the loss of photosynthetic

88

capacity is usually insignificant. However, if the level of insect activity is increased for any reason, the amount of defoliation on the tree may cross a threshold whereby the tree must devote significant resources to replacing the foliage. Under such circumstances, the stress becomes Important. It is extremely difficult to identify when a tree is subject to important levels of a particular stress. A possible way in which this can be done is in relation to the amount of new wood produced in a year. When the annual volume increment is reduced, a causal agent can be inferred. However, even this is difficult to apply as the annual increment of trees is extremely variable. Substantial year-to- ear variations in radial increment are normally apparent (Figure l),and these can usua ly be attributed to annual fluctuations in climatic conditions. However, at what stage can climate be considered to be a stress? It is very rare to find no growth, which means that a continuum exists from poor to good. Identifying a threshold at which growth is good or poor is extremely difficult. Long-term avera es might be used, but these are inappropriate due to long-term trends in radial growti associated with tree age and management practices such as thinning. In addition, there is good evidence that the average monthly weather conditions have varied in most parts of the world over the last 50 years, and long-term averages are therefore of questionable value.

r

Scot. pins Lemon8joen Noway

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1800

1820

1840

1860

1880

1900

Dote

1920

1840

1960

1980

(AD)

Figure 1. Tree-ring series from a Scots pine growing at Lemonsjoen, north Norway. The high levels in the early years of the tree are characteristic of young trees, but substantial year-to-year variations persist through the tree’s history.

89

There appears to be no easy way over detecting when a tree is bein adversely affected by a particular stress. This has led to an alternative a proach with empaasis being placed on establishing the condition of a tree rather than the Lctors affecting it.

3. ASSESSMENT OF TREES, STANDS AND FORESTS

Surveys of forest health are undertaken throughout Europe, and their results are iven wide publicity, particularly in policy circles. Assessments of forest health are largely %ased on the assessment of the crown condition of a number of trees in a stand. Assessments are normally restricted to defoliation and discoloration, with considerable em hasis being given to the former. A particular feature of the international programmes of [ealth assessment is the emphasis that is placed on the harmonised nature of the assessments. Such an emphasis is misplaced. The assessment of defoliation is extremely complex and depends on a wide variety of factors. One of the most important, from an international point of view, is the standards that are used to assess trees. Defoliation is normally assessed by comparing the amount of foliage in the crown of a tree with that of a reference tree. Problems arise over the reference trees. In some countries, such as France, a local reference is used. This has the advantage that the effects of adverse environmental conditions can be taken into account. For example, a number of different forms of Norway spruce occur, and the application of reference standards for one type to, another may be inappropriate (Fi re 2). However, the use of sub'ective, variable standards means that changes in the congion of the reference trees wi 1 affect the overall assessments. Similarly, the long-term use of particular trees as standards may result in a subtle change in the assessments through time.

r'

Figure 2. Different crown forms in Norway spruce.

90 The alternative approach is to use photogra hic standards. While these may be fixed in time, the have the disadvantage that on! a limited range of the different morphological pes can be portrayed. This, however, does not appear to be as big a problem as has een made out. Experienced assessors can often apply the photographic standards for one species to other species: the assessments appear to be fixed more on experience than any hard-and-fast rules. Obviously, in such a situation, continuity of assessors is an important prere uisite; although some countries achieve this (over half of the assessment teams used in $e UK in 1991 were the same as in 1987), many rely on newly-qualified foresters to undertake the assessments. The importance of adequate trainin becomes obvious. In the UK, experience has shown that more than one week is neede%to familiarise fully a surveyor with assessment procedures. The problems over the assessment of individual trees, has tended to take Precedence over other issues. However, the sampling design that is used is of critical importance when interpreting the results of national surveys (Innes 1990). The samplin scheme used for the European assessments of forest health is a systematic (16 x 16 kmy grid. However, such a system is only applicable to countries whose forested area is substantial. Where forests are sparse, such as in many industrial regions, the number of sample plots will also be restricted, yet such re ions are often where the greatest problems associated with acidification and other forms ofpollution might be expected. There are also problems with the manner in which the data are presented, Substantial variations occur in the condition of trees, not only on a plot-to-plot basis, but even between adjacent trees. Extrapolating results to the stand level, forest level or the forest growth re ion therefore requires considerable care. Normally, the number of trees assessed on a b t varies from 20 to 30, yet the extent of variation within and between stands means tiat this sample size is often inadequate (Innes and Boswell 1990). As a result of these problems, it is very difficult, if not impossible, to compare reliably the extent of defoliation between countries. Similarly, the association between the deeree of defoliation and the degree of stress cannot be determined from surveys alone. This places a major limitation on the interpretation of data related to forest health surveys.

1

4. ISOLATING INDMDUAL STRESSES

Defoliation is a non-specific phenomenon that cannot be related to any sin@e stress without sup lementary information. It is important at this stage to distinguish between direct ancfindirect stresses. Direct stress involves a readily discernible impact on the tree. Such stresses include defoliation by hytophaFous insects, fungal infections such as the needle cast fungi, rusts and mildew anxdirect injury by a variety of different agents, including wind, hail, acute levels of air pollution and other management operations such as thinning. Indirect stresses are those that affect a tree by influencing the level of other forms of stress. Examples include the effects of a mild winter, which may enable the buildup of insect populations and acidic de osition, which may result in the leaching from the soil of essential base cations such as ca cium and ma nesium to deficiency-levels. In many cases, it is relatively sim le to identify the cause of defoliation or ill-health in a tree. There are numerous pictori8guides to visible damage and many diseases and insects cause very specific symptoms. However, even in such cases, it is difficult to rule out the potential effects of other stresses. The widely applied scientific principle of Occam’s razor, whereby the simplest tenable hypothesis is accepted in favour of more complex hypotheses, may not always hold. For exam le, a marked deterioration of high-altitude Scots pine stands occurred in the Southern bplands of Scotland in the mid-1980s. The

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91

decline was attributed to the combined effects of needle-cast fungi, the pine shoot beetle and Brunchorstia dieback (Redfern et d. 1987). However, this explanation does not explain why only high-altitude stands in a s ecific area were affected, since all three problems can occur throughout Britain. The ikelihood that the problem was purely the result of a chance combination of the three stresses seems small and an additional factor or suite of factors is likely. The majority of trees that have died in areas known to be severely affected by air pollution have not been killed b air pollution itself. Instead, it is normally an insect or pathogen that is responsible for t i e ultimate death of the tree. Thus, in California., ozonestressed Ponderosa pine are killed by the mountain pine beetle (Dendroctonusponakrmue) and ollution-stressed Norway spruce in southern Poland and the northern part of the Czec! and Slovak Federal Republic are killed by the larch bud moth (Zeirupheru dinianu) Baltensweiler 1985). The simplest e lanation is that the trees have been killed by the kid agent but, in reality, this has merry provided the coup de grace, and a more complex explanation is required.

5. RECOGNISING A COMBINATION OF STRESSES

In reality, stresses rare1 occur on their own. Normally, several act together, resulting in the deterioration o the condition of the tree. The interactions between the stresses can vary considerably. At the simplest level, one stress added to another may double the total level of stress that the tree is under. However, such additive stresses are relatively rare. More often, the stresses interact with each other. Such interactions may be synergistic, whereby the effect of two or more stresses is greater than the sum or they may be antagonistic, whereby one stress reduced the impact of another. A articular1 good example of the latter is the interaction between drought and ozone &.g. Do& and Schutt 1990). Ozone episodes are frequently associated with periods of drought, but the response of trees to drought is to close down the stomata, simultaneously reducing water loss and preventing the uptake of ozone. To complicate the situation, prior exposure to its response to drought. ozone may affect the tree's stomatal control, thereby e model proposed by Sinclair The roles of different stresses can be related Stresses can be divlded into (1964) and elaborated b Houston (1973) and three main groups: pre isposing, tnggering and accelerating. Predisposing factors operate over the long-term, making trees more susceptible to other stresses. In themselves, they will not cause tree mortality but, in their absence, short-term stresses might have a much smaller impact than they otherwise do. Triggering stresses operate over the short-term and have a major effect on the condition of a tree, to the extent that they are so weakened that secondary biotic agents (accelerating factors) can eventually kill the tree. While conceptual1 neat, this hypothesis can be difficult to appl . A particular stress may be either pre isposing or triggering depending on its nature. &milarly, some tri ering stresses can, under certain circumstances, be classified as acceleratin . Air pofution is a typical example. Long-term chronic air pollution is probably best classiged as a predisposing stress. However, peak concentrations, while still not causing acute in'ury, may have a sufficient short-term impact to be classified as triggering. At very Jhigh concentrations, the pollutant may actually kill the tree without involving any secondary biotic a ents. fn recognising the most important stresses operating on a given tree, it is im ortant to distinguish between primary and secondary factors. Primary stresses are those t at can adversely affect healthy and vlgorous trees. There are relatively few important pathogens that consistently fall into this category, although some of the rusts are good examples. More often, the effects of a pathogenic stress are secondary, the tree having previously

7

d

d

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92

been weakened by some other factor. This may be a specific stress or may be an inherent factor such as the aging of the tree.

6. STRESS COMBINATIONS AND FOREST DECLINE

While the concept of forest decline being caused by a combination of stresses has received increasing acceptance by scientists, many policy makers appear to have found it difficult to endorse. For example, the annual reports of the United Nations Economic Commission for Europe International Co-operative Programme on the assessment and monitoring of forest health contain statements from certain countries to the effect that air pollution is the cause of forest decline. Interestingly, the number of administrations supporting this view has steadily declined, and the only European Community country still to adhere to it is Germany. The evaluation of the role of air pollution has been hindered by the view that forest decline is a sin le henomenon. This concept was shown to be inadequate several years ago (FBW 1988). L t least five different ty es of Norway spruce decline exist, some more easily explained by others. The decline orsilver fir has not been described to the same extent, but four different forms have been recopised in Romania (Barbu 1987). Forest declines have also occurred which bear no relation to air pollution, but which are caused by an e ually complex web of interacting factors. %he concept of a different suite of stresses acting at different sites (and even for different trees) has been dismissed by some on the grounds that it cannot be tested (Blank et ul. 1988). However, this is not so as a number of case studies have demonstrated that it is possible to hypothesis the links between different stresses and then test each one ex erimentally. Although this is a painstaking task (and one that it is likely to absorb a suEstantia1 amount of money), it is the only way in which the tangle of different interrelationships can be unwound. A number of case studies have been undertaken which have demonstrated the complexity of some of the disorders. One that does not involve pollution at all is the dieback of eucalypts in rural Australia. Landsber and Wylie 1988) have produced a complex model to explain the decline Figure 3), w ich robably as applications to many other agricultural areas. Some of t e proposed lin!s have been demonstrated by experimental work and field observations, others remain more speculative. In these and the other models presented here, it is important to remember that it is not necessary to invoke all the processes described in the models. At any particular site, a combination of stresses will be present, and these may be sufficient to trigger a decline sequence in the trees. In North America, the decline of high elevation red spruce in the northern Appalachians remains a controversial subject. The various processes thought to be operating are summarised in Figure 4. As in Europe, roblems have existed because of the initially poor descriptions of symptomatology. Detai ed observations have indicated that the decline of red spruce in the southern Appalachians is very different from that in the north, and there also appear to be differences between high-altitude and low-altitude declines. While it is widely accepted that atmospheric pollution plays a role in the decline, its importance relative to other stresses remains far from clear.

6

6

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f

93

Figure 3. Conce tual model for rural dieback of trees in Australia (from Landsberg and Wylie 1988). BPacked in pathways are based on the results of research, uncoloured pathways are more speculative. Broken lines indicate feedback pathways. In Europe, a model produced by Schulze et al. (1989) to ex lain the decline of Norway spruce in the Fichtelgebirge is equally complex (Figure 57. In this case, the emphasis is very much on pollutants, and factors such as tem erature-driven changes in nitrogen mineralisation in soils (e.g. Emmer and Tietema 1990fduring drought years have probably been underestimated. However, the model provides a good indication of the complexity of the pollution-rated stresses and their effects on the metabolism and vigour of Norway spruce. A very different set of stresses affects forests in the Netherlands and in parts of northern Germany. Here, the problem can be more definitely linked to pollution, with the development of various symptoms bein attributed to the deposition of excessive amounts of nitrogen. Although emphasis has t e e n placed on the role of ammonium in soil acidification, it is also important in relation to its direct effects on trees. The rocesses operating are very different to those acting in the Fichtelgebirge (Fi re 6), a n f i t would be quite incorrect to consider the two types of decline as being r e l a t e r

94 Acidlc cloudwaler

Nitrogen uptake

Climatoc Increased winler de8iccation black spruce

Armlllarla lnlecllon

Figure 4. Model showing possible inter-relationships between stress factors and the decline of red spruce at high elevations in the northern Appalachians. (Derived from numerous sources).

Y Org. Pnlkldm

Ozone

I Damape 10 CUIIcIe

Sol1 Acldlllcdlon

sol1 SOldIon

?

Needle Vellowlng

Reduced Growtn

Figure 5. Sequential processes leading to the decline of Norway spruce in the Fichtelgebirge of Germany (from Schulze er al. 1989).

95

Figure 6. Possible inter-relationshi s between the various stresses causing decline in Scots pine and Douglas fir in the Peey area of the Netherlands. (Derived from numerous sources). These four examples illustrate the complexity of the processes operating in forests.

As other papers in this volume have also stressed, the effects of acidification and other

forms of a r pollution cannot be readily simplified. While acidification may be important in some ecosystems, this is not universally true, even in those ecosystems generally considered to be sensitive to acidification. This has a number of important policy implications. The interactions between different factors make it difficult, if not impossible, to set critical load values for anything other than very small areas of forest (the work of de Vries (1988) is an exam le of the successful application of the concept). While attempts have been made to do tfis, both for deposition and for gaseous concentrations, it is clear that such attempts have little value. Forest decline is not solely due to air pollution and, in many cases, air ollution is not involved at all. Again, generalisations about the role of air pollution shoulfonly be made with great care. Because of the interactions between factors and the similarity in the end result (tree decline or death , a reduction in the ollution load may not necessarily cause an improvement in tree hea th or a reduction in t e extent of forest decline. For example, in pristine areas of the Pacific Northwest, sul hur deficiency is a major problem for foresters Blake er d.1988, Cole and Johnson 1877). Sulphur deficiencies have already been recor ed in some agricultural crops in Europe, and the complete removal of anthropogenic sulphur could result in deficiencies in some European forests. Similarly, nitrogen is an essential nutrient, and small quantities derived from the atmosphere are probably beneficial to forests. Forest production is currently higher than ever, partly as a result of improved silvicultural techniques but also almost certainly because of nitrogen deposition.

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96 7. CONCLUSIONS

A number of conclusions of relevance to policy makers can be drawn from the current state-of-the-art in the understanding of combined stresses. Assessments of forest decline are much less reliable than is made out, and international comparisons are extreme1 dubious. The link between visual estimates of efoliation and tree vigour, as expressed by volume increment, are com lex and are yet to be fully evaluated. Although many countries Rave reported high levels of defoliation, this does not necessaril represent a problem, nor does it indicate that pollution is having an adverse e&ect on the forests. More than one type of forest decline exists and models in which causes are hypothesized should only relate to particular types. Explanato models are on1 valid if they incorporate different levels of stress and recognise x e importance o!! the interactions between natural stresses and those of anthropogenic origin. Each h pothesized link in the explanatory models needs to be explored and fully tested Before it is accepted with confidence. However, in accepting a particular pathway, the role of synergistic and antagonistic effects need to be incorporated, as do negative and positive feedback mechanisms.

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8. REFERENCES

Baltensweiler, W. Z. Angew. Ent. 99 (1985), 77. Barbu, I. Rev. Padurilor 102 (1987), 195. Blake, J., Webster, S.R., Gessel, S.P. Soil Sci. SOC.Am. J. 52 1988), 1141. Blank, L.W., Roberts, T.M., Skeffington, R.A. Nature 336 (1 88), 27. Cole, D.W. and Johnson, D.W. Water Resour. Res. 13 (1977), 313. Dotzler, M., Scutt, P. Eur. J. For Path. 20 (1990), 59. Emmer, I.M., Tietema, A. Plant Soil 122 (1990), 193. Forshungsbeirat WaldschadenLuftverunreinigungen.Rapport 2. Karlsruhe, 1986. Houston, D.R. Int. Shade Tree Conf. Proc. 49 (1973), 73. Innes, J.L. Proceedings, Division 2, XIX IUFRO World Congress (1990), 380. Innes, J.L., Boswell, R.C., Can. J. For. Res., 20 1990), 790. Landsberg, J., Wylie, F.R. GeoJournal 17 (1988\, 231 Manion, P.D. Tree Disease Concepts, Prentice-Hall, Englewood Cliffs, 1981. Redfern, D.B., Gre ory, S.C., Macaskill, G.A., Pratt, J.E. Report on Forest Research 1987. HMSd, London (1987), 42. Schulze, E.-D., Oren, R., Lange, 0.-L. Forest decline and air pollution. Springer Verlag, Berlin (1989), 459. Sinclair, W.A. Cornell Plant. 20 (1964), 62. de Vries, W. Water, Air Soil Poll. 42 (1988), 221.

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T Schneider (Editor). Acidification Research Evaluation and Policy Applications 1992 Elsevier Science Publishers B V

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EFFECTS OF INCREASING NITROGEN DEPOSITION AND ACIDIFICATION ON HEATHLANDS. J.A. Lee', S.J.M. Caporn' and D.J. Readb 'Department of Environmental Biology, The University, Manchester, M 13 9PL, United Kingdom bDepartment of Animal and Plant Sciences, The University, Sheffield, S10 2TN, United Kingdom ABSTRACT Heathlands occur on nutrient-poor soils, and their vegetation is adapted to systems of low nutrient availability. Large changes to response to atmospheric deposition have been reported in the Netherlands resulting in the conversion of heathland to grassland. However, generalisations about the long term response to, for example, nitrogen deposition across the heathlands of Europe are not possible because of the wide variety of pollution and physical climates and differences in management. Attention is drawn, in this paper, to the upland heaths of the north and west of Britain which can experience large inputs of nitrogen and sulphur in concentrated forms (e.g. several millimolar) as occult precipitation. Experimental applications of large amounts of ammonium nitrate (4-20 g m'* y.') over nearly three years to an upland, managed stand of Calluna vulgaris have, so far, not resulted in detrimental effects. It is concluded that only long term experiments, using realistic simulated pollutant inputs will reveal the true response to long term acidification. 1.

INTRODUCTION

Dwarf shrub communities cover over 8 % of the land surface of the United Kingdom, and in Scotland heather, calluna -, is a major component of the vegetation of more than half of the 4.8 million hectares grazed by freely ranging sheep, cattle, red deer and red grouse (Miller, 1979). Much of this vegetation cover is in the uplands, but heather is an important component of the remaining small areas of lowland heath in southern and eastern England. Similarly, heather is an important component of lowland heaths in other European countries, notably in Denmark, France and the Netherlands. Studies in the Netherlands demonstrate that callunaheaths are at risk from acidic deposition, particularly from the increased atmospheric nitrogen deposition. Schneider & Bresser (1987) showed a great and continuing (over the last seven years of their study) expansion of grasses on these heathlands, and Berendse d.(1987) showed that this change was associated with an increase in the mineralisation of soil nitrogen and level of soluble inorganic nitrogen. During the last century NH, has increased approximately 4 fold in the Netherlands, with a 2 fold increase since 1950 (Asman A,,1988). Similar data are available for NO, deposition.

98 2.

TRANSITION FROM HEATH TO GRASSLAND

Fertilizer experiments have confirmed that increases in atmospheric nitrogen deposition could have caused the changes in species composition in Dutch heathlands. Heil & Diemont (1983) and Berendse & Aerts (1984) showed that nitrogen fertilizer addition favoured the at the expense of ericaceous shrubs. growth of DeschamDsla . f l e s c u w and Molinia Roelofs (1986) concluded that the spread of grasses in these heathlands had occurred as the result of the greatly increased nitrogen inputs in recent years. Dueck (1990) demonstrated that heather seedling survival under experimental conditions was three times higher in charcoal filtered than in ambient air in central Netherlands, and that addition of 53 pg m.’ NH, and 92 pg m-, SO, to ambient air reduced survival even more, but the growth of surviving seedlings was stimulated by NH,treatment. However, replacement of heather by grasses may not be directly mediated by acidic deposition. Thus the demise of heather may or other insects be accelerated by the activities of the heather beetle (Jachmaea perhaps responding to physiological changes in the host plant induced by the increased nitrogen supply (Brunsting & Heil, 1985; Berdowski & Zeilinge, 1987).

m)

A decline in heather could also be mediated through effects on its mycorrhizal symbionts. In unpolluted regions dwarf shrubs are heavily dependent on mycorrhizal infection for the absorption of mineral ions from the impoverished soils which support them (Read, 1983; Read & Bajwa, 1985). There are a number of studies which suggest a decline in ectomycorrhizal activity in nitrogen polluted regions. Arnolds (1985) and h o l d s & Jansen (1987) showed a decline in mycorrhizal fruiting bodies in the south of the Netherlands. Boxman a al. (1985) showed that the growth of 14 ectomycorrhizal fungi was reduced by 6040% in pure culture with NH,: K ratios greater than 10 compared to controls with NH,: K ratio of 1. In an experiment designed to examine the role of ericoid mycorrhizal infection in ammonium uptake by V a c c i n m (Stribley & Read, 1976) it was observed that while increases of NH, concentration up to 8 p.p.m. in solution led to an increase in the amount of infection (visibly scored), the infection was greatly inhibited at higher concentrations. At 56 p.p.m. NH, it was reduced to very low levels. While some clear responses of mycorrhizas to nitrogen treatment have been found in the laboratory, there is a need to discover more about the effects of nitrogen on the symbiosis in the field. The consequences of inhibition of mycorrhizal infection are likely to be severe. The host plant has been shown to be dependent upon its fungal associate for the exclusion of toxic metals (Bradley a d., 1981, 1982) and for the metabolism of potentially toxic organic acids (Read, 1986). Stresses arising from both of these sources are potentially high in any heathland soil because of its inherent acidity, but are likely to be of even greater importance under circumstances of increased acidic deposition. In Britain, the northern Peak District has been the most polluted rural region for most of the last two centuries (Lee d d.,1987). In this upland region Anderson & Yalden (1981) showed that there had been a decline of 37% in the area occupied by dwarf shrubs since 1913. This decline cannot simply be interpreted in terms of a response to acidic deposition because of land-use changes, e.g. increases in grazing pressure, which could have had a marked effect on communities of these plants. However, the pollution climate was heavily

99 sulphur dominated for much of this period (Ferguson & Lee, 1983) and there may have been a role for atmospheric pollutants in this vegetation change.

3.

NITROGEN DEPOSlTION AND THE GROWTH OF CALLUNA IN UPLAND BRITAIN

There is a marked gradient in the deposition of atmospheric nitrogen in upland Britain, with the northern Peak District currently receiving the highest deposition whereas areas of similar altitude in North Wales and Scotland, more remote from major urban and industrial conurbations, receive considerably less (United Kingdom Review Group on Acid Rain, 1990). Concentrations of NO, and NH, in the British uplands are generally low ( < 10 and 6 ppb respectively) although uptake of the latter is high because of very high deposition velocities. Measurements by Fowler and colleagues at I.T.E., Edinburgh over moorland have given deposition velocities of 1-2 mmd for NO2 and 30-50 mms' for NH, (U.K.R.G.A.R., 1990). Wet deposition of organic and reduced nitrogen is, however, likely to dominate over dry inputs over upland heathlands and with an increase in altitude there is a greater deposition of nitrogen because of the prevalence of orographic cloud and the effects of the seeder-feeder process. The concentrations of major pollutant ions in orographic cloud usually exceed those in rain by a factor of 1.5 to 8 (U.K.R.G.A.R., 1990) and levels of NO,, NH, and SO, of several millimolar have been recorded in the Pennine region of northern England. The shoots of heathland plants in this area will, therefore, be bathed for long periods in solute-rich cloud vulparis, in particular, offers a very large and highly wettable and mist-water. surface area across which direct uptake of nitrogen compounds has been demonstrated (Bobbink ad.,1990). This route for nutrient capture which by-passes the conventional soilroot pathway may be a factor contributing to the high nitrogen content of in polluted regions. Figure 1. shows that heather plants in the Peak District have higher shoot total tissue nitrogen concentrations at all altitudes than plants from the less polluted North Wales and Scottish mountains. There is also a marked increase in tissue nitrogen with altitude in Peak District plants which mirrors the increase in deposition (Conlan, 1991). At the very least, these data are suggestive that atmospheric nitrogen deposition may be influencing dwarf shrub communities in this polluted region even if there are not direct detrimental effects of the deposition on the plants. For example, increases in tissue nitrogen concentrations may have profound effects on insect herbivores and decomposition processes.

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4.

EXPERIMENTAL ADDITION OF NITROGEN TO HEATHLANDS

In less polluted districts the upland soils may be impoverished in nitrogen, and Calluna may show a growth stimulation in response to nitrogen fertilizer addition (Miller, 1979). In an experiment which has been running since May 1989, nitrogen has been added to a heather moor in the late building phase of growth in North Wales at 14-21 day intervals so as to increase the natural deposition (1-2 g N m-2y.') by 4, 8, 12 and 20 g mJ y". There was no growth response in the first year of application, but in year two extension growth in the high nitrogen treatments was markedly stimulated (Figure 2.). This response was maintained in the third year of treatments and flowering was markedly increased. Since the site was

I00

selected to be a near monoculture of heather competitive effects with other species were not observable, but an inference of this experiment is that in the short-term enhanced atmospheric nitrogen deposition in this district might increase the competitive ability of €a!ha. Figure 3. shows that after two years of treatment there is no apparent decline in mycorrhizal infection as the result of the nitrogen treatments. This result, which is contrary to expectations, is supported by experiments in the Danish heathlands of Johansson (University of Copenhagan), who also found no decline hut, indeed, an increase in mycorrhizal infection in nitrogen fertilised heather (personal communication). The lack of an effect of the nitrogen supply on the apparent mycorrhizal infection at our site in Wales may be because this heathland is still strongly nitrogen-limited. The positive shoot growth response that is still being observed after two and half years suggests that this is the case. There are also marked increases in shoot total tissue nitrogen and free amino acid contents (Figure 4.), but no significant effects on the shoot contents of other minerals, such as potassium and magnesium. 5.

EFFECTS OF NITROGEN SUPPLY ON FROST TOLERANCE

An increase in shoot growth in response to pollution treatment need not necessarily represent a beneficial effect. For example, increase in shoot growth without concomitant increase in root growth might adversely affect plant water relations. Similarly if the pollutant affects the timing of growth and the length of ‘dormant’ periods, there may be the potential for growth stimulation to result in increased sensitivity to early or late frosts which are prevalent in upland districts. At our field site in Wales we have observed a range of responses to nitrogen which illustrate the complexity of the problem. In the first autumn and winter after the commencement of treatments in May 1989, the increased nitrogen input raised the hardiness of shoots of to simulated frosts (Figure 5.). During the second winter this effect disappeared, and plants from the nitrogen treatments and the control were equally sensitive to frosting treatment. Following both winters, observations during May indicated growth of new shoots was further advanced in Calluna receiving the greater inputs of nitrogen. This new growth could be especially vulnerable to late spring frosts. Data in support of such a possibility comes from a misting experiment in which calluna plants were subjected to mists containing either 1 mM ammonium or nitrate or 0.5 mM of both ions. The plants were misted for two days per week in an unheated glasshouse over the summer and autumn. Freezing treatment in spring revealed increased frost sensitivity of nitrogen treated plants (Figure 6.) because these had developed new shoots, while the controls, which were less injured, had not. Other experiments in which plants have been fumigated with mixtures of 40 ppb SO, and 40 ppb NO, over long periods (up to 11 months) have also revealed a marked increase in the frost sensitivity of Calluna although the same plants showed an increase in growth. These concentrations of gases are high compared with present-day rural concentrations in the U.K., but high concentrations of at least SOz (>50 ppb as an annual mean) certainly existed in the northern Peak District earlier this century (Ferguson & Lee, 1983); and it is possible that increased susceptibility to other environmental stresses induced by air pollutants played some part in the reduction of dwarf shrub communities in this region.

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6.

CONCLUSIONS

There is a great variety of present-day pollution climates in western Europe, and similarly there has been a great variety in the past. Unlike the responses of annual crop plants, the behaviour of semi-natural ecosystems to acidic deposition may be mediated over long time periods. An excellent example of this is the work of Van Dam (1990) who showed in a study of chalk grassland in the Netherlands, that changes in atmospheric nitrogen input could take centuries to produce a new steady state in terms of nitrogen output in drainage waters. Thus it is very likely that many European semi-natural ecosystems are responding to both past and present pollution events. Thus it should not be expected that similar experimental treatment5 in different regions at one instant in time will produce similar results. Further complications in making generalisations of the effects of acidic deposition on Calluna are first changes (or differences) in land use in different regions (e.g. presence or absence of turf cutting and the frequency of burning), and second large differences in climate. The experiments described above have largely been concerned with the response of Calluna to pollutants in upland Britain where current concentrations of SOz, NO, and NH, are extremely low and below the levels which have been shown to adversely affect the growth of plants in fumigation experiments. However, concentrations of the solution products of these gases can be high ( > 1 mM) in the mists in at least some parts of the British uplands. Thus in these regions, in contrast to the position in the Netherlands and Denmark, wet deposition is more important than dry deposition as a source of pollutants. It is therefore not surprising that a variety of responses have been observed in Calluna dominated communities in experiments designed to increase nitrogen Loading. These range from marked increases in shoot growth and flowering reported here from upland Britain, to little or no response in the Netherlands (e.g. Heil, 1984). A difficulty with all field experimentation designed to investigate increased acidic deposition is the simulation of ‘natural’ deposition processes. In semi-natural vegetation this is especially difficult, and even experiments of the kind reported here in which solutes have been added at regular 14-21 day intervals are relatively rare. A further problem is sustaining the experiments over a long enough periods to make them ecologically meaningful. Ice core data and historical measurements demonstrate for instance marked increases (2-3 fold) in nitrate deposition this century in the northern Hemisphere. To what extent can large additions of solutes over short periods of time (usually less than three years) be said to mimic deposition changes over several decades’? To what extent can short-term responses to these additions be used to assess likely responses to ‘natural’ pollution loading? The answer to both these questions is very uncertain, and caution should be used when interpreting the results of such field experimentation or when using them as a basis for modelling ecosystem change. There is a great need to sustain field experiment using concentrations of solutes which can be measured in deposition over long periods of time ( > 5 years). Thus the responses of Calluna in the experiment reported here may represent only some initial response to a marked increase in pollution loading, and major changes in for example host-symbiont relations, mineral nutrition, frost and drought sensitivity may only be observable by sustained experimentation over many years. This may be less true where dry deposition predominates, and high concentrations of pollutant gases have acute effects on vegetation.

102 Heathland, established on generally nutrient-impoverished acidic soils, is undoubtedly potentially susceptible to perturbation by nitrogen and sulphur pollutants. There is clear evidence that this perturbation has occurred and is occurring in parts of western Europe today. Nitrogen deposition is probably critical in this process. Modelling studies suggest that transformation of heathland to grassland in the Netherlands, mediated partly by heather beetle attack, occurs when nitrogen deposition exceeds 15 Kg N ha'' y-I (Bobbink id., 1990). However, it is doubtful whether this is a critical load which can be applied uniformly for the protection of European heathlands. The diversity of QUunii dominated heaths, their climates, management and pollution histories combine to preclude generalisations as to their responses to present-day pollution loading. 7.

ACKNOWLEDGEMENTS

Financial support for part of this work came from the Natural Environmental Research Council, U.K. We are grateful for the assistance of M. Cuculescu, N. Manley, J. Suter and Dr T.W Ashenden.

8.

REFERENCES

1.

Miller, G.R. (1979). Journal of Ecology, 67, 109-129.

2.

Schneider and Bresser (1987). Dutch Priority Programme on Acidification, report number 00-04,RIVM, Bilthoven, Netherlands.

3.

Berendse, F., Beltman, B., Bobbink, R., Kwant, R. and Schmitz, M. (1987). Acta Oecologia, 8, 265-279.

4.

Asman, W.A.H., Drukker, B., Janssen, A.J. (1988). Atmospheric Environment, 22, 725-735.

5.

Heil, G.W. and Diemont, W.H. (1983). Vegetatio, 53, 113-120.

6.

Berendse, F. and Aem, R. (1984). Oecologia Plantarum, 5, 3-14.

7.

Roelofs, J.G.N. (1986). Experentia, 42, 372-377.

8.

Dueck, Th.A. (1990). Functional Ecology, 4, 109-116.

9.

Brunsting, A.M.H. and Heil, G.W. (1985). Oikos, 44, 23-26.

10.

Berdowski, J.J.M. and Zeilinge, R. (1987). Journal of Ecology, 75, 159-175.

11.

Read, D.J. (1983). Canadian Journal of Botany, 61, 985-1004.

103

12.

Read, D.J. and Bajwa, R. (1985). Proceedings of the Royal Society of Edinburgh, 85B, 317-332.

13.

Amolds, E.J.M. (1985) (Ed.). Verandefinga Wetenschappelijke Mededeling K .N .N .V . N . 1 67.

14.

Arnolds, E.J.M. and Jansen, A.E. (1987). Dutch Priority Programme on Acidification, report 25-01. Wijster Biological Station Communication nr. 334.

15.

Boxman, A.W., Sinke, R.J. and Roelofs J.G.M. (1986). Water, Air and Soil Pollution, 31, 517-522.

16.

Stribley, D.P. and Read, D.J. (1976). New Phytologist, 77, 63-72.

17.

Bradley, R., Burt, A.J. and Read, D.J. (1981). Nature, 292, 335-337.

18.

Bradley, R., Burt, A.J. and Read, D.J. (1982). New Phytologist, 91, 197-209.

19.

Read, D.J. (1986). Proceedings of First European Symposium on Mycorrhiza. INRA, Paris (Ed. by V. Gianinazzi-Pearson and S. Gianinazzi) p. 169-175.

20.

Lee, J.A., Press, M.C.,Woodin, S.J. and Ferguson, P. (1987). In: Effects of atmospheric pollutants on forests, wetlands and agricultural ecosystems. (Ed. T.C. Hutchinson and K. Meema). Springer-Verlag, Berlin.

21.

Anderson, P. and Yalden, D. (1981). Biological Conservation 20, 195-213.

22.

Ferguson, P. and Lee, J.A. (1983). Atmospheric Environment, 17, 1131-1 137.

23.

United Kingdom Review Group on Acid Rain (1990). Environment (UK).

24.

Bobbink, R., Heil, G. and Raessen, M. (1990). Dutch Priority Programme on Acidification, report project 119.

25.

Conlan, D.E. (1991). Effects of Atmospheric pollutants on survival of SDhaenum in the southern Pennines. Ph.D. thesis, University of Manchester.

26.

Van Dam, D. (1990). Atmospheric deposition and nutrient cycling in chalk grassland. Ph.D. thesis, University of Utrecht.

27.

Heil, G.W. (1984). Nutrients and the species composition of heathland. Ph.D. thesis, University of Utrecht.

in de paddestoelenflora.

Department of the

104 18

N Content mg / g DW

0

I

I

I

I

200

400

600

800

Altitude (metres) FIG. 1.

Nitrogen content of shoots of Calluna vulgar& sampled from sites in the Peak District (southern Pennines), Wales and Scotland after the shoot growth season, August 1990.

100

80 -

20

-

I

FIG. 2.

I

Current years shoot growth (1990) after 18 months application of nitrogen to field plots.

105 3

ug Glucosamine/mg DW (mean

"

*/-

8.e)

8

4

0

12

NIT RO GEN A PPL ICAT I0N ( g/ m2 / y 1 Estimate of fungal chitin (as glucosamine equivalents) as an indication of mycorrhizal infection of roots in plots of heather after different nitrogen inputs over two years.

FIG. 3.

(mg/g DW)

F.A.A. (mmol/g DW)

25

20

0.04

0.03

15 0.02 10

0.0 1

5

0

0 0

4

8

12

16

20

NITROGEN AP PLICAT1ON (g/m 21y Nitrogen

FIG. 4.

Amino-Acid8

Nitrogen and free amino-acid contents of heather shoots sampled in May 1991 after two years of nitrogen additions.

106

.

0.06

.

0.04

0.02

0

+4

FREEZING TEMP

wo FIG. 5 .

-20

-15

-10

(C)

NITROOEN (g/m2/y) 134 8

12

Cellular frost injury in heather shoots sampled from the field and given a simulated frost at different temperatures. Typical result from year one (March 1990) showing less ion leakage from the high nitrogen shoots.

r e a n Injury Score

-10

-15

-20

FREEZING TEMP (C) -Control

FIG. 6.

mNH4

O N 0 3

(NH4*N03)

Visible frost injury to new growth after simulated frost at different temperatures using heather previously grown in a cold greenhouse and given mist treatments with different nitrogen (1 mM) forms.

T Schneider (Editor). Acidification Research. Evaluation and Policy Applications @ 1992 Elsevier Science Publishers 6.V All rights reserved

107

The interaction of forest vegetation and soils with the aquatic environment; effects of catchment liming on lakes T.R.K. Dalziela, G. Howellsb and R.A. Skeffington‘ aPowerGen plc, Ratcliffe Technology Centre, Ratcliffe-On-Soar, Nottingham NG11 OEE bUniversity of Cambridge, Department of Zoology, Cambridge, CB2 3EJ “National Power plc, Technology and Environmental Centre, Leatherhead, Surrey KT22 7SE

Abstract Evidence that processes within forests and soils, as well as atmospheric deposition, influence the aquatic environment, specifically the phenomenon of acidification, is reviewed and mechanisms suggested for acidification effects. The Loch Fleet Project (Galloway, southwest Scotland) has collated data on atmospheric deposition, soil drainage and runoff chemistry since 1985, with liming of parts of the catchment in 1986 and 1987. Comparison of conditions prior to and following liming of forested and moorland areas can be compared. Both direct observations and modelling of forest acidity transfers demonstrate that forests can acidify soils and surface waters in the absence of acid deposition and forest soil liming is suggested as a possible sylvicultural management practice to counter such acidification. 1. INTRODUCTION

Empirical observations demonstrate that forest conditions have a marked effect on the aquatic environment through changes in physical conditions (light, temperature, sediments, hydrology) and chemical conditions (inorganic and organic components). Characteristic chemical differences include lower pH, increased aluminium concentrations and reduced calcium concentrations. These changes are considered deleterious because they may reduce or eliminate populations of fish and invertebrates and alter food chains of which these species are a part. Comparison of forested and non-afforested streams in central Scotland (Harriman and Morrison, 1982) and in central Wales (Stoner d., 1984), and observation of acidified waters in forest areas with no significant acid deposition, such as the Amazon basin (Krug, 1991) are consistent. The associated biological effects include the loss or depletion of brown trout in forest streams, death of planted salmon eggs and impoverished invertebrate faunas. Loss of fish and other aquatic fauna in a stream draining a spruce forest in Belgium was attributed to some unknown toxic agent (Huet, 1951). The regional growth of coniferous afforestation in Scotland, especially in upland areas where spawning and nursery streams occur, has been strongly

108

u.,

correlated with declining salmon catches there since the 1950s (Egglishaw 1986). When forests have matured sylvicultural practices also often have adverse effects in the aquatic ecosystem; clear felling is associated with a long sustained higher release of 1977, Feger U., 1990, Adamson and Hornung, 1980). Physical nitrate (Likens d., disturbances include drainage changes with associated soil erosion and turbidity in streams, inadvertent destruction of spawning beds, timber litter transfer to water courses and oil and chemical spills. These potential problems and advice on their avoidance have been incorporated into "guidelines" for safe forestry practice (Forestry Commission, 1988). Although the significance of these adverse effects has been challenged (Nisbet, 1990), a relationship between afforestation and acidification seems to be substantial, perhaps especially evident in UK where reafforestation of upland areas (which had lost forests as long as 50,000 years ago, Pennington, 1981) has been promoted over the last 70 years, and where climate, soils and geology and pollution climate may predispose to forest acidification effects.

2. MECHANISMS FOR THE FOREST ACIDIFICATION EFFECT Several mechanisms can be identified to account for the forest acidification effect 1982; Likens 1977). These are; (Skeffington, 1987; van Breeman

u.,

u.,

- Enhanced capture (scavenging) of acidifying pollutants - Progressive base cation uptake by trees - Accumulation of an acidic litter layer

- Organic acid production in the root zone

- Enhanced evapotranspiration and drying out of soil leading to greater oxidation potential

- Altered water pathways through soil, limiting soil/drainage interactions These have varying degrees of importance with regard to surface water acidification and are discussed in turn below.

2.1 Scavenging of acidifying pollutants Deposition of atmospheric acidifying pollutants to forest is greater than to moorland via canopy scavenging of mist/fog and greater dry deposition of aerosols in the more complex canopy of trees. This is evident in the greater transfer (flux) of solutes with higher concentrations sampled in throughfall beneath trees. Fowler Ual. (1989) has argued that the increased surface roughness and complexity of forests reduces atmospheric resistance, and for gases (HCI, HNO,, NH,), dry deposition will be enhanced. However, experimental exposure of forest canopy to SO, enhanced NH, capture (McLeod 1990), a mechanism which ought to work in reverse.

u.,

More important for British uplands, however, is the capture of cloud and mist which contain high concentrations of solutes, including pollutants, relative to rain. Forests appear to be effective collectors of mist and fog, which are quite commonplace in the U.K., as well as in some other areas, e.g. Adirondacks, U.S.A. Deposition of cloud water also increases with windspeed, a feature of the Atlantic climate. The consequence of

109

enhanced scavenging will also be an increased influx of acidifying agents, and also, in maritime areas, of sea salts - significant in episodic release of acidity and aluminium from acidic soils (Irwin 1989).

a.,

2.2 Base cation uptake It is accepted that cations are taken up into above-ground biomass as growth proceeds; t estimates vary with season, climate, species, expected yield and age of stand. Nilsson g al. (1982) give 500 mEq.H+-year for young spruce and only 4 mEq Hf-year for old pine. Increased forestry has been estimated to deplete basic cations at a rate of up to 0.5 KEq ha"-year at a variety of sites in Sweden (Sverdrup and Warfvinge, 1990). This uptake of cations into biomass results in depletion of the same materials in the root zone soils; Skeffington (1983) found acidified soils well beyond the projection of the canopy of isolated trees, implying an influence other than that of the canopy. Cation removal results in an excess of anions within the soil - the concept of a pool of "mobile anions" which could transfer the soil acidity to runoff, to maintain the charge balance. It is argued that acidification will not occur without pollution and the consequent mobile anions (D.O.E., 1991). However, recently this concept has been strongly criticised (Krug, 1991) since alternative measurements are mutually inconsistent. Further, it is now accepted that anion substitution (organic acid buffering) plays a significant role. There is also well documented evidence that soil acidification has occurred prior to combustion generated pollution and indeed from land management regimes which allow the accumulation of soil organic materials (Jenkinson, 1970).

2 3 Increase in acidic litter layer An increase in the acidic litter layer characteristic of many coniferous forests effectively lowers the soil base saturation by increasing the soil exchange capacity that is occupied by the hydrogen ion (Skeffington, 1897). The effect, as with the depletion of the base reserve in the soil is to mobilise acidity (and aluminium) with additional solute input. The development of an acidic litter layer is particularly important where the predominant flowpaths are through the litter and surface horizons.

2.4 Organic acid production Leachates from the surface horizon of soils are often very acidic because of soluble organic acids. If these leachates then pass through a mineral soil horizon the organic acids tend to precipitate and there is little effect on surface water acidity. Direct runoff from these surface layers into watercourses may be very acidic but toxic inorganic fractions of aluminium are unlikely to be present. The general association of forestry and surface water acidification cannot, therefore, be satisfactorily explained by this mechanism. 2.5 Increased evapotranspirationand drying out of soil It is well established that evapotranspiration from the complex forest canopy is significantly greater than that from moorland or other short vegetation; about 30% compared with 20% in Scottish and Welsh catchments (Calder and Newson, 1979; Hall, 1987). This will lead to greater solute concentrations (and acidity) in the throughfall. Other sylvicultural practices in the U.K. (ploughing across contours, field drainage) will

110

also lead to significant drying out of the soils. Although soil or bedrock S has seldom been reported (but rather estimated from input/output budgets) it is undoubtedly present in many mineral formations and consequently the sulphate input from this source may have been underestimated (Berner, 1984; Krug, 1991). In the acid sensitive catchments in the northeastern U.S.A., 25 to 50% of atmospherically deposited S is retained, much of it in reducing conditions in mires and aquatic sediments, or imperfectly drained, gleyed terrestrial soils. Release of oxidised S (sulphate) from such stores during seasonal drying out has been demonstrated in a variety of conditions (Brown, 1985; Sullivan 1986; Jones 1983).

a.,

u.,

2.6 Altered water pathways through soil Forestry practices often result in substantial changes to drainage pathways, usually with the objective of increasing the rate of runoff. Such alterations often reduce the contact time between drainage water and mineral horizons, when some base materials can react with acidity. However, upland land drainage in the U.K. has also been undertaken extensively and there is little documentation about consequences for runoff water quality. Other hydrological manipulations, such as lowering the water table or diverting less acid water supplies, may be relevant since upland areas in the U.K. are exploited for both water supplies and forests. It is found that at the acid Loch Fleet, a groundwater source rich in alkalinity contributes about 5% of water input to the lake. It has been suggested that only a small increase in flow would restore the lake to preacidified conditions, and that the flow of groundwater was diminished over the past three decades (Cook d.,1991), contemporary with regional afforestation. 3. THE LOCH FLEET PROJECT The Loch Fleet Project started in 1984 and forms part of the Joint Environmental Programme operated by PowerGen and National Power. The Project is also jointly funded by ScottishPower, Scottish Hydro-Electric and British Coal. The objectives of the Project are (a) to demonstrate that the water chemistry of the lake can be brought into a range suitable for trout by one or more of several treatments to the catchment area, including liming, and (b) once suitable chemical conditions have been achieved, to demonstrate the suitability of the water for a self-sustaining brown trout population. Loch Fleet is situated 340 m above sea level in Galloway, southwest Scotland where acid deposition is high, similar to that in northwest Europe, e.g. Norway. The site was considered suitable as a site to undertake the proposed project for a number of reasons. The loch is small (0.17 km’) and it was acid (pH 4.0 to 4.5) when the Project was initiated in 1984 and had once supported a trout fishery. Historical records of the fishery and the regional land use were known and the catchment was considered manageable by virtue of its size. The underlying geology of the catchment is a coarse-grained muscovitebiotite granite low in calcium and magnesium, overlain by organic peaty soils or peaty rankers with some glacial drift material. Vegetation is mostly of rough moorland character dominated by heather (Calluna vulgaris) and flying bent grass (Molinia caerulea), but the lower (SW) part of the catchment was planted with coniferous trees (Sitka spruce, Picea sitchensis, lodgepole pine, Pinus contorta and larch, Lark x

111

eurolepis) in 1963. Climatic conditions at this exposed, upland site are quite rigorous. The predominantly westerly airstream has a speed greater than 11 ms-' for more than 25% of the time and gales for 5% of the time. Precipitation is greater than 2 m per year - the 30 year mean between 1950 and 1980 was 2135 mm per year. At the beginning of the Project distinction was made within the catchment of a number of reasonably distinct experimental areas, or sub-catchments, using Ordnance Survey contours, aerial photographs and on-site surveys (Figure 1). Sub-catchments IV, VI and VII were limed with different dosages of finely ground limestone in April 1986. Full details can be found in Howells and Dalziel (1988). Subcatchments IV and VI received between 20 and 25 tonne ha". O n sub-catchment IV the limestone was applied as a slurry beneath the tree canopy; on sub-catchment VI it was applied as a dry powder, distributed as uniformly as possible over the area. Rain water quantity within the Loch Fleet catchment is measured using a Meteorological Office standard 5 inch gauge with the aperture 30 cm above ground level. Volumes of rainfall are recorded manually on a daily basis and monthly values are computed from the sum of the daily readings. Rain water quality is measured in samples collected in a British Standard Bulk Collector located adjacent to the rain water quantity gauge. Chemical analysis is carried out when there is sufficient sample volume (in practice > 300 ml). Samples are analysed at the ScottishPower chemistry laboratory at East Kilbride for conductivity, pH, sodium, potassium, ammonium, calcium, magnesium, chloride, 1991). sulphate and nitrate. Details of analysis methods can be found in Stewart

u.(

Sub-catchments IV, VI and VII each have a distinct drainage stream (Figure 1). Automatic flow monitoring equipment, employing either v-notch weirs or trapezoidal flumes connected to stilling wells, was established on these streams and on the loch outlet stream. Water sampling sites, from where samples for chemical analysis are collected daily, were established adjacent to the flow monitoring points. Details of sample 1991a). collection and analysis can be found in Dalziel

u.(

Data on deposition inputs to the Loch Fleet catchment and on runoff outputs from the monitored sub-catchments and from the whole catchment have been collected since April 1985, one year before liming. In addition to collecting these data, a separate study was carried out between February and December 1985 by members of the Macaulay Land Use Research Institute, Aberdeen, to investigate, in more detail, the interactions between deposition, vegetation and soils. Experimental plots were established in afforested parts of the catchment below the two main species of conifer, lodgepole pine and Sitka spruce, from where data on vegetation throughfall, stemflow, soil water drainage and ditchflow were collected. A separate bulk deposition collector was used in this study, located in a forest fire break between sub-catchments I1 and 111 (Figure 1). In addition, a further deposition collector was deployed, termed an "interception gauge", which was designed to capture cloud water, mist and fog (collectively known as "occult deposition") as well as true rainfall.

112

4. MODELLING

Modelling, using the MAGIC (Modelling of Acid Groundwaters In Catchments) model (Cosby 1985), modified somewhat to allow more flexible run periods and to incorporate liming, was employed to answer some questions concerning the effects of forests on surface water acidification and the ways in which deleterious effects might be minimised. The runs were performed to simulate the response of a Norway spruce forest

u.,

&\

LOCH FLEET

A "/

OUTFLOW FROM

0

-

-

WATERSHED REASONABLY WELL DEFINED WATERSHED LESS WELL DEFINED LOCH EMBAYMENT ......... SmEAMS ... .,.,... ...... . .. . CONIFEROUS TREES 0 '4-NOTCH WEIR/TRAPEZOIDAL FLUME

I -

Figure 1. The Loch Fleet Catchment

500 METRES

113

growing on poor soil overlying a substratum with a low weathering rate. Excess base cation uptake by trees varies widely with species, growth rate, site conditions, life-cycle stage etc. For Norway spruce in southern Sweden Nihlglrd (1970) gives a rate of 97 mEq m-2-year; Norway spruce in central Germany took up 93 mEq m.'-year and Binkley and Richter (1987) quote a figure of 90 mEq m.'-year for a forest in the USA. The model runs used 90 mEq m"-year, divided 54:21:15 in equivalents between Ca2+:MgZ+:K+.The forest was modelled as growing on an acidic soil with a base saturation of 13%, on a substratum with a weathering rate of 11.3 mEq m-'-year.Charlson and Rohde (1982) estimated that the pH of pristine precipitation was between 4.5 and 5.5, taking into account natural emissions of S and N compounds. In order to predict whether forest growth could, in principle, acidiij surface water if there was no anthropogenic acid deposition a "background deposition" run was carried out using precipitation at p H 5 with an excess (or non-sea salt) sulphate concentration of 11.8 pEq 1.'. Data input for this run simulates the presence of mobile anions from only natural S and N emissions. To investigate what effect deposition containing mobile anions&o from sea salt would have, a run was carried out with no excess sulphate and a precipitation p H of 5.78. The p H is higher than the theoretical 5.65 because of a slight excess of Ca2+ and K', but essentially this is sea salt pH. A third run employed the same input data as used in the "back round deposition" run, but incorporating forest soil liming. The dosage was 25 tonne ha- of limestone (CaCO,), which was assumed to dissociate over 4 years.

B

5. RESULTS 5.1 Loch Fleet Project Data

Table 1 shows volume-weighted mean ion concentrations for major ions in bulk precipitation and in stream water runoff from sub-catchments IV (afforested) and VI (moorland) over the period 1985 to 1991. The considerable maritime influence on the precipitation inputs at Loch Fleet, due to the site's proximity to the Irish Sea, is clear from the high sodium and chloride concentrations recorded. These are particularly noticeable between October and March each year due to the quite severe westerly and south-westerly winds associated with high rainfall over this time. The precipitation volume-weighted mean pH between 1985 and 1991 was 4.72 (19 pEq1-I). Less than 4% of the rain events had pHs less than 4 and more than 50% exceeded p H 5. As would be expected, because of evapotranspiration, there is a greater water deficit on sub-catchment IV compared with that on sub-catchment VI. The effects of catchment liming in April 1986 on runoff water quality can be clearly seen by the decrease in Hf and increase in Ca2+ concentrations from 1986-87 onwards. Over 1985-86, prior to liming, however, the acidity of the runoff from both sub-catchments was significantly greater than that of bulk precipitation; runoff from the afforested sub-catchment IV being slightly more acidic than that from sub-catchment VI. Sodium, magnesium, sulphate and chloride concentrations were all elevated in sub-catchment runoff compared with those in bulk precipitation and in all cases concentrations from sub-catchment IV were greater than those of subcatchment VI.

114

Table 1. Volume weighted mean concentrations of major ions in bulk precipitation and in runoff from sub-catchments IV and VI at Loch Fleet, 1985-91. Units are pEq 1-1

19t15-91

M48

177

1

50

97

22

206

115

Table 2 shows the volumes of precipitation and concentrations of major ions collected by a bulk precipitation collector and an "interception gauge" between February and December 1985. The gauge was designed to capture more effectively occult deposition", in a way similar to coniferous trees. The results show that the bulk collector underestimated deposition and ion inputs considerably. The volume collected by the "interception gauge" was more than double that of the bulk collector, indicating the high incidence, at this upland site, of wet deposition other than rainfall. Sea salt ion inputs were increased fourfold in "interception gauge" samples, and non-sea salt concentrations were doubled, compared with samples from the bulk collector. Figure 2 shows the amount of bulk precipitation compared with throughfall recorded beneath lodgepole pine and Sitka spruce stands between February and December 1985. Interception loss was greater in lodgepole pine canopy (35%) compared with that in Sitka spruce (15%). In the same way as shown by the "interception gauge" samples, the throughfall below the forest canopy produced considerably enriched ion concentrations due to foliage-captured inputs of mist and fog (Figure 3). In most cases, ion enrichment was greater in Sitka spruce than in lodgepole pine.

Collector

volm, mn

H+/~H

~ a +

ca2+

ng2+

K+

Bulk collector

2004

29.5/ 4.53

73.5

13.0

16.5

3.8

25.0

"Interception gauge"

4735

37.2/ 4.43

353.0

30.9

56.4

8.4

44.3

NH,,+-N

so4*-s

CI-

19.3

56.7

65.4

45.7

101.6

341.4

N+--N

i

The effects of catchment liming on surface water chemistry are shown in Figure 4, for runoff from sub-catchment IV. Calcium concentration and pH increased rapidly following liming with a dosage of 24 tha-'. Although the calcium concentration has declined since the maximum reached soon after liming, it is predicted that the runoff water quality from this sub-catchment in addition to that from sub-catchments VI and VII, which were limed at the same time, will maintain satisfactory conditions for trout survival within the loch 1991b). at least until the end of the century (Dalziel

a,,

5.2 Modelling The results of the "background deposition" run, in which the effects on surface water acidification of a forest subjected to precipitation of pH 5 and only 11.8 pEq 1.' non-sea salt sulphate are modelled, are given in Figure 5. The results suggest that forests can indeed acidify waters significantly by base cation uptake alone, even in the absence of "acid rain". Over the 60 year growth period, stream water pH declines, aluminium concentration rises to potentially toxic levels and calcium concentration falls steadily as the soil cation exchanger becomes depleted of calcium.

116

0

_i L/POLE

SITKA

Figure 2. Amount of bulk precipitation (BP) compared with throughfall recorded beneath lodgepole pine (LPOLE) and Sitka spruce (SITKA) stands between February and December 1985. (Data of Nisbet and Nisbet, 1991) The results of this run indicate that in spite of the lack of excess strong acid anions in precipitation the trees, by sequestering base cations, effectively create them and the stream water pH falls as a consequence. At these pHs there will be some contribution from HCO, as a mobile anion. There is no aluminium present at these pHs. From this run it seems clear that tree growth can acidify surface waters in the absence of acid precipitation. Figure 6 shows the results of the run which employed a precipitation input of sulphate derived only from sea salt.

117

A c i t l i I,y

200

Ammoniuni

Sodium

61

30

5c 25

150 4C

20

100

30

15

20

10

50 10

5

0

0 L/POLE

S17K4

BP

L/POLE

SITKA

Nitrate

bO

Chloride 701:

00 150

30

100

20

50 10

0

0 L/POLE

SITKA

L/POI.E

SI I K A

BP

L/POLE

SIIKA

L/P0LI

SIlUd

Figure 3. Volume weighted mean concentrationsof major ions in bulk precipitation (BP) and in throughfall under lodgepole pine (LPOLE) and Sitka spruce (SITKA) between February and December 1985. Units are pEq1-l. (Data of Nisbet and Nisbet, 1991)

118

Figure 7 shows the effects of forest liming at 25 tha-' to alleviate the effects of excess base cation uptake by a forest. The improvements in water quality parameters appear dramatic and long-lasting. In the real situation, of course, the other acidification mechanisms would also be operating, so liming would not have such a lasting effect.

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119

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120 6. DISCUSSION

6.1 Loch Fleet Project Data Sub-catchment runoff data prior to liming, compared with bulk deposition inputs, show considerable increases in acidity from both forested and moorland sub-catchments. Surprisingly, runoff from the forested sub-catchment (IV) is only slightly more acid than that from the moorland sub-catchment VI. Some of the increased acidity from the two sub-catchments can be attributed to acid generating processes within the catchment vegetation and soils, but it should be remembered that, as demonstrated by the "interception gauge", the bulk collectors underestimate ion inputs, including acidity, quite considerably.

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122 The data on forest throughfall collected in 1985 by members of the Macaulay Land Use Research Institute (Nisbet and Nisbet, 1991) show more clearly the effects of the forest vegetation at Loch Fleet on modifying deposition inputs. Over February to December the canopy increased deposition acidity, in contrast to results from other coniferous throughfall sites elsewhere in the UK, where rainfall, particularly in summer, was effectively neutralised within the canopy. The Loch Fleet results may be due to the nutrient-poor status of the trees. During the winter and spring it is believed that the main acidifylng processes are canopy interception of cloud water and subsequent washout of this deposited acidity in association with sulphate and nitrate. Other factors may also contribute to the increased acidity of bulk deposition beneath the forest canopy such as canopy leaching of sulphate which may make a significant contribution to sulphate deposition in throughfall, particularly in summer, when the trees are metabolically most active. The foliar utilisation of ammonium ions in deposition as a source of nitrogen at this nutrient-poor site may also be acidifying, since this involves an exchange of H+ ions for ammonium.

6.2 Modelling It may be argued that the attempts to model the effects on surface water acidification of base cation uptake by trees are crude. The importance of the results, however, is that they demonstrate that water acidification due to this process alone, is, in theory, possible taking into account known processes parameterized in a realistic way. None of the other potential effects of forest growth are included in the model runs; it is exclusively base cation uptake. The soil is assumed to start in equilibrium with deposition, so in the absence of forest base cation uptake there would be no change. Thus, it appears that on poor soils forestry will exacerbate acidification. It seems necessary, therefore, to consider the policy implications of this. To not plant forests in these areas is one solution being seriously considered in the UK. However, forestry has other functions which are considered socially beneficial, such as timber production, soil protection, recreation etc. An alternative approach might, therefore, be to modify the effects of sylvicultural practice, for example by liming, to replace lost base cations as demonstrated both by modelling (Figure 7) and from the results of catchment liming at Loch Fleet (Figure 4). One disadvantage of liming which was identified in the model run is that Mg2+ and K+ are displaced from the soil exchange surfaces and may become too low for effective forest nutrition. This, however, may be correctible by fertilisation. Overall, the policy implications seem to be that reductions in acid deposition alone will not be enough to alleviate acidification in areas of slow weathering geology and forest growth. Replacement of base cations taken up by the growing forest and critical thought about other sylvicultural practices will be necessary. Liming is one option that should be considered.

7. CONCLUSIONS There is considerable evidence of adverse effects of afforestation on surface water quality.

123 A number of possible mechanisms can be invoked to account for these effects. These include pollutant capture, but also factors intrinsic to forest growth and to sylvicultural practice. Base cation uptake by trees can be modelled and, even in the absence of inputs of strong acid anions derived from combustion, can be shown by modelling to cause surface water acidification. Liming is a possible measure which may need to be considered in some instances in addition to emission controls.

8. ACKNOWLEDGEMENTS The authors would like to thank Katharine Paterson of ScottishPower and Margaret Proctor of National Power for their invaluable efforts in Loch Fleet data collection and processing, respectively. The authors are grateful for permission to publish this paper from the Loch Fleet Project Management Committee and the Joint Environmental Programme jointly undertaken and funded by National Power plc and PowerGen plc.

9. REFERENCES

(m

Adamson, J.K. & Hornung, M. (1990). The effect of clearfelling a Sitka spruce sitchensis) plantation on solute concentrations in drainage water. Journal of Hydrology, 116, 287-297 Berner, R.A. (1984). Sedimentary pyrite formation: an update. Geochim. Cosmochim. Acta 48,605-615 Binkley, D. & Richter, D. (1987). Nutrient cycles and H + budgets of forest ecosystems. 1-51 Advances in Ecological Research

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Brown, K.A. (1985). Acid deposition: effects of sulphuric acid at pH 3 on chemical and biochemical properties of bracken litter. Soil Biol. Biochem. l7,31-38 Calder, I.R. and Newson, M.D. (1979). Land use and upland resources in Britain strategic look. Water Res. Bull. 16,201-211

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Charleson, R.J. & Rodhe, H. (1982). Factors controlling the acidity of natural rainwater. Nature 295, 683-685 Cook, J.M., Edmunds, W.M. and Robins, N.S. (1991). Groundwater contribution to an acid upland lake (Loch Fleet, Scotland) and the possibilities of amelioration. J. hydrol. 125, 111-128

124 Cosby, B.J., Hornberger, G.M., Galloway, J.N. & Wright, R.F. (1985). Time scales of catchment acidification. Environmental Science and Technology, l9, 1144-1149 Dalziel, T.R.K., Proctor, M.V. & Paterson, K. (1991a). Water quality of surface waters before and after liming. In: Howells, G. and Dalziel, T.R.K. (Eds.), "Restoring Acid Waters: Loch Fleet 1984-1990', Elsevier Applied Science, London and New York Dalziel, T.R.K., Dickson, A. & Proctor, M.V. (1991b). Calcium flux calculations and predictions of catchment liming effectiveness at Loch Fleet, Galloway, Scotland. Lake and Reservoir Management, In Press Department of the Environment (1991). Forests and surface water acidification. Department of the Environment, London, UK Egglishaw, H., Gardiner, R. & Foster, J. (1996). Salmon catch decline and forestry in Scotland. Scottish Geographical Magazine, 102,57-61 Feger, K.H., Brahmer. G. and Zottl, H.W. (1990). Element budgets of two contrasting catchments in the Black Forest (Federal Republic of Germany). J. Hydrol. 116,85-99 Forestry Commission (1988). Forests and Water Guidelines. Forestry Commission, Edinburgh, 28pp Fowler, D., Cape, J.N. & Unsworth, M.H. (1989). Deposition of atmospheric pollutants on forests. Philosophical Transactions of the Royal Society of London, 247-265

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Hall, R.F. (1987). Processes of evaporation from vegetation of the uplands of Scotland. Trans. Roy. SOC.Edinburgh (Earth Sciences) 28, 327-334 Harriman, R. & Morrison, B.R.S. (1982). Ecology of streams draining forested and nonforested catchments in an area of central Scotland subject to acid precipitation. Hydrobiologia, 88, 251-263 Howells, G. & Dalziel, T.R.K. (Eds.) (1988). The Loch Fleet Project. a Report of the Intervention Phase (2), 1986-87. CEGB, SSEB, NSHEB, British Coal Huet, M. (1951). Nocivite des boisements en Epiceas pour certain cours d'eau de Ardenne Belge. Verh. Int. Verein. Theor. Angew. Limnol. lJ, 327-334 Irwin, J.G., Campbell, G.W., Cape, J.N., Clark, P.A., Davies, T.D., Derwent, R.G., Fisher, B.E.A., Fowler, D., Kallend, AS., Longhurst, J.W.S., Martin, A., Smith, F.B. and Warrilow, D.A. (1990). Acid Deposition in the United Kingdom, 1986-1988. Dept. of Environment, Warren Spring Laboratory, 124pp Jenkinson, D.S. (1970). The accumulation of organic matter in soil left undisturbed. Rothamstead Report 1970 (2), 113-137

125 Jones, H.C., Noggle, J.C., Young, R.C., Kelly, J.M., Olem, H., Ruane, R.J., Pasch, R.W., Hyantis, G.J. and Parkhurst, W.J. (1983). Investigations of fish kills in fish rearing facilities in Raven Fork Watershed. Tennessee Valley Authority Report TVA/ONR/WR8319, 60pp b u g , E.C. (1991). Review of acid deposition-catchment interaction and comments on future research needs. J. Hydro]. 128,1-27 Likens, G.E., Bormann, F.H., Pierce, R.S., Eaton, J.S. and Johnson, N.M. (1977). Biogeochemistry of a Forested Ecosystem. Publ. Springer, New York, 146pp McLeod, A.R., Holland, M.R., Shaw, P.J.A., Sutherland, P.M., Darrall, N.M. & Skeffington, R.A. (1990). Enhancement of nitrogen deposition to forest trees exposed to SO,. Nature, 347, 277-279 Nihlgdrd, B. (1972). Plant biomass, primary production and distribution of chemical elements in a beech and planted spruce forest in South Sweden. Oikos, 23, 69-81 Nisbet, A.F. & Nisbet, T.R. (1991). Interactions between rain, vegetation and soils. In: Howells, G. and Dalziel, T.R.K. (Eds.) "Restoring Acid Waters: Loch Fleet 1984-1990". Elsevier Applied Science, London and New York Nisbet, T.R. (1990). Forests and surface water acidification. Forestry Commission Bulletin, 86, Her Majesty's Stationery Office, London, UK Pennington, W. (1981). Records of a lake's life in time: the sediments. Hydrobiologia 79, 197-219 Skeffington, R.A. (1983). Soil properties under three species of tree in southern England in relation to acid deposition in throughfall. In: Ulrich, B. & Pankrath, J. (Eds.), "Effects of Accumulation of Air Pollutants in Forest Ecosystems". Reidel Publishing Co., Netherlands Skeffington, R.A. (1987). Soil and its responses to acid deposition. CEGB Research, 20, 16-29 Stewart, B.R., Paterson, K., Dalziel, T.R.K. and Proctor, M.V. (1991). Deposition input considerations. In: Howells, G. & Dalziel, T.R.K. (Eds.), "Restoring Acid Waters: Loch Fleet 1984-1990'. Elsevier Applied Science, London and New York Stoner, J.H., Gee, AS. and Wade, K.R. (1984). The effects of acidification on the ecology of streams in the upper Twyi catchment in west Wales. Environmental Pollution A, 3, 125-157

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Sullivan, T.J., Christopherson, N. Muniz, I.P., Seip, H.M. and Sullivan, P.D. (1986). Aqueous aluminium chemistry response in episodic increases in discharge. nature 223, 324-327 Sverdrup, H. and Warhinge, P. (1990). The role of weathering and forestry in determining the acidity of lakes in Sweden. Water Air Soil Poll. 2, 71-78 van Breeman, N. ,Driscoll, C.T. and Mulder, J. (1984). Acidic deposition and internal proton sources in acidification of soils and water. Nature 307,599-604

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127

HIGHER ORDER EFFECTS

L. Reijnders: IVAM Universiteit van Amsterdam; Stichting Natuur en Milieu. Utrecht. Abstract Acidification has indirect or higher order effects that seem to be less well studied than direct or first order effects. High order effects may be based on the direct interdependence of species. Such effects may result in the increased or decreased viability of species. This in turn will reverberate in foodwebs and ecosystems to which the affected species belong(s). There may be consequences for geochemical cycles. Higher order effects may also go beyond the direct interdepence of species. Two examples of such effects are discussed. One linking high N-depositions with increased concentrations of the greenhouse gas N,O and associated temperature forcing and another potentially linking acidification with increased deposition of oxidized S-compounds. Introduction Direct or first order effects following from the acidification of waters and soils are by now relatively well researched. However acidification has also higher order or indirect effects that seem to be less well studied. This may mean that effects on higher trophic levels, reverberations in food webs, and effects on fluxes of substances are currently underestimated. In this contribution I will outline a number of potential and all-to-real higher order effects of acidification and related atmogenic changes. In doing so there is no pretence of being exhaustive. My aim is primarily to point out that there are higher-order effects, and that such effects may have wide ranging consequences. Hiaher order effects based on interdependence of species Acidification of soils and waters, combined with associated atmogenic stress-factors like high deposition of N-compounds, leads to many changes in the presence and speciation of substances in soils and waters. These changes have direct (first order) effects on a number of species. First order effects, influencing for instance viability or content of trace elements, may in turn cause higher order effects on other species. These effects will in principle reverberate in food webs and ecosystems to which the affected species belong(s). Such higher order effects often follow from the direct interdependence of species. Interdependence of species may take several fo m s . In line with this there are several types of higher order effects. Table 1 briefly summarizes known types of second order effects associated with different types of interdependence.

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Table 1. Types of second order effects caused by acidification based on direct interdependence of species. first order effect

second order effect

stress effect on, or decreased viability of, one (or more) species

increased success of (an)other species decreased viability of (an)other species

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increased viability of (or more) species

ecreased viability of (an)other species

increased viability of (an)other species

changed concentrations of trace compounds in one (or more) species

decreased health or viability in (an)other species

Firstly stress effects on, or the decrease of viability in, one (or more) species due to air pollution, may lead to increasing populations of (an)other species. Stress induced increases in seed production by spruces have benefited the cross bill (Loxia curvirostra) and the citril finch in Germany (1). Populations of woodboring birds have (temporarily) increased parallel with the increase in standing dead trees. These include the white breasted nuthatch (Sitta carolinensis) in Canadian maple forest and the three toed woodpecker (Pircoides tridactilus) in Germany (1). We also see increased success of species, as a second order effect in the transition of heath land (dominated by Scotch heather, Calluna vulgaris) to grassland (dominated by wavy hair grass; Deschampsia flexuosa). This follows from the increased vulnerability of heather to frost, drought and heather beetles, caused by a combination of acidification and high N-deposition. This gives Deschampsia a competitive edge (2). Even more dramatically we see such changes in advanced cases of forest dieback, such as have occurred in the Harz (Germany) and Bohemia. Here one finds the replacement of forest by open woodland. And this in turn has higher order effects, as is clear from large changes in bird populations. In Bohemia and the Harz populations of robin, chaffinch, firecrest and blackbird have plummeted, whereas the number of willow warblers, tree pipits, redstarts and the rare ring ouzels have increased (3). Secondly the decline of one (or more) species may cause the decline of associated species. The association may be based on relations such as providing water, food or cover. A first example of such a second order decline relates to the association between Douglas fir (Pseudotsuga menziesii) and mycorrhiza fungi. The

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latter are probably essential for an adequate water supply to the fir. Because mycorrhiza are negatively affected by acidification, vulnerability of the Douglas fir to drought tends to increase when soils become acidified ( 2 ) . The changed availability of food for other species associated with species decline has been shown to be a major cause of second order effects. In Scandinavia populations of piscivorous birds like red throated divers and mergansers underwent a marked decline due to reduced fishstocks in acidified lakes ( 4 ) . Similarly in the Netherlands there has been a decline of the great crested gerbe in (the vicinity of) small acidic lakes (5). Ospreys dependent on acidified lakes produce fewer young ( 4 ) . Birds (e.g. dipper populations) that feed on semi-aquatic insects have been found negatively affected by acidification, due to reduced insect abundance in Scotland and Wales ( 4 ) . Another interesting second (and higher) order effect follows from changes in cover. There is suggestive evidence that in Canadian maple woods populations of Vireo olivaceous and Epidonax minimus have decreased due to loss of cover caused by atmogenic stress (5). Loss of needles is implicated in the decrease of large (2.5 mm) spiders in Swedish and Danish forests. This decrease may in turn negatively affect the goldcrest (Regulus regulus) that is dependent on the availability of such spiders ( 4 ) . Because of the negative effects of acidification on overall productivity of plant and animal life, increased success of a species is much less common as a first order effect than decreased viability. However there are cases of increased success and these may in turn have higher order effects. An example thereof may be found in Scots pine forests. Here soil acidification and N-deposition have led to a massive increase of wavy hairgrass (Deschampsia flexuosa) (6). This in turn benefits some aphid and caterpillar species ( 4 ) . However the success of wavy hairgrass is probably the reason for the decrease of red ants. Red ants need bare soils to build their mounts. The reduced presence of red ants in turn is probably the cause of population decline of the green woodpeckers in Dutch forests subject to high N-deposition (6). A further category of higher order effects based on the direct interdependence of species is associated with changed levels of trace-elements such as calcium, aluminium and heavy metals. In acidified soils one may note a major loss of cations from the upper strata, and this is reflected in the biota on such soils. In acidified waters one may find the mobilisation of aluminium and heavy metals and a decrease of Ca- and Mg-concentrations and this in turn influences species dependent on those lakes (7). Preliminary research in Scandinavia ( 4 ) suggests a decrease in diversity and abundance of snails in acidified areas. This is probably caused by reduced amounts of calcium in plants eaten by snails. Kingbirds in acidified Canadian wetlands and dippers breeding near streams with a low pH have been reported to lay eggs with thinner shells (8). Again the probable cause is lowered calcium-levels in food.

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Reduced numbers of snails carrying houses and/or reduced levels of calcium in insects have been identified as the probable cause of a reduced reproductive success of great tits in acidified Dutch woodlands. The birds concerned have been found to lay fewer eggs. Moreover an increasing percentage of their eggs was found to have thin and porous shells (6). There is suggestive evidence that increased levels of aluminium in insects may be linked to a decrease in the populations of some insect eating birds (9). Increased amounts of heavy metals accumulating in food chains have been correlated with acidification. In Canada increased levels of metals linked to acidification have been found in Ontario mink (Mustefa vison) and otter (Lutra Canadensis) (18). In acidified wetlands in Sweden mercury levels in juvenile Goldeneyes were found to be so high that effects on behaviour were not unlikely (11). Increased mercury levels in great northern divers breeding in acidified lakes may have been the reason for lowered reproductive success (12).

All changes in the relative abundance and viability of species caused by acidification may have reverberations in the foodwebs and ecosystems, to which these species belong. It is clear that this will have an impact on ecosystem composition. A matter arising in this context is whether acidification-induced change in ecosystems also influences the non-living environment. This pertains to among other things the ability of living nature to correct major man-made perturbations in the environment and possible effects of acidification on geochemical cycles in which living nature participates. The first matter will be dealt with in another paper at this conference. A s to the second matter, it may be noted that there is suggestive evidence that acidification may influence geochemical cycles. It is known that acidification does negatively affect lichens involved in nitrogenfixation (4) and this in turn will have an impact on the nitrogen cycle. Another impact will emerge later in this paper, where among other things the impact of the deposition of N-compounds on the release of N,O is discussed. In both cases first order effects on species are involved. However, there is no obvious reason why higher order effects on species may not lead to changes is the non-living environment. Hiaher order effects beyond the direct interdeDendence of sDecies So far higher order effects described were tied to the direct interdependence of species. However, there are also other ways in which acidification and associated atmogenic stresses may cause higher order effects. In such cases effects go beyond the direct interdependence between species. Research on such higher order effects is very limited. Nevertheless I would like to discuss two examples. A first example is schematically outlined in figure 1.

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Figure 1. Schematic representation of higher order effects of acidification and associated atmogenic stresses linked with increased release of N,O. soils subject to depositions of high amounts of N-compounds

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1 increased release of N,O 1 increased temperature forcing

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changes in rainfall

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increased temperature stress on species

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1 water related stresses on species This example includes both effects on species that go beyond direct interdependence between species and effects on the nonliving environment. There is suggestive evidence that high depositions of nitrogen containing compounds due to their impact on (de)nitrification may significantly increase the emission of N,O (nitrous oxide) (13). Nitrous oxide has an estimated atmospheric lifetime of 170 years. Its atmospheric concentration currently grows with a yearly rate of 0 , 2 - 0,3 per cent. N,O is a 'greenhouse gas'. Increasing atmospheric concentrations thereof will have a temperature forcing effect. N,O emissions between 1950 and 1990 contributed about 5 per cent to the overall temperature forcing effect by increased concentrations greenhouse gases over that period (13). To stabilize atmospheric concentrations, worldwide emissions of N,O should probably be reduced by 70 - 90 per cent (14). Acidification however probably has the opposite effect increased emissions of N,O. This will increase and may accelerate temperature forcing by greenhouse gases. There is a time delay between increased atmospheric greenhouse gas concentrations and actual temperature increases at the earth surface (15). However there is no serious doubt that (ceteris paribus) in due course such temperature increases will occur (16). These in turn may have a multitude of effects, including sea level rise, changes in precipitation patterns, and water- and temperature related stresses on natural species (16). Species that may be affected by increased temperatures do overlap species that are currently affected by high N-depositions. Thus for instance higher order effects of N-deposition may increase atmogenic stress on species such as trees (15). Moreover increases in temperature may also influence the microbial production of N,O, thereby giving rise to a cyclic relation.

132 A second example of higher order effects that go beyond the direct interrespondence between species is essentially speculative. Part of it consists of a hypothesis advanced by Sangfors (17) linking acidification with eutrophication. This example is schematically represented in figure 2.

Figure 2. Hypothetical cyclic higher order effects of acidification. acidification of soils

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increased leaching of cobalt

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increased cobalt levels in coastal waters

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incieased deposition of oxidized sulphur compounds

increased production of dimethylsulfide by algae

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increased primary production of algae, due to increased availability of vitamin B12

This hypothetical case of higher order effects builds on the increased leaching of cobalt from acidified soils (18). Increased leaching of cobalt is supposed to lead to increased levels of cobalt in coastal waters and in turn to increased availability to algae of the cobalt-containing vitamin B 12. This is hypothesized to induce increased primary production of algae. A s this is potentially associated with an increased generation of dimethylsulfide (la), it is hypothesized that following atmospheric oxidation (18) this may add to the deposition of oxidized sulfur compounds, which in turn may contribute to (further) acidification of soils. I have chosen this example because it exemplifies a cyclic relation in which acidification may induce further acidification. Such an acidification-cycle would by highly worrying indeed, and this provides a good reason for a serious test of the hypothesis advanced. Conclusion In the foregoing it has been demonstrated that acidification and the related phenomenon of high N-deposition do have real and potential higher effects. This demonstration has been based on limited research. Much more research is needed to obtain a more complete and less hypothetical picture of these higher order effects. In the field of effects research this is probably the major challenge of the nineteen nineties.

133 Des Granges, J.L.; in The Value of Birds, IBCP Technical Publication 6 (1987) 249-257; also in Mens en Vogel, (mei 19891, page 87-92 Eerden, J.M. van: IPO Annual Report, Wageningen, (1990), page 9 -17 Oelke, H.; Beitr. Naturk. Nieders., 42 (1989), page 109 128; Stastny K., V. Bejek; in Birds Census Work and Atlas Studies, Taylor, K. et al. eds; BTO Publications, page 243-253

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Toarstensson, P., L.E. Liljekind; Flora- och faunafBrkindringar i terrestra miljber, Rapport 3604, Naturvdrdsverker, Solna, Sweden, 1989 J. Graveland; Experientia 46 (1990), page 962-970 Schuurkes, R., P. Sturmans; Vogeljaar 1 (1987)' page 57-64 Graveland, J.; Experientia 46 (1990). page 462-970 Heij, G.J., T. Schneider; Acidification Research in the Netherlands, Elsevier, (1991) Omerod, S.J., K.R. Bull, C.P. Cummins, S.J. Tyler, J.A. Vickery; Environmental Pollution 55 (1988), page 10-121 Glooschenko V., P. Blancher, J. Herskowitz; Water, Air Soil Pollution 30 (19861, pages 353-367 Nyholm, N.E.I.; 371

Environmental Research 26 (1981), pages 363-

(10) Wren, C.D., P.M. Stokes, K.L. Fisher; Canadian Journal Zool, 64 (1986), pages 2854-2859 (11) Ericksson, M.O.G., L. Hendrikson, H.G. Oscarson; Arch. envir. contam. Toxic. 18 (1989), page 155-160 (12) Barr, J.F.; ( 1986 )

Canadian wildlife Service ,Occasional Paper 56,

(13) Robertson, K.; Ambio 20 (1991), page 151-155; Magaritz, M.; in International Environment Reporter 0149-8738, (1991), page 47 (14) Reijnders, L., C. Kroeze; in CLTM, Het Milieu, Kerkebosch, Zeist, (1990), page 285-304 (15) Krause, F.; Energy Policy in the Greenhouse, IPSEP, El Cerrito, (1989) (16) UNEP/WMO; Climate Change - The IPCC Scientific Assessment, Cambridge University Press, (1990) (17) Sangfors, 0.; Ambio 17, (1988). page 296 (18) Reijnders, L.; H,O 23, (1990), page 430-432

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135

ACIDIFYING EFFECTS CN GFCUNWATER

Jouko Soveri National Board of Waters and the Environment P.O. Box 436, SF-00101 Helsinki, Finland Abstract

In recent years, several reports concerning the utilization of acid groundwater, in water supply have emphasized corrosion and health aspects. Acid rain may dissolve harmful elements from soils and indirectly affect water supply distribution systems. Groundwater acidification occurs also in areas where lake acidification is reported. Acid groundwater is encountered with increasing frequence in Scandinavia, northwestern Europe and in the northeastern parts of the American continent, where the bedrock and the soil mainly consist of acidic crystalline rocks, such as granite and gneiss. In these catchment areas, the buffering capacity is generally low and the sensitivity to acidification rather high. The extent of groundwater acidification is still largely unknown. 1. INTRODUCTION

Acidification of groundwater and surface water has been considered to be one of the most serious environmental problems of the future. Waterway acidification is already a generally recognized phenomenon in almost every region where fossil fuels are used. Air pollutants have been shown to cause significant changes in the state of the environment over extensive areas in Scandinavia, Canada and North America. Future prospects also appear depressing unless we can decisively reduce sulphur and nitrogen emissions. We can assume that acidification will further increase in certain areas despite the reduction of emissions. Acidification has been shown to have affected the ion ratios of groundwater. Special concern is being expressed about the increasing mobilization of toxic metals (Pb, Cu, Cd and Al) from both soil and drinking water pipes, as well as their possible harmful effects on health. And in the future, the water supply will be obtained, to an ever-increasing extent, from groundwater sources. The greatest threat of groundwater acidification is in shallow groundwater aquifers. The water supply in rural areas is often derived from such sources especially in the Nordic countries and in Canada. Groundwater constitutes normally about 30-70% of the municipal mains water in Europe. In Denmark, however, the amount is almost loo%, but in Norway less than 10%.

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Signs of the acidification of groundwater have been reported by many authors, for example in Norway (l), Finland ( 2 1 , Germany (3), Sweden ( 4 ) , England ( 5 ) and in Canada (6). 2. AREAS SUSCEPTIBLE TO ACIDIFICATION The risk of groundwater acidification is strongly related to acidic and intermediate bedrock and soil type composed of acidic hard rocks, such as granites, gneisses, sandstones and quarzites. These rocks and soil types are rather resistant to weathering. Which means that the buffering capacity in such catchment areas is generally low and the sensitivity to acidification rather high. The susceptibility of groundwater to acidification is therefore closely associated with the acid neutralizing capacity (ANC) of the soil. In glaciated regions with thin soils, the location of sensitive groundwater can roughly be determined from bedrock geology maps e.g.,in Europe (Figure 1 and 2). As can be seen, nearly all of the aquifers in the Nordic countries are categorized as highly sensitive. Other regions of high sensitivity include northern Scotland, northwestern Spain, and parts of Central Europe. Northern and mountainous regions with thin soils and low weathering capability are more sensitive to groundwater acidification, whereas deep-soiled agricultural areas show the least sensitivity.

Figure 1. Acidic and intermediate acidic hard rocks in parts of Europe (7).

Figure 2. Qualitative indication of groundwater sensitivity in parts of Europe after Holmberg et.al. (8).

137

At the International Institute for Applied Systems Analysis (IIASA), a method for evaluating the sensitivity of European groundwater to acidification, using aggregation matrices was developed by Holmberg et. al., (8). Various factors important to groundwater acidification are compiled on a European grid: soil type, depth, and texture: aquifer size; mineral composition: and water available for recharge. The risk of groundwater acidification is evaluated by assessing to what extent physical and chemical soil and aquifer properties of a certain region will contribute to the neutralization of acid deposition. (Figure 3).

Figure 3. Aquifer sensitivity or risk and assessed indicators after Holmberg et al. (8). Groundwaters differ from surface waters as they have normally longer residence times in soil. These combine to make solid/solution reactions which are of major importance in determining groundwater acidity and alkalinity. The presence or absence of carbonates affects the susceptibility of an aquifer to acidification. The pH in carbonate-containing aquifers rarely falls below 6.5. The buffering in these aquifers is mostly due to the presence of solid calcium carbonate rather than the dissolved bicarbonate. Groundwater in regions with granitic bedrock, thin soils and a humid climate, may show comparatively high concentrations of hydrogen ions. In Nordic countries no decreasing trends in the pH-values have been observed, although changes in bicarbonate, calcium, magnesium and sulphate concentrations have been noticed. In general, the groundwaters in the Nordic countries have relatively low pH-values in the small aquifers of glaciofluvial and till deposits.The reason for this is partly the composition of the bedrock and the overburden, and the shallowness of the aquifers. The shallow groundwater is poorly protected by the overlying soil and its residence time in soil is short. The groundwater therefore reacts easily to changes in the environment.

138

3. SUBSOIL PROCESSES AND GROUNDWATER ACIDIFICATION

Groundwater acidification may be caused by both natural and manmade processes. Gradual acidification of the environment has been taking place for thousands of years as a result of natural processes. However, man's activities have clearly speeded up acidification during the past few decades. Groundwater quality is determined by the geohydrological conditions prevailing in the area where it is formed and hence the composition of the groundwater usually also reflects the mineral composition of the soil and bedrock in the area. Acid rainwater and meltwater effectively dissolve out those materials in the soil which slowly change the mass balance of the groundwater. 3.1 Impact of Acid Deposition on Groundwater

The consequences of acidification of soils and groundwater are threefold. Acidification is considered to begin when the base cations on the exchange sites of the soil particles are replaced by hydrogen ions. During the first stage small amounts of ions such as calcium, magnesium, potassium, sodium and sulphate are leached out into the groundwater (Figure 4 ) . At this time there are still large amounts of buffering material present and the pH remains stable or may even increase. A slight increase of this sort thus often indicates the onset of acidification.

Figure 4 . Sum of cations (broken line) and anions (full line) during the period 1975-1988 at a groundwater monitoring station in Finland (9). In the second stage the amount of available buffering material decreases and acidification increases. In the third stage the buffering capacity has disappeared completely (alkanity=O) and the pH has fallen below 5. Dissolution of aluminium increases sharply at this stage (Figure 5).

139

O

0

O

0

0 0

O

800-

oooo

0

-

0

0 0

400 -

0

-

0~

I

O Q

Figure 5. Aluminium concentration and pH in groundwater in Finland in area with high sulphate deposition (10). Areas northern winters. The most

-

affected by acidic precipitation are often located in latitudes where snow accumulates during long, cold Soil acidification may be caused by several processes. important ones are the following:

acid rain, in the form of wet and dry deposition, nutrient uptake by vegetation, oxidation of sulphur and nitrogen compounds in the soil, oxidation and hydrolysis of ferrous iron in the soil, soil respiration giving carbonic acid (11).

3.2 Soil Processes and Leaching

Many of the processes are reversible, hydrogen ions secreted to the soil solution when plants take up cations, are neutralized when the plant is decomposed. Likewise the reduction of sulfate and nitrate is a sink for hydrogen ions. However, in general, these processes tend to be more acidifying than neutralizing (11). Nature is able to protect itself from acidification because the mineralogy of the soil plays an important role in regulating the acidity balance. As the acidity of the soil increases, the amount of hydrogen ions in solution increases and the amount of base cations decreases. The length of time through which exchangeable ions can maintain a state of equilibrium depends on the buffering capacity of the soil. Base cations in the soil are transferred, as a result of ion exchange reactions, from exchange sites to the soil solution, and from there into the groundwater. The pH of the groundwater may even rise in the initial stages, and only start to fall when the alkalinity decreases. Leaching of metals increases as the rate of chemical weathering of minerals speeds

140

up. Reactions of heavy metals and aluminium in both the waterways and the soil are an important aspect of acidification because they frequently have a toxic effect on organisms. The capacity of soil to neutralize hydrogen ions can be estimated using the so-called acid neutralization capacity (ANC) method. The effects of natural and anthropogenic influences on acid-base chemistry can be described using the def inition of ANC (12): ANC

=

2Ca2+ + 2Mg2* + K+ + Na+ + NH,+

+2Fe + 2Mn + 2A1

-

2S0,'-

-

NO3- -C1- -F-

(1)

From this equation, it follows that reactions or processes that reduce the concentration of the cations listed without an equivalent reduction in the anions will reduce the ANC of water: processes that decrease the anions without a concurrent decrease in the cations will increase the ANC.The buffering capacity can also be determined as the difference between the basic and the strongly acidic components present in the soil. The reactions in the soil which involve the formation and consumption of hydrogen ions have an effect on the buffering capacity (13). The exchangeable base cations, mainly C a 2 + , play a deciding role in buffering acidification pushes, which are caused by the temporal discoupling of the ion cycle. The mobility of different weathering products in soil and in the soil water system may vary considerably. Calcium is mobile to some extent, and sodium is considerably more mobile than potassium, although both occur in almost equal amounts in the primary igneous material (14). The reactions between the soil and the soil water system depend above all on the mineral composition of the soil, the specific surface of soil particles and the infiltration rate. The dark minerals formed at high temperatures, such as olivine, pyroxene or hornblende are more easily weathered than for example K-feldspar, quartz or muscovite, formed at lower temperatures. One typical reaction in Precambrian bedrock and in Quaternary deposits is the disintegration and transformation of calcium silicate (15). CaAl,Si,O, =

+ 2H' +H,O

Al,Si,O,(OH),

+Ca3' (2)

If strong acids such as sulphuric or nitric acid prevail in meltwater, the weathering reactions rain water or in snow proceed as (16): H,SO, + Ca-silicate =

Ca2+ + SO,,- + CO, +H,O + A13*

(3)

Percolating water and dominating natural acids dissolve base cations from the surfaces of the soil minerals, and transport the ions to the groundwater. The weathering of cations and their exchange with hydrogen are the main reasons why the pH of the infiltration water and o f the groundwater is higher than that of the rainwater and meltwater. According to the results of lysimeter and percolation experiments, the pH of meltwater changes from 4.7 to 5.8 and to

141

6.2 (without organic topsoil) when passing through 60-110 cm of organic and mineral soil. The mean level of nitrogen compounds in percolation water samples was much lower than that in the meltwater in areas with organic subsurface matter. The loss of nitrogen was due mainly to its uptake by vegetation. There is a continuous turnover of inorganic compounds into organic compounds and vice versa. On the other hand, in areas without organic topsoil the nitrate concentrations were clearly much higher than in meltwater and in rainwater (17). Nitrogen leaching from natural forest soils is usually small. Inorganic nitrogen is mostly in the form of NO, - , because unlike NH, - , it is not adsorbed on soil particles. If NO, from soils results in a loss of NO,into aquatic systems and, if this is accompanied by H' or A 1 3 ' , acidification will occur (13). The sulphate ion is dominant in the acidification process. It is termed a mobile anion, which means that all the sulphate ions which reach the ground in one area will appear in the runoff water and groundwater in the course of time. When this occurs equivalent amounts of cations must be transported through the same areas. These are ions with the opposite charge, mainly hydrogen, aluminium, calcium and magnesium. Hydrogen and aluminium ions cause acidification of water. The major N compounds present in the atmosphere are nitrogen oxides (NO. ) and ammonia (NH,). Normally,most of the nitrogen compounds added to the soil via precipitation and dry deposition will be taken up by trees and plants. If more nitrogen compounds are deposited via precipitation than the vegetation can utilize then the surplus will seep through the topsoil and overburden into the groundwater. The nitrate ion will then have the same acidifying impact as sulphate. Nitrogen saturation of soils has already been observed as a growing problem in central parts of Europe. The significance of nitrate in the acidification of watercourses may be illustrated by the ratio of nitrate concentrations and the sum of the concentrations of sulphate and nitrate, expressed as ueq/l. This ratio expresses the contribution of the nitrate ion to the acidification of water. For example in Norway the ratio is relatively low at present, from approxmately 0 to 0.2 (18), while after assessments relatively large areas of southern Sweden will be nitrogen saturated within 10-25 years, unless the input of nitrogen is continued (19). There are also many signs in southern Finland of increasing nitrate and sulphate trends in natural groundwater areas, which are not affected by fertilization. In one undisturbed forested catchment in southern Finland, the same type of increase in nitrate has been noticed during the low flows in the years 1966-1988 (20). In the Figure 6, for example, we can see how the sulphur content has increased up till 1987 at the groundwater monitoring station in southern Finland. After this time, the concentration leveled off. This is due to sulphur emissions decreasing significantly in western Europe (more than 50% in Finland) during the last 10 years. On the contrary the nitrate content of the groundwater has dramatically increased towards the end of the 1980s and the beginning of the 1990s. At the same time, nitrogen emissions have also increased. Also groundwater acidity has been seen to be slowly increasing during this time period.

142

PH SO4-S

0

Figure 6. pH-values and increased sulphate and nitrate at a groundwater monitoring station in southern Finland during the years 1975-1991.

4.CONCLUSIONS 1.Acid groundwater occurs in the same geological environment, where acidified lakes appear and are often located in northern latitudes. 2.Acid deposition of nitrogen and sulphur compounds have an impact on the elements of groundwater and are the main cause of groundwater acidification. 3.Groundwater acidification may be caused by both natural and manmade processes. 4.Groundwater in regions with granitic bedrock, shallow soils and humid climate may show comparatively high concentrations of hydrogen ions. 5.The most important ions in buffering acid rain are calcium, magnesium, carbonate and bicarbonate. 6.Changes in bicarbonate, calcium, magnesium, aluminium, sulphate, nitrate concentrations and slowly increasing trends of groundwater acidity have been observed in Nordic countries. 7.Changes in the chemical properties of soils and groundwater and signs of acidification have been reported in Northern Europe, in Canada and in northeastern United States.

143

5.REFERENCES

10 11 12

13 14 15 16 17

18 19 20

A. Henriksen and L.A.Kirkhusmo. Acidification of groundwater in Norway. Nordic Hydrology, 1982. J. Soveri. Influence of meltwater on the amount and composition of groundwater in quaternary deposits in Finland. Publ.of the water research institute, 63. Helsinki 1985. P. Benecke. Process of acidification in soil and groundwater. Air Poll. Research Report 13, United Kingdom 1988. M. Aastrup and G. Person. Utbredning och tidstrender avseende surt grundvatten i Sverige. In: Konferens luft och milje) 84, 1984 Sweden. W.M. Edmunds and D.G. Kinniburg. J. Geol. SOC. London 143, 707. D.J. Bottomley, D. Craig and L.M. Johnston. Neutralization of acid runoff by groundwater discharge to streams in Canadian Precambrian Shield watersheds. J Hydro1 75:l-26,1984. U.V. Bremssen. Acidification Trends in Swedish Groundwaters. Review of time series 1950-1985, SNV Report 3547, 1989. M, Holmberg, J. Johnston and L. Mare. Mapping Groundwater Sensitivity to Acidification in Europe. Impact Models to Assess Regional Acidification, IIASA,1990. J Soveri and T. Ahlberg. Effects of Air Pollutants on Chemical Characteristics of Soil Water and Groundwater. Adidification in Finland. HAPRO 1990. J. Soveri Influence of Air Pollutants on Groundwater Acidification in the Porvoo area, southern Finland. Publ. of the Water and Environm. Institute,Finland, 8.( 1991). G. Jacks, G. Knutsson, L. Maxe and A. Fylkner. Effect of Acid Rain on Soil and Groundwater in Sweden. Pollutants in porous media, 1984. T.J. Sullivan, C.T. Driscoll, S.A. Gherini,R.K. Munson, R. Cook, D.F. Charles and C.P. Yatsko. Influence of aqueous aluminium and organic acids on measurements of acid neut ralizing capacity in surface waters. Nature, 1989. NAPAP, Acid Deposition: State of Science and Technology, Report 10, 1990: G. Matthes, G. The properties of groundwater.USA 1982. G. Jacks. Groundwater chemistry at depht in granites and gneisses. KTH, Stockholm,l978. C.F. Elder. Chemistry of precipitation: Its importance to eastern Canada.ASCE 1981. J. Soveri. Acid percolation and disintegration, transformation and mobilization of some substances in Finnish quaternary deposits. Groundwater Contamination, IAHS, Baltimore, 1989. B. Kvaven and T. Syversen.The Contribution of Nitrogen to Acidification. The National Environmental Monitoring Programme,Report 408/90, Norway. B. Aniansson. The situation surpasses our worst fears. Acid magazine, NEPB. Stockholm 1988. A. Lepiste) and P. Seuna. Hydrological Characteristics Af fecting the Runoff Water Acidity. Acidification in Finland, HAPRO 1990.

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T Schneider (Editor). Acidification Research Evaluation and Policy Applications 1992 Elsevier Science Publishers B V

145

MONITORING FOR THE FUTURE: INTEGRATED BIOGEOCHEMICAL CYCLESINREPRESENTA~CATCHMENTS

T.Paces Czech Geological Survey, Czechoslovakia 1. INTRODUCTION Chemical elements and compounds move through environmental "compartments" such as atmosphere, surface water, soil, groundwater, oceans, sediments, rocks and living matter. These fluxes take the form of a cycle. The cycling processes proceed with variable rates. The slowest cycle is the geological cycle, a medium rate is shown by the hydrological cycle and the fastest rate is found in the biological cycle. All together, these three cycles compose the biogeochemical cycle. Within this cycle a chemical composition in the environmental compartments is found that allows man to live on the earth. The economic and population growth after the Second World War has changed the natural fluxes and has introduced new man-made chemicals into the environment. Such changes have been monitored for a long time in the separate parts of the environment. Monitoring of pollutants in the atmosphere, surface water and groundwater a t international, national, regional and local level has demonstrated, that man is changing the chemical properties of air, water and soil to such an extend that it endangers the future development of mankind. With the predicted population and economic growth, the adverse changes in the chemical composition of the environment will become, next to nuclear war, the most dangerous factor in the human existence. In spite of recent improvements in environmental quality i n the richest industrial countries, the global environmental deterioration and the deterioration in the third wold countries and the communist and postcommunist countries does not show signs of change for the better. Or does it? Our data and predictive models are not always sufEcient to answer this question conclusively. More and more politicians and industrialists realise, that environmental deterioration is a major factor in their decisionmaking. For this purpose they need precise and objective information that is also understandable to the general public, To obtain such information and to gain reliable knowledge on causes and consequences of environmental deterioration is not a trivial task. It is a n expensive and long-term scientific exercise. The reason is th a t environmental deterioration is a multi media, long-term phenomenon with many, as yet unknown feedbacks which often have the character of a n environmental time bomb. Physical, chemical and biological changes in the environment can proceed for a long time unnoticed without obvious harmful effects to become later an irreversible mechanism for a series of environmental changes. The most obvious changes are the influence of acidic atmospheric deposition on terrestrial and aquatic ecosystems, depletion of stratospheric ozone, accumulation of greenhouse gases and release of man-made toxic

146

compounds from river sediments and from sediments in estuaries. Such complicated environmental changes cannot be fully understood when we monitor only concentrations of a single compound in a single environmental compartment. The principles of ecology describing the interrelationships between living communities and their environment teaches us that we have to monitor behaviour rather than parameters, or more precisely, we should monitor behaviour through the integrated monitoring of ecological parameters. 2. HISTORY OF INTEGRATED MONITORING A research and monitoring concept which comes nearest to the integrated monitoring of biogeochemical cycles in representative catchments was introduced by G.E.Likens and F.H.Bomann in 1963 when they started a biogeochemical monitoring system within small hydrological watersheds of Hubbard Brook Valley, New Hampshire, USA (Bomann and Likes, 1967 and Likens e t al., 1977). This type of research was inspired by the concepts of modern ecology as was stated by E.P. Odum (1953).This approach was used by other researchers too. Two most comprehensive studies which use integrated monitoring to explain the functioning of forest and lake ecosystems were conducted in Coweeta watersheds, North Carolina, US (Swank and Crossley, 1988) and in lake Garsjon, Sweden (Andersson and Olsson, 1985). The first description of a n integrated monitoring programme was presented by Barnes et al. (1986). In 1987, an International Workshop on Geochemistry and Monitoring in Representative Basins (GEOMON) in Czechoslovakia brought together many researchers whose methods of investigation were very close to the integrated monitoring concept (Moldan and Paces, 1987). In the same year, the executive body for the Convention on Long-range Transboundary Air Pollution under the UN Economic Commission for Europe initiated a n international Pilot Programme on Integrated Monitoring. The Programme started in 1989.Two international workshops on integrated monitoring were organised in Sweden in 1987 where general outlines of the Programme were decided and in Finland in 1988 where the field, laboratory and data handling procedures were agreed upon. The results of all these international activities were published in the Field and Laboratory Manual, in a Manual for Input to the E C MM Data Bank (Anonymous, 1989)and in the Annual Synoptic Report 1990 (Anonymous, 1990).

3. PFUNClPLES OF INTEGRATED MONITORING Integrated monitoring of the environment is a system of objective measurements which yield data on the state of a selected ecosystem and its temporal development. An important characteristic of integrated monitoring is that it yields information on temporal and spacial fluxes in natural systems. It is not enough to measure concentrations of chemical compounds in individual parts of ecosystem or a distribution of plant species. The goal of the integrated monitoring is to determine the fluxes of matter, their influences on the development of living organisms and the feedbacks between environmental properties and the activity of these organisms. The results of integrated monitoring describe a n ecological metabolism. As in a human body, the metabolism of a healthy ecosystem, maintains a steady state or homeostasis

147

which is reflected by the chemical composition of the individual parts of the system. Atmospheric deposition, weathering of rocks and fertilization represent the 'food" of the ecosystem. Runoff of water and gaseous emissions carry away 'waste". Integrated monitoring is a scientific method to detect not only an illness of the system by also the rules for "a healthy diet" by studying the ecological metabolism. Integrated monitoring of biogeochemical cycles should be carried out in a representative part of the countryside over a long enough period of time to yield reliable and comparable results. Two basic types of representative parts of the countryside are a pristine and a polluted hydrological catchment. The period of observation should not be less than seven years. This number comes from our experience, that such a period usually includes some wet years and some dry years and that the final result is close to a n average climate for a selected ecosystem. 4. BIOGEOCHEMICALCYCLES AND BUDGJiXS OF ELEMENTS

Global cycles of biologically essential elements, such as 0, C, S, N, P and Ca have been investigated by geochemists for the last twenty years (Garrels et al., 1975, Moldan, 1983). These studies showed that "industry man" became a geological factor. His activities have reached such a size that he influences global cycles of many elements, including metals which may be toxic to living organisms (Stumm, 1977,Bolin, 1981). The global cycles integrate the fluxes and reservoirs of chemical elements over the whole earth. The basic data which enable us to put together such a global cycle for a chemical element come from geochemical studies of geospheres and from inventories of natural and anthropogenic processes which produce or consume the element. In table 1, there is a n example of the natural and anthropogenic fluxes which together structure the global cycle of sulphur. In order that we learn more about continental, regional and local segments of the global cycles, we investigate local cycles by measuring the fluxes of elements in well defined smaller parts of nature. An example of local fluxes of sulphur measured and calculated for a less polluted forested catchment and for a catchment with an extensive dieback of forest is given in table 2. While the global cycles are expressed by fluxes which describe cycling in the whole globe (in terragrammes per year), the local cycles are usually expressed in specific fluxes per unit earth's surface (in kilograms per hectare per year). The specific fluxes can be compared for different types of ecosystem, climatic zone, morphological and geological structure, land use, state of biological cover etc. The cycles in a n ecosystem consists of inputs and outputs which link the system with its surrounding and of sources and sinks which represent the internal fluxes of matter. The sum of all the fluxes compose a budget or mass balance of the ecosystem. By comparing local budgets we can make a judgement about status and development of an environment under stress.

148

Table 1 Fluxes within the global biogeochemical cycle of sulphur, in Tg/yr., data by Ivanov, 1983

Nature of flux

Natural flux

Continental part of the cycle: Emissions to the atmosphere from fuel combustion and metal smelting Volcanic emissions Aeolian emissions Biogenic emissions Atmospheric transport of oceanic sulphate Deposition of large particles from atmosphere Washout from the atmosphere, surface uptake and dry deposition Transport to the atmosphere over the oceans Weathering of rocks River runoff to the oceans Subsurface runoff to the oceans River runoff to continental water bodies Marine abrasion of shores and glacier abrasion Pollution of rivers with fertilizers Emuents from the chemical industry Acid mine waters Oceanic part of the cycle: Volcanic emissions Biogenic emissions Marine spray to the atmosphere Washout, surface uptake and dry deposition Burial of reduced sulphur in sediments Burial of oxidized sulphur in sediments

Anthropogenic flux

113 14 20 17.5 20

12 25

47

34.5 114.1 104.1

66 104

9.2

35

6.8 28 28 1

14 23

140

258 111.4 27.8

149

Table 2 Budget of sulphur in two forested catchments with a different level of atmospheric pollution in Czechoslovakia, fluxes in kg.ha-1.yr-1, data by Paces (1985)

Q p e of region

Rural, forested

Industrial, dieback of forest

Wet precipitation Surface runoff by streams Biological output due to lumbering of timber Weathering of bedrock Aeolian deposition Deposition of dry SO2

10.5 9.0

Accumulation in soil and biota

13.0

0.61 1.2 3.3 7.6

19.6 96.0 0.3 3.8 0.2 88.6 15.9

5. SllES FOR INTEGRATED MONPrORING Integrated monitoring is carried out in hydrologically well defined drainage basins (catchments), o r plots. Whenever it is possible, natural catchments are used. The size of the catchment should be between 0.5 to 5 km2. The catchment must be representative of the surrounding countryside. Its geology should be simple with one type of bedrock. Special care should be taken to determine the subsurface flow. It is easier if the subsurface flow is negligible with respect to the surface runoff. Otherwise, its velocity and chemical composition has to be continuously monitored together with the surface runoff. The site should be protected from land use changes so that the monitoring can continue in unchanged setting for at least 7 hydrological years. When no hydrological catchment is available a hydrological plot may be selected. This is usually more costly because a n artificial drainage has to be build into the plot in a way that all discharges from the plot are measurable. Usually, integrated monitoring of forested and lake ecosystems is carried out in artificial plots. However, we have been monitoring two agricultural natural catchments for 15 years with accurate mass balances for the major chemical elements (Paces, 1985,1991). 6. METHODS USED FOR INTEGRATEDM0"ORING

The methods cover chemical and biological monitoring. It is not possible to monitor all inputs, outputs, sources and sinks in the catchments and plots. Some of the fluxes can be measured continuously, some fluxes are determined periodically and some fluxes have to be calculated or modelled.

Table 3 Chemical and biological parameters monitored in Swedish catchments; the figures refer t o the number of measurements made per year; ’Y5” indicates that analyses are made every fiRh year; parentheses indicate limited number of catchments where this observation is made (Barnes et al., 1986)

Precipitation Moss Spruce Mor Enrichment Other soil Soil Ground Running Reindeer/ Rabbits Starlings Fish needles horzion horizons water water water moose absorbance conductivity redox potential

I2

PH Na K Mg Ca A1

I2 I2 I2 I2 I2

HC03

I2

total C NH4-N NO2-N NO3-N organic N total N PO4-P other P total P

(6-8) 1 4 1-4 1/2

m

1/10 1/10 1/10 1/10 1/2 1/2

u2

m m m

1/10 1/10

1/10 1/10 1/10 1/10 1/10 1/10

(6-8) (6-8) (6-8) (6-8) (6-8) (6-8)

1-4 1-4 14 14 14 14 14

(6-8) 1 4 14 (6-8) 1-4

I2 1/10 1/10 1/10 1/10

1/10 1/10

Si

F

1224 1224

1224 1224 1224 1224 1224 1224 1224

14

02

SO43 totals

1224 12-24 1224 1224 12-24

1/10

I2

I2

1224 1224

I2

(6-8) 1-4 V2-3 Y2-3

V5

14

1224 1224

+ v,

Precipitation Moss Spruce Mor Enrichment Other soil Soil Ground Running Reindeer/ needles horzion horizons water water water moose

c1 As V Cr Mn Fe Ni

cu

Zn Cd Hg

w

COD

1/2-3 1/2-3 Y2-3

Y2-3 1/2-3

(12)

14

1/10

1/10

14 14

14 1/2-3 1/2-3 1/2-3 1/2-3 1/2-3 1/2-3

V2-3 Y2-3 Y2-3 1/2-3 Y2-3 1/2-3

1224

U5 1/10

(12) (12) (12) (12) (12)

(6-8)

1/2-3

12

Rabbits Starlings Fish

1/5 1/5 1/5 1/5 1/5 1/5

1/10 1/10 1/10 1/10

14 14 14 14 14

1224 1224 1224 1224 (1)

1

(1)

1

1224

DDT

(1)

FJCB

(1)

1 1

1

1

152

Table 4 Chemical and biological parameters monitored in the integrated monitoring programme of UN ECE (Anonymous, 1989)

Frequency Variables

x 3 6 5

Deposition and litter-fall: Bulk precipitation Precipitation chemistry Metal deposition (mosses) Throughfall (+ chemistry) Stemflow (+ chemistry) Litter-fall (chemistry)

x 52(365) I2 x 1 x I2 X I2 x l(4)

03

Total NOr and total N H 4 +

X

x 52 (365) x (hourly) x 52 (365)

Soil and groundwater chemistry: Soil water chem., B/C-hor x Soil water chem., A/B-hor Groundwater chem., springs x Groundwater chem., tubes x Surface water chemistry: Runoff Vertical lake gradients Runoff water level Bottom fauna, fresh water Soil variables: Nutrient chemistry of soil (0-10 cm) Nutrient chemistry of soil (below 10 cm) Heavy metals of soil (0-10 cm) Heavy metals of soil (below 10 cm) Soil physics Soil temperature

(yr-1)

(yr-1)

General meteorology

Air chemistry (EMEP): Gasses 602,NO,, m03)

Extended programme frequency

X

lJ5

X

52

X

12

X

l/2

I2 I2

l/2

x 1224 x 68 x (contin.) x 1

X

X

x (some metals)

1 v5 X X

x

52-l/2

X X

1

lJ5 l/5 365

153

Variables Biological variables: Epiphytic lichen vegetation Field layer vegetation Bush and tree layer veg. Canopy cover of trees Biomass of the tree layer Nutr. chemistry of needles Micro-nutrients of needles Enzyme monitoring (soil, leaves) Mycorrhiza + fine roots Decomposition Misc.biol.monitoring

Frequency (P-1) X

x X X X

X

Extended programme frequency (yrl)

m lt5 lt5 1 X

5

X

1 1 1 1

1 1 X X

X

Hydrological inputs (precipitation, throughfall) and outputs (runoff, infiltration) should be monitored continuously. Samples for chemical analysis are collected periodically. Electric conductivity and pH of water samples and concentrations of atmospheric gasses (S02,NO,, 03) are sometimes monitored continuously to determine episodical variations in general long-term trends. There are two examples of sets of monitored data, one by Barnes et al. (1986) applied in Sweden (table 3) and another introduced by the International Cooperative Programme on Integrated Monitoring within the Convention on Long-Range Transboundary air pollution of ECE, Anonymous (1989) (table 4). There a r e not too many catchments t h a t are monitored today i n full accordance to the table 4 because of insufficient funding which would be needed to guarantee a long-term monitoring programme. Results of partial monitoring, where only selected parameters are measured, indicate that complete integrated monitoring is necessary in order to understand local, regional and global biogeochemical cycles, in order to determine the effect that human activities on them.

7. R . E S U L ~ O F I " E G R A T E D M O ~ F U N G In general, integrated monitoring should yield information on any influence of a chemical substance on functions of an ecosystem. Today, most information comes from research onto acidification of surface water and soil. Acidification is caused by flux of protons. This flux is the end result of all acidobasic biogeochemical reactions in an ecosystem and the flux of chemicals which yield or consume hydrogen ions. An example of hydrogen - ion production and consumption in two monitored catchments in Czechoslovakia is given in table 5. The proton budgets in mmo1.m-2yr-1 indicate quantitatively how fast individual biogeochemical processes proceed. These rates can be

154

compared i n catchments (ecosystems) under different environmental stress. When integrated monitoring continues for a longer period, the changes in the proton budget indicate causes and consequences of environmental changes. Finally, when biological monitoring is conducted together with the chemical one, changes in proton budget and budgets of biological essential and toxic elements are the most reliable indicators of the stresses on the particular ecosystem. A comparison of the data for the healthy forest and forest damaged by acidification in table 2 and 5,indicate that the dieback of spruce (Picea abies) is related to acidification due to the anthropogenic input of SO2 and NO, . When the budget of nitrogen in the catchments is studied (table 6) however, it appears that the input of NO, and the consequent transformation of nitrogen compounds is not the cause of the acidification. The transformation of nitrogen compounds is a result of the dieback of trees and their inability to fix nitrogen. These two nitrogen budgets indicate that our knowledge of denitrification a ndor accumulation of nitrogen in soil is inadequate and th a t the more complete monitoring of the nitrogen budget is needed. The input of acidifying gases is going to decrease due to the international effort to reduce industrial emissions and car exhausts. Only a long term integrated monitoring system will give us a reliable database to study the possible improvements of representative ecosystems. Even monitoring of simple input - output budgets in forested and lake catchments indicate that the reduction in emissions of SO2 and NO, reduces environmental acidification considerably. However, without the integrated monitoring which includes biological as well as chemical monitoring it will be very difficult to evaluate the economy of such %on-market" environmental effects. Present results of integrated monitoring in Europe are rather incomplete in spite of the fact that 14 countries take part in the ECE programma (figure 1, Anonymous, 1990). An experience of several teams that conduct long-term biogeochemical monitoring is th at such programmes, i n order to be successful, require a single or only a few researchers who keep the programma going through their sheer enthusiasm. As stated by Swank and Crossley (1988)Th ese individuals are the fabric of and provide the continuity of any successful long-term field research program, but they seldom receive credit for their valuable contribution". It will be very useful not only to future generations of scientists but also for the general public if funding and political support is granted to such individuals who wish to carry out a n integrated monitoring system to describe quantitatively the evolution of biogeochemical systems under anthropogenic stresses in separate European countries. Integrated monitoring of pristine ecosystems is necessary for comparative studies. We have learned how valuable early chemical analyses are of rain, of runoff in rivers and of soil analyses which we can compare with present data (Hanamann, 1898 in Paces 1982,Pelisek J., 1984,Kazay, 1904 in Horvath, 1983, Malmstrom and Tamm, 1925 in Tamm and Hallbacken, 1988,Hofman - Bang, O., 1905 i n Lofirendahl, 1990,Schindler, 1988,Gorham, 1982). Such historical data enable us to evaluate environmental changes today. However, it is not enough for setting standards for sustainable development. We need to carry out integrated monitoring of biogeochemical cycles in representative catchments in order to understand anthropogenic effects on nature in future.

55 I

Table 5 Hydrogen-ion production and consumption by known geochemical, biological and anthropogenic processes, mmo1.m-2.yr-1 of H Type of region

Rural forest

37.3 Input of H30+ by wet precipitation -0.01 Output of H30+ by runoff Hydrolysis reactions involving aluminium 7.3 Oxidation by pyrite calculated from the rate of release of S from bedrock according to 3.9 stoichiometryFeS2 + 3.5 0 2 + H B = F e h + 2 SOs4 + 2H+ Precipitation of femc hydroxide calculated from input andoutputs of Fez+ and the amount of Fez+ released from oxidation of pyrite Fe2+ + 0.250 2 + 2.5 H20 = 3.7 Fe(OH)3 + SH+ Dissolution and dissociation of C02 according to stoichiometry COP + H2O = HC03-+ H+ 52.5 Biological fixation of cations and anions (except nitrogen 59.6 species) in biomass removed by harvesting Transformation of nitrogen species according to stoichiometry NH4+ + ROH = RNH2 + H f l + H+;NO3- + ROH + H+ = RNH2 + 202; N&+ + 2 0 2 = NO3-+ H20 + 2H+ where R is 12.7 organic matter Atmospheric deposition of dry SO2 and oxidation according to stoichimetry SO2 + H2O + 0.5 02 = SO$- + 2H+ 47.5 Release of cations and anions by hydrolysis of bedrock and depletion of exchangeable cations in soil calculated from budgets of Na, K, Mg, Ca, C1 and P -151

Z ("excessn production of protons)

74

Industrially damaged forest

45.7 -5.4 -13.0 11.8

11.6 0.0

100 554 -609

116

157

Table 6 Budget of nitrogen in two forested catchments with different levels of atmospheric pollution in Czechoslovakia; fluxes in kg.ha-1.yr-1 of nitrogen derived from data by Paces (1985)

Type of region

Rural, forested

Wet precipitation N-NO3

N-NH4 Surface runoff N-NO3 Biological output due to lumbering of timber Weathering of bedrock Deposition of dry N-NO, Denitrification or unknown accumulation

8. CONCLUSIONS

3.7 4.9 0.58 8.7

Industrial, dieback of forest 5.5 7.5

I2

0 5

2.5 0 15

4.32

13.5

Results of measurements of input - output budgets in small hydrological catchments yield useful data on the metabolism of ecosystems. Such measurements should form a long-term programme in order to determine significant trends in the changes of the biogeochemical cycling of elements. The simple input - output budgets, however, are not enough to evaluate the influence of environmental changes on the performance of living communities. Such a n evaluation can only be made through integrated chemical and biological monitoring. Historical data on the concentrations and fluxes of elements in environmental compartments are used today by scientists to evaluate the longterm trends in the evolution of the environment. These environmental trends have influences on economic and political decisions (The World Commission of Environment and Development, 1987). Such data, which are usually based on concentrations of major chemical elements only, are not sufficient however, for the evaluation of the state of ecological metabolism a t local, regional and continental level. Politicians and industrialists will need more complete data sets in future, when environmental conflicts a t all these levels may be even more severe than today. Results of the integrated monitoring of ecosystems carried out under the auspices of the Convention on Long-Range Transboundary Air Pollution within UN Economic Commission for Europe (Anonymous, 1990) indicate that support given to this programme is not adequate to establish a reliable network of monitoring sites. Many researchers today agree that a comprehensive

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integrated monitoring of biogeochemical cycles in small catchments will yield a set of data that is essential for the establishment of guidelines for a sustainable development. 9. REFERENCES

Andersson F. and Olsson B., 1985,Lake Gardsjon, An acid forest lake and its catchment, Ecological Bulletins 37,Publishing House of the Swedish Research Council, Stockholm, 336 p. Anonymous, 1989,Field and Laboratory Manual and Manual for Input to the ECA/IM Data Bank, Programme Centre EDC, National Board of Waters and Environment, Finland Anonymous, 1990, 1 Annual Synoptic Report 1990, Environmental Data Centre, National Board of Waters and Environment, Finland Barnes C., Giege B., Johansson K. and Larsson J.E., 1986, Design of integrated monitoring programme in Sweden, Environmental Monitoring and Assessment 6,113- 126,D, Reidel Publ.Corp. Bolin B., 1981, Changing global biogeochemistry, Report CM 52, Dept. Meteorology, University of Stockholm Bomann F.H. and Likens G.E., 1967,Nutrient cycling, Science 155,427-429 Garrels R.M., Mackenzie F.T. and Hunt C., 1975,Chemical cycles and global environment, W.Kaufmann, Los Angelos Gorham E., 1981, Scientific understanding of atmosphere - biosphere interactions: a historical overview, In: Atmosphere - Biosphere Interactions: towards a better assessment of the ecological consequences of fossil fuel combustion, Chapter 2, National Academy Press, Washington D.C. Hanamann J., 1988,Die chemische Beschaffenheit der fliessenden Gewasser Bohmens, II., Theil, Hydrochemie der Elbe, Archiv der naturwissenschaftlichen Landesdurchforschung von Bohmen, vol. 10, Kommissions Verlag von Fr.Rivnac, Prag. Horvath L., 1983,Trend of the nitrate and ammonium content of precipitation water in Hungary for the last 80 years, Tellus, 35B, 304 - 308 Hofman - Bang O., 1905, Studien uber schwedische Fluss und Quellwasser, Bull. Geol.Inst.Uppsala 6,101- 159 Ivanov M.V., 1983,Major fluxes of the global biogeochemical cycle of sulphur, Chapter 7 in: The Global Biogeochemical Sulphur Cycle, (M.V.Ivanov and J.R.Freney, eds.), SCOPE 19,449- 463,John Wiley & Sons, Chichester Kazay E., 1904,Chemical analysis of atmospheric precipitations, Idojaras 8, 301 - 306 (in Hungarian) Likens G.E., Bomann F.H., Pierce R.S., Eaton J.S.D. and Johnson N.M., 1977, Biogeochemistry of forested ecosystem, Springer-Verlag, New York, 147 p. Lofvendahl R., 1990,Changes in the flux of some major dissolved components in Swedish rivers during the present century, Ambio 19,210- 219 Malmstrom C. and Tamm O., 1926,The experimental forests of Kulbacksliden and Svartberget in north Sweden, Skogsforsoksanstaltens exkusionsledare XI, Stockholm Moldan B., 1983,Cycle of matter in nature (in Czech), Academia, Praha, 171 p. Moldan B. and Paces T. (eds.), 1987, GEOMON, Extended abstracts, International Workshop on Geochemistry and Monitoring in Representative Basins, Geological Survey, Prague, 253 p.

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159

Odum E.P., 1953,Fundamentals of Ecology, W.B.Saunders, Philadelphia Paces T., 1982,Natural and anthropogenic flux of major elements from central Europe, Ambio 11,206- 208 Paces T., 1985, Sources of acidification in Central Europe estimated from elemental budgets in small basins, Nature 315,31-36 Paces T., 1991, Changes in rates of weathering and erosion induced by acid emissions and agriculture in central Europe, In: Land Use Changes in Europe (Brower F.M., Thomas A.J. and Chadwick M.J., eds.), Chapter 15,317323,Kluwer Academic Publishers Dordrecht Pelisek J., 1984,Changes in the acidity of forest soils of the Orlicke Mts. caused by acid rains, Lesnictvi 30,955- 962,Praha (in Czech) Schindler D.W., 1988, Effects of acid rain on freshwater ecosystems, Science 239,149- 156 Stumm W., 1977,Global chemical cycles and their alteration by man, Dahlem Konferenzen, Abakon, Berlin Swank W.T. and Crossley D.A. (eds.), 1988,Forest Hydrology and Ecology a t Coweeta, Ecological Studies 66,Springer-Verlag, 469 p. Tamm C.O. and Hallbacken L., 1988, Changes in soil acidity in two forest areas with different acid deposition: 1920s to 19808,Ambio 17,56- 61 The World Commission on Environment and Development, 1987, Our Common Future, Oxford University Press, Oxford

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T. Schneider (Editor). Acidification Research. Evaluation and Policy Applications 1992 Elsevier Science Publishers S.V

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The critical loads concept for the control of acidification Jean-Paul Hettelingh, Robert J. Downing, and Peter A.M. de Smet Coordination Center for Effects, National Institute of Public Health and Environmental Protection, P.O. Box 1, 3720 BA Bilthoven, The Netherlands

Abstract "Critical loads" have been defined as the highest deposition of compounds that will not cause chemical changes leading to long-term harmful effects on ecosystem structure and function. European maps of critical loads of acidity, sulphur and nitrogen have been produced to assess the sensitivity of forest soils and surface waters. These maps can be compared with present and projected levels of pollutant emissions, allowing assessment of the effects of various emission abatement strategies. Parts of central and northwest Europe currently receive 20 times or more acidity than their critical loads, thus affecting the long-term sustainability of these ecosystems. It is shown that stringent emission reductions are needed to protect large parts of European forests and surface waters against acidification. 1. INTRODUCTION

The reduction of acidifying emissions of sulphur and nitrogen are the subject of international negotiations in the United Nations Economic Commission for Europe (UN ECE) under its Convention on Long-Range Transboundary Air Pollution (LRTAP). These negotiations have led to international agreements ("protocols") on reducing the emissions of sulphur [l]and nitrogen oxides [21. The sulphur protocol, which took effect in 1987,commits participating countries to reduce sulphur emissions by a t least 30 percent as soon as possible and a t the latest by 1993. A nitrogen protocol in effect since early 1991 requires that by 1994,annual national emissions of nitrogen oxides should not exceed 1987 levels. The protocol also requires that average annual emissions of NO, between 1987 and 1994 should not exceed the 1987 emission level. Both protocols are similar in that they do not explicitly include quantitative considerations about environmental effects. The abatement intentions are based predominantly on technical and economic considerations related to emission reductions. The renewal of these sulphur and nitrogen protocols, in 1993 and 1994, Note: The policy examples presented in this paper do not necessarily reflect the views of the National Institute for Public Health and Environmental Protection but have been introduced by the authors for illustrative purposes only.

162

respectively, will also include consideration of the effects of deposition reductions on European ecosystems. Critical loads provide a measure of the relative sensitivity of ecosystems on a large scale, and thus can serve as a means by which to assess the environmental effects of such deposition reductions. The general definition of a critical load is "a quantitative estimate of an exposure to one or more pollutants below which significant harmful effects on specified sensitive elements of the environment do not occur according to present knowledge." 131 (See also [41 for a general overview of critical load definitions.) In other words, a critical load is a maximum "no-effect''level of a pollutant. If an ecosystem has limited natural capacity to absorb or neutralize pollutants, then the critical load for that ecosystem would be low. Areas which are more able to withstand pollutant deposition have correspondingly higher critical loads. The pollutants considered in this paper are acidity and one of its major components, sulphur. "Harmful effects" are defined as chemical changes in forest soils and surface waters which may cause damage to an ecosystem, and a limited number of key parameters and values have been used for this mapping exercise. The use of critical loads as a measure for assessing the effectiveness and efficiency of emission abatement strategies constitutes the "critical load concept". However, this basic concept gives rise to different interpretations. The critical load concept from a policy point of view includes setting "target loads", which are desired levels of pollutant deposition which consider not only the sensitivity of a n ecosystem, but also other technical, social, economic and political considerations. In theory, these target loads should gradually be reduced to become equal to critical loads in order to prevent continued damage to sensitive ecosystems. The critical load concept, from an environmental point of view, consists of defining the critical loads and includes the estimation of the long-term effects of pollution levels which are higher than critical loads. These "exceedances" of critical loads may induce chemical changes which deplete an ecosystem's capacity to buffer acidity, leading to toxic concentrations of (e.g.1 aluminum in forest soils or to the excess of acceptable soil or water acidity. This process may take many years and can be demonstrated using dynamic models which describe chemical processes in soils and surface waters. Such models may also be used to describe ecosystem recovery once acidification is alleviated 151. This paper emphasizes the critical load concept from a policy point of view. The paper first describes the methods used to obtain European maps of critical loads of acidity, sulphur and nitrogen. Maps of critical loads of acidity and sulphur are then compared with abatement strategies for acidity ("maximum reductions"), and the achievement of nationally specified sulphur target loads. 2. CALCULATION AND MAPPING OF CRITICAL LOADS

The UN ECE Task Force on Mapping developed a Mapping Manual [6] in 1988 to provide guidance in national efforts to calculate critical loads. The RIVM Coordination Center for Effects in the Netherlands produced a Mapping Vademecum [71 and held workshops for national mapping experts to address more specific mapping issues as the work progressed.

163

The European map of critical loads of acidity was obtained by incorporation of data on critical loads for surface waters and forest soils of 14 European countries'. Critical loads for 11other countries were computed using European data on forest soils. Details on national and European critical load mapping can be found in [81, The Steady State Mass Balance (SSMB) method was predominantly used for the computation of critical loads of acidity. This method assumes a timeindependent steady state of chemical interactions involving an equilibrium between the soil solid phase and soil solution [9,101. Similar assumptions apply to in surface and groundwater chemistry. The SSMB method computes the maximum acid input to the system that will not cause excess of the critical alkalinity value. The latter value has been computed from average thresholds for chemical values: pH, aluminum concentration, and aluminum to calcium ratio [8]. A simpler, qualitative method (the "Level 0" approach) was used by the United Kingdom, Ireland, Czech and Slovak Federal Republic, and the Soviet Union, to calculate critical loads. This method uses existing geographical data bases on four ecological factors (bedrock lithology, soil type, land use and rainfall) to assess the sensitivity of ecosystems to acidic deposition to which critical load values may be applied [81. 2.1. Critical loads of acidity The primary equation used to compute the steady-state critical load of actual acidity in forest soils is as follows [8, 91:

CL'(Ac,,)

= BC, - Alkl(,t)

(1)

where: CL'(Ac,,,) = critical load of actual acidity for forest soils BCW = base cation weathering rate Alk,,,,, = critical value of alkalinity leaching All of the above terms are expressed in moles of charge per hectare per year (mol, ha.' yf'). The critical value of alkalinity leaching is defined from critical hydrogen leaching and critical aluminum leaching as follows: AlkI(m0 = - HI(,,, - All,,,,

(2)

where: All(,,)

= critical value of hydrogen leaching (mol, ha.' yr.') = critical value of aluminum leaching (mol, ha.' yr.')

The two exogenous variables in Equation 2 are defined as a function of net precipitation3, base cation deposition (as an approximation of the calcium concentration), base cation weathering, and base cation uptake as follows:

Hl(kt)

= Q * [HI,,

(3)

164

where:

[HI,,

= critical value of hydrogen concentration (= 0.09, approximately

Q

= net precipitation

equivalent to pH=4.0)

and

where: = critical value of aluminum concentration (= 0.2 mol, m-3) R(Al/Ca),, = critical value of the aluminum to calcium ratio (= 1.5 mol, mol;') = seasalt-corrected deposition of base cations (mol, ha" yr-') BC,' = base cation uptake (mol, ha.' yr.') BC,

[MI,,

In Equation 4 the minimum is taken from a function of aluminum concentration and a function of the aluminum to calcium ratio. The critical values assigned to the exogenous variables of Equation 4 are further elaborated in [8,91. Substitution of Equations 2 through 4 into Equation 1yields the following final expression used to compute the critical loads of acidity for forest soils in Europe: CL'(Ac,,,)

+ 0.29.Q ; 2.5.BC, + 0.09.Q

= min ( BC,

+ 1.5.BC,' - 1.5. BC,

1

(5)

The critical load of acidity for surface waters [9, 111 is computed by:

where: CLw(Ac,,) = critical load of actual acidity for surface waters = seasalt-corrected original base cation concentration (mol, ha-' yf') BC,' 2.2. Critical loads of sulphur

The computation of the critical loads of acidity using Equations 5 and 6 does not, however, provide European policymakers with s f i c i e n t tools for evaluating required emission reductions, because current UN ECE protocols are designed to control individual acidifying compounds; i.e., sulphur and nitrogen oxides, rather than on acidity as a whole. Therefore it was necessary to derive a "sulphur fraction" by apportioning the critical load of acidity between the acidifying share of sulphur and the acidifying share of nitrogen. This fraction was obtained using the following assumptions: (a) The share of present sulphur deposition in total acidic deposition is used as a surrogate of the portion of the critical load of actual acidity which can be attributed to sulphur. (b) The share of present nitrogen deposition in total acid deposition contributes to acidification only when it is not taken up or immobilized by the ecosystem.

165

In other words, nitrogen is assumed to be acidifying when the ecosystem is unable to use nitrogen as a nutrient. Assumptions (a) and (b)lead to the definition of the sulphur fraction: Sf =

PL(S0,) PL(S0,) + PL(N0,) + PL(NHJ - Nu - N,(&t)

(7)

where:

Sf PL(S0,) PL(N0,) PL(NH,) NU

= sulphur fraction (mol, ha.' yi') = present load of sulphur (mol, ha.' yf') = present load of nitrogen (mol, ha'' yf') = present load of ammonia and ammonium (mol, ha.' yr-') = nitrogen uptake of managed forests (mol, ha.' yr-') = critical value of nitrogen immobilization (mol, ha.' yi')

Equation 7 has been applied for areas in which the total present load of nitrogen (PL(N0,) + PL(NH,)) exceeds the nitrogen uptake capacity (Nu). The sulphur fraction is assumed to be equal to 1 when nitrogen uptake exceeds the total nitrogen load. In that particular case the critical load of acidity becomes equal to the critical load of sulphur. This is always assumed to be the case in watersheds. From Equations 5 and 7, the critical load of sulphur on forest soils is calculated by multiplying the critical load of acidity by the sulphur fraction: CL(S) = Sf*CL*(Ac,,)

(8)

As sulphur and nitrogen are complementary in the definition of acidity, a derivation of the critical load for nitrogen is obtained from Equation 8: CL(N) = Nu + (1-Sf).CL(Ac,,)

(9)

From Equation 9 it can be seen that the critical load of nitrogen is equal to the nitrogen uptake only when the critical load of sulphur is equal to the critical load of acidity (i.e., when S, = 1). In that case the critical load of nitrogen becomes consistent with its original definition as "the maximum deposition of nitrogen compounds that will not cause eutrophication or induce any type of nutrient imbalance in any part of the ecosystem or recipients to the ecosystem" E91. The relationship between the acidifying and eutrophying potential of nitrogen is not covered in this definition, although described in existing literature (see for example [12]). In fact, Equation 9 states that nitrogen will lead to acidification when a n ecosystem is subject to a supply of nutrients which exceeds its demand. The validity of the assumptions used to derive the sulphur fraction have not been verified in the field and should be made subject to further work. For example, the drawback of assumption (a) is that the relative contributions of sulphur and nitrogen compounds to total acidity in deposition is not constant over time, since these are dependent on national emissions of sulphur dioxide, nitrogen oxide and ammonia.

166

One of the drawbacks of assumption (b)is that it assumes nitrogen uptake to be constant over time. Some indications exist that the ability of vegetation to take up nutrients is dependent on the acid stress 1123. 2.3. Exceedances of critical loads Abatement strategies can be evaluated by comparison of the deposition to the critical loads. However, this comparison does not simply consist of subtraction of critical loads from deposition. The computation of exceedances needs to include chemical compounds in the air and in the soil which affect the level of acidity of deposition. These chemical compounds are (1) base cation deposition, nitrogen uptake and nitrogen immobilization which decrease the exceedance, and (2) base cation uptake which increases the exceedance. Using the definitions of critical loads from Equations 5 and 8, the inclusion of these variables lead to the following formulation of the exceedance of critical loads of acidity: CL(Ac),,

+ PL(N0,) Nu Ni(crit1

= PL(S0,)

+ PL(NHJ - CL(Ac,) + BC, - BC,' (10)

The difference between present load and critical load leads to the exceedance of the critical load of sulphur:

Equations 10 and 11 do not consider the effects of processes (such as forest filtering and throughfall) which tend to produce higher levels of acidic deposition in forested areas as compared to open land (lakes and agricultural areas). These factors have been excluded from the present exercise since available data on a European scale are very preliminary. Results of exceedance calculations which consider such effects can be found in 191. 2.4. Mapping critical loads and exceedances To allow direct comparison between maps of critical loads and present loads of pollutant deposition, the geographic resolution of the two maps must be similar. The Co-operative Programme for the Monitoring and Evaluation of the LongRange Transmission of Air Pollutants in Europe (EMEP), one of the first programs initiated under the LRTAP Convention of the UN ECE, operates a monitoring network over a European grid of approximately 150 x 150 km2 grid cells. Modeling work conducted by EMEP [13] to compute emissions and deposition of pollutants in each grid cell is used in negotiations on emission reduction by parties of the LRTAP Convention. Thus, critical loads have been mapped on the EMEP grid system to achieve consistency with the practice of deposition mapping. The size of an EMEP grid cell is such that it contains many ecosystems with a range of critical load values. In mapping critical loads for each EMEP grid cell, a decision is needed about which value best represents the range of ecosystems contained in it. A cumulative frequency distribution (CDF) of critical loads in each EMEP grid cell has therefore been calculated. A critical load value is calculated for each ecosystem in a grid cell, and a CDF is constructed indicating the

167

percentage of a particular grid cell which has a critical load lower than or equal to a particular value. For instance, if half of a grid cell area has a critical load of 1000 acid equivalents per hectare per year or lower, then the 50-percentile value of the grid cell is 1000 eq ha.' yf'. The critical load of the lowest (i.e., most sensitive) 1 percent of the grid cell area is called the "1-percentile critical load". A level of acidic deposition equal to the 1-percentile critical load of acidity protects 99 percent of the EMEP grid cell area. In this paper the 1-percentile critical load values (which reflect the most sensitive ecosystems) have been used to calculate the exceedance of acidity and sulphur by deposition patterns computed from national emissions. 3. ASSESSING EUROPEAN EMISSION ABATEMENT STRATEGIES

National emissions are linked to grid cell deposition using the Regional Acidification INformation and Simulation (RAINS) model [141 which uses the EMEP source-receptor relationship. RAINS enables the comparison of deposition patterns resulting from a variety of emission abatement options. In this paper, current levels of acidic deposition and sulphur deposition El31 are used as reference scenarios and compared, respectively, with two abatement strategies: Maximum feasible reductions of emissions of acid precursors: In this scenario it is assumed that removal efficiencies of 90 to 98 percent are achieved by applying flue gas desulphurization to large boilers in refineries, power plants and industry. It is also assumed that small boilers are supplied with low-sulphur fuels, and best available techniques are applied to reduce NO, emissions. Achieving target loads for sulphur: Table 1lists the countries which have formulated preliminary target loads for sulphur deposition. National emission reductions for all European countries needed to achieve these target loads a t minimum costs have been calculated using the RAINS model's optimization module. These optimized emissions are then used to calculate sulphur deposition patterns for Europe. While some countries have set specific target loads for individual EMEP grid cells, only the total national ranges are shown here. As mentioned above, these target loads include environmental, political, socioeconomic, and other national considerations. Austria, France, Switzerland, and the Soviet Union have set their target loads for sulphur to be equal to the 5-percentile critical load for sulphur. The resulting effects of these two strategies on critical load exceedances have been evaluated by analyzing deposition patterns in contrast to the references cases of current patterns of exceedances of acidity and sulphur, respectively. 4. RESULTS

Exceedance of the critical load of acidity: The exceedance of present loads of acidity over critical loads under current conditions has been calculated using Equation 10. Values for present loads of sulphur and nitrogen compounds are

168

Table 1 National target loads for sulphur deposition' Target load Country (range) Units' Remarks Austria

161 - 393 eq ha" y i ' 0.71 - 1.21 g m-' yr"

Preliminary target load for sulphur Preliminary target load for sulphur corrected for base cation balance

Denmark

0.5

g m-' yr-'

Preliminary target load for sulphur

Finland

0.2 - 0.5

g m-' yr"

Preliminary target load for sulphur

200 - 2000 eq ha.' yr.'

Unofficial preliminary target loads for sulphur

2400

eq ha" yf'

1400

eq ha" yr.'

For 2000. Target load for acidity of which a maximum of 1600 eq ha.' yr-' is attributed to nitrogen For 2010. Target load for acidity of which a maximum of 1000 eq ha.' yr-' is attributed t o nitrogen

Norway

0.5

g m-' yr"

Unofficial preliminary target loads for sulphur

Sweden

0.3 - 0.5

g m.2yr"

Unofficial preliminary target loads for sulphur

Switzerland 0.71 - 0.94

g m-' yr.'

Preliminary target loads for sulphur corrected for the base cation balance

France Netherlands3

United Kingdom

200 - 2300 eq ha.' yr.'

For 2005. Target loads are set equal to critical loads where they can be achieved by 2005

USSR

3.0 - 20.0 kg ha.' yr.'

Preliminary target loads of sulphur for forest and water ecosystems

1. Adapted from UN ECE, 1991. The critical load concept and the role of best available technologies and other approaches (EB.AIR/WG.5/R.24/Rev.l).Convention on Long Range Transboundary Air Pollution, Working Group on Abatement Strategies, Geneva. 2. eq ha" y i ' = acid equivalents per hectare per year;

g m.* yr" = grams per square meter per year. 3. The Netherlands has specified a target load for total acidity, with allowed maximum values for nitrogen, the remainder being sulphur.

169

based on 1990 EMEP emissions and deposition data C131. At current levels of acidic deposition in Europe, critical loads of acidity are exceeded in approximately three-quarters of the European area under consideration (Figure 1). Areas of the greatest exceedance (more than 2000 eq ha.' yf'), account for approximately 20 percent of the land area and occur primarily in central Europe. Under this scenario, all European countries have some areas in which critical loads are exceeded. Figure 2 shows the expected pattern of exceedances resulting from the implementation of "maximum feasible reductions" of emissions of sulphur and nitrogen. Under this scenario, the area in which deposition is estimated to be less than critical loads increases (from 25 percent in Figure 1) to 68 percent. Generally, the pattern of critical loads exceedance is similar to the current conditions depicted in Figure 1: the areas of highest exceedances occur in central Europe and the United Kingdom. However, the magnitude and scope of the areas receiving deposition higher than critical loads is reduced significantly. The maximum exceedance for a grid cell under this scenario is 2144 eq ha" yf', as compared with a maximum of over 10,000 eq in the current base scenario. European emissions of SO, and NO, are reduced by 81 and 57 percent, respectively, from the levels used to calculate the exceedances shown in Figure 1.

Exceedance of the critical load of sulphur: Figure 3 shows the current pattern of exceedances of the critical load of sulphur, based on the calculations of Equation 11. These exceedances are derived from subtracting critical loads of sulphur [9] from the present deposition of sulphur [13]. Approximately 25 percent of the area receives sulphur deposition less than critical loads. The area of maximum exceedance, covering roughly 10 percent of the area under consideration, is centered in north-central Europe (eastern Germany, Poland, and Czechoslovakia), but also includes large parts of the United Kingdom, the Netherlands and Belgium. Figure 4 shows the expected pattern of exceedances resulting from emission reductions which are optimized to achieve target loads for ten countries. In this sulphur target loads scenario, both the magnitude and geographic extent of critical loads exceedances are reduced significantly. European emissions of SO, are reduced by approximately 60 percent from the levels used to calculate exceedances in Figure 3. The area in which sulphur deposition is less than critical loads increases to 61 percent. The areas of greatest exceedance shift eastward and slightly southward, reaching a maximum of 2144 eq ha" yf' in eastern Yugoslavia. These shifts reflect the uneven geographic distribution of countries, primarily in north and northwest Europe, which have identified target loads for sulphur deposition. Additional countries which identify national target loads in the future would of course change this distribution. 5. CONCLUSIONS

The critical loads concept provides an environmentally based measure by which to optimize the environmental benefits of future emission reductions within

Figure 1. Exceedance of the critical load of acidity due to present emissions of acidic precursors. (1 percentile).

F'igure 2. Exceedance of the critical load of acidity after maximum feasible reductions of SO2 and NOx emissions in Europe (1 percentile).

Figure 3. Exceedance of the critical load of sulphur, due to present patterns of sulphur emissions (1 percentile).

Figure 4. Exceedance of the critical load of sulphur a h r reductions of SO2 emissions to achieve target loads in 10 European countries (1 percentile).

-

4

*

172

Europe. By classifying the relative sensitivities of different ecosystems and relating these values to patterns of pollutant deposition, areas which are most affected by current and projected levels of pollutant deposition can be identified. The scenarios presented here compare the current patterns of exceedances of critical loads of acidity and sulphur with the results of two different emission reduction scenarios: maximum feasible reductions (of acid precursor emissions), and achieving sulphur target loads. The results indicate that even under these stringent pollution control policies, the geographic area in which acidic deposition exceeds critical loads is reduced, but that large parts of central Europe and most of Scandinavia would still receive more acidic deposition than these ecosystems can safely absorb in the long term. Large parts of Scandinavia have critical loads which are close to zero; i.e., these ecosystems can absorb little or no additional acidity without harm. Since a "zero deposition"level could never be achieved practically, different methods to evaluate and protect these areas should be addressed. The efficacy of other options, such as liming, to ameliorate these ecosystems should be further investigated. A number of national considerations influence the calculation of critical loads, target loads, and exceedances. Important factors t o be considered in applying the critical load concept include: Choice of "most sensitive/valuable" receptor. For this first attempt to map critical loads on a European scale, the selection of receptors was limited to forest soils and surface waters. Critical loads for forest soils were mapped for all countries in mainland Europe, while Scandinavian countries mapped surface waters or a combination of the ecosystems. Definition of "harmful effects". The guidelines developed for calculating critical loads for forest soils included a limited number of critical chemical values to be used as cutoff points in defining a harmfbl effect. Selection of percentiles. Use of higher percentile maps (e.g., 5 or 10 percentile) leads to lower exceedances, but at the cost of greater risk to the most sensitive ecosystems in each grid cell. Some countries are now investigating in detail the ramifications of exceedances on individual ecosystems. Definition of target load. Many countries have set target loads based on critical load value; for example, a target load equal to the 5-percentile critical load for a grid cell. This assumes that (roughly) 5 percent of the area within the grid cell would remain above critical loads. Most nationally set target loads exceed critical loads. Current policies and negotiations do not specify for how long target loads will remain operational, which could lead to loss of ecosystems over the long term. While the current critical load mapping effort has been successful in preparing maps for Europe, the preliminary data will be revised in the future as new and better data becomes available. Future work will include the consideration of more chemical criteria and ecosystem receptors. In the present exercise, no biological effects have been considered, but only chemical balances in forest soil and surface water systems. In addition, there is a need to investigate synergies among different pollutants (e.g., acidity and heavy metal leaching). In spite of these uncertainties, the ability to compare and assess projected ecological effects of various emission reduction strategies is important. The

critical loads approach in designing international emission reduction agreements can be more effective than flat-rate "30%"protocols, since emission reductions can be targeted to achieve the most environmental benefit. In combination with various computer models of pollutant transport and deposition in Europe, the critical loads concept offers a method by which to analyze options for national and multilateral pollution control schemes. For example, it may be more effective (and economical) for some countries to provide economic and technical assistance to other countries to reduce their sources pollution. Different types of these "burden sharing" mechanisms are now being examined within the UN ECE framework. While no analysis has been made in this paper concerning the costs associated with such emission reductions, there is a large body of research underway within the UN ECE and elsewhere to assess the economic implications of such Europewide emission reductions. These economic aspects of emission reductions play a key role in the assessment of future European emission control strategies. Using critical loads as an environmental indicator of ecosystem sensitivity to acidifying deposition has gained much support in recent years. While elements of the scientific principles and assumptions and their implementation are undergoing continued refinement, the basic concepts of defining and mapping critical loads on a European scale are the result of a broad scientific consensus among the countries involved. The application of the critical loads concept as a tool for designing and implementing strategies to reduce pollutant emissions on an international scale has also gained widespread support. Critical loads offer a yardstick to measure the efficacy of multilateral pollution abatement strategies, and thus can serve as a usehl tool in international fora such as the UN Economic Commission for Europe's current negotiations. 6. ACKNOWLEDGEMENTS

This research was funded by the Air Directorate of the Ministry of Public Housing, Physical Planning and the Environment of the Netherlands. The work of the National Focal Centers which collaborated closely in the development of the methodologies, data collection and processing, and mapping of European critical loads is gratefully acknowledged. 7. NOTES

1. Austria, Bulgaria, Czech and Slovak Federal Republic, Denmark, Finland, France, Germany, Ireland, Netherlands, Norway, Sweden, Switzerland, Soviet Union, and United Kingdom. 2. Base cations considered include calcium, magnesium and potassium. 3. Net precipitation in soils is defined as precipitation minus evapotranspiration minus surface runoff. Surface runoff is not included for the computation of net precipitation in watersheds. In this paper net precipitation is denoted by Q for both soils and watersheds for notational simplification.

174

8. REFERENCES 1 United Nations Economic Commission for Europe. 1985. Protocol to the 1979

Convention on Long-Range Transboundary Air Pollution on the Reduction of Sulphur Emissions or their Transboundary Fluxes by at least 30 per cent. U.N. ECE, Geneva. 2 United Nations Economic Commission for Europe. 1988. Protocol to the 1979 Convention on Long-Range Transboundary Air Pollution concerning the Control of Emissions of Nitrogen Oxides or their Transboundary Fluxes. U.N. ECE, Geneva. 3 J. Nilsson and P. Grennfelt (eds.). 1988. Critical Loads for Sulphur and Nitrogen. Report from a workshop held a t Skokloster, Sweden, 19-24 March 1988. Nordic Council of Ministers, Milj~rapport1988:15, Copenhagen. 4 K.R. Bull. 1991. The Critical Loadfievels Approach to Gaseous Pollutant Emission Control. Enuiron. Pollut. 69:105-123. 5 J.-P. Hettelingh, R.H. Gardner, L. Hordijk. In press. A Statistical Approach to the Regional Use of Critical Loads. Environ. Pollut. 6 UN ECE. 1990. Draft Manual on Methodologies and Criteria for Mapping Critical Levelfioads and Geographic Area Where They Are Exceeded. Convention on Long-Range Transboundary Air Pollution, Task Force on Mapping, Geneva. 7 Hettelingh, J.-P. and W. de Vries. 1991. Mapping Vademecum. National Institute of Public Health and Environmental Protection, Bilthoven, The Netherlands. 8 J.-P. Hettelingh, R.J. Downing, P.A.M. de Smet (eds.). 1991. Mapping Critical Loads for Europe. National Institute of Public Health and Environmental Protection, Report 259101001, Bilthoven, The Netherlands. 9 H. Sverdrup, W. de Vries, and A. Henriksen. 1990. Mapping Critical Loads: A Guidance Manual to Criteria, Calculation, Data Collection and Mapping. Nordic Council of Ministers, Miljerapport 1990:14, Copenhagen. 10 W. de Vries. 1991. Methodologies for the Assessment and Mapping of Critical Loads and Impacts of Abatement Strategies on Forest Soils. Winand Staring Center Rep. 46. Wageningen, The Netherlands. 11 A. Henriksen, L. Lien, and T.S. Traaen. 1990. Critical Loads for Surface Waters: Chemical Criteria for Inputs of Strong Acids. Norwegian Institute for Water Research Rep. 0-89210, Oslo. 12 T. Brydges and P.W. Summers. 1989. The acidifying potential of atmospheric deposition in Canada. J. Water Air Soil Pollut. 43:249-263. 13 T. Iversen, N.E. Halvorsen, S. Mylona, and H. Sandnes. 1991. Calculated Budgets for airborne acidifying compounds in Europe, 1985,1988,1989,1990. Cooperative Programme for Monitoring and Evaluation of the Long-Range Transmission of Air Pollutants in Europe (EMEP), MSC-W Report 1/91. Norwegian Meteorological Institute, Oslo. 14 J. Alcamo, R. Shaw, and L. Hordijk (eds.). 1989. The RAINS Model of Acidification: Science and Strategies in Europe. Kluwer, Dordrecht.

SESSIONC ACIDIFICATION POLICY

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CANADIAN ACID RAIN POLICY S. Milbum-Hopwood and K.J. Puckett

Atmospheric Environment Service Environment Canada, 4905 Dufferin Street, Downsview, Ont. M3H 5T4, Canada

INTRODUCTION On March 13 of this year, the Prime Minister of Canada, Brian Mulroney and the President of the United States of America, George Bush, signed an Agreement on Air Quality [l]. This agreement enshrines Principle 21 of the 1972 Stockholm Declaration which states that countries are to ensure that activities within their jurisdiction do not cause damage to the environment of another country. This agreement also includes provisions for controlling acid rain. The Agreement on Air Quality followed 11 years of discussion between the two countries and is a significant milestone in the history of Canadian acid rain policy. This paper will begin by describing Canadian acid rain policy and its evolution. The paper will also outline the Canada-United States Air Quality Agreement and the effect of the acid rain provisions on deposition in Canada. Finally, it will consider the future work that must be undertaken to further resolve the acid rain problem. CANADIAN ACID RAIN POLICY Damage due to acid deposition in Canada is most severe in the eastern part of the country, in the provinces of Ontario, Quebec, New Brunswick and Nova Scotia. The extent of the acid sensitive land mass is shown in Figure 1. Approximately half of the 800,000 water bodies greater than .18 hectares in size, in the areas east of the Manitoba-Ontario border and south of James Bay are sensitive to the effects of acid deposition. It is estimated that more than 3 1 ,000lakes greater than 0.18 hectares in size and 14,000 lakes greater than 1 hectare in size are acidic [2]. The damage is mainly caused by sulphur dioxide emissions from smelters and fossil-fuelled power plants in eastern Canada and from power plants in the midwestem and northern United States (see Figure 2). Emissions of nitrogen oxides contribute to the acidity of precipitation; however, they are not currently a major cause of surface water acidification in eastern Canada. The Canadian sulphur dioxide control policy consists of two components: (1) an environmental objective or target loading to protect a specified component of the environment, which in this case is moderately sensitive aquatic systems; and (2) the emission reductions required to achieve the target loading. The target loading was established in the early 1980s, based on the limited information

178

Areas most sensitive to acid deposition

Figure 1. Areas of Canada most sensitive to acid deposition available at that time on aquatic effects [3]. The value chosen was 20 kilograms of sulphate in precipitation per hectare per year (kg/ha/yr) as a maximum deposition level and it was seen to be protective of moderately sensitive aquatic ecosystems. It was realized at the time, that very sensitive basins would not be protected by this loading and that further evaluation would be needed when more information was available. The Canadian federal and provincial environment ministers adopted the 20 kg/ha/yr of sulphate in precipitation target loading as the objective of their control program. To attain this target loading, it was agreed that a 50 percent reduction in emissions was required since the maximum deposition being observed was approximately 40 kg/ha/yr. It was also recognized that emission reductions in the U.S. would be required to achieve the target loading. In 1985, the Canadian sulphur dioxide control program was established, whereby the seven eastern provinces agreed to achieve, by 1994, a 50 percent reduction in annual sulphur dioxide emissions from the 1980 allowable base case value of 4,516 kilotonnes. As part of the 1985 agreement, each province was responsible for passing the necessary legislation or establishing the necessary programs to ensure that the emission commitments would be achieved. The specific reduction requirements for each province are shown in Table 1. The approach taken by the provinces was not to specify the type of control technology required but rather to allow industries and utilities to determine themselves how

179 to achieve specific control orders or objectives. This approach allowed the industries and utilities the flexibility to choose the most cost-effective approach to emission control. Table 1 Provincial SO, emission reduction commitments 1980 Base Case

Reduction

(tonnes)

(tonnes)

Newfcundland

738,000 2,194,000 1,085,000 215,000 6,000 219,000 59,000

188,000 1,309,000 485,000 30,000 1,000 15,000 14,000

mtdl

4,516,000

2,042,000

Province hnitoba

mtario Quebec New BrunsWick EX1

Nova w t i a

Still to be apportioned: 174,000 tonnes 1994 SO, emission objective for eastern Canada: 2,300,000 tonnes Although the 1985 emission reduction requirements focused specifically on eastern Canada, the distribution of emission sources in Canada was such that the 50 percent reduction in emissions in eastern Canada translated into a more than 30 percent reduction in 1980 national sulphur dioxide levels, permitting Canada to sign the Helsinki Protocol of the Economic Commission for Europe's Convention on Transboundary Pollution. The major emitters of sulphur dioxide in eastern Canada are six large copper, zinc and nickel smelters, one iron ore sintering plant and three provincially owned electrical utilities. Between 1987 and 1994 the major emitters will invest about $1,500 million (U.S. dollars) in capital projects to reduce their sulphur dioxide emissions. The average annual investment over the period is $220 million (US.dollars) per year but, during the final four years, the investment in capital projects will be higher at approximately $312 million (U.S. dollars) per year [4]. Some of the companies involved in the control program have indicated that they may be able to further reduce emissions after 1994. In 1990, the federal and provincial environment ministers agreed to expand the acid rain program to all of Canada and permanently cap sulphur dioxide emissions at 3.2 million tonnes by the year 2000 [5]. In implementing this program, the governments will consider the feasibility of using emission trading as a means of controlling emissions in both eastern and western Canada.

180

Figure 2. Eastern North America Sulphur Dioxide Emissions

CANADA-UNITED STATES AIR QUALITY AGREEMENT The main text of the "Agreement between the Government of Canada and the Government of the United States of America on Air Quality" provides a framework for addressing transboundary air pollution problems. Additional annexes dealing with other transboundary air quality problems, such as ground-level ozone and air toxics, will be added in the future. The first annex of the agreement specifies the targets and timetables for the reduction of acid rain causing emissions. Highlights of the annex follow:

181 -annual sulphur dioxide emission will be permanently capped to approximately 13.3 million tonnes in the United States and 3.2 million tonnes in Canada; -emissions of nitrogen oxides from power plants and factories will be reduced over the next ten years in both countries; -standards for new motor vehicles will be further tightened in both countries; -emissions of sulphur dioxide and nitrogen oxides will be closely monitored; -specific actions will be taken to protect pristine wilderness areas in both countries from transboundary air pollution. The second annex of the agreement deals with cooperative scientific activities between the two countries. These activities will assess the effectiveness of the acid rain controls described in the agreement and will provide information for addressing other transboundary air pollution problems. Canadian scientists have undertaken a preliminary analysis of the effect of the combined Canadian and U.S. control programs on sulphate loadings in eastern Canada. Atmospheric models were used to compare sulphate deposition for the period 1982-1986 with predicted sulphate deposition when the Canadian and U.S. control programs are fully implemented. The models are better at predicting relative changes in deposition under different emission conditions than quantifying the actual deposition levels. For this reason the percentage change between the 1980 wet deposition and the future deposition conditions, as predicted by the models, were applied to the observed mean excess sulphate deposition for the period 19821986, in order to prepare a map showing future deposition conditions (see Figures 3 and 4) [6]. The models predict that when the control programs in both the U.S. and Canada are fully implemented, the sulphate load in eastern Canada will drop below the 20 kg/ha/yr target over virtually the entire region.

FUTURE WORK The Canadian acid rain research and monitoring program will continue for an additional 6 years as announced by the Minister of the Environment in September 1991 [7]. The main goal of the program is to assess the effectiveness of the Canadian and US.sulphur dioxide control programs in reducing acid deposition below the target loading to protect human health, forests and very sensitive aquatic systems, and assess the need for further reductions. A significant fraction of the Canadian population resides in areas where some of the highest levels of acidic air pollution have been observed in eastern North America. Within this general population, there is clear evidence from other air pollution studies that there are subpopulations (i.e. children, asthmatics) which can be expected to be sensitive to elevated levels of air pollution, including acidic pollution. Evidence of increased hospital admissions and visits to emergency departments have been correlated with air pollution levels. In addition, transient but statistically significant decreases in the lung function of children have been observed during and after air pollution episodes. Currently there is no air quality objective or critical load for acidic sulphate aerosols. Although national objectives exist for the precursors of acid rain, SO, and NO,, there are insufficient data on dose-response relationships to properly determine the human health risk posed by acidic sulphate aerosols. A comprehensive health research and assessment program is underway to determine the magnitude of the effects, and necessity and extent of a

182

kglhalyr Figure 3. Observed 5 year (1982-86) mean excess sulphate deposition

Figure 4. Model projections of deposition under condition of full implementation of the SO2 control programs in eastern Canada and the United States of America

I83

mitigative program. Since the Canadian Acid Rain Policy came into being, new information gathered over the past few years has been analyzed to determine the "critical load" for aquatic ecosystems. The critical load is the highest deposition of acidifying compounds that will not cause chemical changes leading to long-term harmful effects on the overall structure or function of the aquatic ecosystem. Critical load information can be used along with information on economic and social concerns in the selection of target loads and the design of control programs. Recently, the criterion of pH26 needed to protect aquatic systems, has been used in aquatic models to predict critical load values for the different regions of eastern Canada. These values range from less than 8.0 in Atlantic Canada to more than 20 kg/ha/yr in some of the less sensitive regions of Ontario and Quebec (see Figure 5) [8].

I

Critical Load Values (kg/ha/yr of sulphate in precipitation)

12-16

Figure 5. Critical load values for eastern Canada

184 However, given the uncertainty in the predictions of the effect of the Canadian and U.S. sulphur dioxide controls on acid deposition levels and the critical load values themselves, modifications to the current control programs will not be made at this time. Nevertheless, specific projects will be carried out to monitor the rate and response of aquatic recovery in eastern Canada following sulphur dioxide controls and substantiate the critical load estimates for aquatic ecosystems. Changing air quality has been implicated in forest declines of maple and birch forests in eastern Canada. The complexity and variability in the forest ecosystem and the lack of information on the history of both air pollution and natural stresses makes it extremely difficult to establish cause-effect linkages. The current Canadian acid rain control program is based on the need to achieve a "target" acid loading to provide protection for moderately sensitive aquatic ecosystems. To date there has been insufficient information to develop target loadings to protect Canadian forests. Research will continue to determine the causes of the observed forest decline, to develop critical loads for forest ecosystems and to assess the need for further controls. In summary, the extended acid rain research and monitoring program is designed to assess the adequacy of the Canadian and U.S. acid rain control programs, to determine the rate and extent of aquatic recovery and to develop criteria for protecting forests and human health. The results of this program will be reviewed in the mid 1990s to determine the need for refinements to the control program.

REFERENCES Agreement between the Government of Canada and the Government of the United States of America on Air Quality. Federal/Provincial Research and Monitoring Coordinating Committee, The 1990 Canadian Long-Range Transport of Air Pollutants and Acid Deposition Assessment Report, Part 4, Aquatic Effects, 1990. Bangay G.E. and Riordan C., Impact Assessment-Work Group 1, Final Report, United States-Canada Memorandum of Intent on Transboundary Air Pollution, 1983. Federal/Provincial Research and Monitoring Coordinating Committee, The 1990 Canadian Long-Range Transport of Air Pollutants and Acid Deposition Assessment Report, Part 7, Socio-Economic Studies, 1990. Canadian Council of Ministers of the Environment, Information Release, November 29, 1990. FederallProvincial Research and Monitoring Coordinating Committee, The 1990 Canadian Long-Range Transport of Air Pollutants and Acid Deposition Assessment Report, Part 1, Executive Summary, 1990. Government of Canada, News Release, Green Plan provides $30 million to Acid Rain Controls, Sept. 23, 1991, PR-HQ-091-32 Federal/Provincial Research and Monitoring Coordinating Committee, The 1990 Canadian Long-Range Transport of Air Pollutants and Acid Deposition Assessment Report, Part 4, Aquatic Effects, 1990.

T Schneider (Editor) Acidtfication Research Evaluation and Policv Applicat ons

@ 1992 Elsevrer Sc<encePLblishers 8 V All rights resewed

185

Acidification policy in Finland E. Lumme

Ministry of the Environment, Air Pollution Control and Noise Abatement Division, P.O. Box 399, SF-00121 Helsinki, Finland

INTRODUCTION

Finland is a Nordic country. Nature in Finland is very sensitive to the effects of environmental pollutants. Our climate aggravates any natural stress on the ecosystems. Acidification is still the most serious regional air pollution control problem in Finland. A special national research project on acidification, HAPRO, was finished in 1990. This programme gave a lot of information to Finnish decision-makers. Some facts about the state of our water and forest ecosystems were quite obvious. A certain amount of damage has already been found in our water ecosystems: till date the harmful effects of lake acidity on fisheries are concentrated to the southern part of the country, but areas sensitive to acid deposition occur throughout the country. Changes in soil chemistry threaten the growth and health of our forests. Some negative effects on trees are already visible. Acidification may become an extremely serious threat to the forests of Finland especially in the future, if emissions are not curbed. The acidifying deposition exceeds the critical loads for forest and water ecosystems in nearly all parts of the country. A remarkable part of the deposition is long range transported. It has been calculated that e. g. in 1980 about 35 per cent of the sulphur deposition in Finland was coming from national sources and, correspondingly, in 1987 about 25 per cent. From the beginning of the 1970's, Finland has actively participated in international co-operation, because exclusively Finnish emission control will never be enough to guarantee a healthy future to the different ecosystems in Finland. During the 1980's the Council of State of Finland has given high priority to the reduction of emissions of acidifying compounds in our own country and to negotiations on the reduction of emissions in our neighbouring countries. Finland has supported the critical loads approach as an acceptable basis for further air pollution abatement strategies. Many kinds of emission reductions have been based on the use of the best available technology *

186

EMISSIONS AND EMISSION REDUCTIONS The major cause of acidification in Finland will for a long time continue to be sulphur emissions. In 1980 o u r sulphur emissions were 292 000 tonnes (S) per year, which was about 1 per cent of the total European emissions. Finland signed the minus 30 per cent sulphur reduction protocol in 1985, but two years later the Council of State set the reduction goal at 50 per cent by 1995. The Finnish sulphur emissions have already been reduced by about 55 per cent since 1980 (in 1990 they were about 0,5 per cent of European emissions). Emissions have been reduced by renewals of process technology and structural changes. In the near future fuel gas desulphurization will further reduce the emissions of sulphur dioxide. In 1987 the Council of State made seven decisions on sulphur emissions. They cover the sulphur content of oil products and hard coal, sulphur dioxide emissions from new and old coal fired power stations, sulphur emissions from oil refineries and sulphate cellulose and sulphuric acid plants. At the beginning of 1991 the Council of State made a decision in principle to reduce sulphur emissions by 80 per cent from the 1980 level by about 2000. In 1980 the Finnish emissions of nitrogen oxides were about 260 000 tonnes ( N O 2 ) per year (about 1,2 per cent of the total European emissions). The annual emissions have increased slightly during the 198O's, as they have done in many other parts of Europe. It has been estimated that in Finland, nitrogen oxides emissions can be reduced by 15 per cent by 1998 by applying, in addition to the exhaust regulations for new private cars, the following measures: development of combustion techniques in new and existing energy generation units and flue gas cleaning in new ones: a 50 per cent cut in exhaust emissions from new heavy diesel vehicles. Structural means will at least in theory make it possible to reduce the emissions to the target level (reduction by about 30 per cent) by the year 2000. Until now the Council of State has limited the emissions of vehicles and energy generation plants (the latest decision was made in March 1991). Ammonia emissions in Finland are about 43 000 tonnes (NH,) per year (about 0,5 per cent of the total European emissions). Nearly 70 per cent of the Finnish emissions are due to cattle breeding. It has been estimated that about 20 per cent of the acid deposition in Finland is caused by ammonia emissions. In Finland the acidyfing potential of ammonia emissions is at present approximately on the same level as that of nitrogen oxides. An ammonia working group has recently (report in March 1991) investigated technical ways and means of reducing ammonia emissions and the costs thereof. N o decisions have as yet been made on how to reduce ammonia emissions in the future.

187 CRITICAL LOADS

Finland has actively participated in the recent European work to produce the first maps of critical loads. Mapping critical loads has been and will continue to be an important area of Nordic co-operation under the Nordic Council of Ministers. According to the available maps Finland and the other Nordic countries belong to the most sensitive areas in Europe. In the whole of Finland the load of actual acidity or sulphur should not exceed 20 milliequivalents per square meter per year. The critical loads for Finnish lake and forest ecosystems have also been illustrated in smaller grids, corresponding to three by three subdivisions of the EMEP grids used in the European work. These more detailed maps give more information on the distribution of sensitive areas and on the differences in sensitivity of lakes and forests. In a number of grids throughout Finland the critical load of potential acidity for lake ecosystems ( 5 percentile) is less than 20 meq/m2, but for forest soils the critical load of potential acidity is 50 - 100 meq/m2 in large parts of the country. The European maps of critical loads of actual acidity show that critical loads are exceeded in all parts of Finland. The critical load of actual acidity ( 5 percentile) is exceeded by 100 - 200 meq/m2 in parts of southern Finland and by 20 - 100 meq/m2 in the central parts of the country. The critical sulphur load (5 percentile) is exceeded in quite large parts of southern Finland by 20 - 50 meq/m2 (even over 50 meq/m2 ) . According to the current European reduction plans the critical sulphur load would still be exceeded in the southern part of the country in about 2010. POLICY AND STRATEGIES

In 1984, when issuing air quality guidelines, the Finnish Council of State defined the long-term goals of air quality policy: to protect conifers in wide forest and agricultural areas or nature conservation areas: the annual sulphur dioxide levels outside towns and bigger villages should not be more than 25 ug/m3: and in order to avoid acidification effects the total sulphur deposition should be under 0,5 grams sulphur ( S ) / m 2 . For most of Finland this still seems to be a reasonable goal. On the basis of the national critical loads mapping programme the Ministry of the Environment identified, in May 1991, preliminary national target load values for sulphur. In the most sensitive regions in the arctic and subarctic areas of Finland, the sulphur load should not exceed 0,2 g S/m2/a. In some areas in the western part of Finland the sulphur load should not exceed 0,4. For the rest of the country the target load was set at 0,5, the earlier value for the whole country. When setting these target loads it was realized that for many aquatic ecosystems the critical sulphur load will still be exceeded. For forest ecosystems these values should ensure that the

critical load according to the present knowledge will not be exceeded. Studies on critical loads and levels and on the state of the environment will be continued. Although Finland has been able to cut certain emissions, there is still a lot to be done. In January 1991 the Council of State made the decision in principle that Finland will further reduce the acidifying sulphur emissions by 80 per cent (from 1980 level). Before the decision was made it was estimated that if the other countries also reduce their emissions by about 80 per cent the critical loads would not be exceeded. The Ministry of the Environment has set up a sulphur committee to prepare a ten year programme on how to achieve the 80 per cent reduction in sulphur emissions. The first proposals given by the committee in last September concern the increased use of heavy fuel oil with less sulphur. Finland will also study further possibilities to reduce emissions of nitrogen oxides and ammonia. The Finnish Carbon Dioxide Commission (report dated June 1991) has recently investigated alternative strategies and measures for limiting and reducing the emissions of carbon dioxide and other greenhouse gases. It concluded that it is possible to reduce greenhouse gas emissions by increasing the use of carbon dioxide free forms of energy and natural gas and by more efficient energy generation. It is obvious that the future policy in reducing greenhouse gases will also influence some emissions of acidifying compounds. These policies will be reconsidered in the future. The effects of an 80 per cent reduction in sulphur dioxide emissions, a 30 per cent reduction in emissions of nitrogen oxides and a 25 per cent reduction in emissions of carbon dioxide will be estimated in terms of their influence on Finland's energy policy. The parliamentary energy policy council has presented a report on Finland's national energy strategy at the beginning of October 1991. The most essential goals set to energy policy are to guarantee availability of energy sources and economical energy generation while taking environmental demands into account. Until recently, the general promotion of environmental protection in Finland has relied largely on legal controls. However, additional financial controls (like environmental taxes) have also been introduced during recent years. A working group on environmental taxes presented its report in April 1991. Finland tries to encourage saving and to reorient consumption in a direction more conducive to sustainable development by means of tax reforms. It is believed that if sufficiently high taxes and charges are levied on operations causing pollution, this will guide development in a less polluting and environmentally sound direction more effectively than would laws and statutes alone. Some financial controls were included in the state budget as early as in 1990. It is obvious that the 1993 budget will include more measures aimed at influencing e. g. the emissions of carbon dioxide and nitrogen oxides. The Ministry of the Environment has, in October 1991 set up a high level project to study and to make concrete proposals on the use of economic instruments in promoting environmental protection. Special attention will be given to the demands of international development. The work is leaded by the ministers of the environ-

ment, finance, agriculture and forestry, trade and industry, transport and communication. Finland will also continue to participate in international co-operation to negotiate efficient emission reductions for the whole of Europe. The co-operation between the Nordic countries will be continued. Moreover, Finland has bilateral co-operation with those neighbouring countries which contribute most to our acidifying load. The Finnish Government has drawn up a plan of action concerning co-operation with Eastern Europian countries in the near future. Co-operation in the field of environmental protection occupies a central position. In connection with the plan, e. g . an Environmental Review and Priority Action Programme for St. Petersburg, the St. Petersburg region, Karelia and Estonia has been made (report dated September 1991) to determine the main environmental problems and the main measures to reduce them. Significant air pollution projects are included in the project priority list. To achieve concrete results the Ministry of the Environment has set up a special East-Europe project in 1990. Solving the problems of transboundary pollution calls for deviation from the polluter pays principle, which is a commonly accepted means of promoting generally sustainable development.

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ACIDIFICATION POLICY GERMANY

-

191

CONTROL OF ACIDIFYING EMISSIONS IN

Bernd Scharer Umweltbundesamt, Bismarckplatz 1, 1000 Berlin

33,

Germany

Abstract Since the mid-eighties total annual acidifying emissions have started to decline in West Germany. There was considerable impact on this positive trend in air pollution by the control of SO2 and NO, emissions from large boilers, which were reduced by more than 8 0 % . Corresponding control programmes have been established for other groups of sources as well as other pollutants and - with unification - for East Germany. The driving force behind this development was and still is first of all the legal principle of anticipatory action or precaution which means in practical terms “emission minimization“. This cornerstone of German clean air legislation is the most powerful component of Germanyls “acidification policyt1,as it requires policy-makers to draw up new or review existing regulations for emission reduction based on requirements according to the state of the art and forces operators to apply the most modern ways and means of operation. This paper, as point of departure, describes the system used in Germany to deal with air pollution, with the emission minimization strategy in the center, and the actions against acidifying emissions based thereon. In addition, an outlook on what might be necessary to cope with the challenges of a sustainable development concerning acidification is given. 1. INTRODUCTION

Is there a German acidification policy? Acidification policy is understood to mean here the activities of the political decision-making bodies, the parliament and the government, which have as their aim the improvement of the initial acidification situation by the use of specific measures. This includes the generation of information on the initial situation (pollution) and its assessment, the fixing of goals as well as the adoption and implementation of measures which are suited to bringing about a change in the initial situation and contributing to achieving the goals.

192

German environmental policy evolved in the late sixties. At that time policy-makers were already aware of an acidification problem in Germany: Concern was directed at the llsmoke-induced damage near the emission source1', which one attempted to counteract more by a high-stacks policy than by reducing emissions. Largely constant pHs in precipitation in the seventies prevented the development of comprehensive reduction programs. An acidification policy developed in Germany mainly against the background of dying forests, receiving important impetus from that direction. The warnings and calls for help from Scandinavia, prompted by the acidification of otherwise nonpolluted lakes, observed there as early as the beginning of the seventies, had not yet managed to awaken a specific acidification policy in Germany. It was with reserve that Germany contributed to the political activities launched thereafter on the international level - initially within the OECD, later at the UN Economic Commission for Europe (ECE) in Geneva. And also the findings on problems caused by the acidification of German forest soils, which, as from the mid-seventies, were obtained as a by-product, so to speak, of research into the question of "what forests feed onll, still required some time to gain political relevance. Then, however, with rapid acceleration in the late seventies and early eighties, awareness of the dying of forests - in officialese: new forms of damage to forests - among the public, the media, and policy-makers became so high as to effect tremendous pressure for a change towards a solution to the acidification problem. In 1982/83, within a short time, the Federal Government - initiated systematic surveys of forest damage to be conducted at regular intervals (as from 1982); - convened the Interministerial Working Group "Forest Damage/Air Pollution1@(1983); - initiated a research programme to study damage to forests (1983)

- submitted the action programme IISave the Forests1' (1983); and last, but not least - took effective actions against the air pollutants considered

the cause of the problem, which in terms of the reduction of relevant emissions have met with considerable success. Also in its foreign policy, the Federal Government gave a clear indication of its active role in the solving of transboundary environmental problems by sponsoring a I1Multilateral Conference on the Causes and Prevention of Damage to Forests and Waters by Air Pollution in Europet1(Munich, June 1984). Given the successes achieved by policy-makers in pushing through emission reduction programmes and the reductions in acidification causing emissions already obtained except for NH3, on which attention did not focus until later an assessment of German acidification policy in the eighties would have

-

193

to be quite favourable. It should, however, be borne in mind in such considerations that a basic set of legal instruments to combat acidification processes and the dying of forests did not have to be conceived, but was already on hand, when the acidification problem became apparent. The basis for the measures was the strategy of anticipatory action which was incorporated into German air pollution control law (Federal Immission Control Act) as early as 1974 and which in practice is a requirement for all emission sources to minimize emissions. The emission minimization requirement was the main driving force of German air pollution control policy in that it provided the decisive justification for the further development of control technologies and the establishment of emission limitation requirements. It thus gave impetus to measures aimed at combatting acidification and damage to forests. Due to the fundamental importance of the emission minimization requirement, this basic conceptional element of the Federal Immision Control Act is explained in more detail in the following section before moving on to the description of various emission reduction programmes. 2. CONTROL OF ACIDIFYING EMISSIONS 2.1 The Technology-based Concept of Emission Control

In the beginning of the eighties, Germany systematically started to establish comprehensive cleanup programmes for existing emission sources. Up to that time, mainly new plants had been controlled and existing plants only in areas where ambient air quality standards were exceeded. The new regulations required both new and existing sources to control their emissions according to the state of the art, irrespective of whether air quality standards had been exceeded. The most important driving force behind air pollution control actions was and still is the legal principle of I'anticipatory action1' (Vorsorge) based on the Federal Immission Control Act (Bundes-Immissionsschutzgesetz - BImSchG). In practical terms, anticipatory action means minimizing emissions from all sources as far as technically reasonable. Air quality conservation (principle of protection), which is likewise founded on the BImSchG and aims at maintaining certain immission standards which may not be exceeded (ambient air quality standards AAQS), practically only plays a minor role as immission standards usually are not exceeded. Anticipatory action and conservation form a double-track strategy, that aims at the maintenance of a certain air quality and, independently thereof, at minimizing emissions. To implement the principle of anticipatory action, the Federal Immission Control Act exacts highest standards with respect to a) the development of emission reduction regulations, and b) the responsibility of the sources to be in compliance.

194

Emission standards and other requirements for emission reduction are principally based on technical feasibility, i. e. they have to reflect the state of the art. State of the art in this context means that emission standards must reflect the development of advanced processes, installations or operation practices, which have the practical potential of reducing emissions efficiently and to the greatest extent possible. This legal definition of the state of the art is intentionally directed at pushing technology forward, it thus illustrates that environmental policy tends to be technology-driven. The emission standards, referenced to the state of the art, must always be attained by the regulated party and facilities must be fitted with the appropriate technology, even if emissions do not result in Ambient Air Quality Standards being exceeded in the affected area or even if the measured air quality lies well below AAQS. The instruments designed to implement the state of the art at stationary sources in the Federal Republic of Germany are - construction and operating licenses for new or significantly altered plants, subsequent orders issued by the competent authorities for existing plants, and - ordinances containing provisions that operators have to comply with directly. Installations subject to licensing under the Federal Immission Control Act are all plants with a certain emission relevance. For example, all boilers with a thermal capacity of more than 1 MW require licenses €or both their construction and operation. Larger plants are required to monitor their emissions continuously. The results have to be made available to the responsible authorities for review. Small plants are normally not subject to licensing in accordance with the Federal Immission Control Act, they have to undergo plan approval. However, the requirement of "state of the art1*also applies to small plants. The ways and means of ensuring compliance of stationary sources are based on - ambitious obligations directly imposed on the operators, as well as - a differentiated implementation and enforcement infrastructure. The law imposes ambitious obligations on the operators, which are specified in Article 5 of the Federal Immission Control Act. According to these provisions, sources have be established and operated in such a way that the following four requirements are met: (a) attainment of ambient air quality standards (b) compliance with emission standards according to the state of the art, and regular monitoring (c) prevention, utilization, or at least harmless disposal of residual substances and (d) use of produced (waste) heat.

-

195

While (a) is to ensure that certain ambient air pollution thresholds are not exceeded, (b), (c) and (d) require plant operators to use progressive technologies while taking crossmedia aspects into account, and they thus promote the use of integrated technologies. With respect to automotive emissions, however, the BImSchG is overruled by relevant provisions agreed upon within the Commission of the European Communities; but also in this sector, national efforts are aiming at the introduction of state-of-the-art technology. The following section describes the measures taken to implement the state-of-the-art concept.

Action against Acidifying Emissions

2.2

stationary sources An all-encompassing system of emission standards exists in Germany, covering boilers, waste incineration plants, all relevant industrial processes and smaller area sources. Concerning stationary sources subject to licensing table 1 gives a summarizing overview of important emission standards for NO, and SO2 for new and old plants. These emission standards have been promulgated in the Ordinances on Large Combustion Plants ( 1 3 . BImSchV) and Waste Incineration ( 1 7 . BImSchV) and last but not least in the Technical Instructions on Air Quality Control (TA Luft), an administrative regulation binding on the authorities responsible for implementing legislation. 2.2.1

Table 1 Important emission standards for stationary sources subject to licensing in Germany w/m3 MWth Heat and power generation

> 300 > 50 1

-

-

-

300

50

-

300

100

< 10

Waste incineration Industrial processes

9

50

200 400 500 4001) 8003) 1000 2000

15D 3 00 450 so2 4001) 680 8503) 17004)

SO2

200

5 0 0 2 ) SO2

100 200 200 352) 352) 352) 352)

NO,

5002) NOx

solid fuel; 1: liquid fuel; g: gaseous fuel and 1 5 % max. emission rate for all plants; special regulations for special processes required in implementation and enforcement practice for plants < 5 MW use of light heating oil is required

X) s:

l) 2) 3) 4)

1

No,

> 300 > 100 10

sx)

196

At large combustion plants, the main source of SO2 and NO,, the practical implementation of the emission standards has been effected without any problems within a period of six years, as a result of which power plants operated by public utilities, for example, achieved emission reductions of 8 8 % for SO2 and 73% for NO,. The trends of these emissions, which follow that of the application of FGD and SCR technology in Germany, are shown in figure 1.

1.4

.......................................................................

......................................................................................................................... .........................................................................................

........................................................................................ ...............................................................................

a

Figure 1. Trend in

SO2

and NO, emissions from power plants.

For smaller plants (< 1 m t h ) not subject to licensing under the Federal Immission Control Act, the Ordinance on small combustion plants (1. BImSchV) lays down emission reduction requirements with respect to fuels and combustion technology. Another regulation that is particularly applicable here is the Ordinance on the sulphur content of light heating oil and diesel fuel (3. BImSchV) , which limits the sulphur content to 0.2%.

Ammonia emissions in Germany are roughly estimated at 1 million t NH3/a. 90% thereof originate from agriculture, of which two-thirds arise from the application of manure to farmland. An ordinance is currently being prepared, which prescribes requirements for workha the manure directly into the sail during application 85 well a s for stabie design. In addition, large stables have been included in the list of installations subject to licensing.

197

Mobile Sources As the main sources of NOx emissions, motor vehicles are an important target of acidification policy. In Germany in 1 9 8 3 , the Federal Cabinet adopted a policy goal, according to which emission standards were to be based on the three-way catalyst, which at that time had already been in use in the USA and Japan. It was then decided, however, to agree to the decisions adopted within the EC, in order to promote an EC-wide regulation and to prevent a split of the automobile market. To still be able to achieve emission reductions immediately, that is, before the EC emission standards went into effect, t a x privileges were granted in Germany for environmentally friendlier cars, both for new cars and retrofitted cars in traffic. Thus, with financial support by the government, some 0 . 5 million cars in traffic had been retrofitted with catalytic converters or primary measures and some 4 million environmentally friendlier new cars had been put into traffic before the corresponding emission standards became effective, these numbers accounting for approx. 15% of the vehicles in Germany. A s for heavy-duty vehicles, emission control requirements for NOx will not become effective until model year 1 9 9 2 / 9 3 , with a second stage effective as of 1995/96. Emissions will thereby be reduced by 4 0 % and 2 0 % , respectively, versus previous levels. 2.2.3 Cleanup of E a s t Germany A s a result of German unification, West Germany's environmental regulations are now also applicable in East Germany. This means that new plants there can only obtain a license if they comply with the current emission limits and that existing plants will have to be retrofitted within certain transitional periods, comparable to those applicable for West German installations, or be shut down. The S O z - and NOx-relevant restructuring and retrofitting of East Germany's industry is expected to be completed to a great extent in the second half of the nineties. An earlier date is envisaged for the conversion of small combustion plants to fuels low in sulphur. The renovation of power plants, industrial boilers and heating stations currently underway in eastern Germany comprises shut-downs, conversion to low-sulphur fuels and retrofitting with effective flue gas cleaning systems. Some lignitefired power plants and industrial plants have already been shut down. According to the power industry, the first 3 , 0 0 0 MWel will be retrofitted with flue gas desulphurization systems as early as 1 9 9 3 / 9 4 . 2.2.2

2.3 Efficient use of energy

The energy sector accounts for more than 9 0 % of acidifying and NOx emissions. Efficient, economical use of energy therefore is an essential component of acidification policy. A high standard of energy efficiency has already been attained in West Germany in comparison with many other industrial nations. Specific energy use by the West German economy as related to the gross domestic product dropped by more than 20% over the past 15 years. In addition to price SO2

198

increases for energy and advances in energy technology, a number of regulations and support measures to promote efficient use of energy, e. g. cogeneration of electricity and heat, use of renewable energy sources, provided important impetus in this direction. An ordinance on heat utilization is currently being prepared, which is designed to minimize the heat generated by installations subject to licensing and ensure that the heat, whose generation cannot be avoided, is utilized. Due to the necessity of climate protection, the intensity of energy-related measures has taken on a new dimension. The Federal Government, in a decision adopted by the Cabinet on 13 June 1990, pledged to reduce energy-related C02 emissions versus 1987 levels by 25% in western Germany, and even more in eastern Germany, by the year 2 0 0 5 . Preliminary estimates show that corresponding measures will result in a further reduction in emissions by approx. 20% for SO2 and approx. 7% for NO, versus the levels already reached in 1 9 8 9 . Trend in aaidifying pollution Acidifying emissions as a whole have been on a downward trend since the early eighties. However, considerable differences exist among the various pollutants and, above all, between the trends in eastern and western Germany. Drastic reductions of about 70% were achieved for SO2 in western Germany in the eighties, with an 88% reduction in the emissions from power plants (see figure 1). The reduction of NO, emissions was likewise considerable in this sector, amounting to 74%. Due to an increase in traffic emissions, however, total NOX emissions decreased only slightly. The emissions of N H ~ ,the third major acidifying component, are roughly estimated at 1 million t/a, of which 90% originate from agriculture. According to the latest emission projections of the Federal Environmental Agency as presented in table 2 , the future will bring further emission reductions. For example, the application of West German air pollution control legislation in eastern Germany will effect a 95% reduction in SO2 emissions there. The measures taken to protect the climate ( 2 5 % reduction in C 0 2 emissions by the year 2 0 0 5 ) as well as the increased use of catalyst-equipped vehicles will also bring about reductions in emissions. For the united Germany SO2 emissions are projected to decrease by 88% and NO, emissions by 50% comparing 1989 with 2005 figures. 2.4

199

5 1

Table 2 Trend in acidifying emissions in Germany Year

1980

4300

500

1000

-

1000

2700

1989 5250

Projection W without W with no data

2005

E without E with

no data

W: West Germany; E: East Germany; without: without climate protection measures: with: with climate protection measures The decrease in SO2 emissions has had an effect on pollution levels. Since 1988 , the measurements taken by the Federal Environmental Agency's air pollution monitoring network have shown a marked decrease in the concentration of sulphur dioxide versus the levels ascertained in the years prior to that time. The reductions range from just below 30% in the eastern part of the pre-unification Federal Republic of Germany to 70% in the west, averaging about 5 0 % . This development, however, benefitted from relatively mild winters in the course of which stagnant weather conditions and winds from the east rarely occurred. The decrease in SO2 concentrations is even more pronounced if the times of transport from emitting regions further to the east are excluded from consideration. The drastic decline in economic activity, which occurred in eastern Germany after the Itchangett, is now also becoming apparent in the available air pollution measurement data. Regrettably, for NO, there is no reduction in pollution levels, since the successes achieved in reducing emissions at combustion plants have so far been cancelled out by additional NOx pollution from traffic. The situation will only improve over a longer term, as a result of the emission reductions that are expected for the future. An improvement is also expected for NH3, however it has not yet been quantified. Acid deposition, too, has decreased in Germany in recent years. To deal with this in detail would, however, go beyond the scope of this paper, as the situation is complex and therefore differs greatly, especially because of the great influence of weather conditions.

200

Whether these successes in reducing emissions in Germany are sufficient to ensure effective protection of the environment and restore to health those components of nature that have been damaged, increasingly depends on the extent to which pollutant exports to Germany can be reduced. Recently, in scientifically determining and justifying effect thresholds below which adverse effects are not expected to occur, it was found that the llcritical loads" cannot be complied with in large parts of Germany. To be able to do so, much further-reaching emission reduction measures would be required. These would have to be taken above all in countries exporting acidifying pollutants to Germany. 3. FUTURE CHALLENGES

SUSTAINABLE DEVELOPMENT BY CRITICAL LOADS AND LEVELS?

Critical loads and levels are scientifically founded effect thresholds below which environmental damage is not expected to occur according to present knowledge. The critical loads/critical levels concept would thus meet the demands of a sustainable development and would be in accordance with the principle of protection (hazard prevention) which is incorporated in the Federal Immission Control Act. Compliance with the critical loads as specified in the report "Mapping Critical Loads for Europelll) will not be possible in Germany in a timeframe used in relevant projections. This holds true even if critical loads were applied which would allow the respective area or reference values of an EMEP grid to be exceeded by 10% and even if all countries "exportingn SO2 to Germany were to take measures according to the state of the art, although the latter would of course result in a marked reduction of pollution levels. Scenario calculations have been performed using RAINS (IIASA). In this work, the respective 99%, 95% and 90% critical load values were compared with the deposition levels obtained - when the ECE countries! current reduction plans are implemented ; - if measures according to the state of the art (maximum feasible reduction) were performed throughout the ECE. A comparison with the deposition levels resulting from the "current reduction plansr1 scenario shows that even the 90% critical load values are drastically exceeded in all EMEP grids. A comparison with the deposition levels of the !!maximum feasible reductionll scenario still shows exceedances of the 99% critical load values for large parts of Germany, whereas the 95% values are exceeded only in the EMEP grids along the border to the Netherlands, the North Sea coast and along the border to

l)

United Nations Economic Commission for Europe, Convention on Long-Range Transboundary Air Pollution CCE Technical Report No. 1, July 1991

201

the CSFR in the EMEP grids Chemnitz, where the deposition level is greatly in excess of the critical load, and Fichtelgebirge (mountain range near Bayreuth, Bavaria) with only slight exceedance. For all other German EMEP grids, the application of state-of-the-art technology would result in compliance with the 95% critical load values. Based on the RAINS calculations the Federal Environmental Agency has proposed that the 9 5 % critical load values be used as target loads for Germany. These are attainable (except for a few grids), if measures based on the state of the art are implemented in the countries emitting to Germany. For all other grids, the deposition levels attainable by using measures according to the state of the art are proposed as target loads. The Federal Environmental Agency considers these targets to be realistic. They are not only necessary, but also feasible, and do not amount to asking too much of neighbouring countries. As for Germany itself: As described above, the philosophy of German air pollution control law is to actually accomplish feasible emission reductions (emission minimization requirement). In conclusion, mention should be made of the planning currently underway to clean up an area known as the "black triangle1I, which comprises parts of northern Bohemia, Silesia and Saxony. This region has the highest emissions of acidifying pollutants in Europe. Within the framework of the *@Working Group for Neighbourly Co-operation on Environmental Issuesl@, the Environment Ministers of the CSFR, Poland and Germany initiated an analysis of the current situation and the development of a Ifrecovery and development plan". This hopefully is the beginning of the solution and thus the beginning of the end to the acidification problem.

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T. Schneider (Editor), Acidification Research. Evaluationand Policy Applications

203

@ 1992 Elsevier Science Publishers E.V. All rights raseNed

A d i c i f i c a t i o n p o l i c y i n Hungary

E . KovAcs I n s t i t u t e f o r E n v i r o n m e n t a l P r o t e c t i o n , Aga u . 4 . B u d a p e s t , H-11U Hungary

Abstract

A N a t i o n a l Conference on t h e A c i d i f i c a t i o n o f t h e Environment gave t h e framework o f t h e l a t e s t r e v i e w .

A i r p o l l u t i o n abatement

i s a m a j o r e l e m e n t i n a c i d i f i c a t i o n p o l i c y . A r e m a r k a b l e reduction of

SO2 a n d N O x e m i s s o n s h a s b e e n a c h i e v e d i n t h e l a s t d e c a d e .

I n s p i t e o f t h i s f a c t s o i l s h a v i n g a pHKC1C5,5 i n c r e a s e d t h e i r s h a r e b y 6 % .New l e g s l a t i o n o n a i r p o l l u t i o n a b a t e m e n t comprises i m p o r t a n t e l e m e n t s t o c o n t r o l t h e e m i s s i o n o f a c i d i f y i n g compounds.

1. INTROOUCTION I n 1985 t h e N a t i o n a l A u t h o r i t y f o r E n v i r o n m e n t a l P r o t e c t i o n and N a t u r e C o n s e r v a t i o n a n d t h e H u n g a r i a n Academy o f S c i e n c e s i n i t i a t e d

a r e v i e w c o v e r i n g t h e t o p i c s o f a c i d i f i c a t i o n . E x p e r t teams p a r t i c i p a t e d i n t h e work a i r quality, waters,

c o v e r i n g emission o f p o l l u t a n t s , ambient

d e p o s i t i o n o f a c i d i f y i n g compounds,

s o i l and f o r e s t ,

h e a l t h and on s t r u c t u r e s ,

acidification in

t h e i m p a c t o n t h e e c o s y s t e m , o n human r e s p . h i s t o r i c a l monuments.

The r e p o r t

One o f t h e f i n a l f i n d i n g s o f t h e r e p o r t was i s s u e d i n 1 9 8 7 [I]. was t h a t f u r t h e r c o - o r d i n a t e d r e s e a r c h s h o u l d b e c a r r i e d o u t t o d e s c r i b e t h e v a r i o u s phenomena, t o r e v e a l t h e p o s s i b l e c a u s e s g i v i n g a f i r m background t o t h e measures t o be t a k e n .

Primary i n t e r e s t

s h o u l d b e p a i d a n d f i n a n c i a l means s h o u l d b e a l l o c a t e d t o i m p l e m e n t a t i o n o f c o n t r o l measures.

the

204

In November 1990 a National Conference on the Acidification of the Environment gave the framework of the latest review on that topic. Some 6 0 papers were presented covering the fields ofresearch as follows: modelling; emission of pollutants; ambient air quality; dry and wet deposition of acidifying compoundsat different energy scenarios; emission control of S O x and NOx; correlation between emission, depositions and acidification; aquatic ecosystems;data and the causes of acidification of soils; puffer capacity of soils; mobility of toxic ions in soils; chemical treatment of soils; the impact of fertilizers; forest damages due t o air pollution; impact of acid rain on structures and historical monuments; impact of salting on roads and structures [ 2 ] . It seems to be difficult to give a comprehensive review of this broad field of reserach. However, some results will be presented here.

2 . E M I S S I O N S , CONCENTRATIONS AND D E P O S I T I O N S OF A C I D I F Y I N G CWOUNDS

Acidification policy is overlapping with air pollution abatement, however, it covers a wider range of topics. Nevertheless, the reduction of the emission of acidifying compounds is of primary interest. The main targets guiding the national policy for abatement of air pollution comprise the elements as follows

-

t o reduce air pollution in the heavily polluted regions of the country, as in the capital and in some industrialized areas, with an emphasis on improving ambient air quality in the big cities, - to maintain the ambient air quality of the relatively "clean" regions, - to fulfill the obligations due to international agreements, a s of the protocols on SO2 and NOx. The first target is clearly human health oriented because there exists evidence that serious damages to human health occurs in regions where people suffer expositon to excessive air pollution. Even in terms of financies, damages to human health expressed in

205 e x p e n s e s l e a d s t h e l i s t b e f o r e damages o f s t r u c t u r e s ,

s o i l s and

f o r e s t s due t o a c i d i f i c a t i o n . The i m p r o v e m e n t o f u r b a n a i r q u a l i t y r e q u i r e s t h e r e d u c t i o n o f p o l l u t i o n generated mainly by road t r a f f i c , hydrocarbons,

carbon monoxide, s o o t ,

such as n i t r o g e n o x i d e s ,

l e a d and s u l p h u r d i o x i d e .

The s e c o n d a n d t h i r d t a r g e t s a r e c l o s e l y r e l a t e d t o r e d u c t i o n o f p o l l u t a n t s c o n t r i b u t i n g t o a c i d i f i c a t i o n . However,

interrelation

between t h e s e elements c l e a r l y e x i s t s . E m i s s i o n s o f s u l p h u r d i o x i d e a r e shown i n T a b l e 1. H u n g a r y i s P a r t y t o t h e C o n v e n t i o n on l o n g - r a n g e

transboundary a i r p o l l u t i o n

and t o t h e P r o t o c o l s on t h e c o n t r o l o f SO2,

r e s p . NOx e m i s s i o n s .

The P r o t o c o l on t h e r e d u c t i o n o f SO2 e m i s s i o n s i m p o s e s t h e b a s i c o b l i g a t i o n o f a 30 % r e d u c t i o n o f e m i s s i o n s b y 1 9 9 3 b a s e d o n 1 9 8 0 emission values.

Thus,

SO2 e m i s s i o n s o f 1 . 6 3 3

kt/year

i n 1980 s h o u l d

be r e d u c e d t o 1.143 k t / y e a r b y 1993. T h i s r e q u i r e m e n t w i l l be met due t o s t r u c t u r a l changes i n f u e l m i x , t h e economic r e c e s s i o n .

r e s p . i n t h e i n d u s t r y and

However, measures a r e t o be t a k e n i n t h e

f u t u r e c o n s i d e r i n g t h e p r e d i c t a b l e economic e x p a n s i o n and t h e i m p l i c a t i o n s o f a second s u l p h u r p r o t o c o l based on t h e c r i t i c a l l o a d s a p p r o a c h . Means o f f u r t h e r r e d u c t i o n o f SO2 e m i s s i o n s a r e a s follows:

f u r t h e r s t r u c t u r a l changes i n f u e l m i x ,

e n e r g y conservation

measures, t h e i m p l e m e n t a t i o n o f f l u i d i z e d bed c o m b u s t i o n methods

i n power g e n e r a t i o n on a l a r g e s c a l e and t h e i n t r o d u c t i o n o f f l u e g a s d e s u l p h u r i z a t i o n f a c i l i t e s i n some p o w e r p l a n t s .

The c o s t

e f f e c t i v e i m p l e m e n t a t i o n o f t h e a b o v e m e n t i o n e d m e a s u r e s may r e d u c e c a p i t a l i n v e s t m e n t n e c e s s a r y and o p e r a t i o n a l c o s t s i n a r e m a r k a b l e r a t e . I n v e s t i g a t i o n s i n t h i s m a t t e r h a v e shown t h a t t h e r e a r e r e s e r v e s i n energy u t i l i z a t i o n which s h o u l d be e x p l o r e d f i r s t . The P r o t o c o l o n t h e c o n t r o l o f NOx e m i s s i o n s i m p o s e s t h e b a s i c o b l i g a t i o n o f an e m i s s i o n f r e e z e b y 3 1 December 1 9 9 4 b a s e d o n 1987 e m i s s i o n v a l u e s . Thus,

t h e e m i s s i o n o f 280 k t / y e a r

i n 1987

s h o u l d n o t b e e x c e e d e d a f t e r 1 9 9 4 . The e m i s s i o n s o f NOx a r e shown

i n Table 2.

E m i s s i o n s o f power p l a n t s and o f i n d u s t r y d e c r e a s e d

i n t h e l a s t years,

however,

t h e amount o f t r a f f i c g e n e r a t e d n i t r o g e n

o x i d e s r e m a i n e d p r a c t i c a l l y unchanged. T h i s i s due t o t h e f a c t t h a t t h e i m p a c t o f g r o w i n g c a r f l e e t h a s b e e n c o m p e n s a t e d b y decreasing c a r use.

206

T h e reduction of nitrogen o x i d e s emission i s envisaged o n mobile sources. T h e catalyst p r o g r a m m e planned will reduce emissions o f N O x , HC and CO by a r e m a r k a b l e r a t e . O E N O X equipments for power plants have not b e e n c o n s i d e r e d s o f a r . The ammonia e m i s s i o n s a r e i n t h e range o f 160-180 kt/year. T h e emission o f volatile organic c o m p o u n d s i s in t h e range of 2 0 0 - 2 1 0 kt/year. At three background monitoring stations located in ecological basis a r e a s measurements of t h e l a s t 3 y e a r s have shown results concerning acid deposition a s f o l l o w s [ 3 ] : - dry acid depositions were 7 5 , 1 0 9 and 147 mgH+ /in2. year in hidrogen ion e q u i v a l e n t , - in dry acid d e p o s i t i o n s S O 2 and N O 2 g a s e s h a v e a dominating role. Yearly a v e r a g e s a r e w e l l below t h e l i m i t values considered 3 damaging for f o r e s t s ( S O 2 = 30 )g/m , NO:, = 30 p g / m 3 ) , - wet acid d e p o s i t i o n s a r e in t h e same magnitude (110, 9 2 , 2 1 5 2 mgH+/rn .year) a s dry a c i d d e p o s i t i o n s , - in wet acid deposition s u l p h u r and nitrogen c o m p o u n d s have a 40 % , resp. 60 % s h a r e , - in total acid d e p o s i t i o n t h e s h a r e o f d r y acid deposition i s bigger in winter, r e s p . wet acid deposition i s bigger in summer. T h e yearly a v e r a g e values o f S O 2 and N O 2 concentrations o n a regional s c a l e , their o r i g i n , t h e d e p o s i t i o n s o f sulphur and oxidized nitrogen c o m p o u n d s a n d their acidifying potentialaccording to model c a l c u l a t i o n s a r e s h o w n in T a b l e 3 [4].

3. ACIOIFICATION OF SOILS

According t o investigations [5] there a r e some 2 , 3 million hectares o f acidic s o i l s in Hungary. Out o f t h i s total t h e r e a r e - 300.000 ha s t r o n g l y acidic (pHKC1<4,5) - 800.000 ha a c i d i c (pHKC1 = 4 , 5 + 5 , 5 ) - 1 . 2 0 0 . 0 0 0 ha mildly a c i d i c (pHKC1 = 5,5+6,5) These figures r e p r e s e n t 3 , 2 ; 8 , 6 , resp. 1 2 , 9 % o f the territory of the country.

207

Investigations have s h o w n t h a t s o i l s having a p H K C 1 < 5 , 5 increased their s h a r e by 6 % in t h e l a s t 1 2 years. Decreasing pH values of s o i l s may beexplained by t w o main factors: by the acid deposition and by t h e decreasing r a t e o f c h e m i c a l treatment of s o i l s (liming). F e r t i l i z e r s and m a n u r e c o n t r i b u t e t o a large extent to t h e acidification o f s o i l s , too. According t o r e s e a r c h work c a r r i e d o u t by different institutions acid deposition d o e s n o t play a primary r o l e in forest damage experienced in Hungary.

4 . ABATEMENT MEASURES

As primary c a u s e of a c i d i f i c a t i o n h a s been the excessive emission of pollutants, abatement measures to c o n t r o l damages should f o c u s on the reduction o f c o m p o u n d s having acidifying character. Hungary h a s considered i t n e c e s s a r y t o r e d u c e S O 2 and NOx e m i s s i o n s o n national and o n international l e v e l , t o o . T h u s , fulfilling t h e obligations d u e t o international a g r e e m e n t s is o n e o f the main targets of air pollution a b a t e m e n t s t r a t e g y . T h e new legislation o n a i r pollution abatement due to be issued in 1992 and in c o m p l i a n c e w i t h t h e n e w Act o n t h e Environment comprises important e l e m e n t s c o n t r i b u t i n g to t h i s e n d , namely - directives concerning t h e s u l p h u r c o n t e n t o f f u e l s , - national emission l i m i t s o f t h e most important air pollutants f o r n e w a n d in-use f a c i l i t i e s o f power g e n e r a t i o n , industrial boilers, incinerators, on-road motor v e h i c l e s e t c . , - establishing c r i t i c a l l o a d s f o r t h e s e n s i t i v e ecological s y s t e m s , - a comprehensive m o n i t o r i n g s y s t e m a n d e v a l u a t i o n methods. There exists a well established policy and strategy for air pollution abatement i n H u n g a r y . T h e implementation o f it i s a promising but hard task.

208 5 . REFERENCES Increasing a c i d i f i c a t i o n o f the environment. Research i n e n v i r o n m e n t a l p r o t e c t i o n and n a t u r e c o n s e r v a t i o n , No.7.

N a t i o n a l A u t h o r i t y f o r E n v i r o n m e n t a l P r o t e c t i o n and

N a t u r e C o n s e r v a t i o n a n d t h e H u n g a r i a n Academy o f S c i e n c e s , Budapest, 1987 ( i n H u n g a r i a n ) . Proceedings o f t h e N a t i o n a l Conference on t h e A c i d i f i c a t i o n o f t h e Environment (Balatonfured,

14-16 Nov.1990),

B u r e a u f o r E n v i r o n m e n t a l Management, B u d a p e s t ,

Programme

1990 ( i n

Hungarian). E.Fuhrer

and L .

H o r v i t h , A c i d d e p o s i t i o n on e c o l o g i c a l b a s i s

a r e a s , i n [2]. K.Fekete,

E.KovBcs,

L.Jermendy and D.Szepesi,

Predictable

v a l u e s o f a c i d d e p o s i t i o n i n c a s e on d i f f e r e n t e n e r g y s c e n a r i o s , i n [2]. P . S z a b d a n d J. B e n e s d c z k y , A c i d i f i c a t i o n of s o i l s a n d p o s s i b l e

p r o t e c t i o n measures, i n

[z],

Table 1 Sulphur dioxide emission in Hungary (103 t/year)

Power Plants District heating Industry Agr.,forest and water management Transportation Domestic Services

1980

1985

1986

654,l 33,3 522,2 38,l 49,O 290,6 44,9

504,O 21,9 487,3 29,l 21,l 303,5 36,l

528,8 18,l 479,8 26,9 18,5 273,O 25,O

532,6 461,6 1 6 , l 13,5 405,4 397,5 27,3 25,l 18,l 18,6 260,4 271,2 31,4 29,9

1632,8

1403,6

1370,l

1291,9 1218,O

1084,O 1010,O

1987

1989

1987

1988

1989 436,4 13,3 316,4 23,8 16,5 249,l 27,9 ~~~

Total

1990 423,l 12,4 285,l 22,3 16,5 221,6 29,O ~~

Table 2 Nitrogen oxides emissions in Hungary (lo3 t/year) 1980

1985

Power Plants District heating Industry Agr.,forest and water management Transportation Domestic Services

69,O 4,l 53,3 9,9 111,3 18,2 7,l

61,6 3,8 48,4 8,6 110,5 21,6

Total

272,9

262,2

7,7

1986

1988

62,l 396 53,4

59,l 397 58,5

51,3 3,6 47,6

890 115,3 20,l 691

893 118,9 20,9 636

119,6 20,9 7,3

268,6

276,O

8,O

49,6 3,4 45,5 7,7 116,l 19,9 6,8

258,3 249,O

1990 44,7 3,1 37,O 699

117,O 19,5 698

235,O

%

210

Table 3 Yearly average values o f SO2 and NO2 concentrations o n a regional s c a l e , their o r i g i n , t h e depositions o f sulphur and oxidized nitrogen c o m p o u n d s and their acidifying potential according t o model c a l c u l a t i o n s . Sulphur dioxide Yearly average concentration Originating from sources abroad from inland low sources from inland high sources

Nitrogen dioxide

50 %

78 % 18 % 4%

34 % 16 %

Sulphur compounds Oxidized nitrogen compounds Depositions Acidifying potential Total acidifying potential

2,1 g S/m

.2year

2

130 keqHf/km .year

0,8 N/m

2

.

year

2 60 keqH+/km .year

2

190 keqH+/km .year

T. Schneider (Editor). Acidification Research. Evaluation and Policy Applications @ 1992 Elsevier Science Publishers B.V. All rights resewed

ACIDIFICATION ABATEMENT POLICY G.J.A.

-

21 1

THE NETHERLANDS EXPERIENCE.

A1 and V.G. Keizer

Ministry of Housing, Physical Planning and Environment, Directorate General for Environmental Protection, Air Directorate, P.O. Box 450, 2260 MB Leidschendam, The Netherlands

Abstract This paper presents a description of the way in which the present- acidification programme has been developed in the Netherlands. The current policy has adopted an integrated approach for all acidifying substances on the basis of effect oriented goals (2400 acid equivalents deposition per ha, per year, in 2000 and 1400 equivalents in 2010). Two elements of importance in translating these goals into emission reduction objectives are elucidated: the two-track approach in environmental policy and the involvement of emitter categories (target groups) in policy development. The emission reduction objectives for 2000 and measures related to SO,, NO., NH, and VOC emissions under the 1989 abatement programme, are highlighted. Some changes with respect to present legislation are shown. It is stressed that emissions in neighbouring countries have to be decreased proportional to Dutch reductions, to reach an acceptable level of acid deposition in the Netherlands. Finally the perspectives for implementation of the abatement programme and for policy measures to reach the 2010 deposition goals are indicated.

1. Soma basic data Before sharing with you our experience with acidification abatement policy, it may be elucidative to give you some characteristics of the situation in our country. The Dutch territory is about 37,500 square kilometres. With a population of 15 million the Netherlands is one of the most densely populated countries in the world (on average 440 people per square kilometre). But we also have the world's highest livestock density. We have 5 million cattle, 14 million pigs, 90 million chickens. And with respect to the holy cow of the industrialized world: we have 6 million passenger cars, to be used on 115,000 kilometres of roadways and covering 77 billion kilometres per year.

212 It will come as no surprise to you to learn that we have serious environmental problems, with cars and manure playing dominant roles. But it should be clearly understood, that the Netherlands' case does not differ basically from that in other West European countries. Its special features make the situation just more striking. Recent calculations show that total acid deposition in 1989 was 4,800 acid equivalents per hectare, per year. According to these calculations deposition has been 6,800 acid equivalents in 1980. The decline in deposition results mainly from the decrease in SO, emissions in the Netherlands and neighbouring countries. The contribution of the three acidifying substances to the 1989 deposition is given in Table la and of the different emitter categories in Table lb. Table la Contribution to deposition in 1989

Table lb Contribution to deposition in 1989 other Icountries1 :%re1 46%

33%

1

traffic industry 10%

5%

I

power refi- house plants Inerieslholds 1.5%

3%

1

1.5%

Emissions in other countries contribute to a large extent to acid deposition in the Netherlands. But we are net exporters of acidifying substances. In 1980 we emitted 41x109 acid equivalents while total deposition in the Netherlands amounted to 23x109 acid equivalents. In 1989 these figures were respectively 34x109 and 1 6 ~ 1 0 ~ .

2. Introduction In t h e Netherlands environmental policy regognized acidification as a problem in the early eighties. The progress in abatement since then is the result of a series of developments. Remarkably analogous to the difficulty in distinguishing between all origins of a multi-cause problem like forest dieback, it is not possible to attribute the positive results of our policy to one or two developments. Therefore I will sum up the major elements of our approach which of course are largely interdependent.

213

-

-

Rapid development towards a broad political basis for abatement: Integrated approach for all acidifying substances: Development of clear effect oriented goals and a common nominator for acidification: Development of a both source and effect oriented approach (two-track approach); Early involvement of industry and other emitter groups in research and policy-development (target groups).

The integrated approach for all acidifying substances on the basis of effect oriented goals has already been discussed extensively by Wolters in the opening session. For that reason I will focus on two other items: the two-track approach and the way in which the emitter groups (target groups) were involved in developing our abatement policy. Two-track approach. Pollution abatement in general, in our view, requires an effect-oriented approach based on environmental quality objectives as well as a source-oriented approach based on best available technology. Instead of making a preclusive choice between the two tracks, we consider them of equal importance. The track leading to the lowest pollution determines the policy target. The other track only influences the time path for reaching this goal. Contrary to a mere effect-oriented strategy, this approach allowed us to start with emission reduction targets and measures to reduce emissions as soon as preliminary critical loads were available. It did not force policy makers to wait for the critical loads to be indisputable. Any emission reduction measure, deemed necessary on the basis of preliminary critical loads and not surpassing the level of best practicable technology, should endure, even if the final critical loads turn out to be less stringent. Because under those circumstances the source-oriented approach would require such a measure anyway. Only reductions beyond the level of best practicable technology would require fully scientifically established critical loads. Involvement of target groups When describing the way industry and other emitter groups were involved in policy preparations, it is useful to first consider the reasoning behind national emission reductions. Emission sources in the Netherlands, apart from ammonia sources, are only minor contributors to acid deposition in the country itself. 54% of total acid deposition and only 28% of SO, deposition and 41% of NOx deposition is "Made in Holland". So reduction measures for these substances taken in the Netherlands will result only in relatively small improvements in the Dutch environment, unless they are accompanied by comparable reductions in the neighbouring countries. Since it did not make sense to wait for others to take action for our benefit, it was decided to use a rather unselfish approach. As a point of principle for Dutch SO, and NOx sources, emissions should be reduced by at least the same percentage as

214

calculated to be necessary for reduction of total deposition of each substance. This was decided a so that we would have the moral right to request other countries to reduce their emissions accordingly, and b because of our responsibility under the Convention on Longrange Transboundary Air Pollution towards other countries.

A prerequisite when judging the applicability of available technology was to avoid setting standards that would block the development of cost-effective new abatement techniques. This principle was also used when measures were tightened up in later years. In practice NO, emission standards for combustion plants were set at such a level that application of highly sophisticated low NO, technology would provide an alternative to expensive selective catalytic reduction. Efficiency of flue gas desulphurization (FGD) was fixed on a level (85%) allowing for energy saving processes. Emission standards for SO, from refineries were formulated on an integrated basis. One standard was chosen for the refinery as a whole. In a system with different standards, i.e. more stringent requirements for new installations than for existing ones, modernization of existing installations, which does lead to less emission, may be blocked. Integration of standards prevents such an effect. Traditionally the decision-making in the Netherlands is more consensus than command based. So, before deciding on abatement measures, extensive consultations were held with target groups: the industrial organizations and the Electricity Board, the Automobile Association and road transport organizations and, of course, the other competent ministries. In these consultations industry was represented by the employers' organization, the organization of refineries and other specific branch organizations. The consultations led to a satisfactory level of agreement. Negotiations with the various emitter categories were conducted in parallel. This integrated approach prevented emitters from trying to put the burden of emission reduction onto other categories. Concerning Volatile Organic Substances a detailed scheme of reduction measures, some under certain conditions, was set up by the central and regional authorities and industry collectively.

3. Towards present abatement measures A national approach Apart from fuel quality standards and emission standards for motor vehicles, in the early eighties the central government didn't have instruments to implement a national acidification policy. Licenses for large industry and power plants are granted by the provinces. Only by sending ministerial circulars to the licensing authorities, by giving financial support to demonstration projects and by urging industry to take initiatives to reduce emissions on a voluntary basis, could a start be made with a national abatement policy. In 1983 it was agreed between the parties concerned, to apply FGD and low NO, burners on new coal fired power plants.

215

The first abatement programme In 1984 a first comprehensive abatement programme was published' which was to lead to daring but realistic reduction goals and strict legal obligations with respect to emission control. This programme is still the basis for a large part of our present legislation. The abatement programme contained limits for SO, and NO, emissions from the various categories of combustion plants, gasturbines and gas engines and for non-combustion emissions in some branches of industry. These limits were subsequently converted into the emission standards and new fuel quality standards adopted in our present legislation, which came into force in 1987. The EC was urged to decide on stricter emission limits for motor vehicles. The Netherlands played an active role in speeding up the negotiations. Later on, in the framework of the 1985 Luxemburg Agreement, the emission standards for passenger cars were tightened. The introduction of cars which did meet these standards was accelerated in the Netherlands by means of fiscal incentives. For that purpose the special user's tax for new cars that satisfy the standards was decreased in order to compensate for the price difference with ordinary cars. The abatement measures would result in a NO, reduction of 30% and a SO, reduction of 60% in 2000, both relative to 1980 levels. NH, abatement was lagging behind. Abatement technology still had to be developed. Too little was known about total emissions, but a 50% reduction was considered feasible. An indication of possible measures was given. The existence of a clear objective, applicable to all acidifying substances, made the various reduction measures more acceptable to industry and to the public. Present abatement policy Our present abatement objectives for total acid deposition and for ozone (Table 2), were established on the basis of the final report of the first phase of the Dutch Priority Program on Acidification, November 19882. They were published in the National Environmental Policy Plan3 and the Acidification Abatement Plan'. The scientific background for the choices made has been already explained by Wolters. Table 2 Deposition and concentration objectives.

I acid deposition (eq.H+/ha/y)c ozone (pg 0, /m3 ) max. 1 hr average 8 hr average growing season average

1994

2000

2010

crit.ld.

4000

2400

1400

400

---

240' 160b 100

240 160 100

120

--

maximum 2 exceedances per year. maximum 5 exceedances per year. objective for 2000 is the interim target load

50

216

To comply with these objectives dramatic required. Tabel 3 summarizes these reductions.

reductions

are

Table 3. Overall emission reductions (compared to 1980).

[ 80-85%]

20% 30% 30%

NO, NH,

voc

50% 70% 60%

[55-60%]

80-90% 80-90% 80-90% 80%

.data between [brackets] result from acceleration of measures in June 19905.

a

slight

The emission reduction objectives per acidifying substance were also presented as part of the National Environmental Policy Plan and in the Acidification Abatement Plan. The emission reduction objectives have been derived from the interim target load. Furthermore the Abatement Plan describes in detail emission reduction measures and emission limits on the basis of best available technology for the various emitter categories (e.g. power plants, refineries, other industrial sources, households, mobile sources and agriculture). Additionally, structural measures such as those aimed at energy conservation and slowing down the increase in car use, are needed. Measures to reduce emissions of NH, were laid down in the "Plan of Action: Limiting Agricultural Ammonia Emissions)"6

.

The emission reduction objectives are shown for the respective emitter categories in Tables 4a-d, together with a short description of the major additional measures. Tables 4a-d. Emission objectives by target group for 1994 and 2000, in 1000 tonnes per year and percentages respective to 1980. (notes explained after Table4d)

I

Table 4a

1980

1994

-

234 8 9

164 4 9

NH,

I

Total

agriculture industry households

251

177

-

7

1

70 3 9

70% 70% 0%

1

82

70%

I

NH, measures. - attuning the N-content of fodder to animal needs: stimulating a minerals accounting system per farm: - emission standards and zoning for animal housing: - manure treatment, including emission standards f o r manure processing plants: 6 million tons processing capacity in

-

217

-

1994, or otherwise the number of cattle will be decreased: mandatory covering of manure storage facilities: mandatory use of low-emission spreading techniques: extension of the annual period during which no manure may be spread:

Table 4b

NO,

-

-

power plants industry passenger cars trucks households other sources

Total (incl.extra red.b)

1980

1994

2000

80 86 162 115 25 80

55 57 100 110 17 83

30' 37b 40 72 1l b 6ab

548

422

238-243

62% 57%b 75% 37% 56%b 15%b 55-60%

NO, measures, stationary sources. - energy conservation in cooperation with energy suppliers, by insulation and cogeneration: energy efficiency targets for industrial branches: use of non-fossil or renewable energy sources: - further reduction in the electricity production sector: new emission standard for new coal fired power plants: new emission ceiling for power plants for 2000: - subsidies and subsequent emission standards for new gasfired central heating boilers: - applying latest state of the art low-NO, technology in industrial boilers. NO, measures, mobile sources. - applying US-equivalent standards for new passenger cars by means of fiscal incentives in conformity with EC directives: - tightening by 50% of the NO. emission standard for trucks: financial incentives for rapid introduction of cleaner trucks in anticipation of stricter mandatory standards: - mandatory emission tests of "clean" passenger cars: - limitation of car use by various means: stimulating public transport: increase of fuel excise.

I

1980

Table 4c

-

1994

I

2000

195 121 90 38

30 56 45 30

18. 36b 15b 14

90% 70%b 86%b 63%

21

15

lob

52%b

Total (incl.extra red.b) 465

176

75-90

SO,

-

-

power plants refineries industry traffic households and other sources ~~

80-85%

218

SO, measures. - energy conservation; - further reduction in the energy production sector: new emission standard for new coal fired power plants; emission ceiling for power plants for 2000: - for refineries an emission standard of 1000 mg/m3 will be promulgated as a next stage in the stepwise reduction: - process emissions are being reduced by 50%; - lowering the sulphur content of gasoil and diesel oil; sulphur content of bunker oil to be halved. Tabel 4d VOC

I a

-

Total

industry small businesses households agriculture passenger cars trucks other sources

1980

1994

2000

130 83 31 20 159 46 23

85 68 26 13 80 40 25

45c 40' 15' 6 35 30 25

65% 52% 52% 75% 78% 35% -9%

492

337

194=

61%

agreement with Dutch Electricity Generating Board (Sep), June 1990; subject to conditions that may increase emissions slightly. extra reductions for these sources together (25-30 tons of NO=, 15-30 tons of SO, ) 5 if all conditions are met and all uncertainties removed.

VOC measures. The abatement pertaining to VOC emissions is mainly directed at preventing emissions by lowering the VOC content of products and/or switching to low-emission (production) techniques. Government and industry have agreed to measures on a sector-by-sector basis. VOC sources that could not be eliminated by emission prevention will have to implement add-on control technology7 The measures presented in the 1989 abatement plan are currently being translated into (amendments to present) regulations and legislation concerning emission standards, fiscal stimulation, subsidies, etc. and into guidelines f o r licensing authorities and formal agreements with industrial sectors. At present over 40 of these actions have either been completed or are in a final stage of completion. Thus within two years of publication of the abatement programme almost half of the most important measures have been taken.

219 4. Perspectives A first evaluation study by the National Institute of Public Health and Environmental Protection (RIVM) concerning implementation of the National Environmental Policy Plan, was published recently. It includes the measures under the Acidification Abatement Plan. Although part of the proposed measures could not be taken into account, it was concluded that we did set our expectations a little too high. Reduction of ammonia and of NO, emissions from traffic is developing far from smoothly. To attain the required reduction of ammonia emissions a decrease in the number of cattle may be needed in the longer run. Such measures would threaten the very existence of a large number of farmers and could give rise to big social problems. Recently, the ministers of Environment and Agriculture announced the development of legal instruments to ensure the decrease of manure production required in 1994. The increase of car use should be brought to a halt. But when the use of private cars comes into play the public gets uneasy. The growth in volume of freight traffic plays an even more important role. As a consequence, the decrease of emissions from mobile sources proceeds with great difficulty.

And that concerns only the interim target load. But our policy aims at reaching the 1400 acid equivalents target load in 2010. That requires an even more dramatic decrease in emissions, nationally and internationally. The abatement measures required for the national part of that reduction have to be developed before 1993, when the new National Environmental Policy Plan is scheduled to be put before Parliament. Apart from applying the newest abatement technology, reduction of car use, decrease in the number of cattle and extensive energy conservation will probably be required. New policy instruments have to be developed, like levies on energy use, on emission of ammonia and maybe SO, and NO,. Such measures should be accompanied by a compensating mechanism for unintended social and economic effects and should preferably be taken in an ECcontext. Apart from these national difficulties, the perspective for reductions in the surrounding countries is also precarious. The international part should arise from ECE, or maybe EC, negotiations. Taking into account the required emission reductions of 80 -90% it will be a tremendous task to attain sufficient results in the next 20 years. Although progress has been made, international discussions about the critical loads approach leave me with the impression that too many countries still play the ostrich where the necessary level of national emission reductions is concerned.

220 5. Conclusions Establishing deposition objectives as part of a two-track approach to acidification abatement has so far proved to be very useful. It has allowed a prompt start with emission reductions, even when only preliminary critical loads were available. By providing a common basis for reduction targets for all acidifying substances, from all categories of emitters, it made an integrated reduction of SO,, NO, and NH, possible. This integrated approach prevented one category of emitters from trying to put the burden of emission reduction onto other categories. The deposition objectives made drastic reduction measures plausible, even measures exceeding mere application of best practicable technology, and helped to gain the cooperation of the emitters in implementing the abatement programmes. The far-reaching additional measures needed to comply with national target loads or even critical loads will make very strong demands on such cooperation. Without the support of an overall European strategy like a critical loads approach, a national programme to reach the target load will be extremely difficult to implement.

6. References 1. Indicative Multi-year Programme for Air 1985-1989, Parliamentary Documents XI, 1984/85, 18 605, nrs 1-2. 2. Dutch Priority Program on acidification: Summary report Acidification Research 1984-1988, Publication 00-06, November 1988, RIVM (National Institute of Public Health and Environmental Protection). 3. Parliamentary Documents XI, 1988/89, issued to Parliament May 25rh 1989.

21

137,

nrs.

1-2,

4. Parliamentary Documents 11, 1988/89, 18 225, nr. 31, issued to Parliament July 20rh 1989. 5. Parliamentary Documents XI, 1989/90, 21 137, nr. 20, issued to Parliament June 14th 1990. 6. Parliamentary Documents XI, 1988/89, 18 225, nr. 32, issued to Parliament July 20th 1989.

7. Control Strategy for Emissions of Volatile Organic Compounds, Project KWS 2000: Project group Hydrocarbons 2000.

T Schneider (Editor). Acidificatlon Research Evaluatlon and Policy Applications Publishers B V All rights resewed

0 1992 Elsevier Science

22 1

The Convention on Long-range Transboundary Air Pollution: its Achievements and its Potential H. Wiister

Secretariat of the United Nations Economic Commission for Europe'

Introduction "Calling for negotiation of international treaties is an easy way for governments to divert attention from their own failure to deal with domestic pollution and environmental damage; and once proposals for international negotiation gain acceptance, no government can afford to be left out."' This critical remark, referring to negotiations on a global level, can be seen as an example of the kind of criticism that accompanies i n t e ~ ~ t i ~ n a l cooperation on environmental issues. Even though the potential benefits from international cooperation are well accepted, the question remains whether all the benefits can be fully achieved by international agreements. More than ten years after the signing of the Convention on Long-range Transboundary Air Pollution, it is worthwhile to attempt an evaluation of what has been achieved. This could help to determine whether international cooperation should continue along the lines that seem to have been followed so far. This paper will start with a presentation of the Convention and its protocols on the basis of their history. It is suggested that the Convention has proved to be a dynamic instrument that provides a great deal of flexibility for new developments. This conclusion will be underlined by looking, in section 11, at those past and present activities under the Convention that may not lead, or have not yet led to a protwo1 or any other legal text. The final section will addtess in an exploratory way some of the aspects that may be addressed by future agreements. Here, many questions will have to remain open.

I. The Convention and its Protocols A. Historical Overview The history of the Convention can be traced back to the 1960s, when Swedish scientists demonstrated the interrelationship between SO, emissions outside Swedish territory

' The views expressed are those of the author and do not necessarily coincide with those of the United Nations Economic Commission for Europe Robertson (1990). p. 111.

222

and the acidification of Swedish lakes. This led to a presentation of a case-study on "Air Pollution Across National Boundaries, the Impact on the Environment of Sulphur in Air and Precipitation" by the Swedish delegation at the 1972 United Nations Conference on Human Environment in Stockholm. The other participating delegations showed hardly any immediate interest in the issue.' Skepticism arose obviously from those countries, far away from the Scandinavian lakes, that were implicated in the destruction of these natural resources. In the final text adopted by the Conference, no mention was made of the acid rain issue. The only aspect included in the Stockholm Declaration related to acid rain, or more generally to transboundary air pollution, is included in article 31 that mentions, inter alia, that each state is responsible for the fact that activities undertaken within its territory should not cause any damage beyond its frontiers. Between 1972 and 1977 several studies, particularly those undertaken by the OECD with Norway as the lead country, confirmed the hypothesis that air pollutants actually can travel over several thousands of kilomeires before a deposition and damage occur. This also established the fact that cooperation at the intemational level was necessary to solve the related problems, in particular the acidification problem. At that stage, however, mainly the Scandinavian countries considered themselves victims of transboundary pollution and therefm showed interest in finding a solution. It would still take several years until it was recognized that actually most countries were victims and polluters at the same time and therefore cooperation could be beneficial to most if not all countries. With the first meeting of the Conference on Security and Co-operation in Europe (CSCE) in 1975 in Helsinki, an additional impetus helped to speed the process. The final document of the CSCE contains the elements which enabled the Convention to be prepared. For example, it is emphasized that participating states will use every suitable opportunity to cooperate in the area of control of air pollution: "Desulphurization of fossil fuels and exhaust gases; pollution control of heavy metals, particles, aerosols, nitrogen oxides, in particular those emitted by transport, power stations, and other industrial plants; systems and methods of observation and control of air pollution and its effects including long-range transport of air pollutants"'. Among the forms and methods of cooperation the participating states agreed, inter alia, "to develop through international cooperation an extensive programme for the monitoring and evaluation of the long-range transport of air pollutants starting with sulphur dioxide and with the possible extension to other pollutants",4 which provided the basis for EMEP. In addition, the Participating States advocakxk Itthe inclusion, where appropriate and possible, of the various areas of cooperation into the programme of work of the United Nations Economic Commission for Europe"(UN/ECE). The selection of the UN/ECE as a forum for developing such a programme relies mainly on two elements: first, the UN/ECE already had substantial experience in the field of environment as its f i i activities in this area dated back to the late 1950s. Secondly. which is perhaps more important, the UN/ECE was the unique forum where East and West were represented on equal footing. This process eventually led in 1979 to the signature of the Convention on Long-range Transboundaxy Air Pollution in Geneva. For the European Economic Community (EEC), this Convention was the first occasion to become a N1 contracting

'See Bjorkbom (1988), p.128.

CSCE, Final Act, Chapter 11, Section 5.

' bid.

223

Party of an international treaty with participation from the East? The Convention entered into force in March 1983. This provided the basis for the development of the protocols that added more substance to the general character of the Convention. Already in 1978 a European monitoring and evaluation programme had been established by UN/ECE with the assistance of the World Meteorological Organization (WMO) and the United Nations Environment Programme (UNEP). The financial contributions to this programme were made at first on a voluntary basis. The first protocol to the Convention, the Protocol on Long-term Financing of the Co-operative Programme for Monitoring and Evaluation of the Long-range Transmission of Air Pollutants in Europe (EMEP), adopted in 1984 and in force since January 1988, commits Parties to mandatory annual contributions to the EMEP budget. The next major task was to deal with the atmospheric pollutants involved in the process of acidification. In 1977 an OECD study on the cost and benefits of SO, control gave a clear indication that there were net benefits to be gained by reducing sulphur emissions. In the early 1980s the political climate towards environmental issues had changed drastically in western Europe; particularly public awareness of the effects of pollution, above all the "Waldsterben" (forest die back) had increased. In 1982 at an International Conference in Stockholm on Acidification of the Environment, the 20 participating governments reached general agreement on the effects of long-range transported sulphur compounds and on the need for internationally coordinated action. At the first meeting of the Executive Body of the Convention in 1983 the Scandinavian countries presented a proposal for a reduction of SO, emissions before 1993 by at least 30%, using the 1980 emission level as a basis. In the same year the EEC adopted recommendations on SO, and NO, emission reductions. In 1984 ten western countries met in Ottawa (Canada) at the ministerial level to commit themselves, inter alia, to SO, reductions along the lines proposed by the Nordic countries. A multilateral conference with representatives from all parts of the ECE region held in Munich (Germany) on the Causes and Prevention of Damages Caused to Forests and Water by Air Pollution in Europe led to the breakthrough and the agreement to lay down the 30% reduction commitment in a legally binding text. After a year of negotiations the Protocol on the Reduction of Sulphur Emissions or Their Tmboundary Fluxes by at least 30 per cent was adopted in Helsinki in July 1985; it entered into force in 1987. In the year of the adoption of the Helsinki Protocol, in 1985, negotiations already started for action to be taken on nitrogen oxides. Three yea's later, in 1988, the protocol was ready for adoption: the Protocol concerning the Control of Emissions of Nitrogen Oxides or Their Transboundary Fluxes was adopted in Sofia and entered into force in February 1991. Originally there were intentions to reach an agreement on the actual reductions of NO,, but the protocol commits Parties only to a freeze at 1987 levels by 1994. Twelve of the Parties to the Rotocol additionally committed themselves in a declaration to reduce NO, emissions by at least 30%.

B. The Convention on Long-range Transboundary Air Pollution Even though the Convention does not include any commitments for the reduction of air pollutants, nor does it provide a legal basis for state liability, in 1979 a Convention on transboundary air pollution bringing together 32 ECE member states and the EEC was

224

a major breakthrough. A list of Parties to the Convention with the date of signatory and ratification or accession is pmvided in table 1. Besides laying down the general principles of international cooperation for air pollution abatement, it set up an institutional framework that managed to bring together research and policy. With Article 2, containing the major principle of the Convention, contracting Parties undertake "to protect man and his environment against air pollution" and "endeavour to limit, as far as possible, gradually reduce and prevent air pollution including long-range transboundary air pollution." The ensuing fundamental principles call upon the Parties to exchange information, not only on the volume of the emissions but also on the policies and strategies aimed at limiting these emissions, and to hold consultations, if a Party affected by long-range transboundary air pollution so requests. An "Executive Body" was established by Article 10 as the supreme policy-making assembly in which all Contracting Parties are represented and which meets at least annually to adopt its work-plan and review the implementation of the Convention. Between sessions a Bureau consisting of the chairman and three vice-chairmen deals with matters requiring interim action. The second level of the institutional framework under the Convention consists of intergovernmental Working Groups, established by the Executive Body as standing bodies open to all Parties of the Convention and dealing with specific areas of the work-plan. Under these subsidiary bodies intergovenunental Task Forces can be established either on an ad hoc basis to prepare a specific report, or to supervise a continuing cooperative progtamme. Responsibility for each task force rests with a designated lead country. The topics contained in the work-plan and covered by the subsidiary bodies have been established by the ensuing articles of the Convention. This includes the exchange of infomation and cooperation in research on technologies related to air pollution abatement, on the effect of air pollution, on economic aspects of the activities under the Convention and on national policies and strategies related to air pollution. Although the Convention has been referred to as "the Acid Rain Convention"6there is no reference to acidification; the only pollutant directly referred to in the text, e.g. in the Preamble, is sulphur, but the possibility of an extension to other pollutants is mentioned at the same time. The vagueness of the Convention, which has been subject of much of its criticism, can also be regarded as its strength: it p v i d e d the flexibility that served as the frame for the cooperation in the field of air pollution in the ECE region and the development of the protocols.'

C. The protocols to the Convention 1. Co-operative Programme for Monitoring and Evaluation of Long-range Transmission of Air Pollutants in Europe (EMEP)

Within the framework of the Convention and its successive protocols, EMEP plays a key role enabling Parties to assess the progress made in the implementation of the various aspects of the Convention. Its main objective is to provide Governments with

' Rosencranz (19881, p. 174. ' See Fmenkel (1989), p. 447.

225

TABLE 1

Status of the Convention and the EMEP Protocol

Austria Belarus Belgium Bulgaria Canada Czech and Slovak Federal Republic Denmark Finland France

Germany(5) Greece Holy See Hungary Iceland Ireland Italy Liechtenstein Luxembourg Netherlands Norway Poland Portugal Romania San Marino Spain Sweden Switzerland Turkey Ukraine USSR United Kingdom United States Yugoslavia European Community Total:

Convention Signature Ratification* 13.11.1979 16.12.1982 (R) 14.11.1979 13.6.1980 (R)

13.11.1979 15.7.1982 (R) 14.11.1979 9.6.1981 (R) 13.11.1979 15.12.1981 (R) 13.11.1979 23.12.1983 (R) 14.11.1979 18.6.1982 (R) 13.11.1979 15.4.1981 (R) 13.11.1979 3.11.1981 (Ap) 13.11.1979 15.7.1982 (R)(2) 14.11.1979 30.8.1983 (R) 14.11.I 979 13.11.1979 22.9.1980 (R) 13.11.1979 5.5.1983 (R) 13.11.1979 15.7.1982 (R) 14.11.1979 15.7.1982 (R) 14.11.1979 22.11.1983 (R) 13.11.1979 15.7.1982 (R) 13.11.1979 15.7.1982 (At)(3) 13.11.1979 13.2.1981 (R) 13.11.1979 19.7.1985 (R)(2) 14.11.1979 29.9.1980 (R) 14.11.1979 27.2.1991 (R) 14.11.I 979 14.11.1979 15.6.1982 (R) 13.11.1979 12.2.1981 (R) 13.11.1979 6.5.1983 (R) 13.11.1979 18.4.1983 (R) 14.11.1979 5.6.1980 (R) 13.11.1979 22.5.1980 (R) 13.11.1979 15.7.1982 (RM4) 13.11.1979 30.11.1981 (At) 13.11.1979 18.3.1987 (R) 14.11.1979 15.7.1982 (Ap) 34 32

Notes: R = Reifcation, Ac =Accession,Ap = Approval, At = Acceptance

EMEP Protocol Signature Ratification*

28.9.1984 25.2.1985 4.4.1985 3.10.1984 28.9.1984 7.12.1984 22.2.1985 26.2.1985

4.6.1 987 (Ac) 4.1 0.1 985 (At) 5.8.1987 (R) 26.9.1986 (Ap) 4.12.1985 (R) 26.1 1.I 986 (Ad 29.4.1986 (R) 24.6.1 986 (R) 30.10.1987 (R) 7.10.1986 (R)(2) 24.6.1 988 (Ac)

27.3.1985

8.5.1985 (Ap)

4.4.1985 28.9.1984

26.6.1987 (R) 12.1.1989 (R) 1.5.1985 (Ac) 24.8.1987 (R) 22.10.1985 (AtI(3) 12.3.1985 (At) 14.9.1988 (Ad 10.1.1989 (Ac)

21.11.1984 28.9.1984 28.9.1984

28.9.1984 3.10.1984 3.10.1984 28.9.1984 28.9.1984 20.11.1984 28.9.1984 28.9.1984 22

(1) With declaration upon signature (2) With declaration upon ratification (3) Foc the Kingdom in Europe (4) Including the Bailiwick of Guernsey, the Isle of Man, Gibraltar, the United Kingdom Sovere' n Base Areas of Akmtiri and Dhekhelia on the a l ndof Cyprus (5)TheformerGDRsignedtheeonventionon13.11.1979, ratified ilon7.6.1982 and acceded to the EMEP Protocol on 17.12.1986 with a declaration upon accession

11.8.1987 (Ac) 12.8.1985 (R) 26.7.1985 (R) 20.12.1985 (R) 30.8.1985 (At) 21.8.1985 (At) 12.8.1985 (R) 29.10.1984 (At) 28.10.1987 (Ac) 17.7.1986 (Ap) 30

226 information on the deposition and concentration of air pollutants, as well as on the quantity and significance of long-range transmission of pollutants and fluxes across boundaries. The EMEP programme has three main elements: (1)the collection of data; (2) the measurement of air and precipitation quality; and (3) the modelling of atmospheric dispersion, using emission data, meteorological data and functions describing the transformation and removal processes. The measurement activity of EMEP initially focused on sulphur oxides in air and precipitation, but it has been gradually expanded. It now includes sulphur dioxide, sulphate, several nitrogen compounds and ozone in rural air, and all important ions in precipitation. The Chemical Co-ordinating Centre (CCC) at the Norwegian Institute for Air Research (NILU) is responsible for the chemical part of the programme. The main tasks of the CCC include: collection, assessment and storage of data from national monitoring stations; the organization of laboratory tests to establish and improve the quality of the chemical analyses; and the review and recommendation of sampling and analytical methcds for compounds included in EMEP. The meteorological part of EMEP is carried out by two Meteorological Synthesizing Centres (MSCs), with scientific coordination by WMO. The Meteorological Synthesizing Centre-East (MSC-E) is located at the Institute of Applied Geophysics in Moscow (USSR), the Meteorological Synthesizing Centre-West (MSC-W) is located at the Norwegian Meteorological Institute (DNMI) in Oslo (Norway). The main tasks of the MSCs include: model calculations of the transboundary fluxes and the deposition of air pollutants; verification and development of models for the transport of air pollutants, taking into account ground measurement data received from CCC; assessment and utilization of scientific results concerning transport and transformation processes for the development of new models or improvement of existing ones. The EMEP sampling network provides the database for concentrations and depositions. The network consists of 96 stations in 24 countries in Europe and is based on 24-hour sampling of air and precipitation; t h m are still areas in Europe where the density of stations is insufficient. Canada and the United States also contribute to the programme with reports and interlaboratory comparisons. The CCC has developed technical guidelines for reporting and estimation of the SO,, NO, and VOC emissions in order to harmonize emission inventorying in Europe, taking into account the reporting requirements of the protocols. Emission data have been provided by Parties to the Convention or have been estimated by the Centres on the basis of statistical data or other information. The spatial resolution of the emission data needed is given by the model grid size of 150 km x 150 km and efforts to improve the accuracy of the data are going on.

2. The Protocol on the Reduction of Sulphur Emissions or Their Transboundary Flwea by at least 30 per cent This protocol consists only of basically five operative articles, with a basic provision by which Parties undertake to reduce their annual sulphur emissions or their transboundary fluxes by at least 30% as soon as possible and at the latest by 1993,using 1980 levels as a basis for calculating reductions.* Although the protocol may rightly be criticized for the arbitrarily choosen figure of 30%, that has been considered insufficient, or the arbitrarily choosen base year, that may give countries that have reduced emissions before that date a disadvantage, with this protocol countries have committed themselves for the

* Article 2 of

the Sulphur Protocol.

221

first time to reduce emissions, and as such the Protocol has proved to be a useful yardstick for national policy measures. The Protocol also contains a provision that calls upon Parties . will be to go beyond the 30% goal "when environmental conditions ~ a r r a n t " ~As elaborated below, some Parties to the Protocol have already reduced sulphur emissions by more than 30%. On 9 July 1985 the Protocol was signed in Helsinki, Finland, by 20 of the then 30 Parties to the Convention. A list of Parties to the Sulphur protocol with the date of signature and ratification or accession is provided in table 2. Among those Parties to the Convention that did not sign the Sulphur h t o c o l , are the United Kingdom, the United States and some eastern European countries. Among the main reasons for opposition were the uncertainties over the linkage between sulphur emissions and acidification and the fact that there were other pollutants involved in the process of acidification. Other Parties to the Convention that did not sign the Protocol considered themselves unable to implement substantial emission reductions, either because of the structure of the energy production, or because they could not bear the heavy investments necessary to achieve such reductions. In March 1991 the United States, however, concluded an agreement with Canada, the Air Quality Accord, that, inter alia, deals with the control of sulphur emissions. The remaining operative articles mainly refer to a reporting system covering emissions and national strategies and policies implemented to reach the objective. The role of EMEP is reinforced by the fact that EMEP shall "provide to the Executive Body calculations of sulphur budgets and also transboundary fluxes and deposition of sulphur compounds"1o. 3. Fhtocol concerning control of emissions of nitrogen oxides or their transboundary fluxes.

With its entry into force in February 1991, the NO, Protocol is the latest legal instrument that will provide the basis for the work under the Convention. It was signed in November 1988 in Sofia (Bulgaria) by 24 Parties to the Convention and so far has been ratified by 18 Parties. The fact that more countries have signed this Protocol as compared to the Sulphur protocol may be due to the differences in sources of emissions of both pollutants: for sulphur the important sources for energy are put into question: in the case of nitrogen, as most emissions of nitrogen oxides come from mobile sources, such as transportation, emission reductions will not require investment of a large scale, but small steps, for example, the installation of catalytic converters. As for the Sulphur Protocol, Article 2 of the NO, Protocol containing the basic obligations, starts with a reduction target. In this case Parties undertake to reduce or to control emissions "so that these, at the latest by 31 December 1994, do not exceed their national annual emissions of nitrogen oxides or transboundary fluxes of such emissions in the calendar year 1987"", the year the Protocol was negotiated. Another alternative offered to the party is to choose upon signature any year previous to 1987 as a base year provided that in addition the average national level of emissions between 1987 and 1996 does not exceed the 1987 level. In addition to this emission target the basic obligations call upon Article 3 of the Sulphur Protocol. Article 5 of the Sulphur Protocol. Article 2 (1) of the NO, Protocol.

228

TABLE 2 Status of the Sulphur and the NOx Protocol Sulphur Protocol Signature Ratification* Austria 9.7.1985 4.6.1987 (R) 9. 7.1985 10.9.1986 (At) Belarus 9. 7.1985 9.6.1989 (R) Belgium 9.7.1985 26.9.1986 (Ap) Bulgaria 9. 7.1985 4.12.1985 (R) Canada Czech and Slovak Federal Republic 9.7.1985 26.11.1986 (Ap) 9.7.1985 29.4.1986 (R) Denmark 9.7.1985 24.6.1986 (R) Finland 9.7.1985 13.3.1986 (Ap) France 9.7.1985 3.3.1987 (R)(2) Germany(5) Greece Holy See 9.7.1985 11.9.1986 (R) Hungary Iceland Ireland 9. 7.1985 5. 2.1990 (R) Italy 9. 7.1985 13.2.1986 (R) Liechtenstein 9.7.1985 24.8.1987 (R) Luxembourg 9.7.1985 30.4.1986 (At)(3) Netherlands 9.7.1985 4.11.1986 (R) Norway Poland Portugal Romania San Marino Spain 9.7.1985 31.3.1986 (R) Sweden 9.7.1985 21.9.1987 (R) Switzerland Turkey 9.7.1985 2.10.1986 (At) Ukraine 9.7.1 985 10.9.1986 (At) USSR United Kingdom United States Yugoslavia EuroDean Communilv Total: 20 20

NOx Protocol Signature Ratification* 15.1.1990 (R) 1.11.1988 8.6.1989 (At) 1.11.I 988 1.1 1.I 988 30.3.1989 (R) 1.1 1.I 988 25.1.1991 (R) 1.11.I 988 1.11.I 988 17.8.1990 (Ap) 1.I 1.1 988 1.2.1990 (R) 1.11.1988 1.11.1 988 20.7.1 989 (Ap) 16.11.1990 (R) 1.11.1988 1.11.1988 3.5.1989 1.5.1989 1.11.1 988 1.1 1.1988 1.11.1988 1.11.1 988 1.11.1988 1.11.1988

4.10.1990 (R) 11.I 0.1 989 (At) (3) 11.10.1989 (R)

1.11.1988 1.11.1988 1.11.1988

4.12.1990 (R) 27.7.1990 (R) 18.9.1990 (R)

1.11.1988 1.11.1988 1.11.1988 1.11.1988(1)

24.7.1989 (At) 21.6.1989(At) 15.10.1990 (R)(4) 13.7.1989(At)

26

18

Notes: R = Ratification, Ac = Accession, Ap = AppmvaI, At = Acceptance (1) With declaration upon signature (2) Wlh declaration upon ratification (3) For the Kingdom in Europe (4) Includingthe Bailiwickof Guernsey, the Isle of Man, GibratIar, the United Kingdom Soverei n Base Areas of Akrotiri and Dhekhelia on the Island of Cyprus (5) The tormer GDR signed the gulphur Protocolon 9 July 1985 and the NOx Protocol on 1 November 1988

229

Parties to apply emission standards to major new stationary sources or to new mobile sources, and to introduce pollution control measures for major existing stationary sources. To this end the Protocol contains a technical annex containing a list of best available technologies for different sources and a table of emission standards; the intention is to keep this annex under constant review. The Protocol contains another dynamic element insofar as Parties undertake to "commence negotiations no later than six months after the date of entry into force of this Protocol, on further steps to reduce national annual emissions of nitrogen oxide or transboundary fluxes of such emissions"'z; particular reference is made here to the concept of critical loads that will be discussed below. Another new element introduced in this Protocol is the notion of exchange of technology that Parties undertake to facilitate, through the promotion of commercial exchange, direct industrial contacts and cooperation, including joint venture planning, exchange of information and experience and provision of technical assistance." Furthermore, Parties undertake to make unleaded fuel sufficiently available in order to facilitate the circulation of vehicles equipped with catalytic converters." With respect to reporting, monitoring and exchange of information the NO, Protocol follows the lines of the previous Protocol incorporating some of the experiences gained.

D. The Next Steps whereas with the adoption of the Convention itself and with the first protocols it may not have been envisaged that this would lead to such a dynamic process, certainly this process has been accepted with the NO, Protocol. The history of the Convention has shown that once countries get together to deal with an environmental issue of a regional scale, like acidification, they start to recognize the benefits of international cooperation. It can be concluded that the stepwise approach, that has originated from the Convention and its protocols, has helped to address the uncertainties that existed when approaching this new field of intetnati~nalcooperation some twenty years ago. The elements of flexibility and dynamics of the NO, Protocol have also been included in the draft text of the next protocol. The Draft h t o c o l Concerning the Control of Emissions of Volatile Organic Compounds or their Transboundary-Hues" has been finalized and is expected be adopted and signed at the next session of the Executive Body in November 1991. In 1988 the Executive Body set up a Working Group on VOCs in view of "the importance of damage to the environment in many countries caused by emissions of volatile organic compounds (VOCs) which, by reaction with the oxides of nitrogen, contribute to the formation of photochemical oxidants such as ozone, and consequently stressing the necessity to reduce effectively VOC emissions"'6. The result of the work,the draft protocol, is much more complex than any previous protocol; the English version of the draft is Article 2 (3) of the NO, Protocol. l 3 Article 3 of the NO, Protocol l4 Article 4 of the NO, Pmtocol. EB.m.54 l6 This had already been noted in 1985; see Report of the Fifth Session of the Executive Body, ECE/EB.AIw16, para. 46.

230

about ten times longer than the Sulphur Protocol. This reflects mainly the complexity of the issue, but to some degree also the fact that a more sophisticated approach has been chosen than for previous protocols, for instance with respect to the application of best available technologies. The emission target set out in the basic obligations of the draft protocol is to reduce national annual VOC emissions by at least 30% by 1999, using either 1988 levels as a basis or any other annual level between 1984 and 1990. There will, however, be alternatives to this obligation. One is to include only areas that have been specified as Tropospheric Ozone Management Areas (TOMAS) for the above emission target and ensure that national emissions of VOCs do not exceed 1988 levels by 1999. The other alternative refers to a situation where the national annual emissions of VOCs in 1988 were below 500000 tonnes and 20 kg per inhabitant and five tonnes per square kilometer; in this case the obligation is relaxed to ensure that 1988 emissions are not exceeded in 1999. The draft protocol includes four annexes including emission control technologies for stationary and mobile sotmes and a classification of VOCs according to their photochemical ozone creation potential (POCP) that indicates those substances that should be avoided as a matter of priority. The basic obligations call on Parties to apply, no later than two years after the date of entry into force emission standards based on best available technology to new sources, and to apply best available technology also to existing sources of VOC after five years, where international tropospheric ozone standards are exceeded or where t r a n s h d a r y fluxes Originate. For a second sulphur protocol preparations started this year. The emission target in the Helsinki Protocol sets 1993 as a deadline, even though the Protocol covers the subsequent period by the obligation to further reductions of sulphur emissions when environmental conditions warrant. This has been emphasized by the Executive Body that, in 1989, agreed on the interpretation that this obligation "means that reductions to that should be reached in that time frame and the levels maintained or extent whether [~WO] further reduced after being rea~hed."'~ Nevertheless, 1993 has been envisaged as a possible date for the adoption of a second Sulphur Protocol. There are clear indications that efforts will be made to include in such a protocol also those Parties to the Convention that have not signed the Helsinki Protocol. It also has become clear that attempts will be made to include such innovations as have already been used in the NO, and VOC protocols and perhaps even to go further by incorporating other results from the extensive research that has been conducted in this field. A major innovation would be to set differentiated obligations and departing from the approach of setting equal percentage reductions for every Party that may be more or less arbitrary. An accepted effects-based scientific approach, the critical load concept, may provide the basis for arriving at differentiated obligations. This will be elaborated on in the last section of this paper. The work on further steps to reduce national annual emissions of nitrogen oxides or their transboundary fluxes that, according to Article 2 of the NO, Protocol was to start no later than six month after the date of its entry into force on 14 February 1991, has started with the preparation of a timetable including the elements that are considered to be important in order to be able to reach an agreement that goes further than the first NO, Protocol. As set out in the NO, Protocol, the critical load concept will play a central role. The work-plan that has been proposed includes work on: emission inventories for NO, as well as for ammonia and VOCs; the control techniques for these pollutants; on economic

Report of the Seventh Session of the Executive Body, ECE/EB.AIR/20, paragraph 22.

23 1 instruments and in particular with relation to the critical loads; the models used to link the different elements. In this connection the possibility of combining the different pollutants in a single acidification protocol has been mentioned.

11. The Role of the Convention for Acidification Policy and Research Before turning to a more explorative discussion of the questions that may have to be addressed in the preparation of future protocols, an assessment of the achievements of the Convention with respect to the problem of acidification will be attempted. An obvious way to do this is to look at emission figures. It also seems important, however, to discuss the role of the Convention with its subsidiary bodies in providing a forum for international cooperation on research and policy formation. Certainly a clear indication of the achievements is the evolution of international agreements contained in the protocols, as discussed above.

A. Emissions of SO, and NO, in the ECE region Emission figures only show one side of the coin and probably not the most important one. The other side would be to look at the actual deposition and its effects. So far, however, the main targets specified in the protocols have been referring to emissions and those seem to be the primary targets of national policy. As already indicated, it is likely that future protocols will specify targets in terms of the effects of depositions and in some countries abatement policy has been specified in those terms already. As mentioned above the Convention and its protocols call upon the Parties to report among other things on annual national emissions and on strategies and policies for their abatement. Every four years this information is compiled and, after derestriction by the Executive Body, published as a Major Review of Strategies and Policies for Air Pollution Abatement. The latest one was published in 1991.'' Annual updates are prepared and published; the latest information was received from governments in August 1991.'' Looking at SO, emission figures one can note that by 1990, that is three years before the target year of the Sulphur Protocol, most of the Parties to the Protocol have fulfilled their obligation to reduce emissions by 30% or more compared to the 1980 level. To show the trend nationally submitted 1980 emission figures, as published in UN/ECE (1991), have been set to 100 for every Party (see table 3). Some Parties to the Sulphur Protocol: Austria, France, Germany (the Federal Republic as prior to 1990), Norway and Sweden, have already achieved emission reductions of more than 50%; other Parties to the Protocol are still on the way to reaching the 30% target. On average the achievements of Non-Parties to the Protocol in Europe are clearly below those of the Parties: The average percentage reduction of 1980 sulphur emissions for all of Europe is still below 3W, whereas the average for the European Parties to the Sulphur Protocol is approaching 400/0.

UN/ECE (1991) Annual Review of Strategies and Policies for Air Pollution Abatement, 9 September 1991, EB.AIWR.62. l9

232

Table 3: SOz Emissions in the ECE region based on 1980 figures (UNECE 1991)

233

For Canada and the United States the picture is not as clear; there are no data for the last few years. There is research being done to link these trends to the trends in concentrations and depositions. A study at MSC-W showed that for the period 1979 to 1986 an emission reduction of 16.4% in Europe led to a reduction of 18.9% in the concentration of SO,, and of 15.5% in the concentration of particulate sulphur.mEMEP measurements in general confim the downward trend for air b e sulphur in Europe: measurements for the period from 1982 to 1988 show no trend at 32 sites or a downward trend at 14 sites for sulphur dioxide. A significant downward trend for sulphur dioxide was found in the United Kingdom, France, at some sites in Germany, and in southern Scandinavia. In other parts of Europe (most parts of central Europe, the Baltic area and at one site in Poland close to the Soviet border and in the northern region), there were no significant trends. Measurements of sulphate in particles give very much the same picture." For NO, there is not yet such a clear picture as for SO,. The base year of the NO, Protocol was only a few years ago (1987) and the Protocol itself only came into force this year. Nevertheless, repeating the same exercise as for SO,, taking 1987 (or, where there are no data for that year, the closest previous figure available) as the base year (see table 4) shows a promising picture: most of the Parties to the NO, Protocol have managed to stabilize their annual emissions and in some cases reductions have been achieved. As much as can be determined from the few figures that are available for the years after 1987, the increase in NO, emissions has been stopped or at least slowed down. The measurements of nitrogen concentrations undertaken by EMEP using figures up to 1988 can hardly be used as confirmation for this. The nitrate in precipitation had a decreasing trend at one site only (in Germany). At most other sites no significant trend was found or there were upward trends as in the Baltic area and Yugoslavia.u

B. Research on the Effects of Air Pollution The ultimate aim of cooperating Parties, as laid down in the fundamental principles of the Convention, is to protect man and his environment against air pollution. As one of the first subsidiary bodies, the Executive Body set up a Working Group on Effects that has since extended its coverage from sulphur to other air pollutants. Research is at present being done in five International Cooperative Programmes for the Assessment and Monitoring of Air Pollution Effects under the auspices of the Working Group. The effects of acidification on freshwaters are monitored and assessed by a cooperative programme led by the Norwegian Water Research Institute ("A), which compiles and intmalibrates chemical and biological data from more than 160 catchment areas in 11 countries. A cooperative programme led by the United Kingdom at the University of Nottingham School of Agriculture conducts research to evaluate the effects of air pollutants and other stresses on agricultural crops, through a joint exposure experiment on selected sensitive crops organized annually in 13 countries. Also a major concern are the effects of air pollution on materials, including historic and cultural monuments; these are monitored by a programme under the leadership of the Swedish See Mylona (1989). See UN/ECE (1991a)

bid.

234

Table 4: NO, Emissions in the ECE region based on 1987' figures (UNECE 1991)

235 Corrosion Institute, with sub-centres in the Czech and Slovak Federal Republic (for steel, zinc and aluminum), Germany (for copper and cast bronze), Norway (for paint coatings) and in the United Kingdom (for stone materials); in the period 1987 to 1991 uniform exposure measurements have been performed at 39 sites in 14 countries. Another programme, led by Germany, has developed a manual for the monitoring and assessment of air pollution effects on forests and has, since 1986. carried out annual large-scale surveys of forest damage in 27 European countries. The focus of further research is on intensive studies at permanent observation plots, with a view to determine causeeffect relationships between air pollution and forest decline. Finally a pilot programme on integrated monitoring of air pollution effects on ecosystems, led by Sweden and supported by the Nordic Data Centre in Helsinki (Finland), carries out sampling and analysis (on terrestrial and aquatic biota, soils, groundwater and surface waters) at 33 sites, usually in small catchments or other hydrologically well-defined areas.

C. Research into Economic Aspects Ideally the basis for the formulation of abatement strategies should be some kind of evaluation that shows that the cost of abatement are outweighed by the benefits of reducing emissions. The Group of Economic Experts on Air Pollution, that is now the responsible subsidiary body for economic aspects, originally was a Group of Experts on Cost and Benefit Analysis. A major line of its work has evolved to an integrated assessment of costs and benefits that is now covered by the Task Force on Integrated Assessment Modelling led by the Netherlands. There has been close cooperation with the International Institute for Applied Systems Analysis (IIASA) to evaluate the cost efficiency of different abatement strategies. IIASA has developed a model system, the "Regional Acidification Information and Simulation" (RAINS) model that combines information on emission generation, such as energy use and industrial and agricultural activities, with emissioncontml technologies and abatement cost; it takes into account the long-range transport of pollutants and the environmental effects of acid deposition in different areas of Europe. On that basis, scenarios of different abatement strategies can be analysed or, alternatively, with given objectives the optimal abatement strategy can be calculated.23The work of this Task Force will provide an important basis for the preparation of future protocols. The activities have gone beyond the economic analysis of impacts of emission abatement. There is work going on, within a Task Force led by Norway, to link emission projections to economic development thereby harmonizing the national projections of future emissions that at present are derived at by using very different models and therefore lack comparability. An important area of cooperation is the exchange of information on the application of economic instruments, such as taxes, subsidies, charges or tradable emission permits. The use of these instruments can help countries to achieve given emission targets at lower cost. The same aspect has also been under investigation for application at the international level; the issue of burden sharing to reach a greater degree of efficiency of allocation on the international level will be of importance for the preparation of future protocols and taken up again in the final section of this paper.

See Alcamo, Shaw 8c Hordijk (1991).

236

D. Cooperation in the Field of Emission Control Technologies As mentioned above, the latest protocol on NO, and the draft proto~olon VOCs include technical annexes that contain descriptions of the state-of-the-art technologies for pollution abatement from different sources. There are several subsidiary bodies under the Convention preparing and unpdating this information. There have been Task Forces looking at VOC emissions (led by France and Germany) and on NO, emissions (led by Germany) from stationary sources and the possibilities of their control and there has been a Task Force on the utilization of by-products and management of wastes from fuel preparation and combustion, led by Austria. The latest event sponsored by the ECE with respect to these activities has been the Fifth Seminar on Emission Control Technologies for Stationary Sources held in Nuremburg (Germany) in June 1991; 83 discussion papers were presented under eight different headings. The seminar adopted a set of conclusions, recommendations to the Executive Body and elaborated draft recommendations to the Parties of the Convention.u In 1988 a Task Force on Exchange of Technology, led by Finland, has begun to explore ways and means of facilitating and promoting technology transfers and direct industrial cooperation. Recent changes in the political and economic systems in many central and eastern European countries have profoundly improved the basis for technological cooperation. Together with the United Nations Development Programme (UNDP), the ECE has organized a series of activities to facilitate cooperation, with particular emphasis on the need of countries with economies in transition.

111. Potential Gains from further Cooperation The Convention has addressed a clear case of an international externality: a case where the activity of one country has an effect on another country and this effect is not covered by any kind of agreement, as in the case of trade where market forces govern the quantities crossing boundaries. In far as countries are at the same time polluters of their neighbors as well as victims of pollution originating in other countries, one can speak of a reciprocal externality. Assuming that benefits of emission reduction on the whole outweigh the costs involved, the problem is that in the case of a reciprocal externality every country would like to have a reduction of emissions, but does not have the incentive to reduce its own emissions, as this would mostly benefit its neighbors. This is the kind of situation where theory suggests that a joint effort is likely to emerge. Taking into account, in addition to this, the internal effect that reduced emissions may have in pmctice, this interest can be seen as the driving force behind the first set of agreements on emission reductions following the Convention. Considering the uncertainty associated with such a new type of international agreement, it may also have been sensible to start with some careful steps and in that case there may still be mom for further steps similar to these f i t ones. In theory there may, however, also be states that are only victims or only sources of pollution and therefore this theory of reciprocal externality does not apply. Those states that are only victims have nothing to offer in an agreement, and those that are only ~~

See Annexes I to I11 of EB.AIWSEM.2/3.

237

polluters, nothing to gain. In practice these differences between national interests will limit the extend to which potential gains can be exploited. A Swedish economisP has shown with model calculations that it tends to be countries at the edges of the region and those that are large emitters of pollution that might not gain net benefits from an agreement. In such cases, it has been suggested that with a transferral of resources from those countries that do benefit from an agreement an efficient solution can be achieved. Without these transfers cooperation would still be worthwhile, but potential gains will not be exploited completely. The question that remains open is how to bring about a transfer of resources without rewarding those that are responsible for emissions. The difficulties associated with the above traditional economic approach, that is based on some evaluation of costs and benefits, arises because of unavailability of data. This applies particularly to the benefit side, where attempts to make estimates have proved unsatisfactory?6 A solution to this problem is the adoption of the Critical Load Concept as the basis for a common long-term goal. Critical loads have been defined as "a quantitative estimate of an exposure to one or more pollutants below which significant harmful effects on specified sensitive elements of the environment do not occur, according to present knowledge"27.The basic idea is that emissions are reduced to the level determined on the basis of critical loads; and damages to the environment, as far as they are known, will be avoided. The work on this involved the following steps: (1) critical load maps for Europe have been produced on the basis of research into the local effects of air pollution; (2) deposition measurements have established whether and to what extent these critical loads are exceeded, and this information has been gathered in exceedance maps; (3) using the long-range transport model calculations have shown emission reductions necessary to avoid exceedances of critical loads; (4) with the help of national abatement cost functions the cheapest of these emission reductions has been selected. A link to the traditional economic approach can be made with the assumption that ecological damages, with their effects for futwt generations, are valued so high that they are to be avoided at virtually any cost. This can also be considered as a link to the concept of Sustainable Development, as economic activity that has no damaging effects on the environment can be considered sustainable. So far emission targets have been negotiated as flat-rate percentage reductions that are equal for all the Parties; a first departure may have been made in the draft VOC protocol where there are different options. The aim of the critical load approach is that pollutant emission reductions are negotiated on the basis of the effects of air pollutants, and to this end different reductions may be necessary in different parts of the ECE region. It has become clear that flat rate reductions will not be a cost-efficient manner to eventually reach critical loads as it will entail excessive cost in areas where critical loads at present are not exceeded. A problem associated with the critical load approach is the fact that, even using best available abatement technology everywhere, it is at present not possible to reach critical loads, As an intermediate step it has been agreed to negotiate target loads based on technological as well as economic and social considerations. One can imagine that a series of decreasing target loads that will eventually approach critical loads will be agreed upon.

25

See Maler (1989).

This has been elaborated in UN/ECE (1991b). The Critical Load Concept and the Role of Best Available Technology and Other Approaches, Report of the Working Group on Abatement Strategies, September 1991, EB.AIR/WG.S/R.24/Rev.l. 26

238

Using the critical load approach in the way described, however, does not solve the problem of the international allocation of resources that has been identified in connection with the traditional economic approach at the beginning of this section. In order to enable all Parties to the Convention to join an agreement based on the critical load approach an international transfer of resources may again be necessary. For sulphur, for example, it has been seen that concentration and exceedances over critical loads vary to a great extent. Model calculations have shown that for the whole of Europe emission reductions would have to be nearly 70% to reach critical loads and the distribution of these necessary reductions would be very unequal.a Comparing the cost necessary to reach those reductions with the national economic potential, as for instance expressed by the GDP, shows the differences even more drastically. This ratio of abatement cost and GDP could be taken as an indicator for the burden that a Party has to incur in an agrement based on critical loads. Particularly in view of the economic situation of central and eastern European countries with economies in transition, it becomes obvious that some degree of burden sharing might improve the conditions for reaching an agreement that exploits the potential gains from international cooperation. There are many uncertainties to be resolved on this and related issues. It has for instance been shown that an increased energy efficiency that may be the consequence of economic restructuring in central and eastern Europe may reduce the problem significantly.z9 There are numerous activities going on in different international fora that might help to solve the problems. To conclude with a quote from a report to the Group of Economic Experts, published in the latest version of the Air Pollution Studies:" "The greatest uncertainty is the assessment of the willingness of countries to cooperate so as to share the cost of reaching target loads. Only international negotiations will reveal the value which countries attach to a clean environment throughout Europe."

References Alcamo, J., Shaw, R. & Hordijk, L., Eds. (1991), The RAINS Model of Acidification, Kluwer, Amsterdam. Amann, M., Klaassen, G. 8z Schijpp, W. (1991), Exploring Europan Sulphur Abatement Strategies, IIASA Working Paper. Amann, M. et al. (19911, Economic Restructuring in Eastern Europe and Acid Rain Abatement Strategies, IIASA Working Paper WP-91-026. Bjorkbom, L. (1988), Resolution of environmental problems: the use of diplomacy, in: Carroll, J.E., Ed., International Environmental Diplomacy, Cambridge University Press, pp. 123-137. Chossudovsky, E.M. (1988), "East-West" Diplomacy for Environment in the United Nations, UNITAR, United Nations Publications, Sales No. E.88.XV.ST26. Fraenkel, A.A. (1989), The Convention on Long-Range Transboundaty Air Pollution, Harvard International Law Journal. 30, pp. 447-476.

a 29

See table 4 in Amann, Klaassen & Schopp (1991). See Amann et al. (1991). UN/ECE (1991b)

239 Maler, K.-G.(1989), "The Acid Rain Game", in: Folmer, H. 8c van Ierland, E.C., Eds., Valuation Methods and Policy-Making in Environmental Economics, North-Holland,

Amsterdam. Mylona, S. (1989). Detection of Sulphur Emission Reductions in Europe, EMEPhISC-W Report 1/89. Robertson, D. (1990). The global environment: Are international treaties a distraction?, The World Economy, Vol. 13, pp. 111-127. R o s e n m , A. (1988). The acid rain controversy, in: Carroll, J.E., Ed., International Environmental Diplomacy, Cambridge University Press, pp. 173-187. UN/ECE (1991), Strategies and Policies for Air Pollution Abatement, 1990 Review, ECElEB.AIW27, United Nations Sales No. E.91.II.E.29. UN/ECE (1991a), The fourth phase of EMEP: 1987-1989, in: Air Pollution Study 7, ECF&B,AIW26, United Nations Sales No. E.91.II.E.18, pp. 3-36. UN/ECE (1991b), Economic Principles for Allocating the Costs of Reducing Sulphur Emissions in Europe, in: Air Pollution Studies 7, ECEYEB.AW26, UN Sales No. E.91.II.E.18, p ~ 63-80. .

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T Schneider (Editor). Acidification Research Evaluation and Policy Applications 1992 Elsevier Science Publishers B V

ACIDIFICATION POLICY

24 1

- SWEDEN

Kerstin Lovgren, Goran Persson and Eva Thornelof Swedish Environmental Protection Agency, S-171 85 SOLNA, Sweden

Abstract Three elements are important in the Swedish policy to combat acidification. (i) (ii) (iii)

Research and monitoring to define critical loads for acid deposition and to produce sensitivity maps. Measures to reduce domestic emissions of sulphur dioxide, nitrogen oxides and ammonia in combination with international agreements on emission reductions. Liming of surface water as a holding operation until acid deposition falls below critical loads.

1. CRITICAL LOADS AND TARGET LOADS FOR SULPHUR AND NITROGEN The critical load for sulphur is in the interval 3-8 kg sulphur per hectare and year for most types of forest land and surface water (see Table 1). Table 1 Critical loads, target loads and required reduction in sulphur deposition (total deposition) Critical Target Actual Required load load deposition reduction kg S h d y r kg Slhdyr’ kg S h d y r ’ % Whole country Gotaland (southern Sweden) Svealand (central Sweden) Norrland (northern Sweden)

3-8

5 5 3

10-20 5-10 3-5

~

Average figure for an EMEP grid Square (150 x 150 km)

75 50 40

242

The target load for sulphur deposition has been set at 5 kg sulphur per hectare and year in Gotaland and Svealand and at 3 kg sulphur per hectare and year in Norrland. The lower value in Norrland is justified by the fact that a great deal of acid accumulates in the snow during winter and produces acid surges in the spring. Foreign sources have a predominant influence on the deposition of sulphur in Sweden. More than 90% of deposition is of foreign origin. Evaluations made indicate that different types of natural environment tolerate a deposition of 5-15 kg nitrogen per hectare and year (see Table 2).

Table 2 Critical loads, target loads and required reduction in nitrogen deposition (total deposition) Critical Target Actual Required load load deposition reduction kg N/hdyr kg Nhdyr'kg N/hdyr'% Whole country Gotaland (southern Sweden) Svealand (central Sweden) Norrland (northern Sweden) ')

5-15 10 8 6

10-20 5-10 2-5

50 20

Average figure for an EMEP grid square (150 x 150 km)

The target load for nitrogen deposition has been set at 10 kg nitrogen per hectare and year in Gotaland, 8 kg nitrogen per hectare and year in Svealand and 6 kg nitrogen per hectare and year in Norrland, since growth and accompanying absorption of nitrogen is lower in the north. Critical loads for marine eutrophication have not been taken into consideration in the proposed target loads for nitrogen in Table 2. In the action programme for marine pollution it is stated that anthropogenic nitrogen input to the Baltic and the seas west of Sweden should be reduced by 70430% in order to achieve conditions which could be characterised as unaffected. Air pollution via deposition on the surface of the sea accounts for around 40% of the total input of nitrogen t o the Baltic. I t will be difficult to reduce the direct nitrogen input to marine areas via aqueous discharges by more than 50%. Thus, in order to stop eutrophication of the Baltic and the seas west of Sweden, nitrogen deposition may have to be reduced by more than 50%. Foreign sources have a predominant influence on deposition of nitrogen in Sweden and surrounding marine areas (Figure 3). For nitrogen oxides the contribution is above 90%. If ammonia is included, the figure is 75-80% for total nitrogen deposition. Total deposition in certain types of forest, particularly exposed forest fringes, may be substantially higher than what appears from the values for larger areas and calls for an even greater reduction of deposition.

243

2. EMISSION REDUCTIONS AND THEIR IMPLEMENTATION

Swedish emissions of sulphur dioxide and nitrogen oxides during the last 40 years are shown in Table 3. The present emissions of sulphur dioxide are only one third of the emissions in the early 50s. In contrast, emissions of nitrogen oxides have increased threefold. Emissions of ammonia add another 40 ktonnes of nitrogen.

Table 3 Emissions of sulphur dioxide and nitrogen oxides in Sweden 1950-1989 ~~

1950 Sulphur dioxide (ktonnes as S)

1955

240

Nitrogen oxides (ktonnes as N)

35

1960

1970

1980

1989

320

460

260

87

90

120

115

The national goals are reduced sulphur emissions by 80% by the year 2000 as compared with 1980 30% reduction in emissions of nitrogen oxides by the year 1995 (if possible) and a t the latest by the year 1998 (NO,-Declaration) as compared with the level in 1980-86 25% reduction in ammonia emissions by the year 1995 in southern and western Gotaland. To reach these goals the best available end-of-pipe technologies as well as structural changes will be used. Fuel and emission standards will be combined with economic incentives and disincentives. The economic instruments introduced are the most far-reaching in the world and are outlined below. As from January 1991 a carbon dioxide tax of SEK 250 per tonne CO, (34 ECU per tonne) is levied on Swedish carbon dioxide emissions. In the energy sector, the tax is levied on oil, coal, natural gas and liquified petroleum gas. During a transitional period the regulations in force for energy taxation will also apply to the carbon dioxide tax. This means in principle that no tax will be levied on electricity production and energyintensive industries. In the transport sector, the carbon dioxide tax is imposed on petrol, motor fuel (diesel) and domestic air traffic. The tax on petrol and diesel was raised

244

in January 1990 by a n amount corresponding t o the present carbon dioxide tax. The sulphur tax, which also came into force in January 1991, corresponds to SEK 30,000 per tonne of sulphur emitted (4,100 ECU per tonne). It is imposed on coal, peat and oil. Technically, the sulphur tax takes the form of a fuel tax. If sulphur emissions are reduced, for example by flue gas desulphurization, part of the sulphur tax (the amount corresponding to the emission reduction) is refunded. The revenue from the sulphur tax has not been earmarked for environmental purposes. Parliament has also decided to introduce a charge on nitrogen oxides emissions. The charge will amount to SEK 40,000 per tonne NO, (5,700 ECU per tonne NO,) and will be levied on emissions from large and mediumsized combustion plants. The charge will be returned to the facilities in proportion to the amount of energy produced. The main reason for this is to avoid penalizing district heating plants and other large installations a s compared with individual heating. The charge will come into force in January 1992. As from January 1991, there is a differentiation of the tax on diesel fuel in order to stimulate the use of environmentally superior grades of diesel oil. Diesel oil with "standard' specifications belongs to environmental class 3. Diesel oils with better specifications belong t o class 2 and the best grades to class 1. From January 1992 the difference in tax compared to class 1 is SEK 200 and SEK 450 per m3 for class 2 and 3 respectively. New vehicles will be classified in three environmental classes as from the 1993 model year. Vehicles which only comply with mandatory requirements will be placed in class 3. These vehicles will be subject to a n increased tax a t the time of sale. Vehicles in the most environmentally Giendly classes will qualify for a reduced tax. Fuel markets began to react to the new environmental taxes and charges as soon as the decisions to introduce the new type of taxation had been taken by Parliament. The sulphur tax has encouraged the use of less heavy fuel oils and favoured low-sulphur qualities. Almost without exception, light fuel oil today contains less than 0.1% sulphur. A precise assessment of the effects on the markets for heavier oils is still difficult. However, a large fraction of heavy fuel oils seems to contain less than 0.5-0.6% sulphur. Measures to reduce nitrogen oxides emissions have been intensified, although the charge of such emissions has not yet come into force. The use of motor fuels was reduced by 2% in the first half of 1990, after having increased by 4 6 % each year over the last five years. This could be a short-term effect of the tax increases on such fuels in early 1990. The trend towards reduced use of motor fuels has continued in the latter half of 1990 and early 1991. It is of course hard to tell how much is a result of domestic policy and how much is due to the Kuwait crisis. Coal has become less competitive due to the carbon dioxide tax and the sulphur tax. Within the district heating systems a number of coal-fired installations have switched to other fuels, especially biofuels.

3. LIMING

245

The chemical objective of surface water liming is to raise the pH above 6.0 and the alkalinity above 100 ueqfl, which gives an acceptable buffering capacity. In addition, metals will be precipitated after liming, thus reducing their toxicity. The biological objective is to detoxify the water so that the natural flora and fauna can inhabit or re-colonize limed water. In the period 1977-90 a total of SEK 780 million of government money was spent on the Swedish operational liming programme. An additional SEK 100 million has been invested by the local authorities, anglers associations, and others, since the governmental subsidies as a rule only cover 85% of the total cost for liming operations. Only in special cases is 100% support given. More than 150,000 tomes of limestone (CaCO,) is spread annually. In 1990 government costs were SEK 109 million. The plan is to at least double the scope of the liming programme by 1994, since recent studies have shown that considerable parts of northern Sweden not yet limed are acidified. In addition, large areas of forest soils in southern Sweden need liming in the near future. Up to 1990 over 85% of the acidified lake area and nearly 6,000 of the 16,000 acidified lakes have been covered by the national liming programme. The remaining acidified lakes are small remote lakes, where local interest for liming is low. At present, more than 10,000 km of the 100,000 km acidified running waters are treated repeatedly. Our experience of lake liming can be summarized as follows:

* * *

*

The chemical objective (i.e., pH>6) is easily achieved by lake liming in combination with wetland liming and lime dosers, when so required. Liming increases the amount of carbon and phosphorus in the lake, thereby probably stimulating productivity.

A postive effect on fish fauna is generally seen, and long-term studies have shown a "normalization" with respect to abundance, recruitment, and species dominance. In heavily acidified areas, where species are extinct, a reintroduction after liming might be necessary.

The only fully satisfactory cure for acidified waters is naturally to reduce deposition of sulphur and nitrogen. A fairly moderate reduction would restore watercourses in areas where the soils still have a good neutralizing capacity. Very high reductions will have to be achieved to restore lakes in areas with impoverished soils. Such reductions cannot be achieved in the near future and liming will therefore have to continue as a holding operation.

246 4. CONCLUDING REMARKS

Critical loads and target loads for sulphur and nitrogen related to EMEP grids have been adopted. International cooperation in research and mapping is important to further develop the critical loads approach. Domestic sulphur emissions were reduced by 65% in the period 1980-89, and by more than 80% from 1970. Nitrogen oxides are down by 5% in the last ten years. Economic instruments are used to implement emission reductions. The programme for liming of surface water includes 6,000 lakes and 10,000 km of streams a t a n annual cost of 24 million ECU. Forest soils in Southern Sweden will have to be limed in the near future. The extent of these types of holding operations is dependent on the size and speed of the emission reductions of sulphur and nitrogen in northern Europe, since approximatly 90% of acid deposition originates from foreign sources.

T Schneider (Editor). Acldlflcatlon Research Evaluatlonand Pol~cyAppllcatlons 0 1992 Elsevier Sctence Publishers B V All rights reserved

247

Air Pollution Control Policy in Switzerland B. C. Galli Purghart Federal Office of Environment, Forests and Landscape, Air Pollution Control Division, CH-3003 Berne Abstract There are a lot of reasons why emissions of air pollutants have to be reduced. Acid deposition and its consequences are among them, but they are not the only ones. This paper presents the comprehensive legal basis of air pollution control policy in Switzerland as well as the overall air pollution control strategy. The present situation with respect to air pollution is discussed. A list of implemented and planned measures is given as well as emission trends of major air pollutants from 1950 to 2010. 1. Legal basis of air pollution abatement in Switzerland In 1983 the Swiss parliament decided on a comprehensive law relating to the protection of the environmentl, which came into force in 1985. Its purpose is to "protect human beings, animals and plants, their biological communities and habitats against harmful effects o r nuisances and to maintain the fertility of the soilo'. As a basis for realizing this purpose, the law contains two main principles: the principle of preventive action ("early preventive measures shall be taken in order to limit potential harmful effects or nuisances") and the polluter-pays principle ("anyone who causes measures to be taken under the provisions of the law shall bear the costs"). The control strategy based on the law is both source-orientated (by emission standards) and effect-orientated (by ambient air quality standards). In order to prevent harmful effects on human beings and the environment, a two-stage approach is applied:

-

In a first step the emissions have to be limited on a precautionary basis, even if the effects of existing pollution are far from being harmful. The emissions shall be limited as much as technology and operating conditions will allow, provided it is economically feasible.

-

In spite of the precautionary measures it is possible that environmental pollution reaches levels that are harmful or a

248

nuisance. In this case the second stage has to be applied. The emissions shall be limited more stringently, to the extent necessary to reduce the burden of pollutants below the critical levels. The main legal instrument for the implementation of the twostage approach is the Ordinance on Air Pollution Controlz, in force since 1986. The ordinance contains emission standards corresponding to the first, precautionary, stage (best available technologies which are economically feasible). 21 inorganic constituents of dust (mainly heavy metals), 13 volatile inorganic substances, 103 organic substances and 21 carcinogens are regulated by means of emission standards for stationary sources. The pollutants are grouped into several classes according to their toxicity and ecological impact properties. The list of substances corresponds to a large extent to those contained in the German "Technische Anleitung zur Reinhaltung der Luft" (TA-Luft). Additionally quality requirements for fuel and gasoline are set. With regard to pollution caused by motor vehicles, emission standards for passenger cars, light duty vehicles, heavy duty vehicles, motor-cycles and mopeds are laid down in the Ordinances relating to the Laws on Road Transport, Navigation and Aviation. Several hazardous substances used in products are regulated by the Ordinance relating to Environmentally Hazardous Substances, in force since 1986. The second stage of the approach requires an assessment of air pollution levels by means of ambient air quality standards. For several major air pollutants such standards are set in the Ordinance on Air Pollution Control. Air pollution levels exceeding ambient air quality standards have to be reduced by strengthening emission control measures taken at the first stage. The enforcement of the Ordinance on Air Pollution Control is mainly a task of the cantons. The cantonal authorities have to assess the existing pollution and its future evolution on their territories. If pollution is or is expected to become excessive, the authorities have to work out action plans at the local and regional level and to prepare the appropriate measures to reduce pollution. Excessive pollution means air pollution levels beyond the ambient air quality standards set in the Ordinance on Air Pollution Control. The action plans require an overall assessment of the air quality situation at the local and regional level and a comprehensive emission abatement programme taking into account stationary sources as well as mobile sources. The plans should be realized within eight years after the enforcement of the Ordinance on Air Pollution Control, that means by 1994. 2. Overall air pollution control strategy

As an addition to the above mentioned laws and ordinances the Swiss government decided in 1986 on an overall air pollu-

249

tion control strategy' covering the major air pollutants sulfur dioxide (SO,), nitrogen oxides (NO,) and volatile organic compounds (VOCs). According to this strategy the total amount of SO, emissions in Switzerland shall be reduced to the level of 1950 by 1990 (designated as 'maximum goal'), the NO, and VOC emissions at least to the level of 1960 by 1995 (designated as 'minimum goal'). In the case of SO2, this corresponds to a reduction of 57% compared to the emission level of 1980. In the case of NO, and VOCs, reductions of 69% and 57%, respectively, are needed compared to the emission levels of 1984. In 1989, the report Ozone In Switzerland' by the Federal Commission of Air Hygiene was published. The report confirms the objectives set in the governmental overall strategy. To abate the excessive photochemical pollution it is necessary to reduce the national annual emissions of nitrogen oxides and volatile organic compounds at least to the extent fixed by the strategy. In the framework of the UN/ECE Convention on Long-range Transboundary Air Pollution, several scenarios with respect to critical loads and target loads of sulphur and acidity were recently explored for the Parties to the Convention, including Switzerland. They lead to similar results. To protect sensitive ecosystems the goals set by the Swiss overall strategy represent indeed the absolute minimum to be done. 3 . Implemented measures A lot of measures have already been tant among them are

-

-

taken. The most impor-

strict emission standards for stationary sources quality requirements for fuels (e.g. low sulphur content) tight emission standards for passenger cars, which correspond to the US-83 regulation mandatory annual inspection and maintainance programme for gasoline-driven passenger cars emission standards for heavy duty vehicles, motor-cycles, mopeds and airoplanes reduction of speed limits on highways and country roads measures to displace the transport of goods from the roads to the rail measures in favour of public transport

4 . Present pollution situation

These measures led to a decrease of the total SO, emissions from 126'300 tons to 62'600 tons (50% reduction) between 1980 and 1990. Between 1984 and 1990 the NO, emissions decreased from 214'300 tons to 183'800 tons per year (14% reduction),

250

whereas

the VOC emissions were reduced from 339'300 tons to tons per year (12% reduction) during the same period (see figure).

297'000

As a consequence, the ambient SO, concentrations decreased to a large extent5. The annual SO, standard of 30 pg/m3 is not violated any more. However, the measured levels of air pollutants still exceed the NO, standard of 30 p9/m3 (annual mean) and the 0, standard of 120 pg/m3 (hourly mean) to a large extent. Indeed, photochemical oxidant pollution is considered to be one of the major environmental problems in Switzerland. Attention is also paid to the deposition of air pollutants, since for sensitive ecosystems critical loads of acidity are exceeded in many areas of Switzerland. 5. Additional efforts

The measures taken up to now do not allow to achieve the objectives of the overall strategy for nitrogen oxides and volatile organic compounds. Neither are they sufficient to guarantee that air quality standards will be met by 1994. Additional steps have therefore to be taken. An expertise, elaborated by the Swiss engineering company Elektrowatt Corporation and published in June 1989, has shown that it is possible to achieve the governmental goals by increased efforts6. Up to now reduction of air pollutants meant setting and implementing tight emission standards for all classes of sources. However, it was shown that these technical mesures won't do. The steady increase of road traffic, industrial production and consumption of goods partly offsets the effect of technical measures. Therefore supplementary action will have to include economical instruments, traffic management and energy saving. The Swiss government pronounced its intention to realize the following measures:

-

-

-

stricter emission standards for stationary sources stricter emission standards for heavy duty vehicles and motor-cycles mandatory periodical inspection and maintainance programme for diesel-fuelled passenger cars, light duty and heavy duty vehicles increased fines on violations of the speed limits incentive taxes on the sulphur content of gas oil incentive taxes on organic solvents taxes on CO, emission introduction of a so-called ecobonus system (i.e. passenger car holders have to pay a certain amount for every kilometre driven or the price of gasoline will be risen; the incoming money is equally redistributed among the whole population) raising the taxes for heavy duty vehicles as a function of kilometres driven with a supplement for strongly polluting vehicles construction of new railway-tunnels through the alps, allowing a substantial increase of the transport capacity

25 1

for goods by the rail

- a programme called "Energy 2000", aiming at energy saving and promotion of renewable energies Further measures will be taken by the cantons which have to establish and realize action plans on the local and regional scale to reduce the excessive air pollution in their areas. The measures contained in the plans presented up to now include traffic management, reducing speed limits, applying tighter emission standards for stationary sources. The majority of these additional measures are still far from being realized, because new constitution amendments and federal laws are needed before they may be implemented. Under the condition that all the planned measures will be translated into action as scheduled, the goals set by the Swiss government can be achieved with a delay of two or three years to the initial time-table of the governments overall strategy (see figure). 6. References

Federal law relating to the protection of the environment of 7 October 1983, SR 814.01; available from BUWAL, CH-3003 Berne Luftreinhalte-Verordnung (LRV) vom 16. Dezember 1985, SR 414.318.142.1; available from EDMZ, CH-3003 Berne (german/french/italian) Bericht Luftreinhalte-Konzept vom 10. September 1986, BBI 1986 I11 269, available from EDMZ, CH-3003 Berne (germad french/italian) Ozon in der Schweiz, Schriftenreihe Umweltschutz Nr. 101, ed. BUWAL, Bern 1989; available from BUWAL, CH-3003 Berne (germadfrench) Luftbelastung 1990, Schriftenreihe Umwelt Nr. 148, ed. BUWAL, Berne 1991; available from BUWAL, CH-3003 Berne (germadfrench) Elektrowatt Ingenieurunternehmung AG, Untersuchungen im Zusammenhang mit dem Luftreinhalte-Konzept des Bundesrates und zusatzlichen Massnahmen zur Reduktion der Luftverschmutzung, Schlussbericht, Zurich 1989; available from BUWAL, CH-3003 Berne (german)

252 annual emissions of air pollutants annual emissions of air pollutants taking Into account all proposals (bask scenario, taking Into account by the government all measures already implemented) In tons per year sulphur dloxlde so, IW

so. [Val

lOOOa0i

1950

1960

1910

1980 84

1890

2wO

2010

1950

1960

1970

1980

1990

2000

2010

2000

2010

nitrogen oxides NO. Illal

NO, [ l h ]

200000

~

-

l50DOO

150000

IOOWO

100000.

50DW

50000

mlnlmum goal-level 1960

-

maximum goal- level 1950

0 ,

0

volatlle organic compounds VOC[lh]

VOC

[Ila]

3wm

zw'ow

2M)m

Iw'ow

1OODOO

0 19s

0 1960

1910

1980 84

domestlc

W O

2000

2010

industry

1960

I

1070

1980

1990

Implemented measures intended action

T. Schneider (Editor). Acidification Research. Evaluationand Policy Applications 1992 Elsevier Science Publishers B.V.

253

Acidification Research : Evaluation and Policy Applications A

R . G . De-t

United Kingdom Policy Response

and R . B . Wilson

Science Unit, Air Quality Division, Department of the Environment, 43 Marsham Street, London SWlP 3PY Abstract The United Kingdom environmental research into the mechanisms of the atmospheric transport and deposition of acidity, understand the impacts of that acidity on soils, surface waters, forests, crops and the built environment and the consequences for fishery status, freshwater and soil ecosystems. The Critical Loads Approach opens the possibility of more subtle and sensitive ways of tackling the problems of environmental acidification on the European scale. The United Kingdom is contributing vigorously to the Critical Loads Approach through the mapping exercises, the environmental studies that underpin them and the understanding of the driving deposition mechanisms which lead both to pollutant removal and ecosystem contamination. Future progress with the UN ECE Convention on the Long Range Transport of Air Pollution and the revision of the NO, SO and VOC protocols will rest in very large measure on the shared confidenge wlthin Europe in the knowledge of the underpinning environmental science. The Critical Loads Approach should provide an important policy focus within the international scientific community to set environmentally-based targets for future co-ordinated emission control programmes. 1.

INTRODUCTION

The United Kingdom has a substantial commitment to the support of environmental research into the mechanisms of the atmospheric transport and deposition of acidity, to understanding the impacts of that acidity on soils, surface waters, forests, crops and the built environment, and the consequences for fishery status, freshwater and soil ecosystems. The United Kingdom collaborates fully with international research programmes within the framework of the UN ECE Convention on the Long Range Transport of Air Pollution with these aims in mind. International action agreed so far on acidic depositions has taken little account of the complexity of their impacts. It has focused on the abatement of emissions and applied to this the straightforward instruments of negotiated national percentage reductions and the application of best economically feasible abatement techniques to major sources. These tools have the advantages of simplicity and intelligibility to policy-makers, regulators and industry, which are important at a stage when we are still a long way from achieving our environmental objectives. They are however relatively blunt instruments. Even over their present planned lifetime, they risk entailing some wastage of resources on action which will not secure environmental improvement. Looking ahead to the next stages of an international abatement programme, we must build our actions on scientific consensus about the nature and distribution of the damage being done by acid deposition, and links between observed deposition patterns and sources of emissions. Within the framework on the UN ECE Convention on the Long Range Transport of Air Pollution, we have begun the search for more subtle and sensitive ways of tackling the problems of environmental acidification. Member States have agreed to adopt a rational approach, based on the underpinning environmental science,

254

for the revision of the NO , SO2 and VOC protocols. This is the Critical Loads Approach, of which you wifl hear much at this International Conference. The Environment White Paper, "This Common Inheritance", explains that the UK Government believes that the Critical Loads Approach is central to further progress on the control of transboundary pollution. Last year, the 34 nations of the ECE region agreed at the Bergen Conference on Sustainable Development that: (i)

strategies for the abatement of SO2 and NO, emissions should be designed in the most cost-effective way, and

(ii) that the concept of critical loads should serve as a guideline to formulate these strategies where science has provided the necessary information. 2.

THE CRITICAL LOADS APPROACH

For a deposited pollutant, we define the critical load as the maximum deposition that a target ecosystem can withstand without long term damage occurring. The critical load is thus one point in the damage-response curve for a particular pollutant-target combination and marks the point where, with increasing deposition load, damage becomes discernible. The critical load for a pollutanttarget combination is not a universal constant, it varies spatially depending on a range of meteorological, ecological, biological, geological and hydrological factors. These variations can be understood and estimated and mapped showing their distribution both nationally and internationally. The United Kingdom Department of the Environment has established the United Kingdom Centre for Critical Loads Mapping at the Institute of Terrestrial Ecology, Monks Wood and the work there has begun in earnest to work with many of the research institutes represented at this International Conference to collect the mapping data. Our National Focal Centre will assist the Co-ordination Centre for Effects which has been established at RIVM, Bilthoven, here in the Netherlands to assemble the maps from each European country and to produce the European critical loads and levels map for the complete range of pollutant-target combinations. The Critical Loads Approach is based on a comparison or overlaying of maps; maps of critical loads with those of deposition loads. These two sets of maps are essential components of the critical loads approach. The accurate mapping of deposition loads requires the folding together of several sets of environmental data. Ambient and precipitation concentration data have been provided for a wide range of pollutants for the United Kingdom from the national primary and secondary precipitation networks operated by Warren Spring Laboratory. Detailed information on the mechanisms of dry deposition, occult deposition and seederfeeder enhanced wet deposition has been assembled at the Institute of Terrestrial Ecology, Bush Estate. Together, concentration and mechanisms data have been employed to provide a comprehensive and quantitative assessment of deposition inputs of all acidic pollutants on a 20km grid across the United Kingdom. These have been overlaid with the corresponding critical load maps to produce the critical load maps to produce the critical loads exceedence maps. The third component of the Critical Loads Approach is the means by which deposition loads of each pollutant can be attributed to the sources of their emission into the atmosphere. The long term weather patterns bring different deposition loads to each point in Europe, from each of the major source areas. These weather patterns are tracked by the EMEP Meteorological Synthesising Centre-West of the Norwegian Meteorological Institute and using the EMEP models they are able to provide estimates of how the deposition in one country can be attributed to the emissions in all the other European countries. Such studies provide an important input both to the Critical Loads Approach and to the UN ECE Convention on the Long Range Transport of Air Pollutants. National modelling can be used to increase the spatial resolution of long range transport models but spatial resolution is ultimately limited by understanding of the vagaries of European weather conditions and of the atmospheric physical and chemical processes involved. Once the European Critical Loads maps are available, they can be overlaid with their corresponding deposition load maps to produce critical loads exceedence

255 maps. Long range transport models can then be used to provide a country attribution to the source of any exceedences. This information will be vital to the ongoing policy formation process in the international context. Each member country will be able to compare its commitments to emissions reductions in the future with the implications of the Critical Loads Approach. The United Kingdom is vigorously supporting the Critical Loads Approach through the mapping exercises, the environmental studies that underpin them and the investigation of the driving deposition mechanisms which lead to both pollutant removal and ecosystem contamination. The Department of the Environment has established a series of independent Review Groups to assess the significant of many of the scientific issues raised at this International Conference, to advise and recommend how the environmental science is to be advanced. Last year saw the publication of the third report from the Review Group on Acid Rain and the second report from the Photochemical Oxidants Review Group. In addition, the Critical Loads Advisory Group have made significant progress in assessing the effects of air pollution and acidic deposition in the United Kingdom. Critical loads maps for the acidification of soils have been published and the mapping of the sensitivities of freshwater ecosystems is well advanced. Maps have also been prepared to show how current estimated levels of acid depositions and sulphur depositions, compare with critical loads for soils. Where depositions exceed critical loads, long term environmental deterioration is likely to be occurring. There is a good correspondence between the areas of exceedence shown and areas where acidification damage has been reported, particularly in the upland part of Wales, Cumbria, Galloway and the Scottish Highlands. 3.

A

POLICY RESPONSE

This mapping work, and the similar work which has been carried out in other UN ECE countries, provides an invaluable new input into the domestic and international policy-making process. However, on its own, it cannot determine policy. There are a number of reasons for this. Pirst, damage by acid depositions is not limited to the acidification of soils and freshwaters and the consequent effects on ecosystems. There are effects on forests and on buildings to consider too. While intensive research continues, we have not yet identified dose-response relationships for these effects with sufficient confidence to be able to establish critical loads, and to map them, for forests and building effects. Nor, even in the case of soils and freshwaters, are we yet as clear as we need to be about the roles of nitrogen and ammonia in relation to that of sulphur. Second, a policy of bringing acid depositions down to critical loads in respect of all effects, in all parts of the United Kingdom or the UN ECE region as a whole, would be practicable, if at all, only in the very long term. Priorities therefore need to be identified and fnterim targets set. This process will require the application of economists as well as scientists' tools, to help evaluate damage and compare costs and benefits of abatement, though decisions must at the end of the day be a matter for political judgment. Third, it is only to be expected that individual countries' domestic targets for abatement will prove difficult to reconcile with each other, since emissions are differently distributed internationally from the damage to which they give rise. This will pose severe challenges for the international negotiating machinery, which we are already beginning to encounter. For the United Kingdom, the foundations of present policy, as regards stationary sources, are laid out in the Environment White Paper, as follows:

-

-

the European Community Large Combustion Plants Directive, with its requirements for major percentage reductions in SO2 and NO emissions from existing power stations and other boilers, and stringent emrssion standards for new plant, the NO protocol to the UN ECE Convention on Long Range Transboundary Air Pol1ut:on , requiring total national emissions to be brought back down to

256 1987 levels by 1994, -

the obligation laid on operators of industrial processes by Part I of the Environmental Protection Act 1990, itself derived from the EC Air Framework Directive, to use best available techniques not entailing excessive cost (BATNEEC) for reducing harmful emissions and minimising their impact.

In addition, the United Kingdom fully supports the EC agreement to Directives on emissions from cars and heavy diesel vehicles which will secure major reductions in (particularly) NO, emissions from this sector. The environmental effects of these policies can be forecast only within wide margins of uncertainty. However, projections based on conservative assumptions suggest that they should enable the proportion of the UK surface area exposed to depositions in excess of critical loads for soil acidification to be reduced to around 8% by the year 2005. As regards acidification of freshwaters, only 4% of surface area of Scotland will remain unprotected. These existing policies do not take specific account of the varying relationship between present deposition levels and critical loads in different parts of the UK. However, the mapping work which has now been done will help to inform their detailed application. In particular, the determination by the regulatory authorities of BATNEEC at individual plant level, and consideration of any proposals for major plant modification, will be assisted by knowledge of the likely effect on soils and freshwaters of acid emissions from the plant in question. Research is in hand to understand how market-based instruments can be employed in the development of strategies and policies which deliver environmental objectives. If an emissions trading or permit system could be devised which delivered critical loads, then this system has the important policy advantage of not providing licences to pollute, according to the definition of critical loads given above. 4.

Conclusions

All the maps of critical and target soils described above for the United Kingdom have been submitted to the UN ECE and other UN ECE countries have been preparing similar maps of their own. They will provide the basis for discussion of the proposed new UN ECE agreement to reduce emissions of sulphur dioxide, which is planned to be ready for signature by November 1992. Future progress with the UN ECE Convention on the Long Range Transport of Air Pollution and the revision of N O , SO2 and VOC protocols will rest in very large measure on the shared confidznce within Europe in the knowledge of the underpinning environmental science. The Critical Loads Approach should provide an important policy focus within the international scientific community to set environmentally-based targets for future co-ordinated emission control programmes. This International Conference on Acidification Research at Maastricht will be seen as an important step along the road to a series of agreements to limit the atmospheric emissions of acidic pollutants in Europe.

T Schneider (Editor). Acidification Research Evaluationand Policy Applications 0 1992 Elsevier Science Publishers B V All rights reserved

257

ACIDIFICATION POLICY IN THE UNITED STATES

D. Leaf Office of Air and Radiation, U S . Environmental Protection Agency, 401 M Street, SW, (ANR-445) Washington, DC 20460, USA.

Abstract Acidification policy in the United States has evolved over the past decade from one that emphasized scientific research and technology development to one that emphasizes pollution reduction through new emission limits on electric utility plants This paper discusses this evolution, and focuses on a discussion of the acid rain provisions of the Clean Air Act Amendments of 1990. The U S . - Canada Air Quality Agreement is also discussed. 1.

INTRODUCTION

Acidification policy in the United States has evolved over the past decade from an initial emphasis on scientific research on the causes and effects of acidic deposition and technology developments aimed at reducing pollutants to an emphasis on achieving emission reduction of acidic deposition precursors, specifically sulfur dioxide and nitrogen oxides. The two most significant policy developments of the last year are the signing into law of the Clean Air Act Amendments of 1990, which include a strong acid rain control program, and the signing of the U S - Canada Air Quality Agreement. This paper will briefly discuss the highlights of the last decade, then focus on the Clean Air Act's provisions for reducing acidic deposition. A brief discussion of the US.- Canada Air Quality Agreement is also included.

2.

ACIDIFICATION POLICY IN THE 1980's

In the 1980's there was a stalemate on the acidic deposition issue, based on a combination of economic concerns and scientific uncertainty regarding the scope and magnitude of the problem. During that time, over 70 different acid rain control bills were introduced in the U S . Congress, but none passed. Acid rain controls threatened economic interests in the iidustrial midwestern part of the country where most emission reductions were likely to occur, while most of the perceived effects occurred in the northeastern section of the country. Although it was clear that effects were occurring, we were not able to quantify the extent of the related damage. In 1980, the U.S. Congress created the National Acid Precipitation Assessment Program (NAPAP), and gave it a ten-year charge to research the causes and effects of acidic deposition, and to explore the costs and benefits associated with control policies. NAPAP was primarily a U.S. federal government interagency effort, which

258

included the Environmental Protection Agency, and the Departments of Agriculture, Commerce, Energy, and Interior. Funding for NAPAP research was funnelled through the individual agencies and departments, and totalled over $500 million over the ten year period. NAPAP represents the single most comprehensive study of an environmental issue in the history of the United States. NAPAP published 27 state of the science and technology reports on a variety of subjects, including environmental effects, visibility, health, atmospheric chemistry and transport, and control technologies and costs. NAPAP is about to publish an integrated assessment document that synthesizes the scientific reports, and provides comparative analyses of various pollution reduction scenarios. While the NAPAP effort was gearing up in the early and mid-I980's, the United Stated made a substantial commitment to develop technology aimed at achieving emission reductions. The commitment was made by President Reagan in 1986, following the publication of the report of the U S . and Canadian Special Envoys on Acid Rain. The U S . commitment was to a $5 billion, 5 year program of technology development aimed at achieving two goals: 1) Reduce emissions of acid rain precursors from coal-burning electric utility plants; and 2) thereby allow the United States to use its abundant supplies of high sulfur coal to produce energy. The support for the program also came from dissatisfaction on the part of many people with the then-available flue gas desulfurization systems, commonly referred to as scrubbers. While these systems were efficient in removing sulfur dioxide from exhaust gases, they resulted in large amounts of semi-solid waste, and they consumed a lot of electricity for their operation. The funding for the Clean Coal Technology program is split evenly between the federal government and private enterprise. Some of the projects that have been funded include pressurized fluidized bed combustion and integrated gasification combined cycle. Ongoing scientific research and technology development were set against a background that was often characterized as polar in nature between those who favored acid rain control and those who opposed it. In addition to the domestic aspects of the situation, the Canadian government was actively pushing for the United States to adopt acid rain controls in order to protect Canadian surface waters and other resources. 3. THE CLEAN AIR ACT AMENDMENTS CANADA AIR QUALITY AGREEMENT

OF 1990 AND THE

U.S.-

In February, 1989, President Bush made two pronouncements that would result in major policy shifts for the U.S. government regarding acidic deposition. First, in an address to Congress, he announced his intention to propose sweeping changes to the existing Clean Air Act, including acid rain controls. The President noted that the" time for study alone is over...the time for action is now." Second, in a visit to Canada, the President told Prime Minister Mulroney that the United States would work with Canada to negotiate a bilateral air quality agreement between the two countries.

259 The amendments to the Clean Air Act were debated throughout 1989 and 1990. They included new controls to reduce tropospheric ozone, air toxics, and acid rain. Nearly every part of the country and every economic sector will be affected by the Act. Congress passed the amendments by overwhelming margins, and the president signed them into law on November 15, 1990. Negotiations with Canada on air quality agreement were formally initiated in August of 1990 and ended in December. The agreement was signed by President Bush and Prime Minister Mulroney on March 13, 1991. The acid rain provisions of the Clean Air Act represent a measured response to a serious multi-faceted problem. The results of scientific research played an important role in the development of the program, particulary in terms of defining the scope and magnitude of the problem.. The Clean Air Act calls for a 10 million ton reduction in emissions of sulfur dioxide, which represents a 40 percent reduction from 1980 levels. Most of these reductions will come from the electric utility sector, which accounts for approximately 70% of sulfur dioxide emissions in the United States. Smaller amounts of reductions have already occurred or will occur in the transportation and industrial source categories. The acid rain control program will be implemented in two phases. The first phase commences in 1995, and will affect 261 units (individual boilers) at 110 coalburning electric utility plants in 21 eastern and midwestern states. Phase 2 commences in the year 2000, and affects a total of approximately 2500 units above 25 megawatts that use coal, gas, or oil in the production of electricity. In both phases, affected units will be required to install systems that continuously monitor emissions. Emissions from the utility sector will be permanently capped at just over 8.9 million tons per year, in order to safeguard environmental benefits from being eroded by increased energy consumption. The Act also calls for a 2 million ton reduction in emissions of nitrogen oxides, a reduction that will occur as a result of new emission limits and controls on electric utility boilers and automobiles. The U S . emits approximately 20 million tons per year of nitrogen oxides from a combination of stationary adn mobile sources. The Act provides for the use of an innovative market-based system of allowances in order to achieve the mandated sulfur dioxide emission reductions in a cost-effective manner. One allowance permits its holder to emit one ton of sulfur dioxide during or after a given year. Each affected unit will be allocated allowances of sulfur dioxide based on formulas contained in the legislation. Accounting of affected sources emissions by the government will take place at the end of the year. Each source must hold allowances at least equal to its emissions. The allowances can be either the original ones allocated to the source or ones that are bought from other sources. The law provides for a great deal of flexibility on the part of individual sources on how to meet their emission limits. There are no mandated emissions reduction technology requirements, which means that there is forced scrubbing. Compliance options include scrubbing, switching to low-sulfur coal, buying allowances to cover emissions, and pursuit of energy conservation options.

260 The legislation contains specific allocations for Phase 1 units. Phase 2 units will be allocated allowances by the Environmental Protection Agency based on formulas contained in the legislation. All electric utility sources that have 25 megawatts capacity will come under the program. There are a number of special allowance allocation provisions in the law that resulted from negotiations during debate on the Clean Air Act. The driving principle for the allowance system is that different firms have variable marginal costs of reducing emissions. From a market perspective, overall costs of the program can be reduced if different firms are allowed to trade among themselves to take advantage of the different marginal costs of control. It is estimated, for example, that the costs of the acid rain control program in the United States will be reduced approximately 20% using the market-based system instead of a traditional "command and control" approach that did not allow trading. The costs of the program with trading are expected to be approximately $1 billion a year in Phase 1, and up to $4 billion a year in Phase 2, so the potential cost savings are substantial. Here is an example of how the system will work. Plant A has a marginal cost of control of $800 per ton of sulfur dioxide removed, an allowance allocation of 5000 tons and emissions of 6000 tons. Plant B has a marginal cost of control of $400 per ton of sulfur dioxide removed, an allocation of 5000 tons, and emissions of 8000 tons. Plant B can choose to "over control" to get its emissions down to, for example 3000 tons, and sell the excess emission reduction credits to Plant A. The actual price of the allowances will be negotiated between the buyer and the seller. In theory, any price below $800 per allowance is attractive to Plant A. There are a number of mechanisms in the law to make allowances available to firms that have a difficult time finding allowances in the market place. These mechanisms include EPA-sponsored auctions and direct sales. Emissions will be monitored by means of highly accurate continuous emission monitors in order to assure that the reduction goals are met and that we have the best data available on what emissions actually are. The accounting of emissions and allowance will occur at the end of each calendar year. If the emissions of a source exceed the allowances held, the penalty is $2000/ton and the excess emissions are subtracted from the source's account in the following year. In the example above, if Plant A was allocated 5000 allowances, did not acquire any additional allowances, and emitted 6000 tons, the penalty would be significant. The fine would be 1000 X $2000 = $ 2 million. In addition the allocation in the next year would be reduced by 1000 tons (to 4000 tons) to offset the excess emissions. Allowances can be bought, sold, traded, and banked. Anyone can hold allowances; for example, some coal companies may buy allowances to use as incentives to buy high sulfur coal. The Chicago Board of Trade, a commodities trading market, recently announced its intent to conduct a futures market in allowances, lending quite a bit of credibility to the market-based approach. The Environmental Protection Agency will be in charge of the allowance system, and will perform the "bankers" function by settling up account at the end of the year. EPA does not, however, have to approve individual trades between companies-it just checks the source's accounts for the sufficiency of allowances and records those trades.

26 1 There are a number of provisions in the Clean Air Act relating to continuous emission monitors, permits, and new controls on nitrogen oxide emissions from electric utility plants. Continuous emission monitors will be required to be installed on all affected sources between 1993 and 1995. All units will have to have operating permits. In Phase 1, these permits will be issued by the federal government, which represents a significant departure from the historic practice of having states perform this function. In Phase 2, the permits will again be issued and administered by state governments. The nitrogen oxides control program will be implemented by requiring new emission restrictions for both mobile and stationary sources. The mobile source provisions are in the tropospheric ozone reduction provisions of the Clean Air Act. In the acid rain program, the legislation calls on electric utility boilers to meet new limits on emissions by applying low-NOx burner technology. EPA must promulgate regulations for different types of boilers between 1994 and 1997. Incentives in the law relate primarily to energy conservation and renewable energy, and repowering. Firms that employ energy conservation and renewable energy programs can benefit from a special 300,000 allowance reserve. Firms that repower their boilers with certain new clean coal technologies can qualify for a 4 year extension of the Phase 2 compliance date. Looking toward the future, our immediate goal is to promulgate the regulations that will implement the acid rain control program. Drafl regulations for public comment on allowances, permits, monitoring, and excess emissions were announced by EPA Administrator William Reilly on October 29, 1991. Other rules will be issued over the next one and one half years. The Environmental Protection Agency is responsible for implementation of the Clean Air Act.. The legislation call on the government to assess the effectiveness of the acid rain control program and to publish periodic reports. For example, one section of the law calls for a retrospective study and prospective studies of the costs and benefits of the entire Clean Air Act, including the acid rain provisions. The retrospective study is a one-time requirement, and the prospective studies are required to be performed every two years. The legislation also sets a out a number of requirements for assessing the environmental effectiveness of the program. In the acid rain area, much of the work will be carried out by various federal agencies, including EPA, and the Departments of Commerce, Interior, Agriculture, and Energy. Much of this work will be coordinated through the National Acid Precipitation Assessment Program. The environmental Protection Agency, for example, some of'the work that we will carry out includes the monitoring of wet and dry deposition, surface water chemistry, and visibility; atmospheric modeling; and the refinement of water chemistry models. The US.-Canada Air Quality Agreement that was signed in March of 1991 is another important development in North American acidification policy. The agreement provides the two countries with a practical, flexible instrument to deal with shared air pollution problems. The agreement sets up a bilateral air quality committee that will coordinate and report on the implementation of the general and specific obligations outlined in the agreement. There are provisions for consultations and referrals regarding air pollution problems. There are formal mechanisms for consultations on

262 new and existing pollution sources that may cause significant transboundary air pollution. There are two annexes to the general agreement, one that covers the requirements for reductions ion sulfur dioxide and nitrogen oxides in each country, and another that covers research and monitoring activities. U.S. and Canadian delegations recently met to discuss the structure and functions of the Air Quality Committee, and to discuss plans for writing the first report under the agreement (due in March, 1992). 4.

CONCLUSION

In conclusion, acidification policy in the United States has undergone a major change over the past decade. The U.S. government continues to support research into the causes and effects of acidic deposition and the commercial deployment of pollution reduction technologies. It has recently embarked on an ambitious pollution reduction program, and has negotiated an air quality agreement with Canada. The United States is implementing a market-based approach to pollution reduction on a national scale. It is anticipated that this program will result in cost-effective emission reductions without sacrificing environmental protection. The United States will monitor the effectiveness of the acid rain control program in producing environmental results, and the overall costs and benefits of the program. Finally, the United States and Canada have negotiated an air quality agreement that will lead to air quality and environmental improvements in our two countries.

SESSlOND NEW RESEARCH RESULTS ON THE ACIDIFICATION PROBLEM

This Page Intentionally Left Blank

T Schneider (Editor). Acidification Research. Evaluationand Policy Applications 1992 Elsevier Science Publishers B.V.

265

Setting priorities for the measurement of acid aerosols and gases: 3 examples from Switzerland Pamela Alean-Kirkpatrick and Jurg Hertz Swiss Federal Institute for Snow, Forest and Landscape Research (WSL), CH-8903 Birmensdorf. Switzerland

The ambient concentrations of acid aerosols and gases appear to be of increasing importance in connection with human health studies and calculations of the total deposition of acid and nitrogen. The results of three field studies in Switzerland, performed under different conditions and in various seasons show what information can be obtained with such concerted efforts, information which could possibly lead to a reduction in the overall effort necessary for long-term field measurements with non-continual measuring devices. 1. INTRODUCTION

The measurement of acid aerosols and gases has been awarded prominence in recent years, particularly in connection with human health studies (Lipfert et al. 1989, Lippmann 1989, Spengler et al. 1989,1990). In the US.considerable importance has been placed on the measurement of sulphuric acid (HzSO.4) and its partially neutralized form, ammonium hydrogen sulphate (NH4HS04) both of which are strong acids and occur as suspended particulates in the atmosphere. The negative effect of H2SO4 on lung function has been confirmed in a number of laboratory studies (Koenig et al. 1983, 1989, Ute11 et al. 1983, Spektor et al. 1989). In field studies, particularly on the east coast of the U.S., various "sulphuric acid events" have been characterized (Spengler et al. 19861, and measurements in connection with the "Harvard six-cities study" (Speizer 1989) have focussed on the measurement of "sulphuric acid equivalents" as being the dominant atmospheric acid with possible effects on the lung function of the inhabitants. Whether the same or even similar atmospheric chemistry is present in Switzerland is of current interest in connection with similar human health studies (Swiss National Research Project, NFP 26, Part A: Swiss Study of Air Pollution and Lung Diseases in Adults). Interest in these compounds is not only limited to health considerations: information on the concentrations of atmospheric acid and nitrogen compounds is needed for calculation of total deposition (actual loads), and knowledge of their temporal variation helps in the further understanding of chemical and physical processes. Since mea-

266

surements of acidic aerosols and gases cannot, as yet, be carried out routinely with commercially-available equipment, such undertakings are inherently time-consuming and costly. It is the aim of this paper to present some of the steps which could be undertaken to reduce the effort required to characterize the acidity of the air.

The first question t o be answered concerns which components should be measured. It is questionable whether aerosol (particle) acidity, sometimes termed "airborne acidity" (Ostro et al. 1991), is of relevance in Switzerland. Swiss SO2 concentrations have fallen considerably over the last few years and rarely reach the limits set by the Clean Air Bill (lOOpg/m3, 24h average (BUWAL 1991)). Summertime concentrations of SO2 are especially low (often <15pg/m3), and localized high emissions of SO2 in summer (when its oxidation to H2SO4 would be most favoured) from smelters, for example, are not characteristic of air pollution considerations in this country. It is a known fact that SO2 and H2SO4 can be transported over large distances, but the chances of H2SO4 being imported as such into Switzerland in summer are small; if washout does not occur, there is a somewhat high likelihood of neutralization en route by ambient ammonia, leading to the neutral particle (NH412SO4. This salt is particularly stable, and does not dissociate under ambient temperatures. In addition, studies in Canton Ziirich have shown that sulphate concentrations in wet deposition are higher in summer than in winter (Hertz et al. 1988). This is explained by the fact that the oxidation rate of SO2 to sulphate is highest in summer. Moreover, this oxidation in the liquid phase is pH dependent (the higher the pH, the faster the oxidation), and thus further enhanced by increased ammonia emissions in summer. At a first consideration, therefore, it is unlikely that Swiss concentrations of H2SO4 are high enough to warrant extensive measurements being made. To check this point, field measurements were carried out in Lugano TI during a summer smog episode in July 1991; this site was selected because the highest SO2 concentrations in summer have been measured in Canton Ticino (BUWAC 1991). Two of the more relevant components for determining the acidity of the air appear to be the gases HNO3 and NH3; in what way they affect lung function still has to be extensively investigated, however. HNO3 is produced under photoehemical conditions, originating from nitrogen oxides which have high emission rates in summer as well as in winter. In addition, this conversion of NO2 to HNO3 is rapid under photochemical conditions. NH3, a primary "pollutant", has a neutralizing effect an far as the acidity of the air is concerned, and its measurement simultaneous to that of the atmospheric acids is therefore important. Measurements of HNO3 and NH3 were carried out during the summer smog episode in Lugano, simultaneously at two sites (one urban and one semi-rural) 7 km apart in October and November 1990, and a t an Alpine site in April 1990 during a high-ozone event typical for this region in

267

Spring (Wunderli and Gehrig 1990). Average sampling times ranged from 4 hours to 24 hours, and their various merits are discussed in the results of the individual studies.

Gases: a n annular denuder system, similar to that developed by Allegrini e t al. (1987) was employed to measure HNO3 and NH3 concentrations. An 0.2% solution of NaCl in methanol was used to coat the denuders for HNO3 collection, and a 1.5% solution of citric acid in methanol for NH3 collection. Acid aerosols: the Allegrini system allows a filter pack to be connected after the denuders in order to collect particles. The acid aerosols were collected on a Gelman teflon filter (pore size lpm, fitted with a polymethylpentene support ring), which was extracted in a 10-4M solution of perchloric acid prior to measurement of pH (Koutrakis e t al. 1988). In order to correct for possible neutralization of the aerosols by the ammonia released after dissociation of NH4N03 (a process favoured after the removal of HNO3 and NH3 by the denuders), the teflon filter was followed by two other filters: a n NaC1-impregnated cellulose nitrate filter mounted in the same compartment of the filter holder as the teflon filter (the two filters were separated by the polycarbonate ring) to avoid any of the HNO3 released being adsorbed on the walls of the filter holder, and a citric acid-impregnated filter to collect the NH3 released. Both cellulose nitrate filters were extracted (15 mins. ultrasonic bath) in 10 ml pure water. From the difference in the HNO3 and NH3 concentrations, a correction can be made for the total H+ collected on the first filter (Slater et al. 1988). It is not possible with this method used to distinguish between H2SO4 and its partially neutralized form NH4HS04, since the procedure is based on the measurement of "free acidity" after filter extraction in the 10-4M perchloric acid solution. Chemicals and analysis: all chemicals were of p.A. grade and Millipore (MilliQ) water was used throughout. NO3- was analysed by ion chromatography (Dionex columns AS4A and AMMS-11, Wescan detector) and NH4+ by the indophenol method and FIA (Tecator FIAstar 5020). Pump system: air was pumped through the denuder system and filter pack a t 15 Vmin with a Gelaire CF 20 (Gruppo Flow SPA, Milan). At this flowrate, both HNO3 and N H 3 are collected in the denuders with an efficiency close to 100% (Allegrini 1987).

3.DESCRIPTION OF THE SlTES 1. Lugano: a medium-sized town (population 28,000, measurement site 280 m a.s.l.1, close to the Italian border and about 25 km N of the Po plain, is one of the tourist centres of Canton Ticino. This Canton on the southern side of the Alps

268

consists of several valleys which open towards the north-Italian Po plain, and boasts a dramatic relief, with large mountains on either sides of the valleys 2. Birmensdorf (semi-rural, 550 m a.s.1.) and Zurich (urban, 450 m a.s.1.): Birmensdorf is situated 7 km SW of the measurement site in Zurich and separated from the city by a 2 km-wide wooded hill. 3. Davos: a n Alpine ski resort renowned also for its specialist high-altitude clinics (permanent population c. 10,000, 1540 m a.s.l.1, is situated in a n Alpine NE-SW valley, 24 km east of Chur, the capital of the Grisons. 4. RESULTSANDDISCUSSION 4.1. A summer smog episode in Lugano, TI

From 8 July to 19 July, 24 hour averages (with 2 exceptions: see Figure 1) of particle acidity "H2SO4", HNO3 and NH3 were collected in order to assess the importance of particle acidity ("H2SO4") in relation to HNO3 and NH3. The results are presented in Figure 1.

Mdm3

H2S04 equivalent6 pdm3

HN03 pdm3

pgNH3/m3

Measurement period Fig. 1: Concentrations of "H2SO4", HNO3 and hW3 for 10 consecutive measurement periods. The values are all 24 hour averages, with the exception of Pl (12 hours, probably in the night, repeated electricity cute) and P3 (48hour average). Values of SO2 during this period ranged from 4 to 12 &m3

269

The results clearly show the low concentrations of particle acidity and the minor role played by H2SO4 when compared to HNO3 during this 10-day period. That negative values of H2SO4 were calculated from the analytical results is due to the varying values of the blanks in the filter stack used to measure particle acidity, and reflect the precision of the method; this fact does not, however, change the main result that particle acidity was of less importance than gaseous acidity in this time period. Gaseous acidity, on the other hand, deserves careful evaluation. In addition to the 24-hour sampling, 4-hoursampling of HNO3 and NH3 was carried out during the three last measurement periods. The full set of HNO3 and NH3 data is given in Figure 2, concentrations being given in nmol/m3.

nmol HN03/m3

nmol NH3/m3

nmol/m3 250.00

T

200.00

150.00 100.00

50.00

0.00

a Fig. 2: Comparison of HNOQand NH3 concentrations expressed in nmol/m3: a 1:l ratio represents neutralization. P1 to P10 are the same measurement periods as in Fig. 1 (24h) and those designated "A" are 4h averages run in the afternoon (13.0017.00), parallel to the 24h period with the same number. Ozone values, averaged over the same time periods as the sampling times ranged from 40 kg/m3 (P9)to 200 pg/m3 (P8A).

A comparison of the parallel sampling results in Figure 2 (e.g., 8 and 8A) reflect that the highest concentrations of HNO3 occur in the second half of the day and that the 24-hour average represents only about half the concentration which could be present during a 4-hour period in the afternoon of a day with high photochemical activity. The low 9 N 9 ratio is explained by the fact that a thunderstorm took place just before the start of the 4-hour sampling and rain continued until 15.30. This HNO3 4-hour-24-hour pattern is not seen for the NH3 concentrations: in all cases the 4-hour values are lower than the 24-hour averages, indicating that the peak concentrations of ammonia occur at some other time period than the afternoon. Since no other 4-hour measurements

270

were carried out, the period of peak NH3 concentrations remains speculative. However, the fact that the HNO3 peak and NH3 peaks do not occur a t the same time has implications for the mutual neutralizing effect of these two components. This "neutralizing factor" (only considering the components HNO3 and NH3 and not other components such a s S02,NO2 which also would contribute some acidity to the air) can be clearly seen in Figure 2. All 24-hour averages show an excess of NH3 - theoretically capable of neutralizing the ambient HNO3; all the 4-hour values show the opposite - a n excess of HNO3 over NH3. This study clearly shows the importance of the 4-hour sampling in the afternoon to capture the peak HNO3 concentration and to establish whether it concurs with the peak NH3 concentration or not. 49. An evaluation of local differtmcea: semi.ruralversus urban

In order to establish whether or not two sites (one urban, Zurich, and one semi-rural, Birmensdorf) a s close as 7 km to each other but with different emission situations exhibit similar air chemistry with regard to acid-determining species, simultaneous measurements of HNO3 and NH3 were carried out for 2 months in October and November 1990. ~~

Birmensdorf nmol HN031m3

w Zurich

I

30.00 25.00

20.00 15.00

10.00 5.00 0.00

3.0kt

10.0kt

17.0kt

24.0kt

31.0kt

8.Nov

15.Nov

22.Nov

29.Nov

30.00

15.00 10.00

5.00 0.00

1.Nov

Figs. 3A (upper) and 3B: HNo3 concentrations measured at Birmensdorf and Zurich city for 2 months in 1990 with 24-hour sampling times during the week and 72-hour sampling times at most weekends. (Weekends w i t h no sampling are shown by three "empty" sampling periods on the x-axis). Concentrations are given in nmol/m3. The horizontal bars mark two periods with higher concentrations in Birmensdorf.

27 1

24-hour sampling was performed during the week and for 72 hours over most weekends. A comparison of the HNO3 results from the two sites is shown in Figures 3A and 3B and the NH3 results in Figure 4A and 4B. Birmensdorf

Ziirich

nmol NH3/m3

m.00 200.00 150.00 100.00 50.00 0.00 3.0kt

10.0kt

17.0kt

24.0kt

31.0kt

-

100.00

II

--

50.00

1.Nov

8.Nov

15.Nov

22.Nov

29.Nov

Figs. 4A (upper) and 4B: NH3 concentrations measured at Birmensdorf and Zurich city for 2 months in 1990 with 24-hour sampling times during the week and 72-hour sampling times at most weekends. (Weekends with no sampling are shown by three "empty" sampling periods on the x-axis). Concentrations are given in nmol/m3. The horizontal bars mark the p e r i d where there is too little NH3 to neutralize the HNO3.

Overall, the concentrations measured are low (0.1 to 1.9 pg/m3 HNO3,O.l to 4 pg/m3 NH3). Figures 3 and 4 show that in almost all periods the concentrations of both components are higher in the city than in Birmensdorf (2 exceptions are shown in Fig. 3B), a fact which is particularly interesting with regard to ammonia concentrations; ammonia is usually associated with agricultural activities and not with urban atmospheric considerations, and the NH3 differences between the two sites is even more pronounced than those of HN03. As has already been seen in the results of the Lugano study, the 24-hour averages show a clear excess of N H 3 over HNO3 for almost all measurement periods. The exceptions are marked in Figures 4A and 4B. Two of these periods coincide with the exceptions in Figure 3B, and all three cases are characterized by low concentrations a t both sites. The sizeable differences in the concentrations between these two sites, separated by only 7 km, underline the fact that regional distribution of these

272

components cannot be easily predicted; relationships between two sites have to be established with such a period of measurement. Short, concerted efforts, measuring preferably for one month per season, would establish if general patterns between sites existed, perhaps afterwards reducing such measurements to only one site. The reasons for the differences found in this comparison, and the particular explanations for the exceptions marked in Figures 3 and 4 are not yet fully known. A further study of the meteorological data, together with a consideration of the concentrations of other atmospheric pollutants such as ozone and nitrogen dioxide is planned. 4 5 Temporalvariability in c o d o n with an ozone event in spring

In spring 1991, measurements of HNO3 and NT33 were performed with short sampling times (two 6-hour periods starting at 7.00 and one 12-hour period starting at 19.00) during a two-week period at a high-Alpine site, Davos. Tubular denuders (see Ferm 1986) were also run parallel to the annular denuders but with a one-week sampling period. The results of the first weeks measurements are shown in Figures 5 and 6. Figure 5 shows clearly the diurnal pattern in HNO3 concentrations with the maxima almost always occurring in the afternoon. These maxima cannot be estimated from the weekly average nor from the variability in the 0 3 concentrations. The extent of diurnal variation of HNO3 is greater than that of 0 3 due to the higher deposition velocity of the former, and of particular interest is the similar pattern of concentration changes shown by these two components. Given the low local emissions of NO2, both HNO3 and 03 are almost certainly regionally transported. Whatever the explanation for the occurrence of these high 0 3 concentrations in spring is (tropospheric folding, reservoirs in the upper troposphere), the results indicate that HNO3 could share the same source as 0 3 . On the other hand, the NH3 concentrations shown in Figure 6 do not display a clear diurnal pattern; in many periods the amount of ammonia sampled was below the detection limits of the analytical method. This is hardly surprising considering the low temperatures (average c. 0°C) and prevalent weather conditions (snow and rain setting in from 27 April onwards). The ammonia appears to be delivered in "bursts". A comparison of Figs. 5 and 6 shows that these bursts often occur at different times to the HNO3 peak concentrations and thus for many of the measurement periods, HNO3 is in "excess". Whether this particular situation is typical for an Alpine site in spring is not known, but it shows that the presence of HNO3 (albeit at low concentrations) at such sites under conditions where almost none would be expected on the basis of the emission situation - cannot be excluded.

273 -03

pg/m3

HN03/m3

-pg

-- 2.25

I-

1.00

pg NH3/m3 6.00

I:

4.00

1.00

0 0

0

Fig. 5 (above): HNO3 and 0 3 concentrations measured in Davos 22-29April 1991.The sampling times were two 6 hour periods (day) and one 12 hour (night) The horizontal dashed line shows the average H N O 3 concentration over the week measured by a tubular denuder. Fig. 6 (below):NH3 concentrations corresponding to the same collection periods as Figure 5 . However, such short sampling times require intensive work with the denuders, although it is clear that considerably more information can be obtained by sampling without breaks. I t may be enough to characterize a site with short, intensive field measurements in each season and thereafter only to measure the weekly averages with tubular denuders and 4-hour averages of H N O 3 and N H 3 with annular denuders in the afternoons. Whether this partic-

274

ular suggestion meets the requirements of the study needing this data can only be assessed in each individual case. 5. CONCLUSIONS

The results presented above have been obtained from field measurements i n the last year. Further work on their interpretation is needed. Nevertheless, several conclusions, based on these results alone, can be drawn at this stage: 1. Aerosol (particle) acidity does not appear to be play a n important role in Switzerland, particularly in comparison to gaseous compounds. 2. Concentrations of HNO3 and NH3 in cities usually exceed those in rural areas. 3. Concentrations of NH3 (in nmoUrn3) generally exceed those of HNO3 on a 24hour basis, whereas 4-hour measurements taken in the afternoon when the peak concentration of HNO3 is likely to occur show the opposite. 4. If acidity is an issue in human health studies, HNO3 and NH3 are the component of interest in Switzerland. In addition, if the short-term peak concentrations are needed for the assessment of acute effects the 4-hour measurements in the afternoon should be carried out using annular denuders.

Thanks are due to Judy T a j a n and Liesbeth Peier for helping with field work, and to Daniele Pezzotta, Janka Retzer, Monika Trachsler and Cornelia Weiss for analysing many of the denuder samples. The cooperation of the following organisations t h a t provided field facilities is gratefully acknowledged: Dipartimento dell'Ambiente, Bellinzona (G. Boffa, M. Camani, C. Sartori), Gesundheitsinspektorat der Stadt Zurich - Immissionsschutz (C. Leuenberger), and Landschaft Davos (G.P. Calonder).

Allegrini I., F. De Santis, V. Di Palo, A. Febo, C. Perrino, M. Possanzini and A. Liberti (1987) Annular denuder method for sampling reactive gases and aerosols in the atmosphere. Sci. Total Environ. 67,1-16. Ferm M. (1986) A NazCO3-coated denuder and filter for the determination of gaseous HNO3 and particulate NOS- in the atmosphere. Atmos. Environ. 20, 1193-1201. Hertz J., Bucher P., Furrer G., Keller L., Daniel 0. and T h h i L. (1988) Chemische Untersuchungen der atmosphhrischen Deposition. Chimia 42, 57-67.

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Koenig J.Q., Covert D.S. and Pierson W.E. (1989)Effects of inhalation of acidic compounds on pulmonary function in allergic adolescent subjects. Environ. Health Persped. 79,173-178. Koenig J.Q., Pierson W.E. and Horike M. (1983)The effects of inhaled sulfuric acid on pulmonary function in adolescent asthmatics. Am. Rev. Respir. Dis. 128,221-225. Koutrakis P., J.M. Wolfson and J.D. Spengler (1988)An improved method for measuring aerosol strong acidity: results from a nine-month study in St. Louis, Missouri and Kingston, Tennessee. Atmos. Environ. 22, 157-162. Lipfert F.W., S.C. Morris and R.E. Wyzga (1989)Acid aerosols: the next criteria air pollutant? Environ. Sci. Technol. 23,1316-1322. Lippmann M. (1989)Background on health effects of acid aerosols. Environ. Health Perspect. 79,3-6. NABEL (1991).LuRbelastung 1990:Messresultate des nationalen Beobach tungsnetzes fir LuRfiemdstoffe. Published by the Swiss Federal Office for the Environment (BUWAL). Ostro B.D., Lipsett M.J., Wiener M.B. and Selner J.C. (1991)Asthmatic responses to airborne acid aerosols. Am. J. of Public Health, 81,694-702. Slater J.L., P. Koutrakis, G.J. Keeler, J.M. Wolfson and M. Brauer (1988)Mea surement of atmospheric aerosol acidity: losses from interactions of collected particles. In: Proc. of the 1988 EPNAPCA Symposium on Measurement of Toxic and Related Air Pollutants, 176-181. Speizer F. (1989)Studies of acid aerosols in six cities and in a new multicity investigation: design issues. In: Symposium on the health effect of acid aerosols; October 1987;Research Triangle Park, NC. EHP Environ. Health Perspect. 79,61-67. Spektor D.M., Yen B.M. and Lippmann M. (1989)Effect of concentration and cumulative exposure of inhaled sulfuric acid on tracheobronchial particle clearance in healthy humans. Environ. Health Perspect. 79,167-172. Spengler J.D., G.A. Allen, S. Foster, P. Severance and B. Ferris, Jr. Pierson W.R., W.W. Brachacz (1986)Sulfuric acid and sulfate aerosol events in two U.S.cities. In: Aerosols, S.D. Lee, T. Schneider, L.D. Grant and P.J. Verkerk (Editors), Lewis Publishers, Chelsea, Michigan. Spengler J.D., G.J. Keeler, P. Koutrakis, P.B. Ryan, M.E. Raizenne and C.A. Franklin (1989)Exposures to acid aerosols. Environ. Health Perspect. 79,4351. Spengler J.D., M. Brauer, P. Koutrakis (1990)Acid air and health. Environ. Sci. Technol. 24,946-956. Ute11 M.J., Morrow P.E., Speers D.M., Darling J. and Hyde R.W. (1983)Airway responses to sulphate and sulfuric acid aerosols in asthmatics. Am. Rev. Respir. Dis. 128,444-450.

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T Schneider (Editor). Acidification Research Evaluation and Policy Applications 0 1992 Elsevier Science Publishers B V All rights reserved

211

High Resolution Assessment of Acid Deposition Fluxes W.A.J. van Pul, J.W. Erisman, J.A. van Jaarsveld and F.A.A.M. & Leeuw Laboratory for Air Research, LLO National Institute of Public Health and Environmental Protection, RIVM P.O. Box 1, 3720 BA Bilthoven, the Netherlands

Abstract In the framework of European and National policy making on abatement strategies there is a strong need for mapping actual and future deposition fluxes of acidifying components on a high spatial resolution scale. The method and results are presented of a mapping exercise carried out for the actual deposition on a 5x5km scale in the Netherlands in the framework of the Dutch Priority program on Acidification @PA). The mapping of acidifying components on a European scale is discussed.

1 Introduction In the framework of European and National policy making on abatement strategies there is a strong need for mapping actual and future deposition fluxes of acidifying components on a high spatial resolution scale. The actual fluxes or loads are necessary in the study on the relation between emissions and effects called the causal chain. In describing the effects of acidification on the level of ecosystems, the acid loads should be available at least on the same grid size as the critical loads i.e. generally in the order of a lOxlOkm resolution. No deposition maps on this resolution are available leaving a serious gap between the spatial resolution of the critical loads and the actual loads maps in Europe. Here a method and results are presented of a mapping exercise carried out for the actual deposition on a 5x5km scale in the Netherlands in the framework of the Dutch Priority program on Acidification @PA). The mapping of acidifying components on a European scale is discussed. 2 Mapping of acidifying components in the Netherlands

The acidifying components studied in the DPA were SO,, NOYand NH,. NHx is considered to be acidifying because of the nitrification processes in the soil in which H' is produced. Here it is considered that all components which are deposited will finally produce one

278

equivalent of acid leading to a maximum potential acid load. The total potential acid load is estimated from: potential acid = 2 SO, + NOy + NH,. The actual load of potential acid was estimated by a semi-empirical method based on concentration and flux measurements and meteorological observations. The method is described into detail by Erisman (1991), and is briefly adressed here. Wet deposition measurements, which are monitored at 14 locations in the Netherlands by means of wet-only samplers, were used for mapping wet &position via interpolation of yearly averaged values. The dry deposition flux on a 5x5km grid was estimated from air concentrations and a component specific deposition velocity at 50m height. Ground level observations of concentration and wind speed from the National Air Quality Monitoring Network at 40 (for NO,) to 80 (for SOJ locations were extrapolated hourly to a height of 5Om. At this height the concentration and wind were assumed not to be largely influenced by local terrain roughness. Therefore concentration and wind fields could be compiled in order to obtain data in areas where no measurements were available (i.e. natural areas). Concentrations of components not measured in the network were calculated with the TREND model (van Jaarsveld and Onderdelinden, 1991, for some TREND model features see below) such as for NH, (Asman and van Jaarsveld, 1990). or obtained from other measurements. The deposition velocity per component at 5Om was determined using a resistance model in which the surface resistance, r,, is component specific and depends among others on rain incidence, temperature and global radiation. Because of lack on receptor specific data, the same r, was used for the total model area. The aerodynamic and quasilaminar sublayer resistance were calculated locally using meteorological observations and

Figure I Roughness length, zo, map of the Netherlands on a Ixlkm grid in cm.

279

Figure 2a (lefi) Total potential acid deposition in the Netherlands on a SxSkm grid (smoothed out by the grey code presentation) andfigure 2b (right) exceedence of the median value of critical loads of total potential acid in the Netherlands on a IOxlOkm scale. Both figures for 1989 and in eq H’ ha” y.‘. roughness (i.e. zg) maps which were derived from detailed land use information on a lxlkm grid. So especially the effect of the surface roughness on the flux was modelled. The fluxes were calculated on an hourly basis and finally aggregated towards yearly averages. The wet and dry deposition were summed to yield the total deposition. Results for the roughness length, q,on a lxlkm scale for the Netherlands are shown in figure 1 (Erisman, 1990). The forested surfaces and cities are represented as black areas with a z, of about lm. The deposition of potential acid on the Netherlands is shown in figure 2a. Figure 2a shows that areas with high potential acid deposition coincides either with m a s with a high z,value (and therefore a high deposition velocity) or with areas with a high NH3 emission density. In the central part of the Netherlands, high q, values (forested area) and high NH, emissions occur simultaneously leading to the maximum acidifying deposition. The average fluxes in the Netherlands are listed in Table 1, along with the uncertainties in the total potential acid deposition derived with this method. The uncertainty is 44% on a 5x5km grid and 16% for the Netherlands as a whole. When a 100% correlation between the various errors in the variables is assumed, the total uncertainty is respectively 80% and 50%. This uncertainty on the local scale is dominated by the uncertainty in the NH, fluxes, where as for the Netherlands as a whole this is caused by the NH, and NO, fluxes.

280

Table 1 Deposition fluxes, in eq H+ ha.' y-', averaged over the Netherlands estimated for 1989 using the semi-empirical method. Uncertainties, in 8,in the deposition fluxes per component for 5x5km grids and the Netherlands as a whole; for the 5x5km grid this is split into systematic errors and random errors given between brackets. component

deposition flux

relative

eq H' ha.' y-I

error 8 Netherlands

1340

30 (11)

17

1160

61 (10)

25

2190

37 (87)

29

4700

44 (26)

16

With this method actual loads on ecosystems can be described and can be used in estimating exceedance maps. An example of such an exceedence map based on the median value for the critical loads in the Netherlands (taken from Hettelingh, 1991) is given in figure. 2b. The semi-empirical method described above has the advantage that local conditions which influence deposition fluxes can be accounted for. A disadvantage, however, is that no relation between emissions and deposition is made; this means that influences of local sources can be missed because of the limited density of the network. Moreover there is always a (undefined) risk that mass is not conserved. To check this point and to obtain estimates of contribution of source areas and/or categories, the atmosphere transport model TREND (van Jaarsveld and Onderdelinden, 1991) is applied. TREND is a long term model with a variable resolution in space (the grid size may vary from hundred metres up to hundreds of kilometres) and in time (concentrations averaged over one month up to one year). Meteo data and information on emissions are needed as input for the model. A comparison of the dry and wet deposition of SO,, NO, and NH,,calculated for 1980 with the TREND model and the semi-empirical method are given in figure 3. The deposition data are calculated for 20 acidification areas in which the Netherlands is divided (so called COROP areas, see Heij and Schneider, 1991). From the comparison it was concluded that in view of the assumptions made in both methods, the results obtained were. similar. With the TREND model scenario studies were made for the Netherlands (Schutter and de Leeuw, 1991).

28 1

Figure 3 Comparison of the dry and wet deposition of SO,, NO, and NH,,calculated with TREND and the semi-empirical method, here indicated with DEADM, per acidification area (COROP areas) in eq H' ha" y-'. The 1:1 line is indicated.

282 3 Mapping the deposition of acidifying components on a European scale The above method, based on the combination of model calculations and measurements along with the inference method, might be applied for mapping actual and future deposition fluxes on a European scale or for parts of Europe. As a fmt attempt the necessary concentration fields and wet deposition fields over Europe can be constructed from concentration and wet deposition measurements. However, the spatial resolution of the operational European networks (ECE-EMEP, EUROTRAC)is too coarse to provide the necessary data; on a local scale, national or local networks may provide the concentrationdata. When the local maps are aggregated towards one European map, problems may arise about the inconsistency between the networks and it is foreseen that still not a full coverage of Europe can be obtained. Therefore the concentration fields will be obtained via the TREND model and a possible integration with measurements will be investigated. The required detailed emissions for the model can be obtained from EMEP, COFUNE, GENEMISfiUROTRAC project or others. Note that the spatial resolution of the emission inventory and the spatial scale on which the deposition maps are required must be comparable unless small concentration gradients (that is at large distance from the source areas) are expected. The meteorological input for TREND will be obtained from a W M O synops data base archived by ECMWF which contains among others observations of wind speed and wind direction, global radiation, cloud cover and temperature. z, and r, maps over Europe have to be incorporated into the model domain from which a location dependent deposition velocity is calculated. These z, and r, values are constructed from land use maps over Europe. The z, values are assigned according to the surface type. The r, values are deduced from the surface characteristics using parameterizations taken from dry deposition experiments (i.e. projects in BIATEXEUROTRAC) and from literature. The dry deposition is calculated from the concentration fields and a location dependent deposition velocity which is derived from the z, and r, maps and the meteorological observations. The dry deposition can be calculated on a scale as detailed as the land use data are given. Effectively the dry deposition can be modelled on a more coarse grid using integrated z, and r, maps (say of about 60x60km scale) and where necessary can be zoomed in to describe the dry deposition on a more detailed grid (say 10xlOkm). The calculations of the deposition will be carried out on a monthly basis seperated in daily and nightly values. In this way the main time dependency of the deposition velocity and the concentration is accounted for. The monthly wet deposition field over Europe will be derived from TREND model calculations using the measured rain amounts in the national networks. The total deposition data on a yearly basis is aggregated from these monthly dry and wet deposition values. Preliminary results of deposition maps of total potential acid and total nitrogen on a 1' longitude x 0.5' latitude scale for Europe calculated with TREND are given in Figure 4. These calculations are based on one r, per component for the total model area. The model results have to be validated with long term deposition measurements derived from the various experimental sites over Europe.

283

Figure 4 The deposition of total potential acid (left) and totat nitrogen (right) in Europe for 1988 calculated with the TREND model in eq Ht ha” y-’. Taken from He0 and Schneider (1991). 4 Bottlenecks

One of the problems will be the availability of the concentration measurements and the consistency of the data obtained from the various networks. The same counts for the emission inventories. For the description of long range transport the meteo data is sufficiently available. However on a local scale, for the estimation of locally representative deposition velocities, the coverage of meteorological data is probably -not sufficient for large parts in Europe. As well orografk effects can induce large errors in the modelled deposition since these effects are not yet accounted for in the TREND model. The calculated dry deposition is very sensitive on the used r, value. For none of the above mentioned acidifying components, a full coverage on r, values for all types of surfaces and their time variations, exist. This is expected to be the largest error source. 5 References

Asman, W.and J.A. van Jaarsveld, 1991, A variable-resolution transport model applied for NH, in Europe. Atmospheric Environment in press.

284

Erisman, J.W., 1990. Estimates of the roughness length at Dutch Air Quality Monitoring Network stations and on a grid basis over the Netherlands. RIVM report 723001003. Erisman, J.W., 1991. Acid deposition in the Netherlands. RIVM report 723001002. Jaarsveld, van J.A. and D. Onderdelinden, 1991. An analytical long ~rm'depositionmodel for multi-scale applications. RIVM report 228603009 in preparation. Hettelingh, J.P., R.J. Downing and P.A.M. de Smet, 1991. Mapping critical loads for Europe. CCE technical report no.1. R N M report 259101001. Heij, G.J. and T. Schneider. 1991. Acidification research in the Netherlands, Final Repon second phase Dutch Priority programme on Acidification. Studies in Environmental Science 46, Elsevier. Schutter, M.A.A. and F.A.A.M. de Leeuw, 1991, Acid deposition in the Netherlands: Scenario results for 1994 and 2000. In Dutch with English summary. RIVM report 222101008.

T. Schneider (Editor), Acidification Research. Evaluation and Policy Applications @ 1992 Elsevier Science Publishers B.V All rights reserved

285

Measuring and modelling atmospheric dry deposition in complex forest terrain G.P.J. Draaijers', R. van EIP, W. Bleuten' and R. Meijersb

' Department of Physical Geography, University of Utrecht. P.O. Box 80.115,3508 TC Utrecht, the Netherlands. Environmental Forecasting Bureau, National Institute of Public Health and Environmental Protection, P.O. Box 1,3720 BA Bilthoven, the Netherlands.

Abstract A study is performed on the impact of forest stand structure on atmospheric dry deposition. Next to the impact of internal structure characteristics of forest stands (e.g. aerodynamic roughness, collecting surface area), also the impact of edge effects are studied. In this paper, methods of research, preliminary results and prospects for dry deposition modelling are discussed.

1. INTRODUCTION In long stretches of uninterrupted homogeneous forest, gases and aerosols reach the canopy surface by vertical transport from the air above the canopy. To a large extent, this vertical transport occurs because of turbulence generated at the top of the vegetation which, in turn, is controlled by the aerodynamic characteristics of the canopy. These aerodynamic characteristics are determined by height, density and species composition of the forest stand (hens, 1990). Gases and aerosols transported to the canopy by this kind of turbulence can deposit on needles, leaves, branches and trunks. Actual deposition amounts will depend on the total collecting surface area of these canopy elements and on their collecting efficiency (wiman et al., 1990). In western Europe extensive uniform forested areas are not common. Usually, a patchwork of different kind of forests, clearings, agricultural land and other land uses types make up the landscape. Consequently, many forest edges (i.e. transition zones) exist. Relatively more pollutants are found to deposit in the edge of a forest stand compared to its centre (Hasselrot & Grennfelt, 1987; Draaijers et al., 1988; Beier et al., 1989). This enhanced deposition near forest edges, also called the 'edge effect', may be attributed to several factors. First of all trees in forest edges are more exposed to pollutants due to an extra inflow of horizontal air masses. Secondly, windspeed and turbulence intensities in forest edges are enhanced compared to the centre of the stand. Thirdly, trees situated in forest edges may have a higher leaf surface area because more light and space is available. Finally, trees in the centre of a stand receive air masses with lower concentrations of air pollutants because these air masses have already been scavenged by trees in the forest edge. Deposition

286

increment in forest edges is believed to be determined to a large extent by the wind pattern and turbulence characteristics in and around the forest edge which, in turn, will be influenced by the height of the roughness transition and by the crown foliage density. Furthermore, actual deposition rates in a specific forest edge will depend on (prevailing) wind directions. A study is performed on the impact of forest stand structure on atmospheric deposition. A discrimination is made between the effect of the internal' structure of a forest stand (aerodynamic roughness, collecting surface ana etc.) on dry deposition amounts and the effect of the situation of a forest stand within the landscape (edge effects). Methods of research, preliminary results and prospects for dry deposition modelling are discussed in this paper. 2. METHODS OF RESEARCH

Dry deposition onto forestdforest edges is studied by sampling water dripping from canopies during rainfall (throughfall) and open field precipitation fluxes. This method relies on the assumption that precipitation removes previously dry deposited contaminants from the foliage. The major advantage of throughfall analysis is that dry deposition fluxes integrated over entire canopies and over long periods can be obtained. The method is relatively inexpensive and easy to implement. Furthermore, the analysis may be applied to non-homogeneous canopies and complex terrain which is important when, for example, studying dry deposition gradients in forest edges. In this case, other dry deposition measurement techniques such as eddy correlation and gradient methods are less suitable because they require uniform vegetation with adequate fetch @avidson & Wu, 1990). Major disadvantage of the throughfall method is that the difference between throughfall and precipitation flux (net throughfall) results from both dry deposition and canopy exchange. If pollutants are taken up irreversibly or leached out by the canopy foliage, estimates of dry deposition may be obscured. Several methods are developed to estimate the magnitude of these canopy exchange processes. An overview of these methods is given by Van Ek 8c Draaijers (1991). Based on "S experiments and comparisons of throughfall data with model prediction of dry deposition on a large number of sites throughout the United States (Integrated Forest Study), Lindberg et al. (1990) concluded that foliar leaching of SO,'- contributes less than 5% to the net throughfall in heavily SO2 polluted areas. From the same study, it was concluded that on average approximately 40% of all nitrogen (NO, and NH,) input to forests was taken up by the vegetation, whereas 60% of the nitrogen deposited was found back in the throughfall data. It was suggested that anorganic substances taken up by the foliage may partly be released again as organic nitrogen through canopy leaching. When extrapolating these results to western Europe, it should be kept in mind that air concentrations of nitrogen containing pollutant in the United States are usually far less compared to westem European levels. It may be expected that a tree species occurring in different pollution climates and ecological settings will show different responses with respect to canopy exchange. For this reason, Van Breemen et al. (1988) and Ivens (1990) suggest that for many forests in western Europe canopy exchange of both sulphur and nitrogen will be relatively small compared to atmospheric input.

287

To study the relationship between atmospheric dry deposition and the internal forest stand structure, an intensive throughfall monitoring program is performed in the middle (excluding edge effects) of thirty different forest stands (Figure 1). Nine measurement sites are situated in Douglas fir (Pseudotsuga memiesii Mirb. Franco) stands, ten sites in Scotch pine (Pinus sylvestrisL.) and eleven sites in Oak (Quercw robur L.) stands. In addition, throughfall is monitored in eight forest edges at respectively 5 , 10, 20, 30, 40, 60,80 and 150 m distance form the forest edge. Forest edge measurements are conducted in three European Larch (Lurix decidua Mill) stands, three Scotch pine (Pinus sylvestris L.), one Corsican pine (Pinw nigra var. maritima) and one Norway spurce (Picea abies L.) stand. All edges have south-west exposition which is the prevailing wind direction in the Netherlands. More or less undisturbed fetch over heather, grassland or arable land is found in front of the edges.

Figure 1. Location of the thirty measurement sites All forest stands and forest edges investigated are located in a forested area in the central part of the Netherlands, called ’Utrechtse Heuvelrug’. Large source areas for SO2 and NO, are located 200 km SE (industrial Ruhr-area) and 1 0 0 h SW

288

(Rotterdam port) from the Utrechtse Heuvelrug. NO, also originates from roads which cross the Utrechtse Heuvelrug. The forested area is enclosed by two agricultural NH, source areas, namely the 'Gelderse Vallei' and the 'Kromme Rijn' area. It is expected that the thirty measurement sites have approximately the same air pollution load as they are situated within a radius of 1.2 km of each other. Throughfall is sampled during one full year by means of 5 m long gutters. Bulk precipitation is sampled in several clearings by means of funnels. No stemflow measurements are performed as stemflow contributes less than 10% of the total input to the forest floor for the tree species under examination (Ivens, 1990; Jansen, 1991). A detailed description of field and laboratory procedures can be found in Draaijers et al. (1991a). Around each throughfall gutter an inventory of relevant tree and stand structure characteristics is made to determine parameters which reflect the aerodynamic roughness and collecting surface area of the canopy. An important parameter often used to characterize the aerodynamic roughness of a canopy is the so-called roughness length (23. This parameter can be derived from solution of steady-state wind profile equations above the canopy, but can also be estimated from empirical relationships which exist with canopy structure characteristics. Often the roughness length is directly related to the mean tree height (e.g. Jarvis et al., 1976). This relationship is clearly too simple as it can be hypothized that the size and density of the trees also determine the aerodynamic roughness of a stand. For this reason, Lettau (1969) and Lovett & Reinders (1986) suggest an equation in which next to mean tree height also the mean silhouette area of trees (= crown projection area normal to the wind direction), number of trees per ha and a species specific drag coefficient (depending on the porosity and flexibility of the crown) are incorporated. The most direct measure for the collecting surface area is the needle/leaf area index (LAI). Needledleaves normally contribute more than 85% to the total surface area (Halldin, 1985; Hutchinson et al., 1986). For the coniferous tree species, needle area is measured through the functional relationship with stem conducting tissue (sapwood area) of the trees. In case of deciduous trees, leaf area is determined by sampling leaves with litter traps in autumn. An indirect measure for the collecting surface area is the crown cover. Crown cover directly above each throughfall device is determined through field estimates and through scanning of panchromatic images using ERDAS digital processing techniques. The collecting efficiency of individual needles or leaves depend on their stickiness, size and shape. A good wettability and the presence of hairs is found to enhance the collecting efficiency. Furthermore, small 'needlelike' surfaces appear to be aerodynamically more effective in filtering gases and aerosols than larger 'leaflike' structures (Shuttleworth, 1977; Goodman, 1985; Wiman et al., 1990). For this reason, information is gathered on the presence of hairs and on mean diameter (width) and total surface area of individual needles and leaves. In case of forest edges, special attention is paid to characterizing edge structures. Relative height and porosity of the forest edge will determine flow circulation and turbulence patterns to a large extent. Gradually increasing tree height with distance to the forest edge is found to be quite common, especially in the first 20 m of forest edges. This may also influence edge aerodynamics to some extent. With respect to porosity a discrimination is made between trunk and canopy layer porosity. Porosity is determined by estimates in the field and by scanning of panchromatic images made by photography from inside the forest looking outside.

289

3. PRELIMINARY RESULTS Net throughfall of acidifying compounds is found to vary greatly between the thirty forest stands. Measured maximum net throughfall fluxes of SO,", NO,- and were 3.1,7.4 and 5.5 times higher compared to minimum fluxes, respectively. Variability between the thirty measurement sites was rather consistent throughout the measurement period, except for SO,'., MgZ+,K+ and PO,'. in oak stands which show relatively high net throughfall fluxes in the summer period. The extremely high summer phosphate fluxes in Oak stands (16.1 meq/m2.0.5yr) have not been reported earlier and may be an indication that Oak trees were suffering some kind of stress. Indeed leaves were in relatively bad condition as they were infected by oak mildew, a fungal disease caused by the fungus Microshaera aliphilitoides. Foliar wash experiments performed on infected oak leaves have shown that this mildew infection is indeed responsible for enhanced leaching of sulphate, magnesium, potassium and phosphate (Van Ek & Draaijers, 1991). In the winter period, on average Douglas fir stands were found to have the highest net throughfall fluxes for the acidifying compounds SO:-, NO,' and NH4+. Oak stands showed the lowest and Scotch pines intermediate fluxes. In the summer period, for nitrate and ammonium the same sequence was found, but due to already mentioned enhanced leaching, relatively large sulphate fluxes were recorded in Oak stands in this period (Van Ek & Draaijers, 1991). As canopy exchange is believed to influence throughfall fluxes to a lesser extent in the winter period, relationships with canopy structure characteristics are studied using winter net throughfall fluxes. More than 70% of the variation in winter net throughfall of SO,'., NO3-and NH.,+ could be explained by differences in aerodynamic roughness length of the canopy (Table 1). With respect to net throughfall of NO,. this can be explained by assuming dry deposition of HNO,(g) and NO, aerosol to be important contributors to the flux as dry deposition of these compounds is believed to be determined to a large extent by atmospheric transfer to the receptor surface (Fowler et al., 1989). Net throughfall of SO:- and NH,+ in the Netherlands is

m+

Table 1 Linear correlation coefficients between winter net throughfall fluxes of acidifying components and several (winter) forest structure characteristics (n=30; ***: p < 0.001; **: p c 0.01)

sod2Roughness length Crown cover LA1 Mean tree height

0.84'" 0.69"' 0.59"' 0.6 1*.*

0.84"' 0.78"' 0.77"' 0.50 **

0.88"' 0.77"' 0.76"' 0.51 **

290

controlled to a large extent by dry deposition of SOz(g) and NH,(g), respectively. Especially dry deposition of SOz(g) is generally believed to be controlled by the stomatal or surface resistance (Fowler et al., 1989). Therefore, it was rather surprising that such strong relationship was found between net throughfall of SO,'- and the roughness length of the canopy. We suggest that the wet and alkaline pollution climate in the Netherlands is responsible for a significant reductibn of the surface resistance of SO2, causing deposition rates to be controlled to a large extent by the aerodynamic characteristics of the canopy (Figure 2).

200

SO4 net throughfall (meq/m2'0.5yr)

Scotch plne

0

1

2

3

4

5

roughness length (m) Figure 2. Relationships between winter net throughfall of SO,'- and aerodynamic = 34.49*[~0]+20.13;r=0.85; n=30). roughness length (za) of the canopy ([SO:-] Roughness length computation is based on an emperical formula, first proposed by Lettau (1969), using several forest stand structure characteristics. Parameters reflecting the collecting surface area of the canopy (LAI, crown cover) also correlated significantly with winter net throughfall of SO:-, NO3- and NH,'. An uni-modal relationship between dry deposition and LA1 has been suggested by e.g. Lovett & Reiners (1986) and Ivens (1990), but in this study the relationship between net throughfall of acidifying compounds and LA1 turned out to be linear in the LAI range of 0 to 12. Due to a strong interrelation between collecting surface area properties and the aerodynamic roughness length of the canopy, the magnitude in which both properties influence atmospheric dry deposition seperately, is hard to quantify. More research on this topic is necessary for a better understanding of the underlying processes.

29 I

The collecting efficiency of Douglas fir and Scotch pine needles are found to be comparable to each other and higher compared to Oak leaves (Van Ek &

Draaijers, 1991). However, the impact of these differences in the overall deposition process of acidifying gases seems to be negligible. Collecting efficiency is found to be more important in case of deposition of aerosols. More information on assessing collecting efficiencies and on the impact of the collecting efficiency of individual canopy elements on dry deposition of gases and aerosols to forest canopies is given by Van Ek & Draaijers (1991).

Scotch pine

0

2

4

6

*Corsican

8

10

pine

12

distance to edge / edge height

Figure 3. Relationship between net throughfall of Na+ and distance to the leading edge as found in a Scotch pine and Corsican pine forest edge. Next to the collecting efficiency of individual canopy elements, also edge effects are found to have large impact on aerosol deposition amounts. This is illustrated by pronounced net throughfall gradients for Na' (Figure 3) and C1- found in forest edges. Net throughfall of these ions is believed to be controlled by dry deposition of sea-salt aerosols. Impaction and interception are known to be the dominant deposition mechanisms for aerosols. Both mechanisms are strongly influenced by windspeed (Davidson & Wu, 1990), a parameter strongly enhanced in forest edges. Net throughfall of SO:-, NO,' and NH4+ show similar but less pronounced gradients in forest edges. Net throughfall of these ions is controlled to a large extent by gas deposition. In the Netherlands, aerosols are found to contribute less than 25% to the total deposition of acidifying substances (Erisman, 1991).

292

Net throughfall is found to decrease with distance to the forest edge, but in

case of dense canopies show a slight temporary increase further downwind (Figure 3). This slight increase is most probably caused by a turbulent wake reaching the ground at several edge heights out. Closer to the edge, there is a ’quiet zone’ occupied by a recirculating eddy. Mean wind speed and turbulent downward transport of pollutants are suppressed in this zone and enhanced in the ’wake zone’. The initial decrease is more pronounced in case of dense edge vegetation (Draaijers et al., 1991b).

4. PROSPECTS FOR DRY DEPOSITION MODELLING In present-day research on atmospheric deposition onto forests either great attention is paid on understanding deposition processes at few forest research sites or deposition is regarded on a large scale (regional, national, continental) in which forest is one of the land use types on which deposition can occur. So far little attention is paid to extrapolation of results from point measurements. Furthermore, large scale models are too general approach to yield accurate estimates of deposition to specific forest stands. Results from our research indicate that dry deposition of acidifying compounds to specific forest stands is determined to a large extent by internal forest stand structure characteristics like e.g. aerodynamic roughness length of the canopy and total leaf area and by edge effects. Therefore, incorporating more detailed information on these forest structure characteristics in present-day deposition models will offer prospects to estimate deposition fluxes to individual forest stands or catchment areas. These more precise models may be useful to assess the effects of atmospheric deposition on soil and groundwater quality and may be used to determine critical loads more accurately. Detailed spatial information on forest stand structure characteristics can be obtained through forest inventories or by using remote sensing techniques. In the Netherlands, for example, approximately every 15yr a detailed forest inventory is made through which information on tree species, mean tree height and crown cover is available on stand level. Present applications of remote sensing in air pollution research have focused on directly visible effects of pollution, such as defoliation and yellowing of vegetation (e.g. Westman & Price, 1988; Vogelmann & Rock, 1986). Reasonable relationships are found between Landsat Thematic Mapper (TM) images and LAI (Peterson et al., 1987). Also tree species and mean tree height can be detected from satellite data (Meijers et al., 1990). Potentially, remote sensing seems to be a useful tool for determining stand structure characteristics relevant from viewpoint of deposition modelling, especially when a combination of multi-temporal Landsat and Spot-images is used (Meijers et al., in prep.). However, the usefulness of satellite images to detect stand/forest edges has to be established yet. Deposition increment due to edge effects is found to depend strongly on variables like the length and exposition of existing edges, the height difference between the different edges and the ratio between the edge area and the total surface area of the stand (Meijers et al., 1990). Application of a Geographical Information System (GIS) makes it possible to use the actual position of a stand within the landscape, necessary to calculate above mentioned variables.

293

5. REFERENCES

Beier, C. & P. Gundersen (1989) : Atmospheric deposition in a spruce forest edge in Denmark. Environmental Pollution, 60,257-271. Davidson, C.I. & Y.L. Wu (1990): Dry deposition of particles and vapors. In: S.E. Lindberg, A.L. Page & S.A. Norton (4s): Acidic preparation: sources, deposition and canopy interactions, Springer Verlag, 103-216. Draaijers, G.P.J., W.P.M.F. hens & W. Bleuten (1988): Atmospheric deposition in forest edges measured by monitoring canopy thoughfall. Water Air and Soil Pollution, 42, 129-136. Draaijers, G.P.J.,R. van Ek & R. Meijers (1991a): Research on the impact of forest stand structure on atmospheric deposition: methods and preliminary results. Environmental Pollution (in press). Draaijers, G.P.J.,R. van Ek & W. Bleuten (1991b): Research on atmospheric deposition gradients in forest edges and impact on soil and groundwater quality. J. of the Austrian Hydrological Service (in press). Erisman, J.W. (1991) : Atmospheric deposition in the Netherlands. National Institute of Public Health and Environmental Protection, The Netherlands, report 723001002. Fowler, D., J.N. Cape & M.H. Unsworth (1989): Deposition of pollutants on forests. Phil. Trans. SOC. London B324, 247-265. Goodman, J. (1985): The collection of fog drip. Water Resources Research, 31, 392394. Halldin, S. (1985): Leaf and bark distribution in a Pine forest. In: B.A. Hutchkinson & B.B. Hicks (eds): The forest-canopy interaction. Hasselrot, B. & P. Grennfelt (1987): Deposition of air pollutants in a wind exposed forest edge. Water, Air and Soil Pollution, 34, 135-143. Hutchinson, B.A., D.R. Matt, R.T. McMillen, L.J. Gross, S.J. Tajchman & J.M. Norman (1986): The architecture of a deciduous forest canopy in eastern Tennes~ e e ,USA. J. Of Ecology, 74,635-646. Ivens, W.P.M.F. (1990): Atmospheric deposition onto forest. Ph.D. thesis, University of Utrecht. Jansen, L. (1991): Stemflowvolumes, -concentraties en -fluxen in relatie tot boom-, opstand- en kiimaatspecifieke factoren (in dutch). Ms.C. thesis, University of Utrecht. Jarvis, P.G., G.B. James & J.J. Landsberg (1976): Coniferous forest. In: J.L. Monteith (ed) Vegetation and the atmosphere. Academic press, London. Lettau, H. (1969): Note on aerodynamic roughness-parameter as timation on basis of roughness-element description. J.App1. Meteorology, 8, 828-832. Lindberg, S.E.,G.M. Lovett, K. Knoer & H. Ragsdale (1990): The integrated forest study: Atmospheric deposition and canopy interactions of sulphur, nitrogen and cations. Proceedings NAPAP 1990 International Conference on 'Acidic deposition: State of Science and Technology', Hilton Head Island, USA. Lovett G.M. & W.A. Reiners (1986): Canopy structure and cloud water deposition in subalpine coniferous forest. Tellus 38B, 319-327. Meijers, R., G.P.J. Draaijers & R. van Ek (1990): Predicting atmospheric deposition onto forests using remote sensing techniques: a local scale model. Poster presented at the International Conference on 'Acidic Deposition: its nature and impacts', Glasgow, 16-21 September 1990.

294

Meijers, R., G.P.J. Draaijers & R. van Ek (in prep.): Assessment of forest structure characteristics from monotemporal landsat TM for application in atmospheric deposition modelling. Remote Sensing and Environment. Peterson, D.L.,M.A. Spanner, S.W. Running & K.B. Teuber (1987): Relationship of Thematic Mapper Simulator data to leaf area index of temporate coniferous forests, Remote Sensing and Environment, 22, 323-341. Shuttleworth, W.J. (1977): The exchange of wind-driven fog and mist between vegetation and the atmosphere. Boundary-layer Meteorology, 12,463-489 Van Breemen, N., W.F.J. Visser & Th. Pape (1988): Biogeochemistry of an oakwood land ecosystem in the Netherlands affected by acid atmospheric deposition. Pudoc, Wageningen, the Netherlands. Van Ek, R. & G.P.J. Draaijers (1991): Atmopsheric deposition in relation to forest stand structure. Dept. of Physical Geography, University of Utrecht, The Netherlands, report AD 199141. Vogelmann, J.E. & B.N. Rock (1986): Assessing forest decline in coniferous forests of Vermount using NS-001 Thematic Mapper Simulation data. Int. J. of Remote Sensing, 7, 1303-1321. Westman, W.A. & C.V. Price (1988): Detection air pollution stress in southern California vegetation using Landsat Thematic Mapper band data. Photogrammetric Engineering and Remote Sensing, 54, 1305-1311. Wiman, B.L.B., M.H. Unsworth, S.E. Lindberg, B. Berghist, R. Jaenicke & H.C. Hansson (1990): Perspectives on aerosol deposition to natural surfaces: interaction between aerosol residence times, removal processes, the biosphere and global environmental change. J. Aerosol Science, 21, 313-338.

T. Schneider (Editor). Acidification Research. Evaluationand Policy Applications 0 1942 Elsevier Science Publishers B.V. All rights reserved

295

T h e t r a n s p l a n t a t i o n of f o u r s p e c i e s of L o b a r i a l i c h e n s t o demonstrate a field acid rain effect

A.M.

Farmer.,

J.W.

B a t e s b and J.N.B.

Bellb

^ E n g l i s h N a t u r e , N o r t h m i n s t e r H o u s e , P e t e r b o r o u g h , PE1 l U A , UK.

b D e p a r t m e n t of B i o l o g y , I m p e r i a l C o l l e g e , S i l w o o d P a r k , A s c o t , B e r k s . S L 5 7PY, U K .

Abstract T h e four B r i t i s h L o b a r i a lichen s p e c i e s ( L . amplissima, L . pulmonaria, L . s r r o b i c u l a t a and L. v i r e n s ) w e r e transplanted f r o m a s i t e in W S c o t l a n d w h e r e all a r e a b u n d a n t t o t w o s i t e s in SW S c o t l a n d a n d N W E n g l a n d w h e r e t h e s e s p e c i e s had s h o w n a d e c l i n e in a b u n d a n c e a n d w h i c h h a v e l o w l e v e l s o f g a s e o u s p o l l u t a n t s but h i g h l e v e l s o f w e t d e p o s i t e d a c i d i t y . Previous w o r k had s h o w n a d v e r s e e f f e c t s o f l o w p H o n t h e s e l i c h e n s , but c o u l d n o t e l i m i n a t e a p o s s i b i l i t y t h a t h i g h SOz l e v e l s a t a n e a r l i e r period m a y h a v e c a u s e d t h e d e c l i n e . T h e lichens transplanted t o t h e s e s i t e s showed a d e c l i n e in health, t h o u g h s p e c i e s d i f f e r e d in s e n s i t i v i t y . C o n d i t i o n s a r e s t i l l n o t favourable t o g r o w t h of t h e s e lichens and it i s concluded that w e t d e p o s i t e d a c i d i t y r e m a i n s t h e m o s t likely e x p l a n a t i o n f o r t h e i r d e c l i n e , p a r t i c u l a r l y a t t h e s i t e in NW E n g l a n d . 1. INTRODUCTION A i r pollution h a s caused t h e d e c l i n e o f many lichen s p e c i e s in B r i t a i n . Plmong t h e m o s t s e n s i t i v e l i c h e n s p e c i e s a r e t h o s e of t h e Lobarion pulmonariae a l l i a n c e ( J a m e s et a l . , 1977). H i g h SOz l e v e l s h a v e c a u s e d t h e loss o f s p e c i e s f r o m t h i s c o m m u n i t y f r o m many p a r t s o f E n g l a n d , W a l e s a n d s o u t h e r n Scotland. Concern h a s a l s o been expressed that t h e Lobarion i s d e c l i n i n g in a r e a s r e m o t e f r o m h i g h l e v e l s o f g a s e o u s a i r p o l l u t a n t s , but r e c e i v i n g h i g h i n p u t s o f w e t d e p o s i t e d a c i d i t y ( G i l b e r t , 1986; Rose, 1988; H a l l i n g b a c k a n d M a r t e n s s o n , 1 9 8 7 ) . It i s suggested that t h e Lobarion lichens a r e being replaced in t h e s e a r e a s by a c i d - t o l e r a n t s p e c i e s o f t h e P a r m e l i o n l a e v i g a t a e alliance. This paper considers three contrasting s i t e s in N W B r i t a i n , w h i c h d i f f e r in t h e s t a t u s o f t h e L o b a r i o n and l e v e l s o f w e t a c i d i c d e p o s i t i o n . It d e s c r i b e s a transplantation study which, together w i t h o t h e r work carried out at these sites, gives strong support t o t h e hypothesis that

296 wet deposited acidity is involved in the decline of these species. 2. METHODS

Characteristics of the three sites are given in Table 1. Arderey Wood at Loch Sunart supports very healthy populations of the Lobarion. P l t Old Garrock Wood, Glenlee, L. pulmonaria is the only species of Lobaria found, while at Seatoller Wood, Borrowdale, Lobaria species no longer occur on Q. petraea (Day, 1985). Table 1

Details of the sites used for the transplant study. ~

~~

~~

~~

Loch Sunart

Glenlee

Bor rowd a 1 e

Wood 1 and

CIrderey

Old Garrock

Seatoller

Grid Ref.

17/75.62

25/5 9 . 8 2

35124.14

Mean Annual SO= p p b +

<2.0

12.0

2 .O-4.0

Mean Annual N0z p p b l

<4.0

4.0-8 .o

4.0-8.0

Wet deposited 0.02-0.03 annual acidityL g H' m-=

0.04-0.05

0.04-0.05

Mean bark PHZ

4.68

5.06

4.27

Mean stemflow pHz

5.66

5.37

4.27

+From the Warren Spring National Monitoring Network, Williams e t a l . (1989). =From Farmer e t a l . ( 1 9 9 1 b ) Specimens of all four British Lobaria species were collected from Q. petraea trees at Loch Sunart in November 1988. One thallus o f each species was fixed (using "Super Glue" gel) to the receiving tree (all 0. petraea) at the corner of a square design at breast height (Figure 1). This method o f attaching lichens has been previously shown to be non-damaging (Richardson, 1967; Hallingback, 1990; Gilbert 1991) and allows the thalli to be in intimate contact with the bark environment of the receiving tree. The positions of the species in the configuration and the trunk aspects were randomised on each

297 tree. Transplants were made t o four control trees at Loch S u n a r t , w h i c h a l r e a d y s u p p o r t e d L o b a r i a spp., and t o e a c h of f o u r r e p l i c a t e t r e e s a t t h e G l e n l e e and B o r r o w d a l e s i t e s . Regular visits were made t o each site until J u n e 1991. O n e a c h o c c a s i o n t h e t h a l l i w e r e w e t t e d w i t h de-ionised w a t e r and any d i s c o l o u r a t i o n r e c o r d e d a s a p e r c e n t a g e of t h e total thal lus area.

F i g u r e 1. A t y p i c a l t r a n s p l a n t c o n f i g u r a t i o n . Species are ( c l o c k w i s e f r o m t o p left): L o b a r i a s c r o b i c u l a t a , L . a r n p l i s s i m a , L . pulmonaria and L . v i r e n s .

3 RESULTS T a b l e 2 p r e s e n t 5 t h e f i n a l h e a l t h s c o r e s o f t h e t h a l l i of e a c h L o b a r i a s p e c i e s f o r e a c h s i t e , t o g e t h e r w i t h t h e day o n w h i c h d i s c o l o u r a t i o n of t h e t h a l l i w e r e f i r s t n o t e d . The t r a n s p l a n t s t o t h e c o n t r o l t r e e s a t L o c h S u n a r t r e m a i n healthy and t h i s c o n f i r m s t h e p r i o r a s s u m p t i o n t h a t t h e t r a n s p l a n t a t i o n t e c h n i q u e itself w a s n o t damaging.

298 At Glenlee and Borrowdale L. amplissima and L. scrobiculata had high final scores. However, at Glenlee L . virens only showed m i l d damage and L . pulmonaria almost none. At this site only L . pulmonaria occurs naturally (Farmer e t a l . , 1991a). L . pulmonaria was damaged at the Borrowdale site, but t h i s took over a year to develop. L . virens still remains healthy at this site. It is interesting to note that this is the only species that remains i n Seatoller Wood, although it is now limited to old, pollarded ash trees with a h i g h bark pH (Farmer e t a l . , 1991a). Table 2 The final damage scores ( % of thallus surface showing chlorosis or necrosis) for four Lobaria species transplanted to three sites. The number of day for which the study was undertaken together with the day on which damage was initially noticed are also given. A l l scores are means, 5 S . E . , n=4. Loch Sunart

Glenlee

Borrowdale

930

938

0.0 50.0

77.5 522.5

97.5 52.5

--

124

262

0.0 50.0

2.5 52.5

52.5 220.6

--

349

348

3.0 22.4

70.0 214.7

92.3 27.5

260

124

128

0.0 5.0.0

13.8 56.9

1.7 51.7

--

349

938

Length of Study ( d ) 943 L.

L.

L.

amplissima Final Score Day of First Damage pulmonaria Final Score Day of First Damage scrobiculata Final Score Day of First Damage

L virens

Final Score Day of First Damage

4.

DISCUSSION

Previous studies had shown a strong effect of wet deposited acidity on the tree bark chemistry at Seatoller Wood (Farmer e t a l . , 1991b). Here the wood grows on nutrient poor soils (Farmer e t al., 1991a) and the bark has become h i g h l y leached and is of low pH (Table 1). However, at Glenlee the bark retains a high pH (Table 1 ) in spite of acidic pollutant inputs, possibly due to the nutrient rich lowland soil (Farmer

299 e t a l . , 1991a). S e a s o n a l s t u d i e s of lichen t i s s u e n u t r i e n t s h a v e s h o w n a p o s s i b l e a d v e r s e r e s p o n s e t o low p H e p i s o d e s in t h e s t e m f l o w ( F a r m e r e t a l . , 1991b). Farmer e t a l . (1992) have a l s o found t h a t a r a n g e of L o b a r i o n l i c h e n s a r e s t r e s s e d by low p H t r e a t m e n t s u n d e r laboratory c o n d i t i o n s , w h i l e P a r m e l i a C u r r e n t c o n d i t i o n s o f low bark p H and low l a e v i g a t a i s not. s t e m f l o w pH a t S e a t o l l e r Wood w o u l d b e e x p e c t e d t o w e a k e n t h e L o b a r i a s p e c i e s but n o t t h e a c i d o p h i l e s l i k e P a r m e l i a laevigata. This transplant study confirms the suggestion that it i s t h e c u r r e n t c o n d i t i o n s and n o t a h i s t o r i c a l a s p e c t of t h e s i t e w h i c h h a s c a u s e d t h i s d e c l i n e in t h e lichen flora. A c i d i c d e p o s i t i o n i s probably n o t t h e o n l y c a u s e o f lichen decline. T h e conditions at Glenlee a r e obviously not f a v o u r a b l e t o t h e g r o w t h of L . a m p l i s s i m a and L . s c r o b i c u l a t a , but bark pH and s t e m f l o w pH r e m a i n s high. This wood maintains a d e n s e canopy and u n d e r s t o r e y and t h u s w o o d l a n d m a n a g e m e n t may b e an i m p o r t a n t f a c t o r in d e t e r m i n i n g t h e s t a t u s o f t h e L o b a r i o n here. In many p a r t s o f E u r o p e L o b a r i o n l i c h e n s a r e key i n d i c a t o r s of a n c i e n t w o o d l a n d s ( R o s e , 1992). T h i s i s in part d u e to their slow ability t o colonize new trees and woodlands. T h u s if s p e c i e s a r e lost f r o m a s i t e , f o r w h a t e v e r r e a s o n , it i s n o t p o s s i b l e t o a s s u m e t h a t t h e lack of r e t u r n to t h a t s i t e i s d u e t o u n f a v o u r a b l e c o n d i t i o n s p r e v a l e n t there. Recent s t u d i e s o f lichen r e c o l o n i z a t i o n w i t h f a l l i n g SOz l e v e l s s u g g e s t w i d e v a r i a b i l i t y in t h e c o l o n i z i n g a b i l i t y o f d i f f e r e n t e p i p h y t i c s p e c i e s ( H a w k s w o r t h a n d M c M a n u s , 1989). T r a n s p l a n t a t i o n s t u d i e s can p r o v i d e a g o o d i n d i c a t i o n o f w h e t h e r u n f a v o u r a b l e c o n d i t i o n s a r e present. This is p a r t i c u l a r l y i m p o r t a n t in acid r a i n s t u d i e s w h e r e h i s t o r i c a l r e c o r d s of g a s e o u s p o l l u t a n t s may n o t b e a v a i l a b l e . In a changing pollution climate, transplantation i s a means of s e p a r a t i n g r e s p o n s e t o c u r r e n t p o l l u t i o n f r o m t h a t of t h e c u m m u l a t i v e p o l l u t i o n history of a site.

5. ACKNOWLEDGEMENTS

W e thank J . 0 . H e n d e r s o n , R . Roper-Caldbeck and t h e N a t i o n a l T r u s t f o r p e r m i s s i o n t o work in t h e t h r e e w o o d l a n d s . T h i s work w a s f u n d e d by a g r a n t from t h e N a t u r a l E n v i r o n m e n t R e s e a r c h C o u n c i 1.

6. REFERENCES

D a y , I.P. ( 1 9 8 5 ) . L i c h e n s in B o r r o w d a l e and P o l l u t i o n . R e p o r t t o the Nature Conservancy Council, Peterborough. F a r m e r , A.M., B a t e s , J . W . and B e l l , J.N.B. (1991a). C o m p a r i s o n s of t h r e e w o o d l a n d s i t e s in NW B r i t a i n d i f f e r i n g in r i c h n e s s of t h e e p i p h y t i c L o b a r i o n p u l m o n a r i a e c o m m u n i t y and l e v e l s of w e t a c i d i c d e p o s i t i o n . Hol. Ecol., 14: 85-91.

300 F a r m e r , A.M., B a t e s , J.W. a n d B e l l , J.N.B. (199lb). Seasonal v a r i a t i o n s in a c i d i c p o l l u t a n t i n p u t s a n d t h e i r e f f e c t s o n t h e c h e m i s t r y o f s t e m f l o w , bark a n d e p i p h y t e t i s s u e in t h r e e o a k w o o d l a n d s in N.W. Britain. N e w P h y t o l . , 118: 441-451. F a r m e r , A.M., B a t e s , J.W. a n d B e l l , J.N.B. ( 1 9 9 2 ) . Ecophysiological e f f e c t s of acid r a i n o n b r y o p h y t e s and lichens. In: B r y o p h y t e s a n d L i c h e n s in a C h a n g i n g E n v i r o n m e n t , J.W. B a t e s a n d A.M. F a r m e r (Eds.), O x f o r d University P r e s s , Oxford, In Press. G i l b e r t , O.L. ( 1 9 8 6 ) . F i e l d e v i d e n c e f o r a n a c i d r a i n e f f e c t o n lichens. Env. Poll. (Ser. A ) 40:227-231. G i l b e r t , O.L. ( 1 9 9 1 ) . A s u c c e s s f u l t r a n s p l a n t o p e r a t i o n i n v o l v i n g Lobaria amplissima. L i c h e n o l o g i s t , 23: 73-76. H a l l i n g b a c k , T. (1990). T r a n s p l a n t i n g Lobaria pulmonaria t o n e w l o c a l i t i e s a n d a r e v i e w o n t h e t r a n s p l a n t i n g o f lichens. W i n d h a l i a , 18: 57-64. H a l l i n g b a c k , T. a n d M a r t e n s s o n , P.O. (1987). T h e r e t r e a t o f t w o l i c h e n s , Lobaria pulmonaria a n d L. scrobiculata in t h e d i s t r i c t o f G a s e n e (S.W. S w e d e n ) . W i n d h a l i a , 17: 27-32. H a w k s w o r t h , D.L. a n d M c M a n u s , P.M. (1989). L i c h e n r e c o l o n i z a t i o n in L o n d o n u n d e r c o n d i t i o n s o f r a p i d l y falling s u l p h u r d i o x i d e levels, and t h e concept of z o n e s k i p p i n g . Bot. J . Linn. soc., 100: 99-109. J a m e s , P.W., H a w k s w o r t h , D.L. a n d R o s e , F. (1977). L i c h e n c o m m u n i t i e s i n t h e B r i t i s h Isles: a p r e l i m i n a r y c o n s p e c t u s . In: L i c h e n E c o l o g y , M.R.D. S e a w a r d (Ed.), pp. 295-413, A c a d e m i c P r e s s , L o n d o n . R i c h a r d s o n , D.H.S. (1967). T h e t r a n s p l a n t a t i o n o f l i c h e n t h a l l i t o s o l v e s o m e t a x o n o m i c p r o b l e m s in Xanthoria parietina ( L . ) Th. Fr. L i c h e n o l o g i s t , 3: 386-391. Rose, F . ( 1 9 8 8 ) . P h y t o g e o g r a p h i c a l a n d e c o l o g i c a l a s p e c t s o f Lobarion c o m m u n i t i e s in Europe. Bot. J . Linn. S O C . , 96: 69-79. R o s e , F . ( 1 9 9 2 ) . T e m p e r a t e f o r e s t m a n a g e m e n t . In: B r y o p h y t e s a n d L i c h e n s in a C h a n g i n g E n v i r o n m e n t , J.W. B a t e s a n d A.M. F a r m e r (Eds.), O x f o r d U n i v e r s i t y P r e s s , O x f o r d . In P r e s s , W i l l i a m s , M.L., Cltkins, D.H.F., B o w e r , J . S . , C a m p b e l l , W., I r w i n , J . G . a n d S i m p s o n , D. ( 1 9 8 9 ) . A p r e l i m i n a r y a s s e s s m e n t o f t h e a i r p o l l u t i o n c l i m a t e o f t h e UK. W a r r e n Spring Laboratory Report L R 7 2 3 (AP), Warren S p r i n g Laboratory, Stevenage.

T Schneider (Editor) Acldlflcatlon Research Evaluation and Policy Applications 1992 Elsevier Science Publishers B V

301

Acidification research activities in Poland W.A. Mill

Institute of Environmental Protection,ul. Kossutha 6, 40-832 Katowice, Poland

Abstract The major activities in acidification research in Poland are carried out under the aegis of the Convention on Long-Range Transboundary Air Pollution.The research programme started in November and is aimed at the computation and mapping of critical loads of acidity and sulphur.The first attempt involved the surface waters acidification study mainly in the Tatra Mountains watershed-the most sensitive to acidification.Subsequently the computations of forest soils critical loads were performed.The Institute of Environmental Protection in Katowice,acting as the Polish National Focal Center,is responsible for the acidification research programme in Poland.This paper summarizes the first results of critical loads mapping.Some specific Polish aspects of the critical loads concept and its policy applications are discussed.

1. INTRODUCTION The major activities in acidification research in Poland are carried out under the auspices of the Convention on Long-Range Transboundary Air Pollution (LRTAP). The research programme started in November 1990 aimed at the estimation and mapping of critical loads of acidity. The institution responsible for the project implementation is the Institute of Environmental Protection in Katowice,acting as the National Focal Center of the LRTAP Convention Because it was the first approach to acidification problems on such a broad scale in Poland,the studies started with a preliminary sensitivity analysis of natural ecosystems to acidic deposition. The analysis was based on the guidelines provided by Chadwick and Kuylenstierna (1990). Results of this analysis indicate clearly that the most sensitive aquatic ecosystem in Poland is the Tatra Mountains catchment, what is a direct consequence of the specific bedrock lithology, soil type, land use and rainfall characteristics. The recently published water quality data from Sudety Mountains lakes and streams suggest the occurrence of acidification problems.This data are currently studied. The surface waters in other parts of Poland are much less sensitive. The preliminarily

302 calculated high critical loads for Katowice and Bielsko regions (southern part of Poland) together with the pH values much exceeding 7 in the majority of Polish surface waters prove the above conclusions to be right. The second very sensitive receptor to acidification are forest and forest soils. There is no need for any scientific evidence to state the obvious fact. The view of the Sudetian forests in the south - western part of Poland speaks for itself. Finally it was decided to perform the calculation and mapping of critical acidity and sulphur loads for the aquatic ecosystem of Tatra Mountains and for forest soils for the whole country.

2. CRITICAL LOADS FOR SURFACE WATERS Critical loads and their exceedance were calculated by use of the Water Chemistry Method (Henriksen et.al., 1990; Sverdrup et.al., 1990). This method comprises some experimental parameters estimated for the Norwegian lakes, and their application to Tatra surface waters may results in some errors. Thus the task for the nearest future is to calibrate Henriksen's model to be adaptable for Polish conditions. A 5 x 10 subdivision of 1' longitude x 1/2" latitude grids is applied. Coordinates of the lower left grid comer are defined. All input data required by the model origin from own measurements, historical data resulting from research works carried out by the Academy of Agriculture in Cracow (Oleksynowa and Komornicki,1969,1970),the Chemistry Institute of the University of Torun (Polakiewicz,1989) and the Institute of Meteorology and Water Management. A very significant contribution to data acquisition introduced the Research Section of the Tatra National Park in Zakopane. The calculation results summarizes Table 1.These results reveal a differentiation in the sensitivity to acidity and sulphate deposition of the Tatra Mountains area. The three following zones can be distinguished: (i) the pure crystalline zone of low mineralized water, where critical loads of low values ranging from 0 to 500 eq ha-'yr-' (sites TPO1-TPO6, TPO8,TPlS) (ii) a mixed crystalline - sediment zone of medium mineralized water with critical loads of about 2000 eq ha-lyr-' (sites TP10-TP12) (iii) a typical sediment zone with domination of calcareous and dolomite rocks with aquatic ecosystems practically insensitive to acid deposition with critical loads for sulphates higher than 5000 eq ha-'yr-' (sites TP07,TP09,TP13 and TP14) The studies on surface waters acidification in Poland will be continuously conducted. Recently a joint Polish-Norwegian project under the scientific supervision of Dr Henriksen started. The project aims at the updating of acidification data and parameter estimation for Polish local conditions.

303

Table 1 Calculation results of critical loads and their exceedances in streams and lakes of Tatra Mountains.Sampling period 12-14.July.1991.(keq ha-' yr')

TPOl TP02 TP03 TP04 TP05 TP06 TP07 TP08 TP09 TPlO TPll TP12 TP13 TP14 TP15

0.26 0.01 0.58 0.29 0.24 0.39 4.46 0.41 4.77 1.99 2.44 1.79 12.22 12.07 0.49

0.05 -0.16 0.29 0.05 0.00 0.15 4.45 0.41 4.08 1.72 1.87 1.23 11.49 11.74 0.13

0.20 -0.04 0.50 0.21 0.16 0.32 4.72 0.36 4.73 1.98 2.34 1.70 12.14 12.00 0.41

I: ANC,=O peq dm-' 11: ANC,,=20 peq dm-' 111: ANC,=O peq dm-'

IV: ANC,=20 peq dm"

-0.01 -0.20 0.21 -0.03 -0.08 0.07 4.40 0.36 4.04 1.71 1.78 1.14 11.41 1I .67 0.05

-0.15 -0.40 0.17 -0.12 -0.17 -0.02 4.35 0.00 4.36 1.58 2.08 1.38 11.81 11.66 0.08

0.05 -0.16 0.29 0.05 0.00 0.15 4.45 0.41 4.08 1.72 1.87 1.23 11.49 11.74 0.13

-0.21 -0.45 0.17 -0.20 -0.25 -0.09 4.31 -0.05 4.32 1.57 1.93 1.29 11.73 11.59 0.00

-0.01 -0.20 0.29 -0.03 -0.08 0.07 4.40 0.36 4.04 1.71 1.78 1.14 11.41 11.67 0.05

BC,=O keq ha-' yr-' BC,=O keq ha-' yr-' BC,=0.41 keq ha-' yr-' BCD=0.41 keq ha-' yr-'

3. CRITICAL LOADS FOR FOREST SOILS The calculation method used was the steady state mass balance method (SMB). A detailed description of this method is given in Sverdrup et.a1.(1990), Hettelingh and de Vries (1991) and Hettelingh et.al. (199 1). Using this method the following forest soils acidification parameters for Poland were computed: - critical loads of actual acidity, CL(AC,,J given by: CL(Ac,J = BCW+0.09*Q+0.2*Q CL(Ac,,J = BCW+0.09*Q+1,5*(BC,'+BC,-BC,) were: = weathering of base cations (mol ha-lyr") BC, = runoff (m'ha-'yr-I) Q

(1)

304 = seasalt-corrected base cation deposition (mol ha-lyr-') BC,' = uptake of base cations (mol ha-lyr-I) BC, The lower of the two values respectively calculated by Equations 1 and 2 should be used.

- critical loads of sulphur, CL(S) given by: CL(S) = SpCL(AC,,J (3) were: Sf = sulphur fraction CL(AC,J = critical load of actual acidity (mol,ha-lyr-') A 5x5 subdivision of I" longitude x 1/2" latitude is applied. Coordinates of lower left grid comer are defined. Data that are needed to map critical loads and exceedances are the forest coverage of the land, weathering and uptake rates of elements, precipitation surplus and the deposition. Forests coverage: This group of data, containing the spatial distribution of forest with the subdivision to coniferous and deciduous, was provided by the Forest Management Bureau in Warsaw (FMB). Weathering rates: Base cation weathering rates were estimated accordingly to the method reported by Hetteingh and de Vries (1991). These 23 soil categories, of the 1:1,000,000 soil map of Poland were assigned to the five weathering rate classes. Base cations and nitrogen uptake: The annual average growth rate from the FMB and values for the element content in stems and branches (proposed by Hettelingh and de Vries, 1991) were multiplied. Precipitation surplus: has been assumed as to be the runoff values and were extracted from the data bases of the Institute for Meteorology and Water Management. Deposition: Because of significant inconsistencies in the national data, the deposition values of base cations, sulphur, ammonia and nitrogen were taken from the EMEP calculated values. Critical loads of actual acidity and sulphur were calculated for each subgrid and 1,5 and 50 percentile values for the basic grids were derived and mapped.The results are shown in Figure 1. The following conclusions can be drawn from the results: i) the 1 and 5-percentile maps indicate that for 92% of the area critical loads of acidity range between 200-1000 eq ha-lyr-l ,what means that the majority of Polish soils can be classified as of medium sensitivity to acidification in comparison to other European countries, ii) according to the maps the most sensitive area is located in central part of Poland,whereas the southern part is the less sensitive.The last observation is rather controversial in view of the fact that this is a mountains region with very high precipitation rates.This problem was already raised by other countries and pointed out as a subject to a modification of the Simple Mass Balance Method, iii)it is the first attempt to produce maps of critical loads for the whole country.Not all of the input data originate from the national data bases and some alterations to the calculation method used are necessary.It implies that the research programme should be continued to give a reliable scientific background for developing potential abatement strategies for sulphur and nitrogen.

4. SOME COMMENTS ON IMPLEMENTATION OF THE CRITICAL LOADS CONCEPT TO ENVIRONMENTAL POLICY IN POLISH CONDITIONS The commonly accepted definition describes the critical loads approach as "a procedure for developing optimized abatement strategies by which emission reductions are carried out on the basis

Figure 1. Critical loads of acidity and sulphur for Poland.

W

8

306

of scientifically derived critical values. " Apparently,the procedure is simple.At the first step,only the setting of the interim target loads is needed,taking into account not only the environmental sensitivity, but also technical, social, economic and political considerations.At the next step,some energy scenarios should be derived on the basis of cost optimization.Such energy scenarios,which comply with the established target loads,can be introduced by a given country to the negotiations on the SO2 Protocol. The above described procedure could possibly work in a country with stabilized economy,being unsuitable for a country under political and economic transformation conditions,like Poland.The principal issues that make the environmental policy so difficult,can be summarized in the following way: i)the struggle with inflation,the first step in the process of transition into the free market economy,resulted in economic recesion .The recesion affected mainly large, stateowned economic units, concentrating almost all Polish heavy industry, which is particularly energy-consuming, therefore being the main atmospheric pollution source. The latest drop in industrial production, exceeding 20%,caused proportional reduction of pollutant emission. However the reduction has not been the result of ecological policy planning, thus it is difficult to be simulated or predicted using rational categories; ii)the structure of fuel consumption has recently been disturbed,due to changes in fossil and gaseous fuel deliveries from the Soviet Union-Poland's biggest supplier.Unti1the new delivery conditions are negotiated under long-term agreements,the input data for development of energy consumption scenarios will be uncertain; iii)financial cost,as the basic parameter for effectiveness evaluation of the considered abatement strategies,is in case of Poland loaded with a certain error margin.The error results from the fact that the Polish currency (zloty) is still not fully exchangeable yet. Irrespective of the above mentioned limitations,the Polish community is considerably interested in undertaking consequent and professionally planned activities aimed at environmental protection,within the available money,as well as technical means and intellectual potentiaLPoland is particularly open to co-operation with other countries,which is regarded as one of the major driving forces for the activities in the field of environment protection.

5. REFERENCES Chadwick,M.J.and J.C.I.Kuylenstierna. 1990. The Relative Sensitivity of Ecosystems in Europe to Acidic Depositions. Stockholm Environment Institute,Stockholm. Henriksen,A.,L.Lien,and T.S.Traaen. 1990. Critical loads for surface waters:Chemical criteria for inputs of strong acids. Norwegian Institute for Water Research Report 0-89210.NIVA,Oslo. Hettelingh,J.-P.and W.de Vries. 1991. Mapping Vademecum.National Institute of Public Health and Environmental Protection,Coordination Center for Effects,Bilthoven,The Netherlands. Hettelingh,J.-P.,R.J.Downing and P.A.M.de Smet. 1991. Mapping critical loads for Europe: CCE Technical Report No. 1,RIVM Report No.259101001, Bilthoven, The Netherlands, Oleksynowa,K.,T.Komomicki. 1969. Materialy do znajomoici w6d w Tatrach cz& I-X. Zeszyty Naukowe WSR w Krakowie. Oleksynowa,K. 1970. Charakterystyka geochemiczna w6d tatrzahskich. Acta Hydrobiol. PWN Oddzial w Krakowie. Polakiewwicz,T. 1989. Chemiczne zanieczyszczenia Srodowiska przyrodniczego TNP. Praca wlasna. Uniwersytet im. M.Kopernika w Toruniu. Sverdrup,H.,W.de Vries,and A.Henriksen. 1990. Mapping Critical Loads: A guidance manual to criteria,calculations,data collection and mapping. In: UN ECE Mapping Manual.

T Schneider (Editor). Acidification Research Evaluation and Policy Applications 01992 Elseviet Science Publishers B V All rights reserved

307

CRITICAL LOADS FOR DUTCH FOREST SOILS W. de Vries, J. Kros, R.M. Hootsmans, J.G. van Uffelen and J.C.H. Voogd DLO The Winand Staring Centre for Integrated Land, Soil and Water Research (SC-DLO), P.O. Box 125, 6700 AC Wageningen, The Netherlands

Abstract Critical loads for N and S on Dutch forests have been derived by using simple steady-state models. Critical loads have been calculated for combinations of 12 tree species and 23 soil types for a 10 x 10 km d. Values thus derived varied between approximately 300 and 1000 mol, ha-' y r - c o r N and between 150 and 1250 mol, ha-' y f ' for S. Amounts by which the critical loads are exceeded varied between approximately 1200 and 6100 mol, ha-' yr-' for N and between 850 and 2900 mol, ha-' yr-' for S, using deposition data based on 1985 emissions. The excess in critical N loads was much less for deciduous forests than for coniferous forests. For S the difference between coniferous and deciduous forests was quite small. A median reduction of approximately 80% will be needed to approach the critical loads for forest in the Netherlands.

1 INTRODUCTION Information on critical deposition levels (loads) are a prerequisite for political decisions on emission reductions. The assessment of average critical loads of N and S for forests on non-calcareous sandy soils formed a n important basis for the development and adaptation of the Netherlands "Acidification abatement policy" (De Vries, 1988; 1991). In order t o gain more insight in the magnitude of - and the spatial varation in critical loads, the Working Group on Effects of the UN-ECE Executive Body on Longe Range Transboundary Air Pollution has installed a Task Force on Mapping. The Task Force is assisted by a Coordination Centre. The Task Force and Coordination Centre have produced a "Manual on Mapping Critical loads" (Sverdrup et al., 1990) and a "Mapping Vademecum" (Hettelingh and De Vries, 1991) respectively, containing guidelines for the derivation of critical load maps. At present, such maps have been produced by most European countries along the guidelines given in these documents and results are summarized in Hettelingh et al. (1991). This paper provides a n overview of the variation in critical loads on Dutch forests and the amounts by which they are exceeded, in relation to tree species and soil type. Attention is also given to the criteria, models and data that have been used to derive these values and the effect of their uncertainty on the resulting critical loads.

308

2 MODELS TO DERIVE CRITICAL LOADS Critical loads for potential acidity have been derived indirectly from critical values for A1 concentrations and/or AVCa ratios in the soil solution using a steady-state soil acidification model according t o (De Vries, 1991): CL (Ac,,)

= BC*dd - BC,

+ BC,, + N,, + Nim(crit)+ Ac,,(crit)

(1)

where CL(Acpot)stands for the critical load of potential acidity, BC*,, is the seasalt corrected dry deposition of base cations, BC, and Np. are the growth uptake (net uptake needed for forest growth) of base cations and nitrogen respectively, BC,, is the base cation weathering, Nim(crit) is a critical long-term nitrogen immobilization, and Acl,(crit) is a critical leachin flux of acidity. The element fluxes in Eq. (1) are all given in mol, ha-' An overview of the various assumptions to derive equation (1) with a justification of it is given in De Vries (1991). The load of potential acidity does not only include free H, but also twice the amount of NH,, assuming that leaching of NH, from the rootzone of forest soils is negligible (Sverdrup et al., 1990). This implies that NH, is not counted as a base but as a (potential) acid that should be added to the acidifying effect of SO, and NO,. In the Netherlands, the (present) load of potential acidity is defined as:

yr-f

PL(Acpot)= PL(S0,)

+ PL(N0,) + PL(NH,)

- BC*,,

(2)

where PL(Ac,,), PL(SO,), PL(N0,) and PL(NH,) stand for tbe present loads of potential acidity, SO,, NO, and NH, respectively and BC dw for the seasalt corrected bulk deposition of base cations. The critical acidity leaching flux has been calculated as the sum of Al-leaching and H leaching. Three options have been used for the calculation of the critical Alleaching flux (De Vries, 1991): 1)a criterion for the Al-concentration in the rootzone: All,(crit) = FW . [All(crit)

(3)

2) a criterion for the molar AVCa ratio in the rootzone:

All,(crit) = mlCa(crit) . (BC*d, + BC,,

- BC,)

(4)

3) a negligible depletion of Al-hydroxides:

Al,,(crit) = r . BC,,

(5)

where FW is the water flux in m3 ha-' yr-l, [All(cr$) is a critical Al-concentration, mlCa(crit) is a critical equivalent AVCa ratio, BC d t is the seasalt corrected total deposition of base cations and r is a stoichiometric equivalent ratio of Al to BC

309

weathering. In the simple mass balance model, the water flux is taken equal to the precipitation surplus leaving the rootzone. An overview of the range in critical A1 concentrations and AVCa (or AVCa+Mg) ratios is given in Sverdrup et al. (1990) and De Vries (1991). Based on the evaluation given in these documents, a critical A1 concentration of 0.2 mol, rn-, and a critical molar Al/Ca ratio of 1.0 has been used. The value used for r equals 2 (De Vries, 1991). The value of Alle(crit) that has been used for the critical load calculation is the minimum value calculated by Eqs. (31, (4)and (5). The critical H leaching flux (Hle(crit))has been calculated as: Hle(crit) = FW . IHl(crit)

(6)

where [Hl(crit) is a critical H-concentration. The critical H concentration has been related to the critical Al-concentration according to: IHl(crit) = ([All(~rit) / GbbY'

(7)

$bb

is the gibbsite equilibrium constant in mol;' m6. For we used a ~ ~ ~ o k mol;' 1 0 m6 2 (=lo8 mo1-2 1'). The value of the critical A -concentration has been determined by the critical Al-leaching flux divided by the water flux. Separate critical loads for N (CL(N)) to avoid eutrophication have been calculated according to (De Vries, 1991): CL(N) = N,

+ Ni,(crit) + NO,,le(crit)

(8)

where N03,1e(~rit) is a critical NO, leaching rate. Calculation of the critical N load with this simple steady-state model, is based on the assumption that any N input above a net N uptake by forest growth and a critical rate of N immobilization and NO,-N leaching will finally lead to unacceptable high N contents in needledeaves and soil organic matter. The critical nitrate leaching flux has been calculated as: NOS,le(~rit) = FW . INO,l(crit)

(9)

where [NO,](crit) is the critical NO, concentration. For [N031(crit) a value of 0.1 mol, rn-, has been used based on data from Schulze et al., (1989). Summation of the critical loads for N and S should give the critical load for potential acidity. Consequently, a critical S load (CL(S)) has been calculated a s (compare Eqs. 1, 8 and 10): c L ( s ) = BC*,, - BC,

+ BC,, + Acle(crit) - NO,,,,(crit)

(10)

3 10 3 APPLICATION METHODOLOGY

Deposition areas have been defined by seeking for an optimum between the number of areas and the spatial variability within each area. For the Netherlands a 10 x 10 km grid has been used because detailed information regarding tree species and soil types exists at this scale. The number of grids containing forests equal 434. A distinction has been made in twelve tree species and 23 soil types. Tree species included were Pinus Sylvestris (Scotch Pine), Pinus Nigra (Black Pine), Pseudotsuga Menziesii (Douglas Fir), Picea Abies (Norway Spruce), Larix Leptolepis (Japanese Larch), Quercus Robur (Oak), Fagus Syluatica (Beech), Populus Spec (Poplar), Salix Spec (Willow), Betula Pendula (Birch), Fraxinus Nigra (Ash) and Alnus Glutinosa (Black Alder). Soil types were differentiated in 18 non-calcareous sandy soils (mainly podzolic soils), calcareous sandy soils, loess soils, non-calcareous clay soils, calcareous clay soils and peat soils on the basis of a recent 1 : 250 000 soil map of The Netherlands. Information on the area (distribution) of each specific forest-soil combination in a grid was derived by overlaying a digitized forest- and soil database. An overview of the percentage of coniferous - and deciduous forests on the various soil types in the Netherlands is given in Table 1. Table 1 Forest coverage on peat, sand, loess, clay and calcareous soils in the Netherlands Soil type

Coverage (%) Coniferous forests ~

Peat soil Sandy soil” Loess soil Clay soill) Calcareous soils2) All soils

0.4 57.9 0.4 0.5 0.8 60.0

~

Deciduous forests ~~

3.1 26.9 1.1 3.3 5.6 40.0

~

All forests ~~

~~

3.5 84.8 1.5 3.8 6.4 100.0

1)Refers to non-calcareous soils 2) Refers to sandy soils (about 1/3) and clay soils (about 2/3)

Table 1 shows that nearly all coniferous forests are located on non-calcareous sandy soils, whereas about 1/3 of the deciduous forests is located on other soils as well.

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4 DATA COLLECTION

Data that are needed to map critical loads and the amount by which they are exceeded, are the deposition, weathering and uptake of elements and the precipitation surplus. Critical N immobilization rates related to the formation of stable C-N compounds have been neglected (cf Schulze e t al., 1989). Dry deposition of base cations has been calculated by multiplying the bulk deposition with a dry deposition factor. The bulk deposition of base cations (Ca, Mg, K and Na) and C1 has been derived from 22 weather stations in the Netherlands (KNMURIVM, 1985) using interpolation techniques to get values for each grid. Dry deposition factors for base cations and C1 on each tree species have been derived from available data on Na in throughfall and bulk deposition a t 42 sites. The total deposition of N and S on forests has been calculated by multiplying average deposition values for each grid by forest filtering factors for SO,, NO, and NH,. Average deposition values are based on calculations with the deposition model TREND (Van Jaarsveld, 1989) for the year 1985. Filtering factors are based on a comparison of average annual deposition and throughfall data. Weathering rates are based on information on base cation depletion rates in soil profiles, budget studies and on column and batch experiments, which have been conducted during five years on the most relevant non-calcareous sandy soils in the Netherlands (Hootsmans and Van Uffelen, 1991; Winand Staring Centre , Internal report). For clay and peat soils indicative values have been derived from literature. For calcareous soils, the weathering rate can become nearly infinite. Here, we used a rather arbitrary high value t o avoid any exceedance in critical acid loads. Uptake rates are determined by forest growth and element contents in stems. Forest growth estimates for all relevant combinations of forest and soil type and contents of the elements N, K, Ca and Mg in stems are based on a literature survey for all tree species included (De Vries et al., 1990). Precipitation surpluses are determined by the precipitation rate minus the sum of interception, evaporation and transpiration (evapotranspiration). Precipitation estimates have been derived from 280 weather stations in The Netherlands, using interpolation techniques to obtain values for each grid. Interception fractions, relating interception to precipitation, have been derived from literature data for all tree species considered (Hiege, 1985). Data for evaporation and transpiration have been calculated for all combinations of tree species and soil type with a separate hydrological model (De Visser and De Vries, 1989). Median values for the model inputs for coniferous and deciduous forests and for the various soil types are given in Table 2 and 3. Table 2 shows remarkable differences between coniferous and deciduous forests with respect t o the various model inputs. Dry deposition of base cations is higher for coniferous forests, whereas uptake of N and BC and leaching of NO, and acidity is less. This is due to higher values for dry deposition factors and lower values of growth rates and precipitation surplus of coniferous forests.

312

Table 2 Median values of input data for coniferous and deciduous forests in the Netherlands. For an explanation of the abbreviations, see section 2. Forest type

Coniferous Deciduous All

Model input data (mol, ha-' yr-') BC*cld

BCwe

BCgu

Ngu

N03,1e

Acle

330 200 230

300 350 350

270 340 315

310 510 420

155 225 215

430 445 440

Table 3 Median values of input data for Dutch forests on various soil types. For a n explanation of the abbreviations, see section 2. Soil type

Model input data (mol, ha-' yr-')

-

BC*dd

Peat Sand Loess Clay Calcareous All

220 230 540 220 240 230

BCwe

200 300 500 1000 10000 350

BCgu

Ngu

255 310 455 395 455 315

570 410 560 510 505 420

190 225 235 185 210 215

305 420 635 535 615 440

Table 3 shows that the uptake of nitrogen (proton sink) is largely compensated by base cation uptake (proton source). Consequently, the overall effect of uptake is small except for peat soils. In non-calcareous sandy soils and loess soils, the critical acidity leaching is the most important proton sink although in loess soils the dry deposition and the weathering of base cations is nearly as important as acidity leaching. The large input of base cations (mainly Ca) on loess soils is due to their locations in the neighbourhood of limestone quarries in the southern part of the Netherlands. For clay soils, the weathering rate is most important. However, this value is rather arbitrary. As with peat soils the weathering rate of clay soils has been assigned on the basis of very few data.

5 CRITICAL LOADS

Variations in critical loads are mainly determined by tree species and soil type. Results for the effect of tree species are given in Table 4.

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Table 4 5, 50 and 95 percentile values of critical loads for nitrogen, sulphur and potential acidity for coniferous and deciduous forests on non-calcareous soils in the Netherlands Forest type

Critical load value (mol, ha-' yr-')

N

S

Acid

5% 50% 95%

5% 50% 95%

5% 50%

235 485 740 455 720 1165 325 640 1030

205 655 1345 145 360 1205 150 505 1250

625 1150 1740 710 1095 2140 700 1115 2055

- ~ _ _ _

Coniferous Deciduous All

95%

The critical N loads for deciduous forests are higher than for coniferous forests because of higher values for N uptake and critical NO, leaching (see Table 3). Contrary to N, the critical S loads for deciduous forests are lower because of a lower input of base cations by dry deposition and a higher base cation uptake (cf Table 2). Furthermore, the range in critical loads appears to be higher for S than for N. Results for the effect of soil type are shown in Table 5 . Table 5 5, 50 and 95 percentile values of critical loads for nitrogen, sulphur and potential acidity for Dutch forests on various non-calcareous soils Soil type Critical load value (mol, ha-' yr-')

Peat Sand Loess Clay All

N

S

5% 50% 95%

5%

50% 95%

235 386 540 235 325

160 150 200 480 150

210 775 475 1010 10201775 11901615 505 1250

665 615 795 690 640

1365 980 1120 1360 1030

-

Acid 5% 50% 95% 660 910 695 1070 9401770 11601940 7001115

1545 1760 2450 2475 2055

Values for calcareous soils have not been included since the critical loads for S and total acid are completely determined by the arbitrary high weathering rate of 10 000 mol, ha-' yr-'. Critical N loads on these soils are comparable to those on non-calcareous clay soils. The critical N loads are generally in the same order of magnitude for all soils. However, it should be noted that the values for peat and

3 14

clay soils are an underestimate, because denitrification is neglected. The effect of denitrification is likely to be high in these soils, with (potential) anaerobic conditions. These conditions are hardly met in sandy soils and loess soils below forests because they are mostly deeply drained. The critical loads for S and potential acid for non-calcareous forest soils increase in the direction peat < sand < loess < clay. The relative high critical S load for clay soils is due to its greater weathering rate (cf Table 3). For loess soils it is due to the effect of dry deposition of base cations.

6 THE DIFFERENCE BETWEEN PRESENT LOADS AND CRITICAL

LOADS As with critical loads, the amounts by which critical loads are exceeded is influenced by tree species and soil type. Results for the effect of tree species are given in Table 6. Present loads are based on 1985 emissions. Table 6 5, 50 and 95 percentile values for the amounts by which critical loads for nitrogen, sulphur and potential acidity are exceeded (mol, ha-' yr-') for coniferous and deciduous forests on non-calcareous soils in the Netherlands Forest type

Excess (mol, ha-' y i l )

5%

Coniferous Deciduous All

S

N 50%

2230 4480 1010 2975 1195 3430

Acid

95%

5%

50%

95%

6900 493 6100

1185 2220 3140 3860 6770 755 1720 2400 2005 4760 860 1855 2860 2370 5330

9600 6905 8670

50%

95%

5%

Table 6 shows that the amount by which critical loads are exceeded is much higher for N than for S. This is due to the high input of NH, in the Netherlands resulting from intensive animal husbandry. Futhermore, the excess in N loads is much higher for coniferous- than for deciduous forests. This is due to both a lower critical N load (see Table 4) and a higher present N load caused by more efficient filtering of dry deposition. Compared to N, the influence of tree species on the excess in S loads is much lower. This is because the higher present load of S on coniferous forests, caused by forest filtering, is partly compensated by a higher critical S load (cf Table 4). An overview of the amount by which the critical loads on the various soil types are exceeded is given in Table 7.

315

Table 7 5, 50 and 95 percentile values for the amounts by which critical loads for nitrogen, sulphur and potential acid are exceeded for Dutch forests on various non-calcareous soils Soil type

Excess (mol, ha-' yi')

S

N

5%

50%

Acid

-~

___-__-

95%

5%

-

50%

95% ~

Peat Sand Loess Clay All

1315 1355 1455 695 1195

2745 3545 2830 2745 3430

5680 6225 5110 5830 6100

970 1040 1330 35 860

1825 1920 1805 1150 1855

2845 2890 2885 2195 2860

5%

50% ~

95% ~

2730 4705 8080 2720 5525 8780 2915 4760 7645 1065 4005 7550 2370 5330 8670

Table 7 shows that the highest excess in critical N loads occurs for sandy soils. This is because the present N loads are higher on these soils, which are mainly located in areas with intensive animal husbandry. The lowest excess in critical N loads occurs on calcareous soils (not shown in Table 8). 5, 50 and 95 percentile values -equal 315, 1325 and 3300 mol, ha-' yr-' respectively. This is because these soils are all located in the western and northern part of the Netherlands with a low NH, deposition. Except for clay soils, the excess in critical S loads is nearly similar for all soil types. The relatively low excess on clay soils is due t o a relatively high critical load (cf Table 5). The high excess in S load on loess soils, which have a critical load that is nearly similar to clay soils (cf Table 5), is because of high present S loads on these soils, which are all located in the southernmost part of the country. The amount by which critical loads are exceeded can also be expressed as a percentage that is needed to reduce the present loads in order to meet the critical load. 5, 50 and 95 percentile values for the reduction percentage to meet the critical acid load are 60%, 82% and 91%. Values are somewhat higher for N (59%, 84% and 94% respectively) than for S (46%, 80% and 94% respectively).

7 UNCERTAINTIES

The uncertainty in critical load values derived before can be large and is mainly determined by the uncertainty in (1) model inputs (data), (2) critical chemical values and (3) model structure (Sverdrup et al., 1990; De Vries, 1991).

316

Model inputs The influence of the uncertainty in model inputs on the frequency distribution of critical loads is moderate. A sensitivity analysis, in which the various model inputs (base cation deposition, base cation weathering, forest growth and precipitation surplus) were varied by plus or minus 50% caused a decreasehncrease in median critical loads for N of 15 to 30% for N and 20 t o 50% for S respectively. The median critical load of potential acidity only changed by 10 to 20%. Reduction percentages that are needed to meet critical loads hardly changed (De Vries et al., in prep.). Critical chemical values Uncertainties in critical values for the Al concentration and AVCa ratio are large due to lack of knowledge about the effects of A1 in the field situation and a natural range in the sensitivity of various tree species for A1 toxicity (De Vries, 1991). The values for [All(crit) and RAlCa(crit) used in the calculations, i.e. 0.2 mol, m-3 and 1.5 mol, rnol;' respectively, are relatively low. A sensitivity analysis in which these values were increased by a factor two showed that the median critical load of potential acidity increased by about 50%. The increase in critical S load was even about 100% whereas the critical N load is not affected. A further increase in [Al](crit) and RAlCa(crit) hardly affected the critical load since the value was then fully determined by the criterion of a negligible depletion of Al-hydroxides (see Eq. 5). The uncertainty in the stoichiometric ratio r is not that large. A value of 3 is likely to be an upper level (De Vries et al., in prep.). Use of this single criterium leads to critical loads that are about 50% higher than those given in Table 4 and 5. The decrease in the median reduction percentage is about 10%. Model structure Uncertainties in the model structure relate to the assumptions that have been made to simplify the "real world". In this context, it is important to realize that use of a one-layer model implies that the critical Al-concentration and AVCa ratio is applied at the bottom of the rootzone. However, most fine roots are concentrated in the topsoil. Consequently, it is more realistic to apply these criteria for the forest topsoil. Values for the A1 concentration and AVCa ratio in the forest topsoil are generally lower because of Ca-cycling, transpiration and Al-mobilization with depth. An indication of these effects on the critical load value has been derived by applying the Al-concentration criterion in the middle of the rootzone and assuming a uniform weathering, transpiration and nutrient uptake with depth. Furthermore, the value for was adapted to lo7 mol-' 1-'. 5, 50 and 95 percentile values for the critical loa of potential acidity on all forests thus derived equal 1485, 2075 and 2560 mol, ha-' yr-' respectively. This is nearly a doubling of the 5 and 50 percentile values given in Table 4, which is due to a much larger critical acidity leaching. The 5 percentile value is in accordance with applications with the dynamic multi-layer model RESAM, which showed that a n acid load near 1400 mol, ha-' yr-' hardly caused any exceedance of a critical Al-concentration or AVCa ratio in the forest topsoil (De Vries et al., 1991). However a t this deposition level, Al depletion will still occur. As stated before, the neglection of denitrification is likely to cause a n underprediction in peat soils and clay soils. An indication of the effect has been

317

derived by assuming that denitrification can be described as a n fraction of the net nitrate input (De Vries, 19911, and using data from Breeuwsma et al. (1991) for the denitrification fraction in clay soils and peat soils. Results showed a more than two fold - and five fold increase in the median critical N loads of clay soils and peat soils respectively (De Vries et al., in prep.). However, the overall increase in the median critical N loads was rather small (about 25%) since most forests occur on non-calcareous sandy soils (cf Table 1). From the results given above, it can be concluded that the uncertainty in critical chemical values, combined with the depth where they have to be applied, is likely t o be more important than the uncertainty in model inputs. However, considering the necessity to avoid A1 depletion, the overall uncertainty is likely to be less than 50%. This hardly affects the reduction percentages that are needed. Sensitivity analysis showed that in all cases a reduction percentage of a t least 80% is needed to meet the critical loads for 95 percentile of the Dutch forests.

8 CONCLUSIONS

Critical loads for Dutch forests are low compared to present loads. Median values for critical loads of N, S and potential acid for forests on non-calcareous soils equal 640, 505 and 1115 mol, ha-' yr-' respectively. Median values for the amounts by which these critical loads are exceeded are 3430, 1855 and 5330 mol, ha-' yr-' respectively. The reduction percentage that is needed t o meet the median critical loads is about 80%. The uncertainty in this number is about 10%.

REFERENCES Breeuwsma, A., J.P. Chardon, J.F. Kragt and W. De Vries, 1991. Pedotransfer functions for denitrification. In: ECE 1991, Final report of the project "Nitrate in Soils", Chapter 5.3. DG XII, European Community, Brussels. De Visser, P.H.B. en W. De Vries, 1989. De gemiddelde jaarlijkse waterbalans van bos, heide en graslandvegetaties in Nederland. STIBOKA, Wageningen, Rapport 2085, 136 pp. De Vries, W., 1988. Critical deposition levels for nitrogen and sulphur on Dutch forest ecosystems. Water, Air and Soil Po11.42: 221-239. De Vries, W., A. Hol, S. Tjalma en J.C. Voogd, 1990. Literatuurstudie naar voorraden en verblijftijden van elementen in een bosecosysteem. Staring Centrum, Wageningen, Rapport 94, 205 pp. De Vries, W., 1991. Methodologies for the assessment and mapping of critical loads and the impact of deposition scenarios on forest soils. Winand Staring Centre, Wageningen, Report 46, 109 pp.

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Hettelingh, J.P. and W. De Vries, 1991. Mapping Vademecum. National Institute of Public Health and Environmental Protection. Coordination Centre for Effects, Bilthoven, The Nederlands, 64 pp. Hettelingh, J.P., R.J. Downing and P.A.M. Smedt, 1991. Mapping critical loads for Europe. National Institute of Public Health and Environmental Protection. Coordination Centre for Effects, Bilthoven, The Netherlands, Technical report no. 1, 183 pp. Hiege, W., 1985. Wasserhaushalt von Forsten und Walder und der Einfluss der Wassers auf Wachstum und Gesundheid von Forsten und Walder: eine Literaturstudie. Studiecommissie Waterbeheer, Natuur, bos en Landschap Utrecht, Report 7a, 193 pp. KNMVRIVM, 1985. Chemische samenstelling van de neerslag over Nederland. Jaarrapport 1985, 124 pp. Schulze, E.D., W. De Vries, M. Hauhs, K. R6sen, L. Rasmussen, C.O. Tamm and J. Nilsson, 1989. Critical loads for nitrogen deposition on forest ecosystems. Water, Air and Soil Poll. 48:451-456. Sverdrup, H., W. De Vries and A. Henriksen, 1990. Mapping critical loads. A guidance to the criteria, calculations, data collection and mapping of critical loads. Miljg rapport 1990: 14. Nordic Council of Ministers, Copenhagen, 1990, 124 pp. Van Jaarsveld, H.A.J., 1989. A model approach for assessing transport and deposition of acidifying components on different spatial scales. In: Changing composition of the troposhere, special environmental report 17. WMO, Geneva, Switzerland.

T. Schneider (Editor). Acidification Research. Evaluation and Policy Applications

0 1992 Elsevier Science Publishers B.V. All rights resewed

319

SCENARIO ANALYSIS WITH THE DUTCH ACIDIFICATION SYSTEMS (DAS) MODEL An example for forests and forest soils A. Tiktak", A.H. Bakema", K.F.de Boer", J.W. Erisman", J.J.M. van Grinsven", C. van Heerden", G.J. Heij', J . Kros', F.A.A.M. de Leeuw", J.G. van Minnen", C. van der Salmb, J.C.H. Voogdb and W. de Vriesb

" National Institute of Public Health and Environmental Protection (RIVM). PO BOX 1, 3720 BA BILTHOVEN, The Netherlands The Winand Staring Centre for integrated land, soil and water research (SC-DLO). PO-BOX 125, 6700 AC WAGENINGEN, The Netherlands

Abstract Within the framework of the Dutch Priority Programme on Acidification, the DAS model has been developed. This model aims a t evaluating the effects of acidification abatement strategies on a number of receptor systems in the Netherlands and describes the entire causality chain from emissions to effects in a regionalized way. Effects of three emissioddeposition scenarios on forest soils and forests are described. The emissioddeposition scenarios are based on political measures as announced in the Netherlands Environmental Policy Plan+ for the period 19902000. For the period aRer 2000, three deposition targets have been chosen. Emission reductions between 1990 and 2000 cause a reduction of the average deposition to the Netherlands of about 50%. This reduction is mainly caused by reduction of Dutch agricultural NH, emissions, although considerable reductions are also foreseen for NO, and SO, emissions in the Netherlands and other European countries. Deposition reductions lead to a fast increase of the pH value and decrease of ratio in the soil solution, decreasing the the A13+ concentration and A13+/Ca2+ risk for indirect forest damage. The exceedance of a critical A13+ concentration of 0.2 rn~l,.m-~ and a critical molar A13+/Ca2+ ratio of 1.0 reduces from about 75% and 65% of the considered forest area a t present to about 40% and 30% in the year 2000. When further deposition reductions are established, the exceedance of these parameters will become negligible in the year 2050. When deposition is ratio will kept constant after 2000, the critical A13+concentration and Al3+/Ca2+ still be exceeded in a considerable area of forest soil in 2050. As a result of the

320

retention of nitrogen in organic matter, continued mineralization and leaching of nitrate may remain a serious problem in areas with high nitrogen deposition several decades after imposing emission reductions. The scenario analysis for forests indicates that the nitrogen content of needles is reduced significantly. However, when the deposition is kept constant after 2000, a critical nitrogen content of 1.8% in needles is still exceeded in forests in areas with high nitrogen deposition. In these areas, the reduction of nitrogen and potential acid deposition leads to a net increase of needle mass and wood production. However, in relatively unpolluted areas, reduction of nitrogen emission will decrease needle mass and wood production. Direct effects of SO, and 0, appear to be less important in this context than soil mediated, indirect effects. The final part of the paper gives a brief description of future model developments.

1. INTRODUCTION In 1985 the Dutch Priority Programme on Acidification (DPPA) was started as a response to increasing concern of policy makers and scientists with the effects of air pollution, in particular on natural ecosystems. One of the most important reasons for this growing interest was the poor vitality of forests in Germany and the Netherlands (Heij and Schneider, 1991). Forest vitality and growth are affected by a large number of factors (Mohren, 1991). Traditional factors such as photosynthetic active radiation, temperature and water and nutrient availability may be limiting t o tree growth (Zahner, 1968; Ford et al., 1987; Kozlowski, 1982). High ambient levels of SO,, NO,, NH, and 0, may result in direct, visual damage to the canopy (Wislicenus, 1908-1916; Skeffington and Roberts, 1985). High nitrogen contents of the foliage may cause increased susceptibility to drought, frost and diseases (De Kam et al., 1989). The deposition of pollutants also affects forest vitality by indirect, soil mediated effects on roots (De Vries and Kros, 1991). The most notable effects are mobilization of A13+ and accumulation of NH,', causing inhibition of base cation (Ca", Mg2' and K+)uptake by roots (Keltjens and Van Loenen, 1989). Furthermore, high A13+ (Murach, 1984) and NH,+(Olsthoorn and Tiktak, 1991) concentrations in the soil solution limit fine root growth and make stands more susceptible to drought and nutrient stress. The complex nature of the problems involved in analyzing the effects of acidification called for an extensive integrated research programme. Findings and conclusions of these projects are summarized in Heij and Schneider (1991). These results also contributed t o the development of the integrated Dutch Acidification Systems (DAS) model (Bakema et al., 1990). This model has been

321

used for estimating and assessing the effectiveness of acidification abatement strategies. The model gives a description of the entire causality chain from emissions to environmental impacts such as soil acidification and effects on forests and heathlands in a regionalized way. This paper summarizes the structure, nature and application of the DAS model. Special attention i s paid to results of scenario analysis. This paper is restricted to the receptors forest and forest soil.

2. MODEL OVERVIEW

The basic conceptual outline of the DAS model is given in Figure 1 (Bakema et al., 1990). The boxes represent modules of the DAS model. The model consists of modules that can be run and developed independently. The model consists of a n emission module, a deposition module (SRM) and a number of effect modules to calculate effects on soils (RESAM), on forests (SOILVEG), on heathland (ERICA and CALLUNA), on heathland pools (AQUACID), on agricultural production (AGRIPROD), on monuments (MONUMENTS) and on construction materials (MATERIALS). A full description of the DAS model is given in Bakema et al. (1990). Figure 2 shows the regionalization used in the model. The division of the Netherlands in 20 areas is a compromise between the desired level of detail and the availability of regional data. DAS focuses on effects within the Netherlands only. Hence, outside the Netherlands, a subdivision of Europe is needed for emission aggregation only. The simulation period is about one hundred years, starting in 1950. Historical data are included to allow for model initialization. The output time step is one year, although some effect modules use a smaller time step. Air pollutants included are SO,, NO, and NH, and their oxidation products. Some effect modules require additional atmospheric input, such as 0,, the base cations (Ca", M C , Na' and K+)and the anion Cl. For 0, the observed yearly increase of 1%was extrapolated until 2000 and kept constant after that year, while the deposition of base cations and chloride were kept constant for the period 1950-2050.

322

activities

measures

emissions

cost & side-effects

agricultural production

a materials

monuments

i

I

effects

Figure 1 Conceptual outline of the Dutch Acidification Systems (DAS) model. Boxes represent modules of the model.

323

The Dutch Acidification

-I

Figure 2 Regionalization of the Netherlands (left) and Europe (right) as used within the DAS model. Within the DAS model, the Dutch areas hnction both as source and receptor for acidifying substances. The other areas function as sources only.

324 2.1. The emission module

In the scenario studies yearly emissions of SO,, NO, and NH, are required for all European countries for the period 1950-2050. As the air transport module requires a distinction in the stack height of emission, this information is provided by the emission module too. Emissions from each economic sector are assumed to be in one of the emission height classes "high', "medium" or "low". For most years, Dutch emission data of NO, and SO, are available on a national scale only. To obtain emission data for each acidification area, results of the TNO emissions inventory (Anonymous, 1985) have been used to build a distribution matrix of emissions over the 20 Dutch Acidification areas. Historical emission data for the Netherlands as well as for the other countries (1950-1990) are based on literature and on estimates (Thomas et al., 1988). The uncertainty in historical emissions can be large, especially when the distance in time becomes large. The scenario for Dutch emissions during the period 1990-2000 is based on policy measures announced in the Netherlands Environmental Policy Plan Plus (Dutch abbreviation: NMP'). Figure 3 shows the estimated yearly Dutch emissions of SO,, NO, and NH, for the period 19502000.

1.o

Dutch emissions (Mtonlyr)

m so, El NO,

0.8

DIlI NH3

0.6 0.4

0.2

0 1950

1960

1970

1980

1990

2000

Figure 3 Dutch emissions of SO,, NO, and NH, (Mton) for the years 1950-2000.

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European emissions between 1990 and 2000 are based on estimates by the Dutch Ministry of Housing, Physical Planning and Environment (Anonymous, 1988; 1989). In this study, it is assumed that the reduction targets will be met. The expected emission reductions for the year 2000 are summarized in Table 1. Table 1 Reduction of emissions in the Netherlands and surrounding countries as a percentage of the emissions in 1980 for the year 2000 (Source: Netherlands Environmental Policy Plan')

The Netherlands Belgium Germany France Great Britain

so,

NO,

NH,

84%

56% 50% 50% 50% 50%

67% 25% 25% 25% 25%

80% 80% 80%

45%

For the period 2000-2050 no emission scenarios are available yet. Instead, three deposition targets have been formulated for the years 2010 and 2050 (Table 2). These deposition targets are based on critical load calculations (De Vries and Kros, 1991; De Vries and Heij, 1991). According to this approach, damage to forests can be expected above a critical load of 1400 mol,.ha". Below a critical load of 700 mol,.ha" long-term effects on all natural ecosystems, including heathland pools, are negligible. In all three scenarios the ratios between the compounds which contribute to the total acid deposition were kept the same a s those calculated for the year 2000 (which are based on emission data). Table 2 Deposition targets for a n average Dutch landscape for the period 2000-2050 (molc.ha-'.yf'). The average deposition t o forests is about 20% higher than the average deposition to the Netherlands.

Scenario 1 Scenario 2 Scenario 3

2000

2010

2050

2200 2200 2200

2200 1400 1230

2200 1230 700

326 2.2. The air transport and transformation module (SRM)

The air transport and transformation module focuses on SO,, NO,,, NH, and their derivatives. The deposition and concentration of these compounds are calculated by means of transfer matrices. These matrices include a linear relationship between emissions and concentrations or deposition of a compound. For distances close to a source, the effective source height strongly determines the contribution to the deposition. Therefore, different matrices are available for compounds emitted from low (c50 meters), medium (50-100 meters) and high (>lo0 meters) sources. The transfer matrices have been generated by means of the TREND model (Van Jaarsveld and Onderdelinden, 1990; Asman and van Jaarsveld, 1990). The TREND model is a statistical atmospheric transport model which can describe transport, dispersion, chemical transformation and deposition of pollutants from local sources a s well as from distant sources. In the construction of the transfer matrices, detailed calculations on a 5 x 5 km2 grid were made, In a second step, the transfer coefficients were averaged for the acidification areas. The model has been extensively validated against measured concentrations in air and precipitation. Based on sensitivity studies with the TREND model it has been concluded that the deviations due to non-linearities remain within the accuracy limits of the DAS model, if the emission regime deviates no more than 30% from the regime for which the transfer matrices have been derived. When the deviations become larger, the error becomes larger but on a national scale, the maximum error (excluding possible errors in the emissions) will in general be about 30%. The module generates depositions for an average Dutch landscape. However, the dry deposition to forests is about 20% higher than the dry deposition to an average landscape. To account for this effect, the average dry deposition is multiplied by compound and region specific correction factors which are derived empirically (Erisman, 1990). Different sets of correction factors are derived for forests, heathlands and heathland pools. 2.3. The soil acidification module (RESAM) The long-term effects of acid deposition on Dutch forest soils have been simulated with the process oriented Regional Soil Acidification Model (RESAM) (De Vries and Kros, 1989). RESAM describes changes in soil chemistry, both in the solid phase (minerals, Al-hydroxides and adsorption complex) and in the aqueous phase. Processes included are weathering of primary minerals and Alhydroxides, cation exchange reactions, nitrogen transformations and nutrient cycling by the vegetation. The model includes the major chemical species in forest soils, i.e. H+, A13+,Ca", M2', K', Na', NH;, NO;, SO,; C l , HCO, and RCOO.. Within the framework of scenario analysis, the model is applied to all of the 20 Dutch acidification areas (Figure 2). Forests are represented by seven tree

327

species, i.e. Pinus sylvestris (Scots Pine), Pinus nigra (Black Pine), Pseudotsuga menziesii (Douglas fir), Picea abies (Northern Spruce), Larix leptolepis (Japanese Larch), Quercus robur (Oak) and Fagus siluatica (Beech). Forest soils are represented by the 14 major Dutch acid sandy soils (the most important are the Leptic podzol, Gleyic podzol, Humic podzol, Humic gleysol and Albic arenosol). These soils cover about 80% of the Dutch forest area. Furthermore, these soils are most sensitive to acidification. The considered forest-soil combinations cover about 65% of the Dutch forest area. Model parameters for soils and vegetation are assumed to be independent of the deposition area. The various data are based on field research (Kleijn et al., 1989), laboratory experiments (De Vries, unpublished results), literature surveys (De Vries et al., 1990) and calculations with an independent hydrological model (De Visser and De Vries, 1989). The presentation of model results is restricted to parameters which are ratio and indicators of forest stress (pH, concentration of A13+, molar A13+/CaZ+ molar NH,+/K+ratio). For most of these parameters, critical values have been defined. Values used are 0.2 m01,.m.~for the A13+ concentration, 1 for the A13+/Ca2+ ratio and 5 for the NH,'/K+ ratio (Heij et al., 1991). 2.4. The forest module (SOILVEG)

The forest module SOILVEG (Berdowski et al., 1991) predicts growth and phenology of Douglas fir (Pseudotsuga rnenziesii) in relation to the availability of carbon from photosynthesis and N, Ca, Mg and K from atmospheric deposition and soil processes. The model has been applied to the four major Dutch acid sandy soil types (Leptic podzol, Humic podzol, Gleyic podzol and Albic arenosol) in all 20 Dutch acidification areas. As forest and soil properties are assumed to be independent on the acidification area, atmospheric deposition and air quality are the only parameters which depend on the acidification area. The objective of this kind of regionalization is to demonstrate the combined effects of atmospheric deposition and soil processes per region for a representative tree species. Soil processes are described by an early version of the RESAM module. SOILVEG accounts for direct and indirect effects of acid atmospheric deposition on growth. High ambient concentrations of 0, and SO, reduce the photosynthetic carbon input into the tree. High levels of H' and A13+in the soil solution reduce both fine root growth and the uptake of base cations. The effect of nitrogen on forest growth can be either positive or negative. If growth is not limited by photosynthesis and the concentrations of K+,Ca2+or Mg2' are higher than their minimal values, increased nitrogen availability will enhance growth. However, extremely high nitrogen availability enhances respiration, which decreases the carbon availability for growth. Forest damage due to drought, frost and plagues is not included.

328

The presentation of model results is restricted to stem mass (indicator of forest productivity), needle mass (indicator of forest health), nitrogen content of needles (indicator of susceptibility to frost and plagues), nitrogen content of the litter (indicator of nitrogen saturation) and base cation uptake reduction (indicator of soil mediated, indirect effects).

3. RESULTS

3.1. Deposition trends The total average acid deposition to the Netherlands inferred from measurements shows a downward trend from about 6800 molc.ha'l.yr.' in the year 1980 to approximately 4800 mol,.ha".yr" in the year 1989. Considering an uncertainty in measurements of about 20% (Erisman and Heij, 1991) depositions calculated by means of the TREND model do not differ significantly from these figures. When all national and international emission targets announced in the Netherlands Environmental Policy Plan+ are met, the average deposition of total acid will be 2200 mol,.ha-'.yr" in the year 2000, which means a reduction of more than 50% compared to the present level. The Dutch interim deposition target of 2400 mol,.ha".yr.' for the year 2000 is within the uncertainty limits of the DAS model. It must be noted that the deposition to forests is about 20% higher than the average deposition. Figure 4 shows the trend of the total acid deposition together with the deposition of SO,., NO, and N K . The decrease of the total acid deposition until 1990 can be attributed to the decrease in SO, emissions in Western Europe. However, the trend of NH, emissions has shown a reverse trend during this period. The reduction to about 2200 mol,.ha.'.yr.' in the year 2000 is primarily a result of the expected strong decrease of NH, emissions from 1990 to 2000. Figure 5 shows the regional distribution of deposition levels of total nitrogen and total acid deposition for the year 1990. From the figure it is clear that the deposition of total acid is lowest is the Northern part of the Netherlands. The high deposition levels of total acid in the Southern part of the country are mainly caused by the enhanced NH, deposition. Furthermore, the Southern part of the Netherlands is close to the industrial areas of Antwerp and the Ruhr.

3 29 molcihalvr

molciha'yr NHx

6000

- .

NOx

.-

SO:

-

4000

2000

0

1950

1970

1990 year

2010

2030

205C

Figure 4 Trends in the deposition of total potential acid to the Netherlands with three acidification scenarios aimed a t achieving a deposition of 2200 mol,.ha-'.yi' (Scenario 11, 1160 mol,.ha~'.yr" (Scenario 2) and 700 mol,.ha-'.yr" in the year 2050. Time trends for SO,, NO, and NH, (mol,.ha-'.yr.') are given for scenario 2.

i

a) total acid deposition

;

0<4000

<

I

b) nitrogen deposition

8,

II

j

i 0<2000 I

I I ,

I

,

I

,

I

,

I

,

I

I

,

I I

, ,

I

I

I

,

I

,

,

I,

I

Figure 5 Regional patterns of deposition of total potential acid [molc.ha.'.yr.'] (a) and total nitrogen [mol,.ha~'.yr~'l(b) for the year 1990.

330

3.2. Effects on forest soils Figure 6 shows the area weighted median values of four key soil solution parameters for the three emission-deposition scenarios as calculated with the soil module RESAM (De Vries et al., 1991). The median values were derived from all relevant soil-vegetation combinations in the 20 acidification areas (total of 512 runs). Between 1990 and 2000 there is no difference in trends for the three scenarios because the deposition values are identical (Figure 4).

I

b) Al concentratton

0.2 0.1 I

1990 .

_ _

2010

2030

2050

I

0.0 1990

2010

2030

2050

I I - _ _ _ _ _ _ _ _ _ _ _ _ _ _ _ _ _ - - _ _ _ I _ _ _ _ _ _ _ _ _ _ _ _ _ _ _ _ - - - - - - - - - - - . .

I

Figure 6 Trends in area weighted median values of soil solution parameters in the topsoil in response to the three emissioddeposition scenarios illustrated in figure 4. (a) pH; (b) A13+concentration; (c) A13+/Ca2+ ratio and (d) NH,+/K+ratio.

33 1

In response to the large deposition reduction between 1990 and 2000, the median values of the pH in the topsoil show a large increase in this period (Fig. 6a). Similarly, there is a considerable decrease in the A13+concentration (Fig. 6b) and molar A13+/Ca2+ ratio in the topsoil (Fig. 6c). Both median values drop below a critical value of 0.2 m ~ l , . m .and ~ 1 mol.mol-', respectively, before the year 2000. However, it must be noted that there is still a considerable area of forest soil exceeding the critical values in the year 2000 (see below). A small decrease of the A13+concentration and A13+/Ca2+ ratio is observed for scenario 1 between 2000 and 2050, although the deposition levels remain constant in this scenario. This may be attributed to a slow depletion of the Al-hydroxide pool and to continued nitrification of accumulated organic nitrogen (see below). For scenario 2, in most forest soils the depletion of the Al-hydroxide pool stops between 2020 and 2030 (De Vries et al., 1991). The relative decrease of the A13+/Ca2+ratio is somewhat higher than the relative decrease of the A13+ concentration due to exchange processes. The NH,+/K+ ratio in the topsoil shows a different behaviour than the A1 related parameters (Fig. 6d). This ratio shows an initial increase until 2010 before it drops. The initial increase is due to a relatively small decrease of the NH,+ concentration and a relatively large decrease of the K+ concentration. The slow decrease in NH,+ concentration is due to continued nitrogen mobilization from the forest floor. As a result of the retention of nitrogen in organic matter, continued mineralization and leaching of nitrate may remain a serious problem in areas with high nitrogen deposition several decades after imposing emission reductions. The K+ concentration shows a large decrease due to the fact that the K+ concentration in the topsoil is mainly determined by foliar exudation. This process is modelled as a hnction of NH,' foliar uptake, which in turn is determined by the NH, deposition. The median NH,+/K+ratio is already below a critical value of 5 a t the start of the simulations. Figure 7 shows the trends in the percentage of area of forest topsoil exceeding a critical A13+concentration of 0.2 m01,.m.~(Fig. 7a) and an A13+/Ca2+ ratio of 1.0 mol.mo1' (Fig. 7b). The figure shows that the area exceeding these critical values is 75% for the A13+ concentration and 65% for the A13+/Ca2+ ratio in the year 1990. In the year 2000, the percentage of forest soils exceeding both values is still approximately 40% and 30%, respectively. In the year 2050 this area is negligible for scenarios 2 and 3. For scenario 1 it remains approximately ratio. 25% for the A13+concentration and 10% for the A13+/CaZ+ The influence of tree type and deposition area on the soil solution chemistry is illustrated in Table 3. In this table, the median values of the four parameters are presented for Douglas fir, Scots pine and Oak and for two deposition areas (area "Drenthe" in the North-Eastern part of the Netherlands having a relatively low acid load and area "Venray" in the South-Eastern part of the Netherlands having a relatively high acid load). In 1990 the median A13+

332

concentration and A13+/Ca2+ratio are higher in area "Venray" than in area "Drenthe". The percentage of area exceeding the critical A13+concentration and A13+/Ca2+ ratio are also higher (figure not shown). As the deposition levels tend to decrease, the differences between the areas become smaller after 2010. The A13+ concentration is decreasing in the direction Douglas fir, Scots pine, Oak. These differences are caused by decrease of the yearly evapotranspiration in this direction. A higher evapotranspiration is leading to a higher extraction of water from the rootzone, resulting in higher concentrations.

100

a) Ai-concentration ~

80

Scen.1 Scen. 2

60

Scen. 3

__

40 20

0 1990

\\

-

---a

2010

~---; --.

-

2030

20501

, ,

1990

2010

2030

2050

I

Figure 7 The percentage of forest topsoils exceeding a critical A13+ concentration of 0.2 m ~ l , . m -(left) ~ and a critical A13+/Ca2+ ratio of 1 in the topsoil (right) in response to the three emissioddeposition scenarios. The influence of tree species on soil solution chemistry overshadows the influence of acidification area. This is illustrated in figure 9a, which shows the ratio of 1.0 in percentage of the area of forest soils exceeding a critical A13+/Ca2+ the topsoil for the year 1990. The relationship between the area exceeding the critical value and the deposition of total potential acid (Figure 5a) is obscured by the spatial distribution of tree species.

333

Table 3 Influence of tree species and deposition area on median values of the A13+ concentration and the molar AP+/Ca2+ratio in topsoils in response to scenario 3 for the years 1990,2000 and 2010

Tree species

Douglas fir Scots pine

Oak All All

Deposition A13+concentration 1990 2010 2050 area

1990 2010 2050

All All All Drenthe Venray

3.7 3.1 1.5 3.1 4.3

0.8 0.6 0.3

0.4 0.7

0.2 0.1 0.1 0.1 0.1

0.0 0.0 0.0 0.0 0.0

Als'/Caa+ratio

1.2 1.0 0.4 1.1

1.6

0.2 0.2 0.1 0.3 0.3

3.3. Effects on forests The main objective of scenario analysis with the SOILVEG module is to compare the effectiveness of the three emission-deposition scenarios with respect to forest productivity and nitrogen saturation. As noted before, the SOILVEG module has only been applied to Douglas fir (Pseudotsugu menziesii) on Leptic podzols, Humic podzols, Gleyic podzols and Albic arenosols. The relative contribution of Douglas fir to total forest area is about the same for all acidification areas except for two areas in the Southern part of the Netherlands. Results apply to 60 years old trees. The most direct response of Douglas fir to decreasing air pollution is a decrease of the nitrogen content of needles (Figure 8a). This direct response is a result of needle uptake of gaseous NH,. In areas with high ambient NH, concentration levels in the air, simulated nitrogen contents are above a value of 1.8% in the year 2000, which is believed to be critical with respect to susceptibility to frost damage, drought stress and plagues (Heij et al., 1991; De Kam et al., 1989). Only when a further reduction of air pollution levels in these areas is established (scenario 2 and 31, the nitrogen contents will decrease t o safer levels (1.6-1.8%). In areas with lower air pollution levels (the coastal regions and the Northern part of the Netherlands), the nitrogen contents are already below the critical value in the year 2000 and decrease to levels close to minimum observed values (0.9%) when a further reduction of air pollution is established after the year 2000. Time trends of needle biomass and wood production in relatively unpolluted areas are opposite to time trends in relatively polluted areas (Figure 8b). The general increase of NH, deposition up to 1990 leads to an increase of both needle and wood mass in unpolluted areas, but to a decrease of needle and

334

wood mass in polluted areas. In clean areas, the positive fertilizing effect of increased nitrogen deposition dominates, whereas in polluted areas adverse effects of increased nitrogen deposition dominate, being enhanced soil acidification and maintenance respiration. Similarly, the strong decrease of nitrogen deposition after the year 1990 causes a decrease of needle and wood mass in unpolluted areas, while they cause an increase of needle and wood mass in polluted areas. When nitrogen deposition is further decreased in unpolluted areas after the year 2000, an increasing risk of loss of needle mass and wood production is predicted as a result of nitrogen shortage. For the most reductive scenario, a total dieback of forests is predicted in some coastal areas. For polluted areas, the model does not predict a reduction of growth. These model results are in accordance with the calculated regional distribution of leaf nitrogen contents, showing minimum concentrations in coastal areas (Figure 9b). Direct effects of high ambient levels of 0, and SO, on gross photosynthesis are small. In industrial areas, the maximum photosynthesis reduction due to high ambient SO, concentrations decreases from about 7% in 1970 to zero after the year 2000. On the contrary, reduction of gross photosynthesis by 0, increases from zero in 1950 to 2% in the year 2000. I t must be noted that gross photosynthesis is not only influenced by the adverse effects of high ambient concentration levels but that it is also reduced when the leaf biomass becomes below a critical value of 9000 kg.ha” (Berdowski et al., 1991). As shown above, the leaf biomass decreases strongly in unpolluted areas in response to decreasing nitrogen deposition, while it increases in polluted areas. Sub-critical leaf biomass only occurs in unpolluted areas after establishing a reduction of nitrogen deposition according to scenario 2 and 3. As a result of both direct effects of SO, and 0, and decreased biomass of leaves, gross photosynthesis in areas with high nitrogen deposition remains constant after 2000 but decreases in unpolluted areas for scenario 2 and 3. The most important soil mediated indirect effect of acid deposition is reduction of base cation uptake by roots due to low pH. This reduction can amount to 40% in the most polluted areas, as illustrated for M e uptake in Leptic podzols (Figure 8c). However, it may be expected that uptake reduction due to low pH in acid sandy soils will be substantial also in relatively unpolluted situations. Thus, scenario 2 and 3 eventually lead to uptake reductions of about 20% in the year 2050 (Figure 8c). These reductions are simular to simulated reductions for 1950. Base cation uptake reduction for scenario 1 remains 30% in the year 2050. Base cation uptake reduction due to high A13+concentrations is lower (10% for area “Groningen“ and 20% for area “Venray“ in the period of highest deposition; figure not shown). In coastal areas (e.g. “Groningen”), the effect of base cation uptake reduction on growth is limited due to the high availability of base cations due to high seaspray

335

deposition (Ivens, 1990). In the South-Eastern part of the country (e.g. "Venray"), growth reduction is relatively small due to high nitrogen input, resulting in a higher base cation use efficiency.

Area "Groningen"

Area "Venray"

a) N-content of needles (YO)

a) N-content of needles (%)

Scen. 1 ~

.<

.Scen. .

2 Scen. 3

---<

0.5 1970

1990

2010

2030

2050

0.5 1970

1990

_~

2010 _

_

_

2030

2050

2030

2050

~

b) needle mass (tonlha)

5

5 1970

1990

2010

2030

2050

c) Mg-uptake reduction by pH 40

40

1

"

"

1970

' 1990

2010

Mg-uptake reduction by pH

1 ..P-

30

.-..- _

-

\ _ - - -

.

- - . -m 20 - -L

1970

20

1990

2010

2030

2050

1970

1990

2010

2030

2050

Figure 8 Trends in nitrogen content of needles (a), needle mass (b) and Mg2' uptake reduction by low pH ( c ) on Leptic podzols in response t o the three emissioddeposition scenarios illustrated in figure 4 for the relatively unpolluted acidification area "Groningen" (left) and the polluted area "Venray" (right).

336

<20%

,-.

Figure 9 ratio of 1 (%) The percentage of area of forest soil exceeding a critical A13+/Ca3+ (a) and the nitrogen content of needles (%) (b) for the year 1990.

4. FUTURE MODEL DEVELOPMENTS The DAS model has shown to be a good tool in quantifying the effectiveness of acidification abatement strategies. The basic structure of the model, using modules that can be run and developed independently, has proven to be valuable. However, this approach has also lead to discrepancies between the modules. In the near future, the modules will be adapted in such a way that these discrepancies are eliminated. Results from experimental programmes that were carried out as a part of the Dutch Priority Programme on Acidification have resulted in detailed forestry data, such a s stomata1 response, nutrient uptake and nutrient allocation. As the forest module in DAS uses overall regression equations to account for the impact of environmental conditions (dose-effect relationships), experimental data cannot be direct incorporated. Instead, these dose-effect relationships will be derived from a process oriented dynamic model of forest growth FORGRO that is developed for use a t the stand level (Mohren et al., 1991). The forest module SOILVEG does not yet describe stress imposed by traditional factors, such as drought, frost and plagues. If these factors had been

337

included into the model, the predicted needle mass would have been lower. A simple water balance and dose-effect relationships for damage due to drought, frost and plagues will be incorporated. The occurrence of drought, frost and plagues will be simulated stochastically. At present, the forest module has been parameterized for Douglas fir only. For regional scenario analysis, parameterization of this module for the most important Dutch tree species (Scots pine and Oak) is important. Dose-effect relationships for air pollution impact for the forest module SOLVEG are also derived from the FORGRO model. For this reason, FORGRO will be calibrated for these tree species too. The soil and forest modules use a great deal of computer time. This imposes serious restrictions to the number of soil-forest combinations and scenarios that can be evaluated. For this reason, the summary model EXPECT (Braat et al., 1991) is now under development. This model will also cover other environmental problems. In the future, the DAS model will be used to assess the effectiveness of abatement strategies in detail, while the summary model EXPECT will be used to study large numbers of scenarios, such as environmental policies of political organisations, and to select key scenarios for detailed analysis with the DAS-model. The acidification module of EXPECT will be calibrated to the DAS model.

5. REFERENCES

Anonymous, 1985. Gegevens Emissieregistratie l e en 2e ronde (Data from the Emission Inventory System, first and second inventory). Centre for Applied Scientific Research, Delft, the Netherlands (in Dutch). Anonymous, 1988-1989. To choose or to lose: The National Environmental Policy Plan. Second Chamber, session 1988-1989, 21137, nr. 1-2, the Hague, the Netherlands. Asman, W.A.H. and J.A. van Jaarsveld, 1990. A variable-resolution statistical transport model for ammonia and ammonium. National of Public Health and Environmental Protection (RIVM) report nr. 228471007, Bilthoven, the Netherlands. Bakema, A.H., K.F. de Boer, G.W. Bultman, J.J.M. van Grinsven, C. van Heerden, R.M. Kok, J. Kros, J.G. van Minnen, G.M.J. Mohren, T.N. Olsthoorn, W. de Vries and F.G. Wortelboer, 1990. Dutch Acidification Systems model. Specifications. Dutch Priority Programme on Acidification 114.1-01, RIVM, Bilthoven, the Netherlands.

338

Berdowski, J.J.M., J.J.M. van Grinsven, C. van Heerden, J.G. van Minnen and W. de Vries, 1991. SOILVEG: a model to evaluate effects of acid atmospheric deposition on soil and forest. Volume 1: Model principles and application procedures. Dutch Priority Programme on Acidification 114.1-02, RIVM,Bilthoven, the Netherlands. De Boer, K.F. and R. Thomas, 1991. Emission and deposition scenarios for SO,, NO, and NH,.In: G.J.Heij and T. Schneider (eds.). Acidification research in the Netherlands. Final report of the Dutch Priority Programme on Acidification. Studies in Environmental Science (46):151-168. Elsevier Science Publishers, Amsterdam, the Netherlands. Braat, L.C., A.H. Bakema, K.F. de Boer, R.M. Kok, R. Meijers and J.G. van Minnen, 1991. EXPECT: An Integrated Model System for Scenario Analysis and Environmental Impact Assessment. National Institute of Public Health and Environmental Protection, in prep. Erisman, J.W., 1990. Atmospheric deposition of acidifylng compounds onto forests in the Netherlands: throughfall measurements compared to deposition estimates from inference. National of Public Health and Environmeiital Protection (RIVM) report nr. 723001001, Bilthoven, the Netherlands. Erisman, J.W. and G.J. Heij, 1991. Concentration and deposition of acidifying compounds. In: G.J. Heij and T. Schneider (eds.). Acidification research in the Netherlands. Final report of the Dutch Priority Programme on Acidification. Studies in Environmental Science (46):51-95. Elsevier Science Publishers, Amsterdam, the Netherlands. Ford, E.D., R. Milne and J.D. Deans, 1987. Shoot extension in Picea sitchensis. 11. Analysis of weather influences on daily growth rate. Annals of Botany (601543-552. Heij, G.J. and T. Schneider (eds.), 1991. Acidification research in the Netherlands. Final report of the Dutch Priority Programme on Acidification. Studies in Environmental Science (46) pp. 771. Elsevier Science Publishers, Amsterdam, the Netherlands. Heij, G.J. and W. De Vries, A.C. Posthumus and G.M.J. Mohren, 1991. Effects of air pollution and acid deposition on forests and forest soils. In: G.J.Heij and T. Schneider (eds.). Acidification research in the Netherlands. Final report of the Dutch Priority Programme on Acidification. Studies in Environmental Science (46):97-138. Elsevier Science Publishers, Amsterdam, the Netherlands. Zvens, W., 1990. Atmospheric deposition onto forests. An analysis of the deposition variability by means of throughfall measurements. Ph.D thesis, Utrecht, The Netherlands. Kozlowski, T.T., 1982. Water supply and tree growth. Part I. Water deficits. For. Abstr. (43157-95.

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Van Jaarsweld, J.A. and D. Onderdelinden, 1990. TREND: An analytical longterm deposition model for multi-scale purposes. National of Public Health and Environmental Protection (RIVM) report nr. 228603009, Bilthoven, the Netherlands. De Kam, M., C.M. Versteegen, B.C. wan Dam, J. wan den Burg and D.C. wan der WerL 1989. Effect of nitrogen and potassium fertilization on the development of Sphaeropsis sapinea in Pinus nigra. Acta Botanica Neerlandica (38):354. Keltjens, W.G. and E. wan Loenen, 1989. Effects of aluminum and mineral nutrition on growth and chemical composition of hydroponically grown seedlings of five different forest species. Plant and Soil (119):39-50. Kleijn, C.E., G. Zuidema and W. de Vries, 1989. De indirecte effecten van atmosferische depositie op de vitaliteit van Nederlandse bossen. 2. Depositie, bodemeigenschappen en bodemvochtsamenstelling van acht Douglasopstanden. STIBOKA report nr. 2050, Wageningen, the Netherlands (in Dutch). Mohren, G.M.J. (ed.), 1991. Integrated effects of air pollution and soil acidification on forests. In: G.J. Heij and T. Schneider (eds.). Acidification research in the Netherlands. Final report of the Dutch Priority Programme on Acidification. Studies in Environmental Science (46):387-464. Elsevier Science Publishers, Amsterdam, the Netherlands. Murach, D. 1984. Die Reaktion der Feinwurzeln von Fichten (Picea abies (Karst)) auf zunehmende Bodenversauerung. Gottinger Bodenkundliche Berichte (14l):l-126 (in German). Olsthoorn, A.F.M. and A. Tiktak, 1991. Fine root density and root biomass of two Douglas-fir stands on sandy soils in the Netherlands. 2. Periodicity of fine root growth and estimation of belowground carbon allocation. Neth. J. of Agricultural Science (39):61-77. Skefington, R.A. and T.M. Roberts, 1985. The effects of ozone and acid mist on Scots pine saplings. Oecologia (65):201-206. Thomas, R., W.G. van Arkel, H.P. Baars, E.C. van Ierland, K.F. de Boer, E. Buysman, T.J.H.M. Hutten and R.J. Swart, 1988. Emission of SO,, NO,, VOC and NH, in the Netherlands and Europe in the period 1950-2030. The Emission module in the Dutch Acidification Systems model. National Institute of Public Health and Environmental Protection (RIVM) report nr. 758472002, Bilthoven, the Netherlands. De Visser, P.H.B. and W. de Vries, 1989. De gemiddeld jaarlijkse waterbalans van bos-, heide- en graslandvegetaties. STIBOKA report nr. 2085, Wageningen, the Netherlands (in Dutch).

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De Vries, W. and J. Kros, 1989. Lange termijn effecten van verschillende depositiescenarios op representatieve bosbodems in Nederlands. The Winand Staring Centre for integrated land, soil and water research report nr. 30,Wageningen, the Netherlands (in Dutch). De Vries, W., A. Hol, S. Tjalma and J.C.H. Voogd, 1990. Literatuurstudie naar voorraden en verblijftijden van elementen in bosecosystemen. The Winand Staring Centre for integrated land, soil and water research, report nr. 94, Wageningen, the Netherlands. pp. 205 (in Dutch). De Vries, W. and G.J. Heij, 1991. Critical loads and critical levels for the environmental effects of air pollutants. In: G.J.Heij and T. Schneider (eds.). Acidification research in the Netherlands. Final report of the Dutch Priority Programme on Acidification. Studies in Environmental Science (46):205-214. Elsevier Science Publishers, Amsterdam, the Netherlands. De Vries, W. and J. Kros, 1991. Assessment of critical loads and the impact of deposition scenarios by steady state and dynamic soil acidification models. In: G.J. Heij and T. Schneider (eds.). Acidification research in the Netherlands. Final report of the Dutch Priority Programme on Acidification. Studies in Environmental Science (461569-624.Elsevier Science Publishers, Amsterdam, the Netherlands. De Vries, W., J. Kros, C . van der Salm and J.C. Voogd, 1991. Effects on forest soils. In: G.J. Heij and T. Schneider (eds.). Acidification research in the Netherlands. Final report of the Dutch Priority Programme on Acidification. Studies in Environmental Science (46):169-179.Elsevier Science Publishers, Amsterdam, the Netherlands. Wislicenus, H. (ed.), 1908-1916.Waldsterben im 19. Jahrhundert. Sammlung von Abhandlungen uber Abgase und Rauchschiiden. Berlin, Paul Parey. Reprint VDI-Verlag Diisseldorf, 1985 (in German). Zahner, R., 1968. Water deficits and growth of trees. In: T.T. Kozlowski (ed.). Water deficits and plant growth. Vol. 11. Academic press, New York, pp. 191-244.

T. Schneider (Editor). Acidification Research. Evaluatlonand Policy Applications 0 1992 Elsevier Science Publishers E.V. All rights reserved

34 1

Acid rain abatement in Belgium: lessons of cost-effectiveness studies C. Cuijpers and S. Proost Centre for Economic Studies, Katholieke Universiieit Leuven, Van Evenstraat 2B. 3000 Leuven. Belgium

Abstract Ln this paper a cost-effectiveness analysis is presented for combating emissions of acid precursors. The focus of concern is to reach the environmental quality goal at least cost. Two cost-effective approaches are elaborated. Firstly, the maximum allowable emission of each acid precursor seperately is allocated in a cost-effective way across the economic sectors. Secondly, the maximum allowable emissions of acid precursors are allocated in a cost-effective way across the three considered acid precursors as well as across the economic sectors. It is argued that not only the energy consumption but also the agriculfurul sector could play an imporiant role in a cost-effective strategy by curtailing its ammonia emissions.

1. PROBLEM DESCRIPTION The evolution of the cumulative emissions of acid precursors in Belgium for the period 1979-1989 is shown in Figure 1 [l]. After a considerable decrease in total emissions (35%) in the period 19791985. emission reductions seem to stagnate. The sulphurdioxide emissions in Belgium decreased by about 55% over the period 1979-1989. This was principally realized by the introduction of lower sulphur content fuels and the use of nuclear energy until 1986. Especially the electric power stations have reduced heir emissions drastically by approximately 71%. The nitrogenoxide emissions have hardly been reduced whereas the ammoniu emissions have increased.

0.6

0.5

0 3

,

I 79

_T 80

01

LV

83

a,

85

86

87

BB

a9

Figure 1. The evolution of the cumulative emissions of acid precursors in Belgium for the period 1979-1989 (in million tonnes).

342 Figure 2 and 3 put the emissions of acid precursors in a sectorial perspective. Figure 2 shows that most sulphurdioxide emissions comes from large combustion plants. The figures for 1989 indicate that 47% of the total SO, emissions come from power plants and 30% from industrial sources. Total SO, emissions in 1989 amount to 340,548 tonnes. The transplt sector is the principal emiaer of NO, in Belgium (see Figure 3). 'Ihe total emissions of nitrogenoxide in 1989 amount to 400,674 tonnes; almost 69% came from transport energy sector 4 7 36%

iranspLr 5 69%

reslmnt a 8 others E 35%

1

residential & others

9.62%

29 81%

energy sector 16.34%

Figure 3. The NO, emissions in a sectorial perspective (1989).

Figure 2. The SO, emissions in a sectorial perspective (1989).

The emissions of ammonia in 1989 amount to approximately 123,000 tonnes. Of this amount 90% came from intensive life stock breeding (production of manure); 9% from handling of manure and only 1% from the industry [2]. SO, and NO, have pollution effects across rather long distances, whereas ammonia precipitates relatively closely to the emission source. There is an important transboundary dispersion of acid precursors between Belgium and its neighbouring counuies. For instance, we establish that 51% of the SO, emissions and 62% of the NO, emissions deposit abroad [3]. We conclude that Belgium is a "net exporter" of sulphurdioxide and in particular of nitrogenoxide. It became increasingly clear that internationally coordinated efforts are required for significant reductions of acid rain. We start from a Dutch reduction programme for Western Europe which is represented in Table 1. This table gives the indicative emission level of each acid precursor for each considered country in order to reach a target loud of 1,400 eq/ha/y in 2010 and an interim objective of 2.400 eq/haly in 2000 in Western Europe [4]. Table 1 Indicative emission levels for Western Europe to attain the environmental quality goals.

1985

Yea emissions in 1000 tonnes Belgium Netherlands West-Germany France Great-Britain

SO, 467 274 2400 1845

NO,

NH,

385 120

539 2900 1693 3540 1690

2010

2000

258 535 957 714

SO, NOx NH,

SO, NO,

160 310 90 175 665 1535 710 1300 2555 1365

80 45 335 355 465

100

75 430 770 575

NH,

135 50 165 50 920 230 560 410 585 305

Reprinted from:RIVM, Zorgen voor morgen. Samsom H.D.Tjeenk Willink. Alphen aan den Rijn, 1989. p. 124.

343 Notice that the assigned emission levels of each acid precursor for each country are not set based on cost-effectiveness but principally on technological feasibility. The theoretical possible combinations of SO,, NO, and NH, also take the rransbounhry aspects into account. Furthermore the emission reduction programme assumes inrernurional cooperation. The indicative emission levels for Belgium call for a 77% reduction of SO2 emissions, a 66% reduction of NO, emissions and a 60% reduction of NH3 emissions to be reached by 2010 based on 1989 emissions. The acid deposition level in Belgium in 1989 amounts to approximately 4,681 eq/ha/y.

2. COST-EFFECTIVENESS STUDIES The general rule for determining the optimal level of environmental quality is that resources should be allocated to improvements of environmental quality as long as the marginal benefits are greater than the marginal abatement costs. The problem faced by this general formulation is the quantification of the marginal costs and benefits associated with different levels of reduction. The practical difficulties of estimating costs and benefits are immense (although this is well defined from a theoretical point of view). Therefore. we identify the deposition target from an ecological point of view and use costeffectiveness analysis rather then cost-benefit analysis. The purpose of this analysis is to describe a "least cost" acid emission control snategy for reaching the indicative emission levels of Table 1 for Belgium in order to get the target load of 1,400 eq/ha/y by the year 2010. Sources are controlled in function of increasing marginal abatement costs. until the assigned level of emissions is reached. This apparently simple methodological principle covers a multitude of complexities which must be adressed. These are associated to some extent with uncertainties in the knowledge of the appropriate reduction technologies. the corresponding emission reductions and the corresponding abatement costs. Since the desired abatement goals have been set somewhere in the future, it becomes clear that the technological evolution as well as the evolution of the energy consumption are important but uncertain factors. The starting point is a cost-effective comparison of appropriate reduction rechnologies applied to Belgian emission figures for 1989. Each of the reduction techniques will have a certain potential for meeting the reduction goal at a certain cost. A simple ranking system then prioritizes the various measures in terms of their cost-effectiveness. We apply this method to emission data per sector which are taken as given. We first of aU assume that the reduction in the economic activity generating the emissions (for instance production of electricity) is not a cost-effective way of reducing emissions. This assumption is not justified if the emission abatement is very costly and the demand for the product rather price elastic. The second assumption is that the economic activity in one sector is not influenced by abatement efforts of the other sectors (e.g. higher electricity demand as a result of scrubbing installations in industry). So we use a partial equilibrium approach. Another limitation is the neglect of fuel switching to other fuels with a lower emission of acid precursors. Figure 4 shows the evolution of the marginal abatement costs for SO, emission reductions graphically. We see that the marginal uburement costs are sharply increasing and the marginal abatement cost of high levels of sulphur removal are substantially higher then the marginal abatement costs of small reductions. Belgium is able to reduce the SO2 emissions by almost 84% given the actual available emission reduction techniques. A widespread abatement option is the switch to lower sulphw contenrfuels. The most common proposed abatement method for large combustion plants is flue gas desulphurization (FGD).This technique reduces SO, emissions by approximately 90% at a reasonable cost [ 5 ] . In a similar way as the previous marginal cost curve, Figure 5 shows the evolution of the marginal abatement costs for NO, emission reductions. We see again that the marginal abatement costs of high levels of NO, removal are considerably higher than marginal abatement costs of small reductions. The first 7.5% emission reduction is even costless (engine tuning). As it stands now, Belgium is able to reduce the NO, emissions by approximately 70%. The equipment of motorcars (gasoline) with a three way cafalytic converter and selective camlyric reduction (SCR)to reduce NO, emissions of large

220 210

-

-

200-

190

ieo 170 160 1%

I40

2

-

130-

ri I:: -

b

-

100 90

-

-

-

8070

-

60

-

50 40 30

-

20

-

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0

t m

,

,

,

,

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I

I60

4

0

%

I

I

2M

w

I

I

240

a

r

280

n

*

NO, emission reductions (loo0 tonnes)

Figure 5. Marginal NO, abatement cost function. The costs of NH3 abatement aTe badly known. Nevertheless various NH, reduction techniques Seem to indicate that the reduction costs per kilogram emission are relatively low. This is closely connected with the proposed reduction measures: decrease the mineral content of feed (as far as it may be cost neutral), cheap techniques to adjust accomodations of live stock breeding (abatement cost of approximately 12.000 BEF1989 per tonne), modification of handling of manure (approximately 55,500 BEF1989 per tonne abatement) and storage of manure (approximately 325.000 BEF1989 per tonne abatement). Considerable emission reductions appear to be possible in the agricultural sector.

3. COST-MINIMAL ALLOCATION OF EMISSION REDUCTIONS An appropriate cost-effectiveness analysis concerning the abatement of acid rain has in fact three

345 dimensions: cost-minimal allocation of emission reductions across countries (the importance of transboundary aspects), across pollutants and across economic sectors. We take Table 1 (see section 1) as given and carry out two cost-effectiveness analyses. Approach 1 carries out a cost-minimal allocation of emission reductions per pollutant across the economk sectors. Approach 2 starts from a global emission reduction of acid precursors for Belgium and allocates this reduction in a costeffective way across the pollutonzs as well as across the economic sectors. The time horizon is the period 1989-2010.

3.1. Approach 1 In this first approach, we carry out a cost-effective analysis by dividing the required emission reduction of each pollutant i (i=1,2,3) across the economic sectorsj (j=l ,...,8) in order to minimize the abatement costs. Formally we get:

x,

M%rl

(1)

CIJw

E,-X,-T,

S.t.

'd i-123

Where: marginal abatement cost function of pollutant i in sector j in BEF1989 total emission reductions of pollutant i in sector j in tonnes total emissions of pollutant i in tonnes in 1989 indicative emission level of pollutant i for Belgium in 2010 (see Table 1 section 1)

CijGij) Xij

Ei Ti

By applying this analysis we get for each acid precursor a cosr-minimal allocation across the economic sectors. Figure 6 shows the sectorial Cost-effective emission reduction graphically.

I

100

residential

trmaport

coke owns

tertiary

exnIuhr8duc~

refincrica

e l e c t r i c a t n t indutricr

abmtaplentccda

Figum 6.Costcffective SO2 emission reduction per economic sector.

8.0

346 What stands out from the figure is the important role of the indusm and electric power plants in curtailing the sulphurdioxide emissions. The industry has to reduce the SO2emissionsby 93% in 2010. Flue gas desulphurization installations seem to be of great importance in reducing SOz emissions in large combustion plants.The residential sector and transport play only a minor role in attaining the SOz emission target in 2010. Moving to lower sulphur content fuels is needed. A rough approximation of the total abatement costs to reach the desired emission level in 2010 amounts to 16 billion BEF 1989 yearly. Figure 7 deals with a cost-effectivereduction path of nifrogenoxide emissions.It shows in a similar way as the previous figure the cost-effective NO, emission reductions per sector. A considerable contributionof the transport sector in reducing their NO, emissions is needed. It stands out that 74% of the NO, emission reductions needed to reach the proposed emission level in 2010, should take place in the transport sector. The application of three-wuy caralytic converters plays an important role. The introduction offieb with a lower potential toform NO,seems to be important. A rough approximation of the total abatement costs to reach the desired emission level in 2010 amounts to 19 billion BEF1989 per year. 15.00 14 25

200

190 180

13 50 12 75 1200

170

~

3

160 150

11 25

1050 9 75

140

P

5

a

I30

u-

120

EEP uc

110 100 go 80

750 675

70

525

60 50

3 75

2

40

300

$

30

2 25 150

%

5 I

3

900 8 25

600

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gu

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450

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20

0 75

10

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0 coke ovcna

rcfincriea

rcaidcntial

Grtinry

e n h d o n reduction

induatry

clcctric a t a t trmaport

abatement c o a b

Figure 7. Cost-effectiveNOx emission reduction per sector. Since almost all m n i u emissions come from agriculture, the NH, emission target should be realized wholly in this sector. Thus the agricultural sector have to reduce their NH, emission by 73.000 tonnes by 2010. Modification of accomodations of live stock breeding and modification of the handling and storage of manure makes the attainment of the abatement goal possible at a reasonable cost of almost 900 million BEF1989 per year. 3.2. Approach 2 In this subsection, we attempt to allocate the necessary efforts in a cost-effective way across the pollutants i (i=1,2,3) as well as across the economic sectors j Q=l, ...,8) in order to reach the proposed target load of 1,400 eg/ha/y in 2010.

347 Formally we get:

Where: C,,(X,,) XI, E, T, 4 2,

marginal abatement cost function of pollutant i for sector j in BEF1989 total emission reductions of pollutant i in sector j in tonnes total emissions of pollutant i in tonnes in 1989 indicative emission level of pollutant i for Belgium in 2010 (see Table 1 section 1) exportcoefficient of pollutant i (L=0.74, t,,0=0.86, ~ = 0 . 6 0 ) acid equivalent of pollutant i (z,=3 1.5,~,,,=21.5.~=59.0)

The total emissions of acid precursors in Belgium were 26,599 megaequivalents in 1989. In correspondence with Table 1 (see section 1 ) the indicative total emission target in 2010 amounts to 8.373 rnegaequivalents. The fact that we assume international cooperation to reach the environmental quality goal is reflected by the two restrictions (inequality 4 and 5). If we assume that the surrounding countries make efforts to reach the proposed emission levels of Table 1 and Belgium takes only account of the first restriction (4) by minimizing the total abatement costs (equation 3) then the cooperation is broken. In this case Belgium does not care about a possible increase of export of emissions. In other words, Belgium is not concerned with the deposition level in neighbouring countries. In this case Belgium reaches the quality target but the surrounding countries have higher deposition levels than expected because of the export of emissions from Belgium which will be higher than agreed. In this case 6 98

5 82

4.66

3.49

2 33

116

0 reddentidcoke

0i-s

t a w

niincriclldectric .tat indrutry +culture

Figure 8. Cost-effective emission reduction per sector.

trm.port

348 What stands out from Figure 8 is the important role of the agricultural sector. The agricultural sector have to reduce their emissions by 71%. At the same time we may conclude that NH, emission reductions seem to be very cost-effective. Also NOxemission reductions Seem to be cost-effective. The transport sector needs to reduce its emissions, in terms of acid equivalents, by the largest amount. Almost 28% of the total emission reductions is needed in the agricultural sector (NHJ and almost 29% of total emission reductions is needed in transport (Nod, in terms of acid equivalents. The cost-eflective allocation of the emission reductions across the acid precursors is as follows: a reduction 256.00 ktOnne nitrogenoxide; 246.60 ktonne sulphurdioxide and 87.40 ktonne ammonia in order to reach the deposition objective in 2010. Notice the correspondence of this cost-effective division across pollutants with the indicative emission reductions represented in Table 1 in section 1 (namely 260.55 ktonne SO2, 265.67 ktonne NOx and 73.3 ktonne NH,).A rough approximation of the total abatement cost to reach the environmental quality goal in 2010 is 17 billion BEF1989 per year.

4. MAIN FINDINGS Despite the limitations of this analysis, the methodology described here seems to present a useful way of prioritizing efforts in acid rain abatement. The new insights will be presented briefly: (1) The Belgian abatement programme gives priority to measures against emissions by large comburtion sources and motor vehicles. But apart from these sources, the agricultural sector is also an important emission source. Furthermore it is cost-effective to reduce ammonia emissions. (2) Belgium has already restricted the sulphur content of fuels to 0.2% for gasoil; 1.0% for medium oil and 2.0% for heavy fuel oil. This cost-effective abatement measure has to be tightened. (3) Belgium has uniform SO2 and NO, emission standards. These measures encourage the adoption of SCR units and FGD installations which are cost-effective emission reduction techniques. (4) Since 1990, all new gasoline-fuelled cars with a cylinder capacity larger than 2oooCc have to be equipped with a three way-catalytic converter. The same requirement will be established in 1993 for all gasoline-fuelled cars with a cylinder capacity smaller than 2oooCc. Before these requirements are enforced, the installation of a three-way catalytic converter is encouraged by a subsidy. Besides this cost-effective policy, the introduction of fuels with a lower potential to form NO, is needed. (5) Only a minor intervention is needed in the residential and tertiary sector. This work has been supported by the Belgian Ministry of Economic Affairs. The views defended here are the authors' views and not necessarily those of the Ministry.

5. REFERENCES [ l ] CUIJPERS, C., PROOST. S.. VAN REGEMORTER, D. De recente evolutie van het Belgisch eindenergieverbruik. Opvolging met een behulp van een systeem van indicatoren. Brussel, Ministerie van Economische m e n , 1990, 78 p. [2] MINISTERIE VAN DE VLAAMSE GEMEENSCHAP. Milieubeleidsplan en natuurontwikkelingsplan voor Vlaanderen. Voorstellen voor 1990-1995. s.I., Gemeenschapsministerie van Leefmilieu, Natuurbehoud en Landinrichting, 14 februari 1990. [31 EMEP. Calculated budgets for airborne sulphur and nitrogen in Europe. Technical report no. 86,Det Norske Meteorologiske Institutt, November 1990. [4] NEDERLAND. RIJKSINSTITUUT VOOR VOLKSGEZONDHEID EN MILIEUHYGIENE. Zorgen voor morgen. Nationale milieuverkenning 1985-2010. Alphen aan de Rijn, Samson H.D. Tjeenk Willink. lste druk. 4de oplage. 1989,456~. [5] BEEK, W.J.. STALLEN. P.J.M. Naar een nationaal beleid tegen venuring. Een verslag van een forumproces. 's Gravenhage. Stichting Maatschappij en Ondememing. 1989,125~. [6] LANDELUK MILIEUOVERLEG. Financitle instrumenten voor het Nederlandse milieubeleid. Utrecht. Landelijk Milieuoverleg. 1990, 104p.

T. Schneider (Editor). Acidification Research. Evaluationand Policy Applications 1992 Elsevier Science Publishers E.V.

Base content in soil and problems arising acidification

349

in connection w i t h

L. Werner Institute for Ecosystem Research Berlin-Halle, Magdalenenstr. 17-19, 0-1130 Berlin, Germany

Abstract Current existing dynamic and static acidification models are not very suitable for making large but detailed statements about the sensitivity of soils with comparison to atmospheric acidification processes due to the great demand for data. Conclusions about the acidification endangering of soils can be drawn from its exchangeable base supply which is accessible as V-value. 7 classes for endangering acidification have been classified according to the represented systematic. This methodology was applied for all regions of Thuringia(Germany). 1. Buffering areae of soils

The soils are able to oppose various buffer mechanisms to the internal or external acid production. Under natural conditions several buffer processes are mostly effecting in parallel. Finally, the dominating buffer process results from the chemical soil composition. Following buffer areas are distinguished by ULLRICH (1981). see also DVGW (1988):

pH: 8,3 - 6,2 buffer rate: 2 kmol H+ per ha*a

calcium carbonat

:

silicate

: pH:

exchanger

: pH: 5,O - 4,2

6,2 - 5,O buffer rate: 0 , 2 - 2 kmol H+ per ha*a

buffer rate: 0,2 kmol H+ per ha*a aluminium : pH: 4,2 - 3.8 aluminium and iron: pH: 3,8 - 3.0 iron : pH: 3,O - 2,4 buffer rate: 2 kmol H+ per ha*a (as long as sufficient aluminium hydroxides are existing) Figure

1 shows that the several buffer areas are

characte-

350 rized by a typical base content. Concerning the calcium carbonate, silicate and exchanger buffers the content of exchangeable besee in soil is relatively high and decrease8 rapidly only with pH under approximately 5 , O . Hence it follows that the buffer area of soil and its sensitivity in opposite to acid depositions can be estimated by means of the exchangeable baee content.

Figure 1. Buffer areas in soil (from UHLMANN 1989). 2. b e e content in soil

Generally it is valid that a high cation exchange capacity (CEC), connected with a high base saturation (V), guarantees A good buffer action to internal and external acidification processes. Both values are often specified for the general chemical characterization of soils simultaneously with the pH. Thue they are accessible to general statements. In this connection literature data (MUECKENHAUSEN 1977, MUELLER 198U, LIEBEROTH 1982. SCHEFFER/SCHACHTSCHABEL 1989) were evaluated and a system of fundamental characteristics for typical german soils was elaborated.

35 1 The S-value is the expression for the content of exchangecapaable bases (Ca, Mg, K, Na) whereas the cation exchange city of a soil is given by the T-value. Usually, the calcium ions are clearly dominating. The percental share of bases in the exchanger complex expressed by the V-value is of decisive importance but not the absolute value of the base content for estimating the buffer efficiency.

V =

(S /

T)

*

100

A clear interaction between V-value and pH (KC1) is evident in table 1. The pH only drop under 4,O if the V-values are under 20%. This diaposon corresponds to the transition from the exchanger buffer to the aluminium buffer. According to the literature data the corresponding V-value varies between 5 and 20 %. If this value is still below the buffering is breaking down because of the cation exchange. Soil with a Ca+Mg share in the effective cation exchange capacity of < 15 % is described by MEIWES et al. ( 1 9 8 4 ) as "slightly" elastic in relation to acidification processes and < 5 % as "very slightly". Table 1 shows the variety of the existing soil types. The buffer capability of soil towards acidification processes reaches from very good (for chernozem) to slight (for Podzol, gleyic mollic gleysol). But it has to be considered that it is a question of a potential endangering by acidification bearing in mind only the soil characteristics. Further impordepositant influencing variables , like e.g. land use and tion rates must be taken into account for a general evaluation. Such estimation of endangering is very favourable,

Table 1 Classification of typical german soils after its endangering (up to 0 , 5 m depth)

acidification

____________________-------_-------------------------------Nr . pH(KC1) S T V CaO C/N Clay Cmeq/lOOgl [%I [mg/lOOgl [%I

________________________________________-----------999 18,5 93 236 6,9 13,7 14,7 I I1 111

IV V VI

VI I VIII

6,5 6,l 5,9 5,o 4,6 4,1 3

9

8

IX

3,7

X

3

XI XII

9

8

3,6 397

19,2 14,7 12,l 10,8 4,2 4,6 3,l 2,6 3,8 3,5 2,l

24,8 17,8 18,9 18,5

10,2

23,7 23,4 19,9 31,6 28,4 35,6

83 83 64 58 41 19 13 13 12 12 6

346 270 228 220 46 59 40 22 52 6 22

14,l 12,l 14,O 11,4

-

21,4 21,l 17,9 22,5 28,3 28,8

34,O 17,l 8 9 1 24,l 10,2 7,2 15,6 10,3 10,7 2,6 16,2

________________________________________----------Legend :

I I1 I11 IV

chernozem rendaina (calcic) cambisol (calcaric) gleysol

352

V VI VII VIII IX X

XI XI1

gleyic Phaeozem (orthic) Luvieol Ranker (dystric) Cambieol (dyetric) Luvisol gleyic Podzol Podzol gleyic Podzoluvisols

espescially for foreet sitee while the impacte of acid atmospheric depositions on agricultural intensively used areas can be neglected by liming and fertilizing measuree. The acidification state of a soil can also be estimated by the C/N relation in the surface soil. Uneuitable humus forms with high internal acid production are moetly represented by a C/N-relation > 20. The evaluation of soil aeeociation textures has also been provided by literature data according to the clay content (fraction < 0,002 nun). On the baeie of table 1 a generalization about the base eaturation was carried out and ae a result of thie 7 endangering claeeee were defined. Therefore eoils of the endangering claeeee 1 - 3 are at least in the silicate buffer area and can be regarded a8 not much sensitive towards acid depositions. Soils of class 4 and 5 are in the traneitional stage between exchanger and aluminium buffer area. Insufficient or no buffering in the exchanger buffer meane that the system has already been located in the aluminium or iron buffer and can be regarded as constantly acidified. Relating to the exchanger buffer following claeeee of acidification endangering are exieting:

v

CEC Cmed100 81 1) very good buffering 2) good buffering 3) sufficient buffering 4 ) average buffering 5) weak buffering 6) unsufficient buffering 7) nearly no buffering

> 15 > 15 > 10 > 10 > 10 > 10 > 10

[%I

> 90 90 80

40 20 10 80 60

60

40 20

< 10.

4. Estimation of acidification endangering in Thuringia

The Land Thuringia has a territory of about 15 300 km2 . It comprises the uplands ( forest Thueringer Wald and Thueringer Schiefergebirge) as well as the lowlands (Thueringer Becken). The estimation of the endangering hae to be realized on the baeis of the soil types characterized in table 1 because an area-covering mapping of the base saturation of soils was not available for Thuringia. As initial material the s o i l cartography has been applied in

353

the map of former GDR, on a scale of 1:750.000. 108 different soil associatione are distinguished that are included in the syetematic representation of the endangering classes. Baeed on it a detailed map of soil sensitivity towards acidification processes has been prepared for Thuringia (Figure 2 ) . The same is being planned for the whole territory of the new Bundealaender as well as for the whole Federal Republic of Germany. In any case the direct evaluation by the exchangeable base content is better than the classification by the existing soil types. On principle, these values are existing for forest locations of former GDR, but they still require a considerable deal of work in order to use them by means of EDP. Due to the prevailing soil associations following coarse evaluation about the acidification endangering can be provided

Figure 2. Territories endangered by acidification in Thuringia.

354 for Thuringia. An unsufficient buffering in the exchanger buffer, what goes equivalent with the transition to the aluminium or iron buffer, can be proved in two areas. It concerns the ridges in the forest Thueringer Wald, especially the territory around Eisenach across Oberhof to Gross-Breitenbach and Neuhaus, on the one hand. The predominating soil types are podzol and gleyic loamy cambisol. The other territory comprises the area around Hermsdorf and Neustadt. There are existing especially wet and sandy cambisol. The exchanger buffer is used up in great parts of Thuringia, in particular in the altitudes of Thueringer Wald, in the mountains Thueringer Schiefergebirge, forest Frankenwald as well as Holz- and Oberland. The transition into a sturdy acidified state is feared to be expected when the acid atmospheric depositions are kept unchanged. In contrast to this fact an acidification endangering does not take place in the areas with chernozem and loess of the basin Thueringer Becken and foreland of the mountains. These results are in good accordance with the local investigations. This map represents the first estimation about endangering for large areas compared with the acid atmospheric depositions in the new Bundeslaender. The methodology applied for it has the advantage that it can use the present information for the whole territory. At the same time it is able to provide detailed results. 5 . Base deficit

Due to the above mentioned explainations a comparison between the actual state of soil and the situation striving for can be carried out under the present local conditions and with the available knowledge about the described soil data. The difference between the actual and destination state regarding the base saturation is called base deficit. A target factor must be defined for determinating a deficit. It has been accepted a homogeneous pH (KC1)-value of 4 , 5 as temporary approximate value for the acid tolerance of forest soils in our territory. This value correponds to a base aaturation of about 40% (table 1). A homogeneous value for a territory of about 100.000 km2 can only be evaluated as first coarse target factor which had to be specified and differentiated according to the respective terms of the location in the future. The cartography of soil types for the new Bundeslaender and its classification by table 1 depicts the basis for the calculation of the base deficit by equation (1).

BD = (0,4 * T - S)

*

D

*

M

*

100 (keq/hal)

(1)

BD - base deficit D - dry bulk density of soil (g/cm3 ) M - thickness (0,5 m) Exposed regions result

for the

dystric

sandy soils in

the

355

Figure 3. Base deficit of soils in the new Bundeslaender. lowland as well as for the ridge of the highlands. The good buffer characteristics of soil in the chernozem and loese regions are clearly expressed. Thereby, negative base deficits represent a base surplus referring to the target factor. Table 2 shows the different share of areas with a base deficit in the case of various tolerable pH(KC1)-values. About 75 % of the grid areas have a base deficit in the case of pH 4 , 5 being achieved whilst it is still above 50 % with pH(KCl)= 4,O. Table 2 Share of the areas with base deficit with pH(KCl)-values 5,5, 4 , 5 and 4,O

This fact emphasizes the danger which can result long-term high acid deposition for large regions.

from

of

a

356 References: DVGW:

Geftihrdung der Trinkwasserversorgung in der BRD durch "Saure Niederachllge". DVGW-Schriftenreihe, Wasser Nr. 57, Wirtschafts- und Verlagageeellschaft Gas und Waseer Bonn, 1988. LIEBEROTH, I.: Bodenkunde. VEB Dt. Landwirtschaftsverlag Berlin, 1982. MEIWES, K.-J., N. KONIG, P.K. KHANNA, J. PRENZEL, B. ULRICH: Chemieche Untereuchungaverfahren fiir Mineralbbden, Auflagehumus und Wurzeln zur Charakterieierung und Bewertung der Versauerung in WaldbClden. Berichte des Fo.-zentrums Wald6kosystemeflaldsterben, Universitdt Gejttingen (1984)7, S. 1-87. MUCKENHAUSEN, E.: Entstehung, Eigenschaften und Systematik der Bbden der Bundesrepublik Deutschland. 2. Auflage, DLG-Verlags-GmbH Frankfurt am Main, 1977. MULLER, G.: Pflanzenproduktion - Bodenkunde. VEB Dt. Landwirtachaftsverlag Berlin, 1980. SCHEFFER, F., P. SCHACHTSCHABEL: Lehrbuch der Bodenkunde. 12. Auflage, Enke-Verlag Stuttgart 1989. UHLMANN, W.: Die Wirkung saurer atmosphlrischer Depoeitionen auf Boden und Grundwasser - Eine Abschltzung auf der Basis hydrogeochemischer Modelle. Dissertation A, Techn. Universitlt Dresden, 1989. ULRICH, B.: dkologiache Gruppierung von B8den nach ihrem chemischen Bodenzuetand. 2. Pflanxenern. Bodenkunde 144(1981), S. 289-305. ULRICH, B.: Theoretische Betrachtungen des Ionenkreislaufes in Waldbkoeystemen. Z. Pflanzenern. Bodenkunde 144(1981), S. 647-859.

T. Schneider (Editor), Acidification Research. Evaluation and Policy Applications @ 1992 Elsevier Science Publishers B V. All rights reserved

357

Measurements of Tree Growth and Health in the Liphook Forest Fumigation Project: An Evaluation of Large Scale Open Air Fumigation Experiments M. R. Holland' and P. W. Mueller

Imperial College at Silwood Park, Ascot, Berkshire SL5 7PY, UK. INow at the Energy Technology Support Unit, B149, Harwell Laboratory, Oxfordshire OX1 1 ORA, UK.

Abstract Over 4000 young trees were planted in the open air in a natural forest soil near Liphook in SE England in March 1985. They were fumigated with realistic and fluctuating concentrations of SO, and O3 from 1987 to 1990. More than 40 groups of scientists were active on the site, studying various aspects of the ecosystem. Much was learnt of the strengths and limitations of the open air fumigation technique for forest decline research. This is illustrated with reference to effects on tree growth, frost damage and the appearance of treatment related foliar symptoms. 1. INTRODUCTION

Most ecological experiments involve some degree of compromise between the need for realism and the resources that are available. This is certainly true for pollution effects research, especially when dealing with trees. A number of techniques (closed chambers, open top chambers and field release), varying in the degree to which the environment can be controlled, authenticity of exposure conditions and cost have been developed to investigate this subject. The advantage of a varied experimental approach to the problem of forest decline and the influence of pollution on plants in general has been the subject of a recent review [ 11. Most experiments in which the effects of pollutants on trees have been studied have used seedlings, typically no more than 4 years old, grown in chambers. These differ in many ways to mature specimens, including biomass distribution, needle age structure, extent of rooting zone, water and nutrient requirements and the ability to produce fruit. A further limitation is that seedling trees will rapidly outgrow most chambers. Hence it is usually not possible to use such systems to fully evaluate the impact of pollution on processes that may take a number of years to develop. This is especially important with respect to the long term processes that affect soil integrity, as these are believed to be responsible for many of the declines observed in Europe and elsewhere in recent years [2]. Following the results of several studies performed in fumigation chambers, demonstrating that ambient levels of pollution were capable of causing damage to young trees in pots [3, 41, and the development of sophisticated devices and techniques to monitor

358 and control fumigation levels [ 5 ] , the Liphook Project was designed to investigate the responses of complete coniferous forest ecosystems to pollutants. The purpose of this paper is to describe some of the observations made on tree growth and health during the project and the advantages of this type of experiment. 2. THE LIPHOOK FOREST RTMIGATION EXPERIMENT A total of more than 4000 two year old seedlings of Scots pine (Pinus sylvestn's L.), Norway spruce (Picca abies Karst.) and Sitka spruce (Picea sitchcnsis (Bong.) Cam.)were planted at Liphook in March 1985. They were planted at 1 m spacing in 7 plots, each of which was 30 m in diameter. The soil was a humoferric podzol of pH 4. Prior to planting the soil was rotovated to a depth of 15cm and fertilized with 400 kg ha1 N P K (0:20:20) to improve uniformity. Plots were further fertilized with 30 kg N ha-I in the form of Nitrochalk in September 1985 and 330 kg ha-l agricultural NPK (15: 15: 15) in April 1986. Treatment gases were supplied through a circle of pipes 50 m in diameter, split into 4 quadrants, around each plot. Gas release was related to wind direction to prevent wastage and to reduce to insignificance any transfer of fumigants between plots. Each quadrant contained 16 upright pipes from which fumigants were released at 2 levels, dependent on the height of the trees. This system produced a uniform horizontal distribution of hourly mean gas concentration across the central 25 m diameter area of a plot. Fumigation with SO, started in May 1987 and with 0, in May 1988. The 0, generator was switched off for the winter of 1988/89 in order to fit a scrubber to remove cogenerated N,O,. 3 levels of SO, were used; ambient (with a mean concentration over the fumigation period of 4 ppb), low fumigation (13 ppb) and high fumigation (22 ppb). Levels in treatment plots were computer controlled and followed the pattern of variation observed in a long term data set collected in the English midlands in 1980/81 [6] to provide realistic SO, fluctuation. Gas delivery was stopped when wind speeds were too low ( < 1 m s-1) for the system to function correctly. 2 levels of 0, were used; ambient and 1.5 x the prevailing ambient at any time. Fumigation continued until the end of 1990. Some measurements were continued during 1991. The experimental design was a 2 x 3 factorial, with no replication of treatment apart from the presence of 2 unfumigated plots. By the end of 1990 mean heights for each species were 1.3 m, 2.2 m and 2.6 m for Norway spruce, Sitka spruce and Scots pines respectively. The largest individuals were more than 4 m tall. These are clearly much larger than could be accommodated in conventional fumigation facilities. More than 40 groups of scientists from institutions throughout the UK and elsewhere in Europe were involved in this work, studying many aspects of forest ecology, including soil chemistry and hydrology, pollutant deposition, tree health, physiology and growth, population dynamics of herbivorous insects, mycorrhizal succession, lichen growth and development of phylloplane fungi. This has allowed the study of numerous interactions between pollutant effects. Technical details of the experiment are given in greater detail elsewhere [7].

359

3. RESULTS AND DISCUSSION The 3 species of tree reacted differently to the soil conditions on the site. Scots pine was ideally suited to it and grew well throughout. Sitka spruce was not well suited to the site, and suffered from nitrogen deficiency after 1988. Norway spruce became nitrogen deficient shortly after planting. To prevent heather (Calluna vulgaris, L.) check developing, all heather was cleared manually from the experimental areas in the spring of 1986. Following consultation between the various groups involved, further manipulation was deemed inadvisable and unnecessary, in view of the fact that nitrogen is limiting in many forest areas [8]. This decision proved to be beneficial in a number of cases, including the detection of increased foliar nitrogen in the N-deficient spruces through the codeposition of NH, with SO, [9]. Increased foliar N concentration was not observed in the Scots pines which were not N deficient, although the codeposition process was active within Scots pine stands. Variation within and between plots (which must be of concern in any unreplicated experiment) on the Liphook site did not prevent the detection of interesting and contrasting effects on growth for all 3 species. These effects reflected the degree and type of stress experienced by each species. Analysis using measurements made on trees growing on the site before it was cleared to make way for the experiment demonstrated that these effects were much more dependent on treatment than variation in conditions across the site. An example of the observed growth responses is shown in figure 1 which shows that Norway spruces were affected in the very first year of fumigation. No effect was observed on growth of the less stressed Scots pines until the end of the experiment, in spite of the presence of a number of deleterious symptoms. These results will be reported in detail in a future paper.

,;i

r = 0.m7,p c 0 . 0 ~

280

'-1

.

ieo

Figure 1. Reduction in growth rate of Norway spruce in response to SO, in the first year of fumigation. Points show the mean increment in main stem basal area during 1987 (mm2 y e a r ' ) for each plot. Each point was calculated from measurements on > 100 plants.

360

Frost damage to newly emerged spruce shoots was assessed as necessary in the spring of each year. No such damage was recorded on pines or in the autumn of any year (the time at which most studies of cold stresslpollutant interactions have been performed [lo]) as conditions in the SE of England are typically too mild. On most occasions the pattern of frost damage observed on Norway and Sitka spruce was dominated by site effects; the site was on a hill and the more exposed plots at the top of the site suffered less damage than the sheltered plots at the bottom of the hill. However, on one occasion, in early May 1989, more frost damage was observed in the 2 high SO, plots than elsewhere (figure 2).

Figure 2. Increased numbers of Norway and Sitka spruces damaged by frost in plots exposed to high SO, following a frost event on 6th May 1989. Frost damage was dominated by site effects on all other occasions. This result was not repeated on any other occasion, and so may simply have been a random event. However, the dominance of site effects on all other occasions on which frost damage was recorded makes this explanation appear unlikely. It is possible that elevated foliar nitrogen concentrations led to an advancement in the date of budbreak. It seems likely that the detection of such a pattern was heavily dependent on the date of the frost event. Although this result is certainly not conclusive evidence for an SO,/frost interaction it is of interest because the frost event was not artificially induced and affected trees were growing under natural conditions. Put in context it thus forms a useful addition to those studies in which excised shoots have been subjected to low temperatures in freezing cabinets [lo].

361 Typical foliar symptoms of exposure to SO, were observed on Scots pines in 1988, 1989 and 1990. A range of symptoms were observed on Norway spruce, though these were less determinate than those on the pines. It was thus necessary to record observed symptoms on Norway spruces in great detail. A total of 9 different forms of injury were identified. Some were caused by insects of various types, others appeared to be of fungal origin. Regular monitoring was essential to distinguish between different forms of injury. The distribution of only one type was closely related to SO, treatment in 1988 and 1990 (no such damage was observed in 1989). This took the form of the rapid death of newly expanded needles from the base of shoots towards the tip in the upper half of the canopy, particularly on the leader and uppermost whorl. Whether SO, was the sole cause of this form of injury or not remains uncertain. The total number of plants affected was small, and the % leaf lost as a result was typically only between 1 and 3 % of the total for any plant. This result reflects the advantage of running an experiment with large numbers of trees. It is possible that this response would not have been seen had fewer trees been used. Leaf fall monitoring in young stands of trees is difficult. Use of conventional litter traps, placed at random within a stand, requires a good degree of canopy closure. Pines at Liphook did not close canopy until 1989. Sitka spruce canopies were only partially closed by the end of the experiment and Norway spruce canopies remained open throughout. Several methods were tested at Liphook, including the use of litter traps, analysis of quantities of needles in each leaf age class at annual harvests and simple observations of needle retention in the field. The third of these methods turned out to be the most useful as all trees on the site could be sampled each year. The other 2 methods were used to check that the observations made in the field were reasonable and that observer bias was not significant. Results showed that leaf retention on Scots pine and Norway spruce was significantly affected by treatment in some years. Leaf fall among Sitka spruce was affected primarily by natural infestations of Elarobium abierinum, the green spruce aphid.

4. CONCLUSIONS

Significant effects of treatment on tree growth and health have been detected in the Liphook Forest Fumigation Project. The lack of replication of treatment in the experimental design has been compensated for both by the numbers of trees involved and the duration of the experiment. It is doubtful whether a number of the findings made could have been observed using conventional chamber techniques. This experiment thus makes a very useful contribution to the assessment of the role of pollution in recent forest declines. A number of lessons have been learnt that will be useful in future studies of tree growth of this kind. The detection and quantification of codeposition processes between SO, and NH, on the site is an excellent example of the benefits of large scale experimentation involving scientists from a range of disciplines working on a single site. The first signs of this effect were detected in measurements of foliar nitrogen concentrations. To demonstrate that the observation was caused by codeposition rather than any one of a number of other pathways and to quantify the amount of nitrogen concerned, involved atmospheric chemists, soil chemists and hydrologists, plant physiologists and microbiologists.

362

The absence of replication of treatment prevented analysis of some interactions between ecosystem components. There were clearly only 2 ways in which replication could have been increased; by reducing the number of treatments or by increasing the cost of the experiment, which was not possible. Evaluation of the project at this stage suggests that the decision to increase the number of treatments rather than replicate has proved beneficial.

5. ACKNOWLEDGEMENTS The author would like to thank Drs. A. McLeod and P.J.A. Shaw and Mr. K. Alexander of NPTEC, and Prof. A.J. Rutter of Imperial College for advice and assistance. Work on tree growth was funded under the Joint Environmental Programme of National Power and PowerGen and by the European Community.

6. REFERENCES

I . Unsworth, M.H. (1991) Air pollution and vegetation: hypothesis, field exposure, and experiment. Proc. R. SOC.Edinburgh., 97B, 139-153. 2. E.-D. Schulze (1989) Air pollution and forest decline in a spruce (Picea d i e s ) forest. Science, 244, 776-783. 3. Garsed, S.G. and Rutter, A.J. (1984) The effects of fluctuating concentrations of sulphur dioxide on the growth of Pinus Jylvesfris L. and Picea sitchensis (Bong.) Cam. New Phytol., 97, 175-195. 4. Skeffington, R.A. and Roberts, T.M. (1985) The effects of ozone and acid mist on Scots pine saplings. Oecologia, 65, 201-206. 5. McLeod, A.R., Fackrell, J.E. and Alexander, K. (1985) Open-air fumigation of field crops: criteria and design for a new experimental system. Atmos. Environ., 19, 16391649. 6. Martin, A. and Barber, F.R. (1981) Sulphur dioxide, oxides of nitrogen and ozone measured continuously for two years at rural sites. Atmos. Environ., 15, 567-578. 7. McLeod, A.R., Shaw, P.J.A. and Holland, M.R. (1991) The Liphook Forest Fumigation Project: Studies of sulphur dioxide and ozone effects on coniferous trees. For. Ecol. Manag., in press. 8. Kenk, G . and Fischer, H. (1988) Evidence from nitrogen fertilisation in the forests of Germany. Environ. Pollut., 54, 199-218. 9. McLeod, A.R., Holland, M.R., Shaw, P.J.A., Sutherland, P.M., Darrall, N.M. and Skeffington, R.A. (1990) Enhancement of nitrogen deposition to forest trees exposed to SO,. Nature, 347,277-279. 10. Fowler, D., Cape, J.N., Deans, J.D., Leith, I.D.,Murray, M.B., Smith, R.I., Sheppard, L.J. and Unsworth, M.H. (1989) Effects of acid mist on the frost hardiness of red spruce seedlings. New Phytol., 113, 321-335.

SESSION E RESULTS FROM NATIONAL RESEARCH PROGRAMMES

This Page Intentionally Left Blank

T. Schneider (Editor). Acidification Research. Evaluation and Policy Applications 1992 Elsevier Science Publishers B.V.

365

THE UNITED STATES NATIONAL ACID PRECIPITATION ASSESSMENT PROGRAM P a t r i c i a M. I r v i n g

The National Acid Precipitation Assessment Program Office of the Director, 722 Jackson Place, NW Washington, D.C. 20503, U.S.A. ABSTRACT

The National Acid Precipitation Assessment Program (NAPAP), created by law in 1980 as a 10-year research and assessment program, was recently reauthorized by the 1990 Clean Air Act Amendments. The mandate of the first decade of activities of the Program was to determine the causes, effects, and control of acidic deposition, and to provide advice to the President and Congress on various policy options relating to acidic deposition. In its next phase, NAPAP will continue studies where important knowledge gaps remain and will expand its mission to evaluate the effectiveness of the acid rain controls required under the 1990 Clean Air Act Amendments. Thus, for the first time in the U.S., environmental legislation includes a mechanism for evaluating its effectiveness. About 90 billion dollars is spent each year in the United States to comply with federal environmental regulations. The new Clean Air Act Amendments will add another $30 billion to that cost (about $4 billion per year will be spent to control acid rain). Obviously this substantial commitment involved difficult decisions, especially when considering other important societal goals requiring expenditures. Therefore, to guide future environmental decision-making, the federal government, through NAPAP, wi 1 1 assess the benefits to society from such legislation, including its costeffectiveness. 1.

INTRODUCTION

The Acid Precipitation Act of 1980 mandated a comprehensive ten-year national program in the United States of policy-oriented research and assessment on the causes, effects and control of acidic deposition. This program, known as the National Acid Precipitation Assessment Program (NAPAP), conducted studies on emissions from fossil fuel combustion and their relationship with effects on the environment and human health. The nation now has a new policy for controlling acidic deposition, Title I V of the 1990 Clean Air Act Amendments. The scientific information compiled by NAPAP helped provide the technical foundation for this policy. The success of this new $4 billion annual strategy to reduce the adverse effects of acid rain will have important imp1 ications for upcoming decisions affecting the energy/environment system. Policy choices in the future will benefit from accurate information about the outcome of past choices. Citizens

366

want confirmation of the benefits of policy decisions that have significant impacts on their pocketbooks. Without an active program to monitor emissions, deposition, and effects, and the continuation of long-term research and modeling activities, accurate evaluations will not be possible. The reauthorization of the National Acid Precipitation Assessment Program by Title IX of the 1990 Clean Air Act Amendments represents the first time that environmental legislation has a built-in mechanism for its own evaluation. The following summarizes the results of NAPAP's first decade of activities and describes it's approach to coordinate acid rain related activities of the federal government so they are of the greatest possible use in supporting future decision-making

.

2.

FUNCTIONS OF A NATIONAL PROGRAM

Enhance Cooperation and Facilitate Coordination The U.S. government spends more than $2 billion each year on environment-related research and development designed to support the missions of at least a dozen federal departments and agencies. These highly decentralized programs are not components of a well-defined, integrated system. Yet large-scale environmental issues such as acidic deposition require technical input for decision-making that cuts across many interests such as energy production, natural resource management, pollution abatement, and environmental policy. The National Acid Precipitation Assessment Program not only links the efforts of the federal agencies, but also coordinates them with the research and monitoring activities of the private sector, and state and local governments, and fosters cooperation with the international community. NAPAP research and evaluation activities includes work in four principal areas: Emissions, Controls, and Control Costs . Atmospheric Processes and Deposition . Ecological Effects and Valuation . Societal Effects and Valuation 2.1.

.

A coordinated and integrated program helps ensure the completeness, practicality, integrity, and credibility of the information developed for this broad, cross-cutting national issue. To achieve its coordination functions, NAPAP, through the Office of the Director develops, updates, and implements a Plan and yearly budget for NAPAP research , monitoring, and assessment activities; identifies gaps and redundancies in agency activities; recommends additional research, monitoring, and assessment activities to provide essential information to reduce important uncertainties; organizes peer reviews of NAPAP activities; . facilitates interaction within the federal program and between the federal program and state and local governments, academia, the private sector, and international organizations; and . integrates research, monitoring, and assessment findings in periodic analyses and summary documents.

. . . .

367

Promote Science/Policy Dialogue NAPAP has unique capabilities to advance the understanding of larger national questions at the interface between science and policy, beyond the endeavors of an individual agency. To fulfill its evaluation function NAPAP will engage decision-makers in a continuous dialogue with the researchers and analysts who are assessing the acidic deposition issue. It is often said that a major impediment to effective government is the lack of research knowledge underlying the development, implementation, and assessment of federal policies and programs. The more a decision-maker is involved in the research and evaluation, the more likely it is that the Pork will help support policy questions and the findings used appropriately. However, the interaction must be in both directions to be fully successful -- NAPAP will provide guidance to scientists so that their research and assessments are designed to produce information that will be most useful to policy-makers. 2.2.

Perform National Assessments The assessment process translates the facts developed through the scientific process for decision-making purposes. In other words, the societal implications of technical facts are evaluated. Without an assessment function, NAPAP's scientific findings will have minimal impact on decisions involving national interests. Therefore NAPAP's assessment objectives link components of the acidic deposition phenomenon in an integrated analysis and comparison of policy alternatives and their effects. The NAPAP 1990 IntegratedAssessmentutilized scientific,technological, and economic information to make projections of future copditions and potential benefits under different emission reduction scenarios. Evaluation of these projections helped inform the decision-making process leading to the current acid rain control legislation and will help answer questions about the implementation of the law. The framework used for that assessment will be updated to answer questions pertaining to the benefits, costs, and effectiveness of the legislation. 2.3.

3.

SELECTED PROGRAM FINDINGS

The following are selected key findings of the Program in several principal categories. For a comprehensive discussion of the issues, ref r to NAPAP's State of SciencelTechnology Reports and Assessment Documents.3.f Emissions National anthropogenic emissions of sulfur dioxide (SO2), nitrogen dioxide (NO ) , and volatile organic compounds (VOC) have declined since reaching peai levels in the 1970s and have changed little since 1985 (Figure 3.1.

1).

Under 1985 emission regulations there would be significant uncertainty about whether emission levels would increase or decrease in the future (Figure 2). As a result of this uncertainty a principal benefit of a defined emission reduction target and cap is the assurance that lower emission levels will be reached within a specified time. Sulfur dioxide is the major pollutant involved in acidic deposition in the eastern United States. Nitrogen oxides and VOCs play important roles in acidic deposition and together contribute to the formation of ozone and other oxidants that are associated with significant effects.

368 SO, Emissions (million short tons)

Emissions (million metric tons)

1

30

+so2

+voc

-+NO.

15

1975

1978

1981

1984

-

1987

Figure 1. Recent Trends in Annual Anthropogen ic Emissions of SO , NO and VOC in the United Stat&. (Reference 3, Figure 6.2).

All Sactan I

0

1880

Moo

I 2010

I

2020

2030

Year

Figure 2. Potential Range in U.S. SO2 Emissions under 1985 Regulations. Upper and Lower Bounds Result From Combining Alternative Assumptions on Power Plant Lifetimes and Rates of Market Penetration of Clean Coal Technology (CCT) at Existing Plants. (Reference 3, Figure 6.3). Control Measures and Costs National SO emission reductions in the range of 8 to 12 million tons per year below 1960 levels can be met using existing technology and control measures. However , advanced clean coal technologies which have been demonstrated at the commercial scale or are nearing commercial demonstration have a number of advantages compared to existing control technologies including reduction in sol id waste amounts and/or treatabil ity requirements, significantly improved removal of NO, emissions, reduced emissions of the greenhouse gas CO,, and potentially lower direct and indirect operating costs. The direct cost of emission reductions varies with the amount and rate of reduction that is achieved -- the more and/or faster the emission reduction, the higher the total cost and the marginal cost per ton of SO, removed. On a national basis, total costs are projected to increase rapidly for SO, emission reduction targets of 8 million tons or more below 1980 levels (Table 1). Emission allowance trading, if broadly adopted, may result in significant long term savings in control costs potentially in the range of 15% to 25% (Table 1). The geographical impact of emissions trading is difficult to predict.

3.2

--

369

Table 1 Range in Annual Direct Costs for SO, Reduction by 2000 in 1990 Dollars

Reduction (mi 1 1 ion tons)

SO2

3.3

Annual Cost (bi 1 1 ion $/yr)

Emission Trading

8

1.7

10

3.0

-

10

2.5

- 3.5

Yes

12

5.0

-

no

2.9

no

3.8

no

7.8

Benefits Associated with Emission Reductions

Benefits in all effects categories were demonstrated to result from SO emission reductions in,the range of 25% to 50% as illustrated under the NAPA6 assessment scenarios. Benefits include the following: Recovery and or/remediation of some acidic Adirondack lakes; prevention of further net acidification of sensitive Adirondack lakes and Mid-Atlantic Highland and southeastern streams (Figure 3); reduced stress on high elevation red spruce trees; reduced nutrient leaching in sensitive soils; reduced deposition impacts on exposed materials; less risk of health effects associated with acidic aerosols; and improved visibility in the eastern U.S. Quantitative information on physical measures of benefits associated with emissions reductions is reported in the NAPAP Assessment; however, a comprehensive economic evaluation of the benefits remains a future goal. In most cases, the differences in benefits resulting from an 8, 10, or 12 mi 1 1 ion ton SO, reduction cannot be quantified because the variabi 1 ity resulting from other factors cannot be distinguished from that resulting from the emissions reductions, and because of uncertainty in the response models used to make projections (e.g. see Figure 3 for comparison of aquatic benefits). However, the differences may be observable over the longer term (decades), particularly for cumulative effects, and possible additive effects of acidic deposition and other stresses. The timing of changes in average deposition and air quality will generally coincide with the timing of changes in emissions. For certain effects categories (visibility, acute human health effects, and episodic surface water acidification), deposition and air quality changes will result in immediate changes in effects. For effects categories where damage accumulates over time or interacts with other stresses (e.g., soil and watershed chemistry, cultural and construction resources, long-term health impacts, high elevation winter injury to red spruce), total damage will be reduced with earlier deposition reductions although the benefits of such reductions may not be observed immediately. Half or more of the sulfur deposition resulting from major SO emission sources typically occurs within 500 kilometers from such sources. fherefore,

3 70

the geographical distribution of sources controlled can have a significant influence on the patterns of benefits projected from such controls. However, for reductions above 1 2 million tons, geographical patterns are relatively unimportant because virtually all major sources would need to be controlled to very low emission levels. Likewise, emissions allowance trading is impractical for SO, emissions reductions greater than 10 million tons because achievement of such reductions would require a high level of control for essentially all utilities. (a) Lake5 wilh ANC

< 0 peq/L (percent)

30

2 0

Figure

3.

ANC, pH, and Unsuitable Water Chemistry for Sensitive Species in Adirondack Lakes. Data are based on Projections from the MAGIC model using different emission reduction scenarios. (Reference 3, Figure 5.4-1). s1

No additional sulfur controls beyond those already leg i s 1 ated

53

12 million ton SO, reduction

s4

10 million ton SO, reduction

s5

8 million ton SO, reduction ;

0

A

0

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10

01 1980

1

I

I

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(b) Lakes with pH < 6 (percent) 40,

J

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1

lo0 1980

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(c) Lakes Unsuitable b r Sensitive Fish Species (percent) 30,

2 10 0 1980

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Year Nois Porcenfage01 Adlmndack lake8 wilh (a) ANC < 0 peqqlL. (b) pH < 6. and (c) chsmimry unrulable forrmnrlllve llrh a wclerbased on MAGIC pmlenmr lor 50 years under 11lu81r811wdepositionscenarios

4.0

M I S S I O N AND GOALS FOR NAPAP-11’

Research and Monitoring Strategy Basic science is the foundation for progress in understanding acidic deposition causes and effects. NAPAP however, as mandated by Title IX of the 1990 Clean Air Act Amendments, i s an amlied program focused on improving the basis for decision-making. Therefore, its research strategy will be molded to contribute needed information in the sequence necessary to meet assessment needs, but will be flexible enough to adjust directions, if necessary, on the basis of new information. There are two fundamental force: affecting policy-relevant research, and they often work against each other. First, scientists desire independence and continuity of funding to undertake sustained and challenging investigations that will advance their careers. Second, federal agencies are required to develop data consistent with their missions and policy objectives

4.1

371

defined by the President and Congress and reflected in an annual budget. The NAPAP Office of the Director will strive to achieve a balance between these two forces, acting as an organizational buffer by interpreting information and channeling its flow from the scientist to the policy-maker and from the policy-maker to the scientist. Competing demands for the taxpayer's dollar continue to increase at a rate greater than the amount of available funds. Growth in the number of researchers, tight constraints on federal spending, and an ever-increasing number of problems that would benefit from technologicalssolutions a1 1 create intense competition for federal research funding. NAPAP has a responsibility to assure that research and assessment activities will broadly meet the needs of the user community and that program expenditures will advance national concerns. Therefore, uniform criteria based on a value-ofinformation approach will be established and utilized to set programmatic priorities. For example, program objectives that have a high likelihood of informing energy/environmental policy will be emphasized. Integration and Assessment NAPAP was established to perform both scientific research and policyoriented assessment. Addressing the acidic deposition phenomenon requires a long-term, mu1 t i-agency, mu 1 t i-disciplinary effort using diverse faci 1 it ies and personnel dispersed around the country and even abroad. However, merely gathering and reporting scientific knowledge does not guarantee its use in serving the public welfare, The need for knowledge is important, but there is an even greater need for understandinq. Development of valid and coherent concepts of the energy/environment system requires an interrelating and a synthesizing of knowledge from many different sources. Therefore, the assessment function will be central to the Program (Figure 4 ) . NAPAP's assessment activities will evolve through several phases, each with the goal to deepen our understanding of the energylenvironment system and its interaction with society.

4.2

u SCIENCE

Figure 4. The Intersecting Triangles Represent the Central Role of Assessment for NAPAP-11, Linking Science and Policy.

NAPAP's original charge to evaluate control options for acidic deposition and its current mandate to assess the effectiveness of the new control policy requires an abundance of statistical data fed into mathematical models that simulate economic, physical, chemical and biological processes. The integrity of the evaluation demands that the models and the data supporting them are reliable. However, many models and statistical data supporting federal policy decisions are not validated and/or updated, nor are model predictions that were used to support or reject legislation compared with the actual effect pf the policy once it becomes law.

372

NAPAP will address these issues by making a significant commitment to testing, evaluating, and updating its statistical data, models, and assessment framework. A strong focus on long-term monitoring will provide the necessary "ground truth'' needed to verify the models. The operations associated with an integrated assessment for a phenomenon as complex as acidic deposition are very complicated. The assessment framework will thus be exercised to identify flaws, inconsistencies, missing information, or faulty constructs or assumptions, thereby helping to identify important research needs. In this way, NAPAP's assessment activities will evolve through several distinct but overlapping phases. Judgements, sometimes controversial, are involved in every step of the assessment process. NAPAP wi 1 1 identify key judgements and assumptions so that others can understand the basis for its conclusions. In order to build trust and enhance credibility, NAPAP will continue to cultivate an atmosphere of proactive openness in disclosing its processes and results. In the early stages of NAPAP's second phase, the Program will produce updated scientific and technical information resulting from further research and monitoring studies as well as reports on "special projects" or "miniassessments" dealing with evaluations of issues that will have a special interest to the public and policy-makers. As the acid rain controls become fully implemented, NAPAP's analyses will become more comprehensive and include long-term responses. Eventually cost-benefit analyses, comparative evaluations, and risk assessments will be possible. 4.3

Economic and Valuation Analyses It is estimated that the U.S. currently spends approximately $30 billion

a year in efforts to fulfill its clean air policies. The 1990 Clean Air Act Amendments will almost double that expenditure. Obviously with expenditures of this magnitude there is considerable public demand that substantial benefits are associated with such costs. A primary objective for NAPAP is to evaluate whether the acid rain control policy produces the maximum effectiveness for the costs incurred. Benefit-cost analysis is used by economists to identify, quantify, and weigh the advantages and disadvantages of pub1 ic policies designed to increase society's overall well-being. It has become an integral part of policy analysis in all levels and types of government and is specifically mandated of NAPAP in the Program's re-authorizing legislation. The quantification of costs and benefits rests on the premise that an action has value if someone is willing to pay for it; however, such values are often difficult to ascertain. The use of natural resources (e.g. our air resource) for economic development entails environmental costs that are usually not considered in determining price. Current cost-benefit analysis techniques are not considered adequate for assessing the value of many important factors, particularly those not bought and sold in the marketplace (e.g. ecosystem diversity). A major new emphasis for NAPAP is a commitment to developing methodologies a1 lowing more equitable comparisons of the diverse costs and benefits associated with the acidic deposition controls required by the 1990 Clean Air Act Amendments. This may involve new approaches to valuing ecological and non-ecological effects categories. It must also be recognized that costs are usually incurred early in a control program, whereas benefits are only fully realized over the longer term because of the cumulative nature of some impacts.

373

Conmunication and Education Strategy NAPAP will take several steps to improve information transfer, including a vital communication/education strategy, and a commitment to analyzing the reasons for conflicting scientific results. Because NAPAP has the ability to convene a wide range of thoughtful technical experts and policy specialists, it has a unique opportunity to develop knowledge from many disciplines into informed public policy. But to do that, its results must be accurate, credible, timely, and noticed. To achieve these aims the Program will continue a commitment to external review and proactive openness in disclosing its processes and results and will undertake substantial efforts to communicate them. A variety of methods, such as discussion papers, brochures, and videos, may be employed to convey Program results on special single-issue topics as well as on comprehensive multidisciplinary analyses. The information will be molded in a format to suit different audiences in order to expand the range and success of the communication/education effort. NAPAP's approach to dealing with conflicting results or analyses will be to describe differences in findings, not reduce them to the lowest common denominator. Most important, however, is NAPAP's intent to analyze the reasons for the differences. Understanding the reasons for conf 1 icting outcomes allows for a much deeper understanding of the issue and of society itself. In this way the work of the Program, through use of the acid rain issue as a case study, can lead to change and progress in matters involving energy and environmental pol icy. Americans' concern about the environment has increased substantially in recent years. The new national resolve to protect and improve environmental quality must be accompanied by a solid knowledge base to make effective progress. Scientists and environmental experts can provide the facts about the causes and possible solutions to environmental problems, but in a democratic society, major policy initiatives must be supported by the people. Strategies to improve environmental qua1 ity are becoming increasingly more sophisticated technologically, and consequently more expensive, as well as more dependant on lifestyle issues. Citizens will want to make informed choices about matters affecting their pocketbooks and individual behavior; therefore, a solid understanding of the issues is important. The acid rain issue can serve as an example for broader public discussion and education about the inter-relationships between energy, environment, and society. An ultimate goal of NAPAP's communication and education strategy is to expand public understanding of the complexities and interconnections within and between human society and the environment using the acidic deposition phenomenon as an example. 4.4

374 5.

REFERENCES

1.

Schaefer, M. "The Federal Research Puzzle", Environment, 33 (November 1991): 17-20, 38-42. Chelimsky, E. "On the Social Science Contribution to Governmental Decision-Making", Science, 254 (11 October 1991): 226-231. The U.S. National Acid Precipitation Assessment Program. "1990 Integrated Assessment Report", NAPAP O f f i c e o f the Director, Washington, O.C. (November 1991) 520pp. Irving, P.M. (ed) Acidic Deposition: State o f Science and Technology, National Acid Precipitation Assessment Program, 722 Jackson P1 , N W , Washington, DC. (September 1991) 265pp. The U.S. National Acid Precipitation Assessment Program, Mission, Goals, and Program Plan, Post 1990, Public Review Draft, (November 1991) 59pp. Chubin, D.E. and E.M. Robinson. "Opinion -- Sound Science Policy Requires Better Data Management", The S c i e n t i s t , (16 September 1991):

2. 3.

4. 5.

6. 7.

_11-13. _

Hamilton, D.P. "Policy-Making: Getting Better Data", Science, 253 (23 August 1991): 847.

T. Schneider (Editor). Acidification Research. Evaluationand Policy Applications 1992 Elsevier Science Publishers B.V.

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The United Kingdom Research Programme and Its Implications for Policy, Now and in the Future R B Wilson Department of the Environment, Air Quality Division, Romney House, 43 Marsham Street, London SWlP 3PY, United Kingdom

Introduction Over the last fifteen years there has been a change in perception of air pollution and the problems it can cause in the United Kingdom. To most citizens and politicians, the Clean Air Act 1956 marked the culmination in tackling air pollution problems. By 1974 this Act had been operational twenty years and the improvement in air quality in urban areas was dramatic, with large reductions in smoke and sulphur dioxide [FIG 11. As a consequence, research on air quality in the United Kingdom was geared to monitoring this decline in urban pollution so that in 1978 for example, the expenditure of E0.8 million was spent almost totally on monitoring air quality a t 1100 urban sites.

Long Range Transport of Air Pollution However in the early 1970s, several United Kingdom scientists were concerned by the claims of some Scandinavian scientists that long range transport 'of air pollution could lead to acidification of sensitive receptors in areas far distant from major sources of emissions. They were also convinced that emissions from the United Kingdom could be impacting on other countries and especially Southern Scandinavia. In the mid 1970s, the UK joined twelve other North West European countries in a study of the long range transport of sulphur under the auspices of the OECD. This study was the precursor of the United Nations Economic Commission for Europe EMEP programme and produced the first evidence that emissions from one country could be deposited on a number of other countries. The OECD study produced the first blame matrix i.e. how much each country deposited and received from every other country (FIG 2), and this approach has since been extended and refined by the UN ECE EMEP. Once the principle of the long range transport of air pollution had been established by the OECD study, it gave the impetus in many countries, including the UK, to evaluate the impact of those depositions and this resulted in increased research funding (FIG 3).

376

Research Policy Although the UK, like many other countries, started to h n d research on an ad hoc single project basis, it soon became obvious that the nature of the acidification problem called for a multi-disciplinary integrated project approach. This realisation came about a t the same time that central government was redefining its approach to research funding along the so called "Rothschild" or "customer-contractor principle" which instigated a move towards short term research contracts (no greater than three years and continuous assessment. On this basis, it was decided to re-organise the UK acidification research programme on an integrated and wherever possible joint funded basis so that the more basic process-mechanisms research funded by Research Councils and Universities could run in tandem with the more applied and goal-orientated research required by central government. A number of large integrated projects were formulated. For example the Llyn Brianne Multi-Catchment Acidification Study involved nineteen different research organisations or universities and was jointly funded by the National Environmental Research Council, the Welsh Water Authority, the Welsh Office and the Department of the Environment. Other examples are given a t Figure 4. The growth in research funding from the early 1980s brought with it a n information explosion and the need to assimilate, analyse and dispense this information to policy makers, the media and the public. One vehicle for undertaking this was the setting up of a series of independent peer review groups t o undertake critical assessments of the science in a number of areas related to the acidification pathway (FIG 5). Over the years, these have been published with increasing frequency and have achieved a high scientific and policy standing, both nationally and internationally.

Critical Loads By the late 1980s, evidence was emerging of impacts of acidic deposition on soils and freshwaters in a number of geologically sensitive areas of the UK. The need to develop a national assessment on the scale and intensity of acidification coincided with the introduction of the concept of critical loaddlevels internationally. The integrated design of the UK Air Quality Research Programme allowed it to respond quickly to this new concept and the UK is increasingly using national critical loaddlevels assessments to identify areas of maximum acidification sensitivity and so target h t u r o research programmes. Although initially designed as a tool to assist policy makers t o optimise emission control strategies, the UK feels the critical loads concept can also be used as a powerful tool to guide research policy.

Modelling However the impression must not be given that in research management terms we got everything right first time. One lesson learned the hard way was

377

the need to keep the modelling of atmospheric deposition and impacts in tandem with the monitoring and research programmes. In the UK as in many other countries, acidification modelling came late on the scene. This meant that some existing databases were not immediately suitable for model input and that research resources could have been better allocated had the requirements for model development and validation been known earlier. The lesson learned by the UK was that there is, or should be, a symbiosis between modelling deposition measurement, air quality monitoring and dose response databases and that work in all these areas should proceed in parallel. The importance of long range transport modelling was recognised a t an early stage in the OECD and UN ECE EMEP programmes and the routine sulphur model has played a vital role in quantifying the transboundary transport of acidic sulphur compounds. However, i t has been more recently discovered that other pollutants can contribute to acid rain damage, including NO, and ammonia. Much less is known about the long range transport of these species and their derivatives than of sulphur. Emissions of NO, and NH, are less adequately quantified, and far less harmonised monitoring data are available either for them or for the secondary species formed from them. 'State of the art' models can only be poorly validated for NO, and NH, species a t present and they need more development before they can be relied upon in policy terms. We have also learnt that there are strong links between the atmospheric chemistry involved in tropospheric ozone and hydrogen peroxide formation, sulphur dioxide oxidation in acid rain formation and the oxidation of NO, and hydrocarbons to produce ground level ozone. The modelling tools are not yet ready to link comprehensively all phenomena together and to provide advice in a n integrated and consistent manner. Such models will be needed in the future if we are ever to be able to assess the impact of emission reductions of one species on the formation or reduction of other (secondary) species.

Priorities for Future Research On the positive side, the integrated research programme approach has, through its multi disciplinary approach, allowed the identification of a number of important factors which need to be taken into account if accurate estimates of deposition and/or impacts are to be established. These include such important factors as:

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altitude which leads to increased acidic deposition and extended ozone exposure,

-

occult deposition whereby fogs, mists and low cloud levels lead to enhanced acid deposition,

-

co-deposition where the presence of one pollutant increases the deposition velocity of another e.g. NH, and SO,,

the need to understand the relative contributions of oxidising (e.g. NO,) and reducing species (e.g NH,) to total nitrogen deposition, a need to understand the behaviour of mature trees to air pollutants in field conditions as compared to chamber experiments with saplings or cuttings, the central role of soils in explaining both terrestrial and freshwater acidification effects, the increasing importance of ozone in W Europe, the need to understand the interaction between air pollution and other biotic and abiotic stresses. Finally, the review group assessments and the UK Critical loaddlevels programmes has enabled a broad assessment of sensitivity to acidification to be made in the UK. For the UK (and not necessarily for other countries) the range of sensitivity appears to be soils, freshwaters >> trees >> crops. It is difficult to place materials and buildings including cultural heritage in this hierarchy but all the indications of the sensitivity and materials to air pollution damage would place this category alongside freshwaters and freshwater biota.

CONCLUSIONS

1. Complex problems such as acidification need a n integrated multi-disciplinary research approach. 2. Modelling should commence and run in tandem with research and monitoring a t the start of each programme.

3. The critical loads concept identifies sensitive receptors, allows the definition of sensitivities and research priorities and permits a range of emission control scenarios to be defined to achieve the optimal use of resources. The use of the integrated monitoring, research and modelling approach developed for acidification can be used for evaluating other national, regional or global issues such as the impacts of climate change, land management and sustainable development. It is already being used in the UK for the evaluation of climate change issues. 4.

319

Figure 1. UK emissions of black smoke

Figure 2. Norway and Sweden, S deposition blame matrix

Figure 3. Air Pollution research funding

EXAMPLES OF UNITED KINGDOM INTEGRATED PROGRAMMES AND PARTICIPANTS

Figure 4

PROJECT LLYN BRIANNE ACIDIFICATION

PARTICIPANTS

MULTI-CATCHMENT

19 RESEARCH ORGANISATIONS

IMPACTS OF AIR POLLUTION, INCLUDING ACIDIFICATION ON SOILS AND TERRESTRIAL ECOSYSTEMS

ITE PENICUIK, ITE BANCHORY, ITE MERLEWOOD, UNIV LANCASTER, UNIV NOTTINGHAM, UNIV ABERDEEN, McCAULAY LAND UK RESEARCH INSTITUTE, IMPERIAL COLLEGE LONDON, UNIV CAMBRIDGE

IMPACTS OF NITROGEN DEPOSITION ON NATURAL VEGETATION

UNIV MANCHESTER, UNIV NEWCASTLE, ITE BANGOR, UNIVERSITY COLLEGE LONDON.

NATIONAL PROGRAMME

BUILDING RESEARCH ESTABLISHMENT, BRITISH COAL, NATIONAL POWER, PUBLIC SERVICES AGENCY, UNIV MANCHESTER AND TECHNOLOGY, AVON SCIENCE SCIENTIFIC SERVICES.

MATERIALS

EXPOSURE

W

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Figure 5

REVIEW GROUP REPORTS

UK REVIEW GROUP ON ACID RAIN Acid Deposition in the UK (pub 1983) Acid Deposition in the UK 1981-1985 (pub 1987) Acid Deposition in the UK 1986-1988 (pub 1990)

UK ACID WATERS REVIEW GROUP: Acidity in UK Fresh Waters (pub 1986) Acidity in UK Fresh Waters (pub 1988) PHOTOCHEMICAL OXIDANTS REVIEW GROUP: Ozone in the UK (pub 1987) Oxides of Nitrogen in the UK (pub 1990) BUILDING EFFECTS REVIEW GROUP: The Effects of Acid Deposition on Buildings and Building Materials in the UK (pub 1989) TERRESTRIAL EFFECTS REVIEW GROUP The Effects of Acid Deposition on the Terrestrial Environment in the UK (pub 1988) (Second report to be published soon)

STRATOSPHERIC OZONE REVIEW GROUP: First report 1987 Second report 1988 Stratospheric Ozone 1990 Stratospheric Ozone 1991 UK CLIMATE CHANGE IMPACTS REVIEW GROUP: The Potential Effects of Climate Change in the UK. 1991 NITROGEN DEPOSITION REVIEW GROUP Impacts of Nitrogen Deposition In the Terrestrial Environment. (To be published in 1993)

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T. Schneider (Editor). Acidification Research. Evaluationand Policy Applications

0 1992 Elsevier Science Publishers B.V. All rights reserved

383

Reeecll.ch into forest decline and air pollution in France.W o r finrlinen and relevance for policy applications G. Landmann DBpartement de la Sante des ForGts. Ministhre de 1'Agricultureet de la For& INRA - Centre de Recherche8 Forestihres, F-54280 Champenoux, France

Abstract "he DEFORPA Programme (Forest Decline and Air Pollution) (1984-1991) aimed at identifying the causes of decline in conifers in the French mountains. The yellowing Norway spruce (Picea abies KJ,which reveals a Mg deficiency, results primarily from the long-term acidification and cation depletion of originally poor soils under the influence of acid deposition and harvest. Other air pollution effects include stream acidification in the Vosges area and eutrophication of forest ground vegetation. Ozone effects remain uncertain. Climatic anomalies affecting vulnerable stands in relation with their history were found to be the main causes of silver fir (Abies alba Mill.) dieback. Despite these air pollution effects and visible damage, the productivity of the studied species has increased over the past century; the causes (climate change, increased N deposition and C02 levels, silviculture) are not yet identified. Critical loads are thought to be a valuable tool for defining further emissions abatements to protect the sensitive parts of the forest ecosystem (flora, soils, surface waters).

1. INTRODUCTION HISTORICAL BACKGROUND AND RESEARCH

STRATEGY

Concern about a possible large scale forest decline in France dates back to the summer of 1983, when a group of foresters and forest ecologists identified severely defoliated coniferous stands in the Vosges mountains ( IWFrance). In fact, the year 1983 probably coincides more with a breakthrough in the awareness of a majority of foresters and scientists than with the real beginning of the crown thinning process in this area. The impassioned debate in the neighbouring Federal Republic of Germany played a major role in this context (1,2). A national survey launched the same year within the French Forest services received little feedback at that time, forest decline was not perceived as a nationwide phenomenon. Beside the Vosges mountains, only a few areas, such as the Jura mountains or the central Pyrenees, were considered to be affected.

384

The Ministry of the Environment took the initiative to launch a coordinated research programme entitled DEFORPA (Forest decline and air pollution), shortened to DEFORPA hereafter. This Programme (1984-1991)was sponsored by 3 French ministries (Environment, Agriculture and Forestry, Research and Technology), several state agencies (Air Quality Agency, Electricit4 de France), regional authorities, a private company (Elf Aquitaine) and by the Commission of the European Communities (DG XI1 and DG VI). The administrative structure was composed of 3 bodies: the steering committee with the representatives of the financing agencies, the scientific committee in charge of the evaluation of the projects and the operational group consisting of scientists responsible for the coordination and implementation of the research activities. More than 40 laboratories belonging to 20 scientific bodies, universities, as well as professionals, participated in the programme (3,4). The emphasis was primarily laid on the understanding of the forest decline phenomenon in a few typical situations rather than on the overall assessment of the air pollution effects at national level. The scientific philosophy and strategy of DEFORPA were the following ones: all the potential causes of forest decline were to be "suspected equally until sufficient scientific evidence allowed the discarding of one or another factor; the various potential causes were first evaluated separately then, if no single factor can account for the phenomenon, in combinations of increasing complexity. By comparison, a typical air pollution assessment programme would focus primarily on the effects of the various pollutants, considered separately or in combination, the other stressors being generally considered in a second phase, insofar as air pollution does not offer a fully satisfactorily explanation for the phenomenon. In agreement with the aforementioned philosophy, DEFORPA tried to offer a platform to the various dominant theses (air pollution, climate, silviculture...I. It was hence possible t o avoid the many misunderstandings resulting from situations where alternative hypotheses were evaluated by individual scientists outside the "official" programme. After a brief overview of the approaches developed within DEFORPA, this paper will highlight some relevant results in the context of the international literature. A more comprehensive view of the accomplishments can be found in a recent report (3) and in the literature quoted throughout this paper. The discussion will focus on the likely reasons for the lack of consensus on the causes of forest decline and on the relevance of current scientific understanding of air pollution effects for further policy applications. 2. "HE APPROACHES

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2.1. Pollution climate Inputa in forest ecosystems

When the Programme was launched, the knowledge of the pollution climate in French rural areas was limited to the data of wet deposition in 5 stations of the EMEP network (European Monitoring and Evaluation Programme) and in 5 stations of the BAPMON network (Background Air Pollution Monitoring Network) (4). There were only few data on deposition in forest stands, and none at all on gaseous pollutants in remote forested areas.

385

Since 1985, significant effortshave been engaged in this field. At the national level (not strictly part of DEFORPA), the MERA network (r6seau de MEsure des Retombees Acides) is composed of 13 rural sites (5 belonging to the BAPMON and EMEP networks, 8 newly created ones), where wet deposition and (partly) gaseous pollutants are monitored (4). A nation-wide network of 25 sites in forest areas, where bulk (partly wet) deposition and throughfall will be monitored, is currently being installed as a part of the national "Ecosystem network" of 100 permanent plots. Within the framework of DEFORPA, extensive data were collected for the Vosges area: atmospheric deposition in forest stands measured since 1989 in a network of 8 sites (51,specific studies of aerosols (61, snow (6,7), rain and fog events (6, 81, vertical profiles of gaseous pollutants (including volatile organic compounds) carried out in a forest stand at the "Donon Tower" located 750m a.s.1. (6, 91, P.A.N. (10) and ozone (111, and, last but not least, the W A C ("Nuages Acides") campaign, combining airborne measurements a t small- and meso-scale in clouds with measurements on the ground, in and under the clouds, involving 17 research teams from 4 countries which took place in the Vosges mountains between April and July 1991. 29. Air pollution effects

Some work has been carried out on possible effects of SO2 under controlled conditions (12, 13) and in situ on healthy versus damaged trees (14). However, most physiological work has been concentrated on long-term open-top chamber experiments, especially the Montardon (southern France) experiment, where Norway spruce seedlings were exposed for 5 years to concentrations of 0 3 and SO2 as measured at the Donon site in the Vosges mountains (14). The impact of acidic deposition on soils was investigated at catchment level in the Aubure basin (Vosges) (15, 16) and in the comparative catchments (spruce, beech and grassland) at Mont-Lozbre (southern Massif Central) (15, 17), and at stand level by nutrient budgets in the French Ardennes (181, and at 5 sites (11stands of various species and ages) in the Vosges mountains (18, 19, 20,21). Special attention was recently paid to the forms of aluminium in soil solutions, both in seepage and capillary water (15). The effects of increased nitrogen deposition on conifers were evaluated in fertilization experiments in the Vosges (22,231 and in the Ardennes (22,241. 25. combinedaffeots

In this paper, combined effects refer to the effects of the combination of pollutants and natural or other man-made factors, as evaluated by controlled or semi-controlled experiments, including manipulative studies in situ. Potential interactions between drought and gaseous pollutants were investigated in closed chambers (SO2 x drought) (12, 13) and in the Montardon opentop chamber experiment ( 0 3 x drought) (14). The relative roles of acidic deposition and silviculture in the soil acidification and desaturation were evaluated by detailed nutrient budgets in spruce stands of various ages (20) and by biomass/mineralomass studies (25, 26). The possible influences of the soil microflora and of drought on magnesium deficiency (which characterizes the so-called yellowing of spruce on acid soil) were studied respectively in pot experiments (27) and in situ in a manipulative study in a spruce stand (20).

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Other combined effects include the interaction between increased nitrogen input with drought stress investigated in an old fertilization trial in a mature silver fir stand (23). a.4. ~ n m t eeffecte d

Integrated effects refer here to effects evaluated by integrative measurements on large samples, representative of regional or national levels. Special attention was paid to forest growth, essentially via the dendroecological upprouch (28, 29,30,31).The visual symptoms (defoliation and discolouration) are recorded in systematic networks: the "Blue Network', a 16 km x 1 km grid with 24 treedplot covering mostly the mountainous areas (ca 20% of the forested area), and since 1989 the "European network', a 16 km x 16 km grid with 20 treedplot extended all over the French forest (14.4Mi0 ha) (32). 2.5. other air pollution effecta

Although forest decline per se was at the beginning of DEFORPA at the heart of the research, the effects of acid deposition on the other compartments of the ecosystem were investigated, especially the soil (as already mentioned) and more recently surface waters in the Vosges area (33)and ground vegetation in several forests in the NE of France (34,351, 3. MAJORFINDING8

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3.1. Pollution climate Inputa in fomt ecueystems

The main features of the pollution climate in forest areas may be summarized as follows: ozone is in the Vosges area the only gaseous pollutant which is permanently present at relatively high levels during the vegetation season, with mean concentrations during the summer period reaching 100 pglm3 a t high altitude and peak values above 200 pg/m3(6); the annual profile in the Vosges is characterized by a a sharp increase of the spring monthly means followed by a plateau during the summer or even a second peak, which was ascribed to "local production" (111,contrary to less polluted sites along the Atlantic ocean, where a sharp decrease can be observed over the summer period (36); in the Vosges mountains, sulphur dioxide reaches high concentrations (200pg/m3and more) in "normal" climatic years (e.g 1986 or 1987)during a few periods of several days in the winter period, when emissions, essentially originating from central Europe, are high and the dilution in the atmosphere is decreased because of stagnant weather. Since 1987, no high concentrations have occurred a t all, mainly because the recent winters were very mild (no eastern winds) and possibly partly because of the reduction in emissions in Western Europe. The annual mean concentrations decreased from 15 pg/m3 (1986-87)to 5 pg/m3 (1988-90)(6). Since SO2 has been measured, the concentrations during the vegetation season are with a few exceptions below the detection limit (45 pglm3); in the Vosges area, nitrogen dioxide shows similar temporal patterns to sulphur dioxide (6).Mean annual values are also very low (c 10 pg/m3) but some (moderate) daily peaks (up to 30 pg/m3) occur occasionnally during the vegetation season. Nitrogen monoxide never reaches high concentrations (6);

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-

387

- the pH values in cloud droplets may reach very low values in the Vosges mountains, down to 2.9 in winter when the cloud base is lowered in situations of temperature inversion, but during the vegetation season the average pH is not very acidic (4.5and above) (8); - sulphur deposition shows a distinct spatial pattern, with "high' rates (3040 kg/ha/yr under a spruce canopy in the most recent years in the French Ardennes, 70-90 kg/ha/yr in the early 1980's) (371, "medium" rates (15-30 kg/ha/yr) at high elevation in the Vosges ( 5 , 6 ) and Mont-LozBre (38) and "low" rates (10 kg/ha/yr and below) in protected or low elevation sites (5). In northern France, a marked decrease has been observed over the past years (37); - nitrogen deposition in forest stands varies greatly within France, with (known) mean deposition rates under conifer canopy (considered as the upper limit of the "real" deposition) ranging from 50 kg/ha/yr (213 as N-NH4) in the Ardennes, to 15 kg/ha/yr (U3 as N-NH4) as a typical rate in the high elevation Vosgian sites (1987-19901,and to a few kg/ha/yr in low elevations sites in the Vosges mountains and in southern France (5,15, 37); - the total proton input (wet, occult and dry deposition) in coniferous forest stands, as evaluated by throughfall measurements (with "realistic" hypotheses about cation leaching rates) is generally moderate (1-2 keq H+/ha/yr) in the Vosges and in the Ardennes (5,18). In southern France (Mont-Lozbre), the alkaline minerals in the eolian erosion dust, mainly from the Sahara, neutralizes most of the rain acidity (38). 33. Air pollution effects Although SO2 is potentially harmful to forest trees according to the abundant international scientific literature and some recent French results (12, 131, current levels in remote sites are unlikely to have a major impact on forest health. The increased leaching effect (probably caused by H2S04 and not by SO2 itself) demonstrated in the Edelmannshof (FRG) open-top chambers for example (39) may, in the most "pessimistic" hypothesis, slightly increase the nutritional disturbances. Ozone was found, especially in the Montardon experiment, to be a potential threat for forest species, even for the Norway spruce, which is considered in the literature to be one of the most resistant species. Although the physiological measurements produced somewhat inconsistent results (141,seedlings in the ozone chambers (0salone and 0 9 + S02,the latter being virtually inexistant) exhibited clearly a premature needle loss from the 3rd yr onwards. Chlorotic mottling was also observed mainly in one clone. However, this "ozone effect" is believed to have been exaggerated: such visible symptoms are neuer observed in situ. It is believed to be caused by the excessive temperatures in the chambers during the summer periods, a nitrogen deficiency (irrelevant for the Vosges sites) and the confinement of roots in the pots. Interestingly too, the shoot biomass was only slightly affected by ozone whereas root biomass and mychorizee were dramatically reduced (40). Despite this ozone effect, the physiological and growth parameters were always influenced more by the presence of the chambers and by the clone than by the pollution treatment. Nutrient and proton budgets at catchment and stand level revealed some important features: - the external sources of protons are dominating in the Aubure forest ecosystem situated on a very acid soil on granite; tab. 1 shows the important contribu-

388

tion of occult and dry deposition. In young spruce stands, however, the contribution of cations uptake by trees can contribute t o about 40% of the proton production (ca. 15% in the neighbouring mature stand) (20)); Table 1 Proton sources in a mature Norway spruce stand a t high elevation in the Vosges mountains (Aubure catchment) (kg-eqhdyr).

ht8d

EMeJmal

precipitation dry and occult deposits nitrogen cycle of external origin

0.46

Total

fLsg

1.30

immobilization in wood nitrogen cycle of internal origin

0.36 0.07

0.63

Total

0.43

Adapted from: M. Bonneau, E. Dambrine, C. Nys and J. Ranger, Sci. Sol 29, (1991)125.

- despite the "moderate" acidic load in the Vosges area, the soil solutions are dominated by the strong acid anions (SO4, NOa) which mobilize large amounts of aluminium and significant amounts of base cations; the soils (down to 60cm) in Aubure are unable to buffer the acidity: protons are neutralized and aluminium is precipitated into Al(OH)3 in the regolith, the stream, although of low alkalinity, remains neutral (tab. 2.1, which is often but not always the case for the catchments situated on deep soils and fractured rocks (see 3.5); Table 2 Mean chemical composition of the water a t several levels in the Aubure catchment (in peqh 1 and pH. Mean 1986-88.(tr. = traces). pH Precipitation Throughfall Drainage water at 60cm Stream water

-

H+ Ca++Mg++ K+ Na+ NH,+AlS+ NO,- C1-

4.49 32 3.77 167 4.03 93 6.0 1

11 63

SO4-

4.5 2.2 11.7 19.0 tr. 24.4 15 39 17.9 50.1 47.2 38.9 4 76.4 63 167

71 34.4 44.6 65.6 9.8 !226 165.0 63 78 186 60.2 22.5 84.7 1.2 tr. 35.7 55 218

Reprinted from: M. Bonneau, in G. Landmann (ed.), French research into forest decline, DEFORPA Programme, 2nd report, ENGREF Nancy, 1991.

- the cation budgets at catchment level show important losses at the Aubure catchment and moderate losses at the Mont-Lozbre (15);base cation losses at the soil level (0-60cm) at Aubure are apparently moderate (tab. 31, but

389

nevertheless lead to a rapid decrease of exchangeable cations (the estimated rates are resp. 0.5% and 2% per year for Ca and Mg) (15, 181, because of the extremely low reserve of easily weatherable primary minerals; Table 3 Mean annual losses (drainage + immobilization in wood - inputs from precipitation and dry deposition) of the main chemical elements in the MontLozBre spruce catchment (1981-85) and in the Aubure catchment (oct. 86-sept. 89) and in the soil (-60 cm) at Aubure (1986-88)( k w y r ) . Ca Mt-LozBre (catchment)lO.O Aubure (catchment) 37.1 Aubure (soil) 3.9

Mg

K

Na

N

C1

S

Si

6.0 7.4 1.9

6,7 7.5 9.8

5.0 10.9 0.3

3.5 -0.7 18.2

1.4 -1.5 1.0

- 3.0

17.5 39.6 1.1

10.4 5.0

Adapted from: M. Bonneau, in G. Landmann (ed.), French research into forest decline, DEFORPA Programme, 2nd report, ENGREF Nancy, 1991; A. Probst, D. Viville, B. Fritz, B. Ambroise and E. Dambrine, Water, Air, and Soil Pollution (1991, in press).

- nitrate leaching may be significant (20-40 kg/ha/yr with important interannual variability) under spruce stands with moderate to high nitrogen input, but seasonal budgets in the Vosges (18, 19) and in the Ardennes (24) show that these losses occur mainly in winter and that uptake by vegetation (trees and grasses) is sufXcient even in severely damaged stands to prevent nitrate leaching during the vegetation season. These results do not confirm the hypothesis formulated by HAUHS and WRIGHT (411, who tentatively attributed the increased nitrate leaching recently observed in some catchments to the progression of forest decline; - as a consequence of these seasonal dynamics, aluminium mobilization is much more pronounced in winter, when the nitrate uptake is low, than in summer (191, except for some late summer events, when rehydration follows a drought. On a yearly and long-term basis, the potential toxicity of aluminium for forest trees in situ is therefore still difficult to assess and can not be directly derived from ex situ experiments. A1 may at least diminish Ca and Mg uptake. Recent fertilization trials have evidenced that an additional input of nitrogen, even a t moderate doses, can enhance existing o r latent nutritional disturbances (mainly Mg and Ca deficiences) on very acidic and desaturated soils (22). This result can of course not be extrapolated to more fertile soils. 3.3. Combined effects:experimentaland pmceai-levelevidence Various factors may modify the effects of air pollutants. The following effects could be identified: - a drought stress applied to a subsample of spruce seedlings from the Montardon OTC experiment has significantly inhibited the growth of the trees of the 0 3 treatment as compared with the control (40). This effect may be due

390

mostly to the aforementioned poor root status rather than to an interaction at leaf level (where drought should rather prevent the 0 3 effect); - the long-term base cation removal by harvest is of the same order of magnitude as the loss through acid deposition, as growth is the major factor of soil impoverishment in young stands and pollution acts mainly in old stands (20); the yellowing, of spruce, which varies noticeably with time, is influenced by several factors: an experimental drought applied to a 30-yr-old spruce stand has proved capable of amplifying spruce yellowing (20);when the corresponding soil has previously been pasteurized, spruce seedlings grow green although they first showed yellowing (20, 27). Neither the drought nor the (unidentified) "deleterious microflora" are interpreted as causes of the yellowing, but their relations to mineral uptake are important to understand the short-term fluctuations of the Mg deficiency; - the impact of nitrogen deposition on forest growth and health seems to vary with climatic conditions: the assessment of an old (1969)fertilization trial in the Vosges mountains showed that the growth of a mature silver fir stand reacted positively to nitrogen during favorable climatic periods but showed a depression during a period with rainfall deficits (23).

-

3.4 Intqgrated&eclzepidemi&gicalevidence

Given the established and potential air pollution effects presented above, one would expect some clear trends at the ecosystem level, such as, for example, a decrease in forest growth, a t least in some regions or for specific site conditions, enhanced drought effects in areas with high nitrogen loads or high ozone levels, etc. The "real world' is obviously more complex and very different from what was expected a few years ago. The dendroecological studies have shown that: - a more or less marked growth increase since the 1850's (29,31)for several tree species Olbies alba, Picea abies, Quercus petraea and Q. robur) (fig. 1). The rate and timing of this increase varies according to the species and the ecological conditions without a clear relationship to the current extent of defoliation; - defoliated trees have grown less than non defoliated ones for the past 16, 30 or 60 years according to the region and the species (29,42). These features are mainly interpreted as a complex long-term differentiation process, which seems to be initiated or accelerated by climatic anomalies, such as droughts, as revealed by detailed growth analysis. The "over-density" of the stands and, secondarily, unfavourable physical soils characteristics were found to favour the dieback of silver fir stands, likely by enhancing the drought stress (30); - the yellowing of the spruce (and fir) is obviously not frequent or severe enough to influence significantly growth at regional level. More striking is the observation made for silver fir (42)(and recently for Norway spruce (43))that stands which had recently become yellow were growing better than average over the past decades, which may confirm the importance of nutrient uptake by trees in the overall nutrient balance of the ecosystem (see 3.3); The follow-up of the visible symptoms has shown that - there is no general decline in France but a great variety of evolutions (positive and negative ones) according to the region and the species. Although the evolution since 1986 is coherent for conifers with the major climatic features,

391 A.

I

Radial growth index - Silver f i r

(%I

I

1850

1900

1950

year

B.

I Radial (XI

t

growth index - Sessile oak

0 1850

1900

1950

year

Figure 1 Basal area increment index of (A) silver fir (n=1674, Vosges area) and (B) sessile oak ( n 4 0 5 , Lorraine plain) according to calendar year ("standardized" in order to compensate for the age effect). Sources : A: M.Becker, in G. Landmann (ed.), French research into forest decline, DEFORPA Programme, 2nd report, ENGREF Nancy, 1991; B: M. Becker, T.M. Nieminen and G. GBrBmia, submitted.

392

that of broadleaved species seems more complex, and was overall less favourable over the past years. On the other hand, the yellowing in spruce on acid soils is largely reversible and has decreased significantly (32); - the apparent overall spatial relationship between the pollution climate (highest load in the NE) and the (known) damage (conifers in the NE) has proved invalid since the picture of forest damage has become more complete: the Mediterranean region is for example more affected and the cases of dieback (mortality of groups of trees or stands) recorded in 17 species by the DBpartement de la Sant.4 de la For% are scattered all over France (32).

3.6. Air pollutionefpecte on ground vegetation and on surface wahm Two recent studies (34,35)have established a significant shifi in the species composition of the ground flora over a short time period (2 decades), on a poor acidic soil in the Vosges mountains as well as on well buffered soils in the neighbouring Lorraine region. In both cases, the nitrophilous species have spread out: atmospheric nitrogen input appears the most likely explanation for this eutrophication. This assumption seems to be confirmed by the more marked evolution observed near the forest edges in the study where a systematic sampling was made (35). A recent survey of 39 small upland streams at snow melting and at low water in the Vosges mountains has demonstrated that a significant proportion were acidified and contaminated by aluminium. In fact, the trout disappeared in these streams many years ago. Interestingly, acidification is more pronounced in the catchments mainly covered by conifers, which is thought to be mainly the result of the enhanced acid deposition (33). 4. DISCUSSION

4.l.AirpohtiondFe~band foreat decline: can acollsensusbereached? ASHMORE et al. (44) stated in their recent review on "Air pollution and forest ecosystems in the European Community" that "the problem of forest decline is qualitatively different (from that of surface water acidification for example), in that there is no scientific consensus on a European scale on the role of pollution, the pollutants involved and the mechanisms of pollution damage". On the other hand, however, the representatives of 10 countries at the 5th MARC meeting (Acidification Research Coordinators) held in France in 1990 recognized that "there are far less discrepancies in the perception of forest decline in the different countries than 5 years ago". They agreed on a few key points: (a) the idea of a steadily increasing damage could not be validated, (b) the visible symptoms are not necessarily associated with air pollution, (c) the direct effects of gaseous pollutants on leaves and needles are probably less important than first assumed, (d) the indirect effects on forest ecosystems may be more important: a consensus of opinion has nearly been reached on the role of acidic deposition in the cation depletion; (. ..) nutritional imbalances could become a major mid-term problem, (e) in the eastern European regions like Erzgebirge, the direct effects are very likely to be more important (46). The remaining differences among countries in the perception of forest damage may be due to several factors:

393

- t h e identified effects on tree seedlings can not be extrapolated to ecosystems: some of them may be exaggerated by the experimental conditions; most of them may affect only sensitive individuals or provenances; -most research has been devoted to the "yellowing of the spruce", a spectacular phenomenon, new by its extent and partly by the causes involved, but also a peculiar aspect of the forest damage considered as a whole; - the population and ecosystem levels have been neglected: the potential contribution of the field ecology has been neglected (42); the role of climate, stand dynamics and past forest management has been a matter of controversy but not of much scientific work; thus, the ways by which numerous air pollution effects may be obscured or compensated for by other factors have not been sufEceintly investigated; - the identity of the "novel forest decline" is still confuse: the idea of an unprecedented phenomenon involving many species, based solely on the recent crown assessments, could not be validated by other objective methods; obviously, much of the currently discussed "new types of forest damage" is not entirely new; historical features of forest decline, as reported and discussed in the literature, have been largely ignored (42); - psychological, scientific cultural and political ones. For example, there is a tendency of some foresters and scientists to interpret each new or apparently abnormal feature as a (most) likely impact of acidification without looking for other reasons; besides, political and social pressure on scientists has led in some cases to delay or even totally prevent the publication of results in disagreement with the dominant national perception of air pollution effects. 45. The mlevance of the current scientific knowledge on air pollution effecta

for policy application Forest decline incited a tremendous research effort devoted to forest ecosystems and air pollution in the 1980's.Paradoxically, forest trees seem to be the least sensitive part of the forest ecosystem to air pollution: less sensitive than forest ground vegetation, soils, surface waters, or crops (not dealt with in DEFORPA). The critical loaflevel concept is a highly attractive one for policy makers and a stimulating one for scientists, but here again, political options should not bias the choice of some scientific criteria, still linked with great uncertainties. In the short term, the goal could be achieved reasonably well for surface waters and forest soils, despite several major problems (e.g. weathering rates) but seems rather unrealistic for forest health. As shown in this paper, air pollution obviously affects sensitive parts of the ecosystem, including trees, but the complex interactions between short-term and, even more important, long-term changes in climate, in silvicultural practices and in the pollution climate are far too poorly understood to derive dose-effect relationships for forest health and productivity. Well acknowledged threats for the sensitive parts of the ecosystems and risks for the mid- and long-term linked to some cumulative effects (cation depletion in soils) should be considered as sufficient reasons to pursue emission abatement. Besides, scientists should also clarify the fact that reductions in emissions will not have all the imaginable beneficial profits: many diebacks in Europe may not be significantly reduced by such measures (42); decreased wood production (and C02 fixation) may result from decreased nitrogen deposition at regional level in some parts of Europe (24).

394

6. CONCLUSIONS

Air pollution causes or contributes to: - Eutrophication of forest ground vegetation on acid and well-buffered soils: the main factor is thought to be atmospheric nitrogen deposition; - Acidification and cation depletion of sensitive forest soils: the main factors are acidic deposition and increased cation uptake linked t o several factors (introduction of fast growing conifers, accelerated tree growth linked to climatic changes, N deposition...); - Acidification and Fish mortality in surface waters in limited French areas: acid deposition is likely to be the prevalent factor but cation uptake by conifers should not be underestimated. Air pollution effects on forest trees or stands are likely under specific circumstances: on poor soils, where nitrogen deposition or cation depletion will lead to nutritional disturbances, the yellowing of spruce being the most spectacular case; in species or provenances sensitive to elevated ozone levels. However, except for the yellowing of spruce, most of the forest declines (which usually last only for a few years) due to abiotic factors are likely to be caused mainly by climatic anomalies affecting vulnerable stands: besides "natural" diebacks, which may be considered as part of the functioning of the ecosystem, various anthropogenic impacts (modification of the composition and structure of the stands, introduction of inadequate genetic material ...I act in a complex and not easily recognizable way (32,421. Overall, forest productivity has steadily increased at regional level, which suggests that the negative effects of air pollution are over-compensated by its positive effects (nitrogen deposition) and other factors (long-term evolution of climate, increase of CO2 levels, changes in silviculture practises). For all these reasons, it does not seem realistic to base the further emission abatement strategies on critical loadsflevels for forest health. The most sensitive compartments of the ecosystem should be considered preferentially.

1 M.Bonneau and G. Landmann, La Recherche, 19 (1988)1542. 2 Ph. Roqueplo, La Recherche, 19 (1988)1553. 3 G. Landmann (ed.), French research into forest decline. DEFORPA Programme, 2nd report, ENGREF, Nancy, 1991. 4 M. Bonneau and Ch. Elichegaray, in : A.H.M. Bresser and W. Salomons, (eds.), Acidic Precipitation, vol. 5 (1990)307. 5 Ch. Aschan, E. Dambrine, G. Nourrisson and M. Tabeau, Ann. GBogr. (in press). 6 E. Ulrich, in 3 (1991). 7 J.L. Colin et al., Atm. Environ., 23 (1989)1487. 8 Ph. Derexel and P. Masnibre, in : CEC Air pollution research report 30, Kluwer Acad. Publ., 431,1990. 9 E. Ulrich, Int. Symp. "Ecological Approaches of Environmental Chemicals", Debrecen, Hungary, 1991/04/14-19(in press). 10 P.E. Perros, N. Tsalkani and G. Toupance, Environmental Technology Letters 9 (1988)351.

395

11 A. Proyou, P.E. Perros and G. Toupance, Atm. Environ. 25A (in press).

12 G. Comic, Physiol. Plant. 71 (1987) 115. 13 M. Pierre and 0. Queiroz, Physiol. Plant. 73 (1988) 412. 14 J.P. Garrec, in 3 (1991). 15 M. Bonneau, in 3 (1991). 16 A. Probst, D. Viville, B. Fritz, B. Ambroise and E. Dambrine, Water Air Soil Pollut. (in press). 17 F. Lelong, C. Dupraz, P. Durand and J.F. Didon-Lescot, J. Hydrol. 116 (1990)125. 18 M. Bonneau, E. Dambrine, C. Nys and J. Ranger, Sci. Sol 29 (1991) 125. I9 T. Becquer et al.,Plant and Soil 125 (1990)95. 20 E. Dambrine et al., 1st Eur. Symp. Terrestrial Ecosystems, Florence, Italy, 1991/05/18-24(in press). 21 A.D. Mohamed et al.,submitted to Forest Ecol. Manage. 22 M. Bonneau, G. Landmann and C. Nys, Water Air Soil Pollut. 54 (1990) 577. 23 M. Becker, Int. Symp. "Tree rings and environment", Lund, Sweden, 1990/09/03-09(in press) 24 M. Bonneau and C. Nys, 1st Eur. Symp. "Terrestrial Ecosystems", Florence, Italy, 1991/05/18-24 (inpress). 25 S. Le Goaster, E. Dambrine and J. Ranger, Oecol. Plant. 6 (1991)(in press). a6 S. Le Goaster, E. Dambrine and J. Ranger, Water Air Soil Pollut. 54 (1990/91) 269. 27 D. Estivalet, R. Perrin, F. Le Tacon and D. Bouchard, Forest Ecol. Manage. 37 (1990)233. 28 M. Becker, Can. J. Forest Res. 19 (1989) 1110. 29 M. Becker, in 3 (1991). 30 M.Becker, G. Landmann and G. Uvy, Water Air Soil Pollut. 48 (1989) 77. 31 M. Becker, T.M. Nieminen and G. GBrBmia (submitted). 32 G. Landmann, in 3 (1991). 33 A. Probst, J.C. Maasabuau, J.L. Probst and B. Fritz, C.R. Acad. SCi., SBr. II,311 (1990)405. 34 M. Becker, M. Bonneau and F. Le Tacon (submitted). 35 A. Thimonier and J.L. Dupouey (submitted). 3f3 F. Aranda, Etude du comportement de l'ozone dans les regions rurales et naturelles en France. Rapport DEA, Univ. Paris VII, 1991. 37 E. Ulrich, in preparation. 38 P. Durand, C. Neal and F. Lelong, J . Hydrol. (in press). 39 G. Seufert and F.K Evers, Int. Congress Forest Decline Research, Friedrichshafen (FRG), 1989/10/02-06,vol. 2 (1990) 649. 40 J. Bonte, in preparation. 41 M. Hauhs and R.F. Wright, Water Air Soil Pollut. 31 (1986) 463. 42 G. Landmann, in : D. Mueller-Dombois and R.F. Huettl (eds.), Springer Verlag (in press). 43 E. Ulrich, M. Becker and J. Bouchon, in preparation. 44 M.R. Ashmore, J.N.B. Bell and I.J. Brown, CEC Air Pollution Research Report 29,1990 45 G.J. Heij, W. De Vries, A.C. Posthumus and G.M.J. Mohren, in : G.J. Heij and T. Schneider (eds.). Acidification Research in The Netherlands, Elsevier (1991)97.

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T Schneider (Editor) Acidification Research Evaluation and Policy Applications 1992 Elsevier Science Publishers B V

397

BACKGROUND, RESULTSAND CONCLUSIONS OF THE DUTCH PRIORFTY PROGRAMME ON ACIDIFICATION G.J.Heij and T.Schneider National Institute of Public Health and Environmental Protection (RIVM), The Netherlands

Abstract The second phase of the Dutch Priority Programme on Acidification was carried out from 1988-1991. The programme included about 50 projects carried out by 30 institutes and university departments. The main research areas were ammonia emissions, the deposition of SO,, NO, and NH, exposure-effect relationships for forest ecosystems (on dry sandy soils) and heathlands, and effectiveness of control measures. The last mentioned topic included the evaluation of emission-deposition scenarios and the assessment of critical loads. In 1989, 54% of the acid deposition in The Netherlands originated from the country itself. NH, is the main component in the acid deposition (46% in 1989, 80% from NL). Agriculture contributes the largest portion to the NL fraction of the acid deposition (62% in 1989). These figures illustrate the important role of ammonia in the acid deposition in The Netherlands. Nevertheless, sulphur is the most important factor in soil acidification (contribution about 65% versus nitrogen 35%). Soil acidification is the main effect of acid deposition in The Netherlands, resulting in depletion of the A1 buffer and deterioration of the groundwater quality. At the same time disturbance of nutrient balances occurs (also caused by excess nitrogen), resulting in a n increasing risk of damage caused by traditional stress factors like drought and pests. In The Netherlands, no mono-causal relationship exists between acid load and forest vitality in terms of needle loss and discolouration; acidification plays a certain role (a role that is very difficult to quantify) but it is not the only factor. In general, its impact is a weakening of the resistance of the tree against attack by other stress factors. The Dutch heathlands are seriously threatened by the overload of N deposition. Reduction of the N load to 700-1100 m o l h d y r combined with frequent sod cutting can improve the situation. From calculations with the Dutch Acidification Systems model, it follows that a scenario with a n average deposition of 1400 mol H+ per ha for the Netherlands as a whole in the year 2010 and an average deposition of 1400 mol H+ per ha for forests in the year 2050, seems sufficient to minimize the impact of acidification, a s one of the factors of combined stress important for the health of forests and heathland.

398

1.

INTRODUCTION

This paper presents an overview of the second phase of the Dutch Priority Programme on Acidification. First, the history of the programme as a whole is described, as well as the main research areas and some management aspects of its second phase. The most important research topics will be discussed in more detail: - ammonia emissions - concentration and deposition of acidifying compounds - effects on forest soils and forests - effects on heathlands Finally, some major conclusions are presented. In other papers, detailed information is given on scenario analyses with the Dutch Acidification Systems model (Tiktak et al.) and on the assessment of critical loads (De Vries). 2.

THE DUTCH PRIORITY PROGRAMME ON ACIDIFICATION

2.1. Hislnryofthepmgmmme

In 1985 the Dutch Priority Programme on Acidification was started, in order to give a solid scientific basis to the increasing interest of policy-makers in the effects of air pollution on ecosystems in particular. Since 1985 almost all the research on acidification in The Netherlands has been co-ordinated within the Dutch Priority Programme on Acidification. The first phase of this research programme (1985 - 1988) aimed to answer the following questions: - Which substances are responsible for the damage caused by "acid rain" and to what extent? - How (by what means and in what way) is this damage inflicted? - How effective are abatement measures? To a certain extent the answers to these questions were given in the Evaluation Report on Acidification (Schneider and Bresser, 1988). On the basis of an interim evaluation, carried out in 1987 , the Steering Committee for Acidification Research (which commissioned the programme) decided to start a Second Phase Programme, which was carried out during the period 1988 - 1991. This Second Phase Programme was focused on further quantifying of the findings of the First Phase Programme, so that effective policy measures could be recommended. 2.2. Main research areas (SecondPhase)

The main research areas of the Second Phase were ammonia emissions, the deposition of SO,, NO, and NH,, exposure-effect relationships for forest ecosystems (on dry sandy soils) and heathlands, and effectiveness of control measures. The last mentioned topic included the evaluation of emissiondeposition scenarios with the Dutch Acidification Systems model (DAS) and the assessment of critical loads. Within the programme much emphasis has been put on the integrated effects research, at the main (forest) research sites in Speuld and Kootwijk (the ACIFORN - Udification Research in EQEests in The Netherlands - project) and at the heathland site Assel, studying the

399

functioning of these types of ecosystems and the role of the stress factors air pollution, soil acidification and nutrient imbalances. The DAS model is intended to describe the quantitative relationships between emissions (in a large part of Europe), abatement techniques and associated costs, transport and transformation, concentrations and deposition, soil acidification and related effects on crops, trees, natural vegetations, small surface waters and materials (in The Netherlands). Under consideration are the acidifying substances SO,, NO, and NH, and also 03. The first version of the DAS model has already been used in the evaluation of the First Phase. The main emphasis in the Second Phase has been on the incorporation of the effect modules in the DAS model, and on the improvement of the input-output aspects (flexibility in use and presentation of results), on sensitivity, uncertainty analysis and validation, and on the use of the model in the evaluation of research in terms of policy recommendations. International cooperation (within the framework of the UN Economic Commission for Europe) is going on, to produce maps of critical loads and levels for Europe, in order to create a sound basis for policy measures. Especially the soil modelling work in the Dutch Priority Programme on Acidification has been closely related to the work on mapping of critical loads.

2.3. orgamah'onandbudget The overall management of the Dutch Priority Programme, Second Phase, has been carried out by the National Institute of Public Health and Environmental Protection (RIVM). The Steering Committee for Acidification Research has been responsible for the programme and for the funds available for subsidies. Responsibility for the scientific guidance of the programme was in the hands of the Programming Committee. With the Programme Director, a Project Group of up to 13 researchers has been responsible for the coordination and evaluation of the research projects. Members of this group each covered a specific research subject. The programme included about 50 projects, carried out by 30 institutes and university departments. In Table 1the sources of the subsidies available in the Second Phase of the research programme are presented. Table 1 Available budget in the Second Phase programme (subsidies; in Dfl millions) Source

period: 1988 - 1991

Ministry of VROM (environment) Ministry of LNV (agriculture, nature preservation) Ministry of EZ (economy) SEP (electrical power companies) SHELL (refineries) Ministry of V&W (public works, traffic)

total

3.1 3.1 3.1 1.2 1.2 0.9

12.6

400

The contributions from the research institutes themselves were substantial. Table 2 presents and overview of the total budget for the programme. Table 2 Total programme budget Second Phase (in Dfl.millions) total

subsidies - manpower - materials and subcontracting

9.2 3.4

contributions of institutes and other programmes - manpower - materials and subcontracting

12.6 3.4

total

28.6

3. AMMONIA EMISSIONS The determination of emissions of ammonia from relevant sources have been incorporated in the programme; emissions of SO2 and NO, have only been included in the integrated assessment study. The research area of ammonia emissions was selected because the uncertainty in the emission estimates of ammonia was too large to adequately calculate transport and transformation i n the atmosphere and the subsequent deposition. At the same time, measurements of the concentrations and deposition of ammonia and ammonium were almost absent, leaving a large area of uncertainty about the actual input of these substances. The livestock industry is the main source of ammonia emissions in The Netherlands. Factors, affecting emissions of ammonia from manure are depicted in Figure 1. The research on ammonia emissions was focused on the development of methods to measure and quantify ammonia emissions from different sources. Together with the Research Programme on Animal Manure and Ammonia, this led, for a number of source categories, to better insights into the extention of ammonia emissions and the possibilities of reducing them. The major source of ammonia emission through volatilization is the surface application of manure. Most ammonia is emitted during the first hours after spreading the manure. Emissions can be reduced considerably if the manure is quickly worked into the soil, through ploughing back o r injection. Application of more manure than can be taken up by the crop however, will cause problems with regard to the quality of the soil and groundwater. If manure is injected, the total amount of fertilizer and manure can be adjusted.

401

spreading

Figure 1.Factors affecting agricultural emissions of ammonia from manure.

So far it has not been possible to make a sufficiently accurate estimate of total ammonia emissions from the livestock industry in The Netherlands on the basis of measurements alone. Therefore, in order to obtain a n emission figure for The Netherlands as a whole and to present a n overview of the spatial distribution over The Netherlands, emission factors were used for the different sources. These factors, based on the method of the nutrient balance, had been compiled already by a n interdepartemental working group (De Winkel, 1988). On the basis of these emission factors, the most recent figure of total annual emission in The Netherlands is about 250 kton NH3. Since 1980 the emission figure for The Netherlands a s a whole has hardly changed. The estimated emission figures in this report for the total emission in The Netherlands and its spatial distribution are consistent with measurements of concentrations and with results of model computations for the pathway from emission via air transport to concentratioddeposition, and also with deposition and throughfall data. Differences are found for heavily loaded areas, where measured concentrations are systematically higher than the estimates. 4.

CONCENTRATIONAND DEPOSITION OF ACIDIFYING C O M P O ~

4.1. Introduction

The most important acidifying compounds are sulphur dioxide (SOz), nitrogen oxides (NO and NOz), collectively known a s NOx), and ammonia (NH3), together with their reaction products. These reaction products are acids ( H N 0 3 , HzS04) and airborne particles o r aerosols (NH4+, Nos-, SO42-). Emission of SO2 results mainly from combustion of sulphur-containing fuels, such as oil and coal, primarily in the processing industry (refineries) and in power stations, which are relatively high-stack sources. NO, is also produced in combustion, by oxidation of the nitrogen in the air to nitrogen oxide and

402

nitrogen dioxide. Important sources of nitrogen oxides are traffic, power stations and heating. The primary sourcea of NH3 are the production and the spreading of animal manure. 4.2. concenbratonlevels Figure 2 shows the spatial distribution of SO2 concentrations over The

Netherlands for 1989. Since 1980 the average concentrations of SO2 have shown a downward trend throughout the country (Annual Air Quality Report, 1989). Figures 3 and 4 show the spatial distribution of average concentrations of NO2 and NO, over The Netherlands for 1989. The average level of nitrogen oxides in The Netherlands has remained more or less the same since 1980. Figure 5 shows the calculated spatial distribution of NH3 concentrations over The Netherlands for 1988. The distribution of concentrations has not been calculated for other years. Since 1980, however, the level of total emissions of NH3 has been fairly constant, since the number of livestock has hardly changed. NH3 concentrations in the years from 1980 onward can therefore be expected to be similar to the concentrations shown in Figure 5.

Figure 2. Average SO2 concentration in The Netherlands in 1989 (pg.m-3).

403

Figure 3. Average NO2 concentration in The Netherlands in 1989 (pg.m-3).

Figure 4. Average NO, concentration in The Netherlands in 1989 (ppb).

404

Figure 5. NH3 concentration in the air in The Netherlands in 1988 (pg.m-31, calculated with the TREND model. 4.3. Depositionvalues

Total deposition is the sum of dry, wet and occult deposition. Occult deposition is less important as in The Netherlands, but can be very important in mountainous areas. In the beginning of the Second Phase, attempts have been made to estimate the fluxes of acidifying compounds. During the execution of the Second Phase, it appeared that stronger emphasis was necessary on atmospheric input fluxes to different receptors. Therefore, in 1989 a project has been started (in fact a small programme in itself) aimed at the determination of deposition fluxes onto forest stands and heathland. Since acid deposition consists of various substances, it is necessary to find a term which covers all these substances. In The Netherlands the term * * , expressed in terms of quantity of acid (mol H+halyrO -x + NO, + NH,. In determining the potential acid

405

deposition, the various substances are considered to contribute to acidification as follows: - the deposition of oxidized sulphur compounds (SO,), includes dry deposition of SOz, and sulphate aerosol (SO4), as well a s wet deposition of sulphate (SO4). 1mol SO, can lead to the production of 2 mol H+; - the deposition of oxidized nitrogen compounds (NO,) includes dry deposition of NO, NO2, HN02, HNO3 and nitrate aerosol (Nos), as well as wet deposition in the form of NO3.1 mol NO, can lead to the production of 1 mol H+; - the deposition of reduced nitrogen compounds (NH,) includes the dry deposition of NH3 and NH4 aerosol, as well as the wet deposition of NH4.1 mol NH, can lead to the production of 1mol H+. Acid deposition to specific surfaces is determined by source characteristics, distance from the emission source, physical and chemical processes in the atmosphere and type of receptor.

..

countrv-averamd deDSince 1980 the annual average total potential deposition has shown a downward trend: from about 6800 mol H+ per ha in 1980 (Figure 6) to about 4800 mol H+ per ha in 1989. About one third of this 4800 mol consists of wet deposition and two thirds consists of dry deposition. One of the major causes of this downward trend is the decrease in SO2 emissions in Western Europe, including The Netherlands. Apart from these emission reductions, the meteorological conditions are also important.

8000

2

D S O X mol ha

'

~ - 1 0 , mol ha

NH, mol.ha

'

-pot

acid mol H+.ha

5000 4000

X

3

3000 2000 1000

1980

1981

1982

1983

1984

1985

year

Figure 6.Deposition trend 1980 - 1989.

1986

1987

1988

1989

406

The contributions of the various compounds to total potential acid deposition in 1989 were: NH, about 46%, NO, about 24%, and SO, about 28% (Figure 7). The origin of the potential acid deposition i n The Netherlands in 1989 and the contribution per sector to the Dutch portion are shown in Table 3. In 1989, almost 55% of the total potential acid deposition in "he Netherlands was caused by emissions in The Netherlands itself.

SOX wet 9%

\

NH, wet 13%

NH, dry 33%

Figure 7. Composition of total potential acid in The Netherlands in 1989. Table 3 Origin of potential acid deposition in The Netherlands in 1989 (per compound and total), and the contributions from the various sectors to The Netherlands fraction (%) total acid

SO, NO,

NH,

UK + Ireland France Belgium Germany The Netherlands Eastern Europe Rest TOTAL

8 2 100%

17 11 2 9 1 3 5 17 8 5 l o 1 9 5 28 41 81 1 6 5 2 2 3 1 100% 100% 100%

Refineries Power stations Road traffic Industry Agriculture Households TOTAL NL

5 3 18 9 Q 3 100%

2 4 2 0 1 3 6 0 2 7 7 7 0 3 5 9 2 1 2 9 1 1 4 4 100% 100% 100%

9 8 10 10

54

401

The spatial distribution of the total acid deposition in 1989 is shown in Figure 8.

Figure 8. Total potential acid deposition in 1989 (mol H+ha-lyr-1). es in rouPhneSQ The influence of large-scale differences in roughness has been accounted for in the estimated figures for regional deposition. These differences lead to a higher dry deposition rate for forests, heathland and heathland lakes than for an average Dutch landscape. As a result of this influence, combined with the location of forests in relation to source areas, the average dry deposition on large stretches of forests and heathland in The Netherlands is estimated to be higher by about 20% and 10% respectively. Changes in roughness and other features of the vegetation have not been taken into account in these estimates. At the local level, for instance in the case of forest edges, inclusion of these

408

aspects can lead to a higher deposition figure than would be the case for a large stretch of forest. throughfid With regard to throughfall (the rainwater which falls through the canopy on to the forest floor), measurements adjusted for sea salt and neutral aerosols were compared with deposition estimates. At two research sites throughfall measurements were compared with deposition values derived from micrometeorological measurements. Both comparisons showed that there are still considerable differences between throughfall and atmospheric deposition, in spite of adjustments being made. Further process-oriented research is required to find a direct causal relationship between throughfall and atmospheric deposition. 6.

EFFECTSONFOREsTSOIISANDFOREsrs

5.1Introduction

The study on effects on forests has been restricted mainly to Douglas fir on well-drained, nutrient-poor sandy soils. Douglas fir has been chosen as the “model tree“, because of its assumed resistance to pests and because, unlike most other coniferous species in The Netherlands, Douglas fir has needles with ages of several years, which is important in determining exposure-effectrelationships. Investigated subjects are: - hydrology of the atmosphere-tree-soil-system; - tree physiology; - morphology of needles and wood; - rootslmycorrhizae; - soil chemistry; - exposure-effect research; - indirect effects (via the soil); - modelling. Controlled laboratory and field experiments on needles, leaves, branches, and entire plants and trees have already shown that air pollution and acid deposition can affect the growth and development of vegetations in different ways. The results of such experiments, however, cannot easily be applied to field conditions, under which an ecosystem is exposed to all kinds of other stress factors. Usually there is not a monocausal relationship between the acid load on forests and their health, in terms of needle density, needle discolouration (observed under field conditions as “vitality”) and growth. There is rather a combined and complex influence of various biotic and abiotic factors. Acidifying deposition generally reinforces the impact of traditional stress factors (frost, drought, disease and pests) on forest health (greater risk).

5.2. Direct and indirecteffects,at present atmospMc acid loading llirect effech With regard to the direct effects of air pollution in The Netherlands, the short-term effects are relatively unimportant apart from the visible damage near local sources (NH3, at low temperatures) and during episodes of high

409

concentration 0 3 , and probably also a reduced activity of photosynthesis at low temperatures and high humidity (S02).A t current daily 0 3 concentration values, a slight reduction of the photosynthesis activity was also observed in mature Douglas trees in the field. In general, such a change in photosynthesis should not have effects on annual growth. So far, there are insufficient data on possible long-term effects of air pollution (during a period of more than one year), because there have been few long-term experiments. There are, however, strong indications that long-term effects are important. The wax morphology and amount of wax on needles of Douglas fir has been studied in the field and in fumigation experiments with young trees. These experiments showed that the needle wax morphology cannot be considered to be an indicator of damage from air pollution. Increasing nitrogen deposition over a period of several decades has led, first of all, to a removal of N deficiencies and increased growth and, secondly, to nutrient imbalances as a result of Mg, K and P deficiencies. More and more forest ecosystems are moving from a situation of nitrogen deficiency to a situation of nitrogen saturation. A t the moment about 15%of the Dutch forests soils is N saturated. At the same time, the nitrogen input, together with SO, deposition, is causing considerable soil acidification. The present contribution of nitrogen to actual soil acidification is about 35%, and that of sulphur about 65%. The combined action of increased N availability and soil acidification has led to a decline in cation availability (nutrient deficiency). This is resulting in a greater risk of damage to forests caused by pests, frost and drought (Figure 9). As regards the effects of acidifying deposition in The Netherlands, soil acidification is the greatest risk factor. A major concern are virtually irreversible changes in the soil caused by depletion of the Al buffer, and the consequences of these changes for soil pH. It is not possible to specify precisely, the long-term consequences of a decline in pH (to between 2.8 and 2.9) associated with aluminium depletion, which (in the event of unchanged deposition) is the expectation for Dutch dry forest soils. In any case, large changes in the soil and thus in the conditions of forest stand locations, are to be expected. This could lead to changes in vegetation and soil fauna. If the current deposition trend continues, groundwater quality will further deteriorate. Research on the composition of shallow groundwater at 150 different locations (forests and heathland) has shown that the nitrate content was higher than the drinking water stand and (0.8 mmol.1-1) in almost 30% of the coniferous forest sites investigated. In the case of deciduous forests this was the case in 13%of the sites. The Al content of the shallow groundwater at the above-mentioned 150 sites is above the drinking water standard (7 pmol.1-1)for almost 90% of the coniferous forest sites investigated and for 70% of the deciduous forest sites. Problems can be expected with A1 and NO3 in case of presence of shallow private wells in forested and heathland areas. Finally there is the danger that the increased amount of stored organic N may be released more quickly as nitrate, if there is increasing mineralization and nitrification (in the event of a temperature rise) or a temporary reduction

410

in N uptake (if forests disappear as a result of disease or large-scale felling). In terrestrial ecosystems in which N is the growth-limiting factor, the input of extra N will lead primarily to increased production. Further increase in the N load will lead to a change in the composition of these ecosystems. An increase in the availability of N strongly stimulates the growth of nitrogen-loving grasses and herbaceous plants, particularly in forests where the canopy is open as a result of needle loss. On the basis of long-running experiments with high nitrogen doses, it is to be expected that this increased growth of nitrogen-loving plants will continue.

i

-

/\

N EXCESS

S

I

/

35% 5% SOIL ACIDIF :ATION

1

1 *, DISTURBED NUTRIENT

BALAPICE

AI-BUFFER <

*

pHc

SITE CONDITIONS

EUTROPHICA AT ION^

uptake of Mg, K, Ca strongly inhibited at Al / Ca 1 and high NH4 I cation - ratio's

Figure 9. Indirect effects. 5.3. Evaluationofde~ticmI.eductionscenarios

From calculations with the Dutch Acidification Systems model, it follows that a scenario with an average deposition of 1400 mol H+per ha for The Netherlands as a whole in the year 2010 and an average deposition of 1400 mol H+per ha for forests in the year 2050,seems sficient to minimize the impact of acidification, as one of the factors of combined stress important for the health of forests and heathland. The critical values for Al concentration, AYCa ratio and NO3 leaching in forest soils are exceeded in the year 2010 in less than 20% of the forest soils investigated.This scenario results in hardly any exceedances of these critical values in the year 2050. A scenario with a deposition of 1400 mol ha to forests in the year 2010,and a deposition of 700 mol per ha for The Netherlands as a whole in the year 2050,achieves a negligible exceeding of all critical values 10 to 15 years earlier than the first mentioned scenario.

41 1

6.

EFFECTSONHEATHLANDS

Heathland research included the heather itself, heathland species and competition with grasses. Heathland can be roughly divided into two types: wet heathland (which is influenced by groundwater for at least some part of the year and in which the dominant species is Erica tetralix), and dry heathland (which is not influenced by groundwater and in which the dominant species is Calluna vulgaris). On a quantitative basis the dry heathland type is by far the most important in The Netherlands. The area of heathland has rapidly decreased over the years, due mainly to cultivation and afforestation. In the remaining heathland the dominant species have been increasingly replaced by grasses. At the same time rare heathland types such as Arnica montana, Antennaria dioica (mountain everlasting), and Viola caninae (heath dog violet) have almost disappeared altogether. An inventory by means of satellite imagery has shown that about one third of the heathland in The Netherlands is still vital (> 70% covered by heathland species), about one third contains large amounts of grass and will probably change into grassland within the next 3-5 years, and about one third has already changed into grassland. So, the Dutch heathland is rapidly changing into grassland. Although many possible causes of this degradation are reported (such as ineffective management, lowering of groundwater levels and stress from excessive recreation), it is obvious that air pollution and the resulting soil acidification and N eutrophication are key factors in this process. The research carried out as part of the Dutch Priority Programme on Acidification was focused on direct and indirect effects of SOx, NO,, and NH, on udominantnheathland species (Calluna and grasses) and on urare” species (Arnica montana, Viola caninae). No research was carried out on the effects of ozone on heathland vegetation. The most important findings of the heathland research are:

-

At the ecosystem level, nitrogen input ultimately leads to the elimination of slow-growing species by fast-growing species, but Calluna will not be crowded out by grasses at nitrogen deposition levels up to 150 kg N ha-lyr-1if its canopy remains closed. Opening of a Calluna canopy can be caused by stress factors such as frost, drought, heather beetle plagues, or by natural ageing. Under normal conditions in The Netherlands, the canopy will hardly ever be opened by natural ageing. The critical nitrogen load for replacement of open Calluna canopy by grasses is about 10-15 kg N ha-lyr-1 (700 - 1100 mol, ha-lyr-1). At this critical deposition level, vital heathland can be maintained with a sod-cutting frequency of once every 50 years. Both experimental research on Calluna and modelling work on Calluna (in competition with Deschampsia) and Erica (in competition with Molinia) indicate the same critical load of 10 - 15 kg N ha-lyr-1.This is illustrated in Figure 10. With grazing and very frequent sod-cutting (once every 10 years) a vegetation of Calluna or Erica, though without rare species, can be maintained at nitrogen deposition levels up to about 30 kg N ha-lyr-1. (Present N deposition on Dutch heathland is approximately 35 - 40 kg N ha-

412 1yr-1.).

- At the individual plant level, nitrogen input (as NH3 or (NH4)2S04) causes growth stimulation even at low dosages. In rare heathland species however, changes may occur that make them more sensitive to frost, drought, and plagues. A critical level of N H 3 cannot be exactly defined, but is probably in the range of 5 - 10 pg.m-3 (long-term).

- The decline of rare heathland species is probably due to direct effects of gaseous SO2 and soil acidification. Adverse effects of SO2 on more than 5% of th e heathland species can be expected a t long-term average concentrations above a critical level of 8 pg.m-3. Effects on dominant species (Calluna and grasses) will probably not occur at the current SO2 levels in The Nether1ands.Crowding out of Violion caninae by grasses can also take place, but is probably only important at nitrogen deposition levels above the current ambient level in The Netherlands. However, this level may affect the reproduction or establishment of Violion caninae.

atmospheric nitrogen deposition 20 k~ N ha ' y r

C'D=5:1

5 60 8 a-D

.

40 20

0

0

--4

#

8

12

yea,

IS

20

40 r.

20

\

24

0

1

8

22

. IS

20

2,

YBBl

Figure 10.Model results of interaction between Calluna and Deschampsia at two levels of atmospheric nitrogen deposition. C:D=ratio between Calluna and Deschampsia at the start of the simulation.

413

7.

FINAL CONCLUSIONS

The impact of acidification on forest ecosystems in The Netherlands cannot directly be expressed in terms of (a percentage of) forest decline; there is no monocausal relationship between acid load and forest vitality in terms of needle loss and discolouration. Acidification plays a certain role (a role that is very difficult to quantify) but it is not the only factor. In general, its impact is a weakening of the resistance of the tree, making i t more vulnerable to other stress factors. Changes in the nutrient status of the tree play also a key role here. Consequently, in the Dutch situation the occurrence of visual damage related to acidification has to be expressed in terms of risk and cannot be predicted with dose response relationships. Abiotic changes in forest ecosystems in The Netherlands have been demonstrated, but its resulting impact on forest health is a risk problem. The greater the exceedance of the critical loads of the ecosystem and the longer it lasts, the greater the risk of damage by attack of traditional stress factors like frost and drought, pests and plagues. 8.

LITEXATURE

Schneider T. and Bresser A.H.M., 1988 Summary report; Acidification research 1984 - 1988, Report nr. 00-06 Winkel Kde, 1988 Ammoniak-emissiefactoren voor de veehouderij Publikatiereeks Lucht, nr. 76, Ministerie VROM, 1988 Annual Air Quality Report, 1989 (in Dutch) Rapport nr. 222101006, Laboratory for Air Research, National Institute of Public Health and Environmental Protection, 1990

This Page Intentionally Left Blank

T. Schneider (Editor). Acidification Research. Evaluationand Policy Applications 0 1992 Elsevier Science Publishers B.V. All rights reserved

415

ACIDIFICATION RESEARCH IN SWEDEN HBkan Staaf and Ulla Bertills Swedish Environmental Protection Agency, Research Department, S-171 85 Solna, Sweden

Abstract A number of acid rain research programmes have been conducted in Sweden since 1978. The total cost for these programmes has amounted to about 250 million SEK, and during this period an additional 950 million SEK has been used to finance practical countermeasures, mainly lake liming. Acid deposition has caused damage to soils, lakes, groundwater, flora & fauna, buildings and materials. The role of acid rain in causing forest damage is not yet fully elucidated. However, there is strong evidence suggesting that ongoing soil acidification and nutrient imbalances associated with it pose the major threat to Swedish forests. Current ozone levels are damaging trees on the physiological level, but the effects of ozone on forest production is unknown. Liming is an efficient means of counteracting the negative effects of acidic deposition on forest soils, lakes and watercourses. 1. BACKGROUND

Acid rain research in Sweden has played a n important role in Swedish environmental history. Swedish scientists identitied the acidification problem a t an early stage, and the political system responded quickly. The connection between emissions of acidic substances and effects on the environment was established as early as 1967-68. As a result, environmental authorities took steps for reducing sulfur emissions, starting in 1969. Initially, most research was directed towards surface water acidification, since it caused obvious damage in the form of declining fish populations. Studies made between 1970 and 1975 showed that liming could counteract surface water acidification. Thus, in 1976 a large-scale lake liming experiment was started. Since 1982 a liming programme for lakes and watercourses f h d e d by state grants has been fully operative. The first integrated acidification research programme, started in 1977178, dealt with ecological effects on soil, surface water, forest, flora and fauna. Since then, several other programmes have been initiated. The total expenditure for these programmes during the period 1977178-1990191 has amounted t o about 250 million SEK (40 million US$). During this same period a n additional 950 million SEK has been spent on monitoring and countermeasures, notably lake liming. Research programmes are currently running in the following fields:

416

* * * * *

*

Ecological effects; soils, groundwater, forests, crops, flora & fauna Surface water acidification - effects and remedial measures Corrosion of underground structures and installations Liming and revitalisation of forests Protection and restoration of historical monuments Effects of soil and water acidification on human health

PO

Foresl.soi1 and groundwalel

H Surface water H Ljming experiments0-0

- 14

soil and groundwater

Buildings and maleria1s

/I / I / I

..i\ I

I

79

81

-8 -6

c77

:lo

I -

83

85

87

89

9lyear

Figure 1. Allocation of funds to W e r e n t areas of acid rain research in Sweden, 1978-1990. 2. DEPOSITION OF ACIDIFYING SUBSTANCES

Sweden stretches about 1 500 km from north to south and represents a considerable deposition gradient. Wet (bulk) deposition of H varies from about 0.05 keqhdyr in the far north to 0.5 keqhayr in the southwestern part. Dry deposition of sulphur and nitrogen compounds to different ecosystems has been mapped using EMEP-data, calibrated against throughfall measurements from about 80 forest sites. GArdsjon, a mixed deciduous forest area on the Swedish west coast, represents an area with high deposition load in Sweden. Here, the deposition of potential acidity is 2.8 keqhdyr, divided on the following components: - SO&30, 1.4 keqhdyr (48%) - NOx/NO, 0.7 " (26%) -" H i 0.7 " (26%)

417

An analysis of air and rain chemistry recordings made a t five EMEP stations in Sweden showed that from 1979 t o 1987 concentrations of sulphur dioxide and particle-bound sulphate generally decreased and NO, concentrations remained about the same, while the wet deposition of sulphate, ammonia and nitrate increased. Although total sulphur deposition seems to have decreased during the 1980s) a t least in southern Sweden, unfavourable climatic conditions counteracted the positive effects of reduced emissions (Figure 2). Throughfall measurements in southern Sweden have revealed a considerable spatial variation in the deposition of sulphur and nitrogen compounds. Exposed hills and forest edges receive the highest loads. Nitrogen deposition has been found to decrease markedly when moving inland from coastal areas. For sulphur, geographic gradients are less steep. Instead, tree stand characteristics are important in determining deposition rates. Thus, the total deposition of sulphur in deciduous forest and Scots pine stands generally increases by a factor 1.2-1.5 in relation to wet deposition, while for Norway spruce the corresponding factor is 2-3. Foreign sources play a major role in both sulphur and nitrogen deposition. On a national level, the Swedish contribution is 10-15%. In southwestern Sweden the domestic contribution to nitrogen deposition is higher (20-25 %) and is mainly in the form of ammonidammonium (Table 1). Table 1 Deposition of acidifying substances in the southernmost part of Sweden (province of SkAne), in 1988, according to EMEP data (Anonymous 1990) Total sulphur Deposition, kg N h d y r

Total nitrogen

NHJNH,

17

11

5

10

10 16

20 19 17 10 7 6 4 4 10 3

38 9 24 4 5 5 2 3 10

100

100

100

Proportion contributed by countries (%) Sweden Germany, West Denmark Great Britain Germany, East Poland France The Netherlands Others Unattributed Total

9 23 8

17 7

0

418

20

-

10

-

RORVIK 20

BREDKALEN

-

10 -

2 1

2 ' -

-

0 1979

1981

1983

1985

1987

0 1979

1981

1983

1985

1987

Figure 2. Estimated sulphur deposition at two Swedish EMEP stations from 1979 to 1988 at Riirvik (southwestern Sweden) and Bredkalen (north-central Sweden). Estimates were made with the EMEP model based on the following assumptions: 1) actual emissions, identical climate every year 2) actual climate, constant SO, emissions in Europe (After Lovblad 1990) 3. ENVIRONMENTAL EFFECTS

3.1. Soil acidification National Forest Survey data for 1983-1987 show that forest soils are least acid in northern Sweden and that the degree of acidity increases southwards. Analyses show that large areas of forest in southern Sweden have a pH value in minerogenic soil under 4.4,i.e. the level at which free inorganic aluminium begins to appear in large quantities in soil water. In total, the area affected is around 650,000 ha, of which approximately 300,000 ha is found in southwestern Sweden. We now have clear evidence that uncultivated soils in southern Sweden have become acidified t o a considerable extent over the last few decades. The strongest evidence for this has been obtained from studies where previously examined soil profiles have been re-examined. Reported reductions in the pH of forest soil over the last 10-55 years in southwestern Sweden are mostly around 0.3-1.0 pH units, but in individual cases, reductions as high as 1.5-2.0 pH units have been observed (Figure 3). In other parts of southern and central Sweden the reductions have been lower and in the northernmost parts of Sweden no acidification caused by air pollution has been found. Soils with a high pollution load seem to be sulphate-saturated to great depth, and leaching of sulphate has resulted in a 30-70% decrease in exchangeable stores of base cations in southernmost Sweden over the last four decades. Chemical balance calculations also suggest that easily available stores of calcium and other base cations are decreasing by 1-2 per cent annually.

419 p l i chunge

0

o &ech forest Other deciduous forest A Noway spruce forest

0

.

h.0: 0 0 0

0 0 0 0

*no0

0 .

I

Figure 3. pH changes in the A horizons of 104 forest and pasture soils in the southernmost part of Sweden over a period ranging from 14 to 35 years. The final sampling was made 1984 or 1985. (After Falkengren-Grerup 1987) The greatest pH reductions have been found in soils of medium fertility with relatively high original pH values, i.e. in the ion-exchange buffer range. The acidification front currently lies a t several meters depth in calcium-poor moraine soils in southwestern Sweden, where it has reached superficial groundwater. Further to the north and to the east, acidification is more superficial. About half of the acidification of the surface layer of forest soils that has occurred since the 1920s can be attributed to biological acidification caused by forest growth and timber harvesting. Acidification of deeper soil layers can only be explained by the deposition of acid, mainly sulphuric acid. Thus far, nitric acid has contributed only slightly to acidification of mineral soil, groundwater and surface water in Sweden. Nitrogen leaching from Swedish forest soils is generally low, normally below 1 kg Nhdyr. Long-term monitoring of watercourses has not revealed any increase in nitrate leaching in northern Sweden during recent decades. However, increases in nitrate concentrations in soil water and small streams have been recorded a t forest sites where nitrogen deposition in throughfall exceeded 10 kg NO,-Nha'yr, indicating that nitrogen saturation is approaching in certain areas in the southernmost part of Sweden. The reasons for the increased nitrate leaching are still unclear. Nitrification in the soil profile of a spruce forest a t SoderAsen (province of SkPne) that showed extensive nitrate leaching was very low. Thus direct leaching of deposited nitrate would appear to be the most probable mechanism.

420

3.2. Groundwater Acidified groundwater has been reported from large parts of Sweden throughout the 1980s) particularly in areas where soils and lakes are acidified. Acidification is usually worst in shallow groundwater and diminishes with depth of the water table. Shallow wells are particularly affected in calcium-poor areas of southern Sweden and in certain coastal areas in the north. The problems are most pronounced for private water supplies. About 400,000 wells are used by permanent residents, and from 1985 to 1988 government grants were available for de-acidifying household water from private wells. Evidence that acidification is increasing has been obtained from inventories of private water supplies carried out between 1984 and 1986 and from data on municipal water supplies, as well as from groundwater measurements made by the Swedish Geological Survey and the Swedish National Monitoring Programme. We now foresee an increase in the acidification of groundwater in large areas of Sweden. Particularly in areas receiving a heavy load of acidifying substances, markedly increased aluminium levels appear in shallow groundwater. As soils are acidified to greater depth, an increase in the transport of aluminium and cadmium to the groundwater can be expected. As a result of the corrosion of plumbing systems, sharply raised levels of copper have been found in drinking water from many private wells throughout Sweden. During the last decade, copper in drinking water has been suggested to be a cause of diarrhoea in small infants and children in Sweden. However, the relationship is still unclear, and further studies in this field are underway. 3.3. Lakes and watercourses

Biological and chemical analyses of water have shown that most Swedish lakes and watercourses have been affected by acidification. About 16,000 of Sweden’s 85,000 lakes are so badly affected by acidification that sensitive species have drastically declined in number or disappeared completely. Almost 6,000 of these lakes have been limed. The acidification situation for Swedish watercourses is less well known. Estimates suggest that at least a quarter of the total length of watercourses would be seriously damaged by acidification if it were not for liming. The acidification of lakes and watercourses takes place primarily as a result of chemical and physical changes in the catchment area. A good general understanding of acidification history has been achieved by combining results from long-term monitoring of individual lakes, paleolimnological studies and modelling. The biotic changes associated with acidification have also been reasonably well documented. This is especially true for the inhabitants of open water and associated food chains. The response of littoral communities in lakes has been less well studied. Large areas in southern Sweden have chronically acidified surface waters with low pH, no alkalinity and high aluminium levels the year around. Here, soil acidification has penetrated so deeply that even deep-flowing groundwater supplies very little alkalinity to the waters. In northern Sweden numerous watercourses are periodically acidified - usually after heavy

42 1

rainstorm events in autumn and during snow-melt in spring. Results from a catchment study in northern Sweden indicate that dissolved organic acids can play a major role in these acid events, and that this acidity originated within a few meters from the stream. Most small streams in northern Sweden are surrounded by moist, humus-rich zones which appear to be important for short-term surface water acidification in northern, humid areas. Acid events can be very pronounced even if the soils are largely unaffected by acid rain. The acidification situation in the southern and coastal parts of northern Sweden has not changed to any great extent since the mid-1970s. By contrast, in mountain and submontane regions in northern Sweden, it has continued t o deteriorate over the last 10-15 years. Damage to minor watercourses in the southern part of the mountain region and in adjacent areas increased considerably in the 1980s. Both low pH and high aluminium concentrations seem t o be harmful to biota in acid lakes and streams. For benthic invertebrates the pH itself seems to be more important than aluminium concentrations in explaining the response and distribution patterns, but species resistant t o low pH generally tolerate elevated aluminium concentrations as well. However, for certain fish species, e.g.salmonids, aluminium is considered to be a main factor limiting their survival in acid waters, Generally, it appears as though biotic changes in response to surface water acidification can be ascribed to biotic interactions in the food-chains combined with abiotic stress on a long-term or short-term basis, giving different physiological and behavioural disturbances, but also favouring certain species. Nu mr d species

m a

60 50

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40

30 20

'

10 '

o

h

I 1

4.5

.

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.

5.0

5.5

6.0

6.5

7.0 pH

Figure 4. Number of phytoplankton species in lakes in southwestern Sweden in relation t o their pH (summer values) (After Eriksson et al. 1983)

422 3.4. Forests

Forest damage, in the form of needle loss and crown thinning, became noticeble in southern and central Sweden in 1983, particularly in coniferous trees. In national inventories made between 1984 and 1990 crown thinning (more than 20% needle loss) was evident on an average of 24 per cent of the Norway spruce trees and 13 per cent of the Scots pine trees. Only about 1 % of the trees were classified as seriously damaged (suffering more than 60 per cent needle loss). Birch, oaks and beech also showed considerable crown thinning. An analysis of the various surveys made show that forest damage varies considerably a t both regional and local levels. The degree of damage increases with stand age, but in most cases no straightforward relationships have been found between forest damage and emission sources or soil conditions. In the case of Norway spruce, the most extensive damage is found in inland areas of northern Sweden, probably because these forests tend to be very old and the climate is harsh. Differences between northern and southern Sweden are considerably smaller for Scots pine. The causes of forest damage in Sweden are still uncertain, but most scientists favour the multi-stress hypothesis. Mean monthly concentrations of sulphur dioxide and NOx in background areas of southern Sweden are generally below 5 ug/ms air, even in winter, which is considerably below damaging levels.

1

.\

\

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-NPK

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.

11

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900

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Figure 5. Basal area growth of young Scots pine forest from 1972 to 1984 in plots given different doses of sulphuric acid. A field experiment in central Sweden (After Tamm and Popovic 1989)

423

Ozone concentrations, on the other hand, are near the critical level thoughout Sweden. Studies on Norway spruce subjected to different ozone levels in open-top-chambers for 5 years showed that the current ozone level in southern Sweden causes ultrastructural damage to needles and reduces the rate of photosyntesis. Indirects effects of acid deposition, especially soil acidification and nutrient imbalance, are probably the most severe threats to the long-term vitality of Swedish forests. Long-term studies indicate that nutrient concentrations in conifer needles in southwestern Sweden have changed over the last twenty years - i.e. nitrogen concentrations have increased while those of most other nutrients have decreased - which might be indicative of an approaching nutrient imbalance. Low base cation stores and increasing concentrations of Al3+ in the soil solution might impair the uptake of phosphorus, potassium, calcium and magnesium, and eventually some of these elements could become limiting for forest growth. However, there is still no example of tree growth responding to a supply of any of these nutrients in a Swedish fertilizer experiment, unless added together with nitrogen. The initial response of forest trees to soil acidification may be an increase in growth. This is suggested by two acidification field experiments in which sulphuric acid was added to the soil for eight years. Plots given moderate doses of acid (up to 600 kg H,SO,/ha) showed increased tree growth, while higher doses of sulphuric acid did not result in a growth increase. These experiments were performed in north-central Sweden, where the acidic load is relatively low, and although the treatments caused losses of available calcium and magnesium from the soil they did not induce any nutrient changes in the trees. To study how trees growing on an acidified soil respond t o stress, a long-term field experiment was started a few years ago at Skogaby in southwestern Sweden. Thus far, both drought stress and the addition of ammonium sulphate have resulted in slightly reduced growth. We do not know the extent t o which air pollution contributes to the forest damage observed in Sweden. Air pollution affects the forest both positively and negatively, and to date it seems as if heavy nitrogen deposition has increased forest growth. However, it seems probable that soil acidification and ozone have been partly counteracting this growth-enhancing effect.

3.6. Crops Crops may be affected directly by gaseous and particulate air pollution and indirectly via soil acidification. Ozone is probably the atmospheric pollutant causing the greatest damage. Levels of ozone in Sweden during the summer months commonly approach or exceed threshold concentrations above which crop damage can occur. Field studies have shown that current ozone levels reduce the yield of spring wheat by approximately 10 per cent in western Sweden, whereas barley seem to be less sensitive. Economic losses associated with anthropogenic ozone-caused yield reductions have been estimated to be 1.4 billion SEK per year for the period 1986-88. Injury to ley, oats, potatoes, winter wheat and spring wheat accounted for most of the losses. Air pollution in the form of acid deposition, together with harvesting, fertilisation and leaching, contributes to the acidity of agricultural land. The

424

relative contributions of these acid sources to acidification have been estimated to be: acid deposition 12 %, harvesting 20 %, use of fertilisers 32 %, and leaching 36%. Compilations of land survey data from the National Board of Agriculture (1958-1961) show that the acidity of soils has not changed much in recent decades. The optimal pH value for arable soils is usually 6.0-6.5, which is considered to provide satisfactory protection against absorption of heavy metals, particularly cadmium, by crops. About 45 per cent of the arable hectarage in Sweden has a pH lower than 6.0 and approximately 15 per cent has a pH value lower than 5.5. These acidity levels are not strongly related to the amount of acid deposition, since the worst lime/calcium situation is found in the two northernmost counties, where deposition is low. The pH of arable land is determined primarily by the type of soil, the crops grown and the amount of liming carried out. 3.6. Flora and fauna It has long been known that acidification of lakes and watercourses causes great changes in aquatic plant and animal life. Knowledge of the effects of air pollution on terrestrial flora and fauna is, however, still very patchy. Air pollution has afTected plant and animal life directly through the effects of gases, acid rain and acid water, and indirectly as a result of changes in nutrient availability and elevated levels of toxic metals in the ground. In addition, secondary effects have resulted from changes in competition, food availability and physical properties of the habitat. Extensive changes in epiphytic mosses and lichens have been observed in southern Sweden. Many species have disappeared over the last 40 years, while others are presently endangered. Nitrogen-f~ng lichens containing blue-green algae are most threatened. This group comprises about 130 species, or 6 per cent of all Swedish lichen species. Lichen disturbances can be ascribed mainly to the direct effects of nitrogen compounds and acid rain. The vascular flora in southernmost Sweden, including trees, herbs, grasses and ferns has changed over the last 15-35 years, particularly in deciduous forests. Populations of many species preferring neutral or only weakly acid soils have declined, while others with a preference for nitrogen-rich environments have tended to increase. The same tendency has been observed for hngi. These changes are probably due to soil acidification and increased nitrogen deposition. Air pollution effects on the terrestrial fauna are of an indirect nature and are generally most pronounced for species preying on aquatic organisms or reproducting in aquatic habitats. This category includes many insect species and some species of birds. As acidification increases, the amount of available calcium in the soil is reduced. This has had severe effects on populations of land snails in several parts of southern Sweden. There are also signs that the number of overwintering coniferous forest birds has diminished in areas of damaged forest in southwestern Sweden. This decline may be linked to changes in spider populations (a source of food for the birds) resulting from the the thinning out of tree crowns. Thus, it appears as though plant and animal life in the forests of southern Sweden, as well as in marshes and in other uncultivated areas are

425

changing as a result of acidification. The terrestrial fauna is probably changing more slowly, with the exception of species dependent on lime-rich environments. Air pollution mainly affects the fauna by causing reproductive disturbances, and such effects take a long time to surface in the form of population changes. There is therefore a significant risk that faunal changes are taking place, even though we have not yet detected them.. 3.7. Human health Air pollution poses a threat t o human health in two quite different ways: First, potential risks arise when people are exposed directly, by inhaling acid air pollutants. Second, acidification will change the level of exposure of several metals and metaloids via food and drinking water. In the latter field, only limited research has been performed. However, these aspects are currently included in two research programmes that started earlier this year (1991). One of these aims a t developing models and methods for evaluating health risks caused by air pollution. The other programme will focus on the indirect effects, mainly via effects on solubility of metals in the environment. 3.8. Buildings and materials The first integrated research programme with the aim of studying the effects of acid rain on Swedish cultural heritage was started in 1988. First, surveys were performed t o estimate the extent of damage to various objects, notably stone buildings, bronze statues and monuments, mediaeval glass paintings, sensitive textiles, runic stones and bronze age rock-carvings. Thereafter the restoration and conservation of threatened objects in all the above mentioned categories is t o begin. Projects of more basic character have also been initiated. The largest of them concerns weathering mechanisms in sandstone and limestone materials. Other projects are directed towards the study of mineral weathering in granitic rocks and the effects of lichens on decomposition of rock material. Lichen growth on rock-camings and runic stones appears to be an increasing problem: Research is also underway on corrosion of installations in contact with acid soil and water. Corrosion affects the longevity of pipes, cables, cisterns, foundations and other structures in the ground. Zinc is the metal most sensitive to acidification, and corrosion of zinc strucures increases as the pH and alkalinity of water decreases. Lead is also considered to be sensitive to acidification, as are copper, cast-iron and carbon steel. Among other materials, concrete is affected negatively by acidification, whereas plastics generally resist acidification well. Water pipes and road culverts, etc., are often exposed simultaneously to acid soil and acidified water. Approximately 60 per cent of the municipal water supply network is made of cast-iron, while copper is the most common material used for pipes in indoor plumbing systems. The occurrence of corrosion damage in water pipes in Sweden is generally most serious in acidified areas. The Corrosion Institute carried out a study on the situation in the early 1980s, a t which time it was estimated that approximately one third of all indoor corrosion damage was due to acidification. It is not

426

possible at present to estimate the total economic loss resulting from such acidification-caused damage. 4. AMELIORATION OF

DAMAGE

4.1. Forest liming

A Swedish programme to counteract forest soil acidification was started in 1983. Initially a number of old liming experiments were reinvestigated and evaluated. In addition, a new set of field experiments on forest liming were started. Tree growth, soil chemistry and soil biology, as well as effects on surface water and flora & fauna, are being monitored in these experiments. Results of the old experiments demonstrate that the application of 5-10 tonnes limeha to forest soils results in long-lasting (more than 50 years) effects on pH, base saturation, exchangeable base cations and aluminium concentration. In the new experiments about 3 tonnes limeha was added in most cases. As a result many large soil animals, such as earthworms, molluscs, diplopods and isopods, were favoured while small ones, e.g. mites, nematodes and enchytreaeids, decreased in number. Active hngal hyphae and bacteria were not affacted to any great extent nor was any damage to mycorrhiza noted. On dry, nutrient-poor soils liming led to a depression in tree growth lasting 10-20 years, followed by a slight enhancement of growth. On more fertile sites this initial growth reduction did not take place, and in some cases a positive growth effect was noted soon after applying the lime. Laboratory studies on soils with different nitrogen contents were set up to explain these observations. These studies showed evidence that the growth response of the trees was determined mainly by the effect of lime on nitrogen mineralization. In nitrogen-poor soils liming induces a decrease in net nitrogen mineralization, thereby reducing the nitrogen supply to trees. A single lime dose of 2-5 tonnesha is recommended for acidified forest soils, but the knowledge base needed for making detailed recommendations is still incomplete. Liming is considered as a long-term measure for soil amelioration. Thus far, Swedish experiments have only produced a few results indicating that tree vitality can be improved by treating soil with limestone, dolomite, wood ash or industrial slags. Additional field experiments have recently been established to evaluate combinations of lime and various fertilisers in terms of their potential for counteracting future acidification-induced nutrient imbalances in trees. 4.2. Liming of wells and groundwater supplies

Different methods for neutralizing acid groundwater supplies have been tested in practical experiments. Good results have been obtained for; (i) recirculating well water over a bedding of sand and lime placed in the ground near the well, and (ii) placing a mixture of lime and sand in a ditch around or beside the well. With these methods positive changes of the well water with regard t o pH, alkalinity and content of iron and aluminum were generally achieved within one year. Other methods have been less successful, e.g. placing lime on the ground

427

over all or part of the watershed of the well. About 10 experiments of this type have been performed, and although large doses of lime were used, up to 50 tonnesha - sometimes applied as a slurry, the lime apparently penetrated the soil very slowly. Although the water quality in the wells has been monitored for up to 7 years after treatment no great effect has been noted so far. 4.3. Liming

of surface waters The Swedish liming programme is practically oriented. Thus, much of the research activies have been directed towards follow-ups to see if the practical liming operations gave the desired positive effects on water quality and aquatic life. Research on mechanisms have thus far been rather limited. Most liming operations are performed by applying finely ground calcium carbonate from a helicopter o r boat directly to the lakes. This procedure is often supplemented by wetland liming and installations of dosers in streams. Generally the desired chemical targets for surface water, i.e. pH above 6, alkalinity above 0.05 meq/L, and reduced levels of toxic aluminium, are easily achieved by lake liming using fine-grained limestone. In addition to these changes, colour increases in limed clear-water lakes and increases in phosphorus levels often occur. Thus, liming results in increased nutrient supply and detoxification of the water. Improvment of the chemical conditions in limed lakes and watercourses generally results in great changes in the lake biota. Although there is often initially a drastic rise in the abundance of certain opportunistic species, biological diversity soon increases. A progressive increase in populations of acid-intolerant species, followed by recolonization, generally takes place. However, certain relatively immobile species like crayfish, snails and mussels normally recolonize very slowly, and this is sometimes also true of fish. Reintroduction of fish species may be necessary in severely acidified areas where acid-sensitive species have been eliminated over large areas or whole water-systems. 5. CRITICAL LOADS

The critical acid load for surface water in Fennoscandia has been determined using the steady-state water chemistry method (Henriksen et al. 1990). In Sweden the mapping was based on data from 4 018 lakes sampled during a national lake survey made in 1990. Although a wide variation exits in sensitivity both within and between mapping grids (50x50 km) it is clear that the present deposition exceeds the critical load of the most sensitive lakes throughout the country. The critical load, based on a n ANC limit = 0.50 meqil, is exceeded for 46% of all lakes in the country. In southwestern Sweden the critical load is exceeded for up to 75-100% of the lakes, while the correspondig figure for the northern part is around 25%. The results also show that even with a 50% reduction of the acid deposition, about 15% of the lakes, mainly in southern Sweden, would still remain acidified, and that a reduction of more than 80% is needed to protect more than 95% of the lakes.

428

The critical acid load of forest soils has been determined for 1392 individual sites in Sweden using the PROFILE model (Sverdrup et a1 1990) based on geochemical data from soil samples collected by the Swedish Forest Survey. The results of the calculations confirm that Swedish forest soils are very sensitive to acid deposition. Of the Swedish forest area with moraine soils, 85% receives acid deposition in excess of the critical load, and sensitive soils occur throughout the country (Figure 6). A minimum deposition reduction of 83% is required to protect 95% of the forest area from being damaged in the long run. Such substantial reductions in acid deposition imply that both sulphur and nitrogen deposition must be reduced.

Figure 6. Critical load of acidity and exceedance of critical load for moraine forest soils in Sweden (5 percentile). The calculations were made with the PROFILE model and mapped using 50x50 km grids (Sverdrup, Warfvinge, Ros6n and Melkerud, unpublished).

429

6. REFERENCES

Anonymous. Air Pollution '90. Swedish Environmental Protection Agency Informs. Solna. (1990). Eriksson, F., Homstrom, E. Mossberg, P. and Nyberg, P. Ecological effects of lime treatment of acidified lakes and rivers in Sweden. Hydrobiologia 101 (1983) 145-164. Falkengren-Grerup, U. Long-term changes in pH of forest soils in southern Sweden. Environmental Pollution 43 (1987) 79-90. Henriksen, A., Kiimiiri, J. Posch, M., Lovblad, G. Forsius, M. and Wilander, A. Critical loads t o surface waters in Fennoscandia. Nordic Counsil of Ministers, Copenhagen. Environmental report 1990:17.(1990). Lovblad, G. Concentrations and deposition of air pollutants in back-ground areas - temporal and spacial variations. Swedish Environmental Protection Agency, Solna. Report 3812. (1990). (In Swedish) Sverdrup, H., de Vries, W., Henriksen, A. 1990. Mapping critical loads. Nordic Counsil of Ministers, Copenhagen. Environmental report 1990:14 (1990). Tamm, C.O. and Popovic, B. Acidification experiments in pine forests. Swedish Environmental Protection Agency, Solna. Report 3589. (1989).

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T. Schneider (Editor). Acidification Research. Evaluation and Policy Applications 1992 Elsevier Science Publishers B.V.

43 1

Finnish Research Programme on Acidification (HAPRO) 1985-1990 Pekka E. Kauppi Finnish Forest Research Institute, Unionink. 40 A,

SF-00170Helsinki. Finland (Director of HAPRO in 1984-1990 a t the Ministry of the Environment, Finland)

Abatxact A programme for research on acidification in Finland (HAPRO) was organised in 1985-1990.Sulphur deposition, estimated using a n atmospheric transport model a n d checked a g a i n s t m e a s u r e m e n t s , decreased from s o u t h to north. The same spatial gradient was observed for sulphur concentrations in lake water samples. This spatial correlation indicates a causal link between (European) sulphur emissions and t h e presence of excess s u l p h u r i n t h e Finnish environment. Nitrogen concentrations in lake water were insignificant indicating uptake of nitrogen in terrestrial systems. Various biological effects were detected in aquatic ecosystems and in forests. However, forest resources in terms of growing stock and wood growth were not affected. Policy implications of the programme with regard to the use of natural resources, air pollution abatement and research management a r e discussed in this chapter. 1.INTRODUCTION Before 1985,there had been little research in Finland on acidic deposition, in comparison to t h e neighbouring Nordic countries Norway and Sweden [see, however, 1-31. A reason for this was t h a t there were no alarming large scale impacts similar to, for example, t h e fish kill in Soerlandet in Norway. The F i n n i s h t o p o g r a p h y i s s m o o t h , w i t h o u t s i g n i f i c a n t m o u n t a i n ranges. Upstream lakes which are sensitive to acidification have only sparse and scattered trout populations. Other fish species are less interesting to the fisherman and, moreover, a r e more tolerant of acid conditions. In Finland, as in Norway and Sweden, fishing is popular and important but is practised almost exclusively in large downstream lakes and in the Baltic. These ecosystems are relatively insensitive to acidification damage. Forests and forest industries are of the utmost importance in Finland and much information about forests has been collected through the decades, especially a s regards timber production and utilization [4].About one half of the ex-

432

port revenue of t h e Finnish economy is from t h e forest sector. Forests and lakes are the main characteristics of the Finnish landscape. In 1983-1984the impacts of acidic deposition on aquatic ecosystems in southe r n Scandinavia were thoroughly understood. The role of t h e long range transport of sulphur compounds had been established. Alarming news were released on the status of German forests. There was much concern in Finland a s in many other countries on t h e s t a t u s of t h e environment. There were widespread opinions urging additional research. 2.

THE HAPRO PROGRAMME

In April 1984 t h e government of Finland decided to s e t u p a special programme for acidic precipitation research. A time frame and a budget plan were established. A General Plan for the programme was released in early 1985,with objectives a s follows 151: "The programme is concerned with the development of acidification caused by sulphur and nitrogen emissions and, more generally, with t h e problems associated with air pollution. The aim of t h e HAPRO programme is to study cause-and-effect relationships in air pollution and, on that basis:

-

-

to determine the extent of regional effects of air pollutants in Finland to study whether the harmful effects of a i r pollutants are on the increase to determine which a r e a s and components of t h e environment are being especially threatened t o assess what measures would most effectively, and a t the lowest cost, reduce the harmful effects of air pollutants."

Only a short time was allowed for establishing t h e research, and in 1985 about 50 full-time and 100 part-time scientists were already engaged in 38 individual projects. The volume of the work remained approximately a t this level until the year 1990. The programme secretariat consisted of the director, three members of the professional staff and two members of the technical staff. The government established a n advisory board which, assisted by a technical ad hoc group, supervised t h e programme, examined and confirmed t h e major decisions, especially those concerning finances. The main programme report, t h e book "Acidification in Finland", was released in August 1990 [61. There are 62 individual research articles by a total of 145 authors contained in one volume of 1237 pages with 440 figures. An international committee of experts reviewed all t h e material offered for the book during a one-week meeting in June 1989.The members of the programme secretariat, on the basis of this material, prepared a n assessment report in collaboration with two programme scientists [71. T h e m a i n outcome of t h e HAPRO programme is summarised i n this chapter.

433

3. PRINCIPAL FINDINGS

3.1. Pollution climate Annual mean concentrations of sulphur dioxide in rural areas were 6 to 12 pg/m3 in southern Finland a n d 2 to 6 pg/m3 in northern Finland. Seasonal peaks occurred in winter, when the concentrations in southern Finland were 5 to 10 times higher than the lowest values measured in August. Monthly mean concentrations during the growing season were between 1 and 2 pg/m3. Three quarters of the sulphur in the air was in the form of S 0 2 , the rest being sulphate. The concentrations of NO2 in the rural parts of southern Finland varied between 4 and 7 pg/m3 i n winter a n d between 2 and 4 pg/m3 in summer. Ozone concentrations peaked in March-April ( a t 70 to 90 pg/m3) and were the lowest in December-January (35-50 pg/m3). In southern Finland, high concentration episodes of all the pollutants were mainly associated with south-easterly, southerly or south-westerly winds [81. The deposition of sulphur varied between 0.7 to 1.2 g/(m2a) i n southern Finland and 0.2 to 0.4 g/(m2a) in northern Finland according to measurements taken by monthly sampling of bulk deposition collectors [91. The same deposition gradient but slightly higher absolute values were estimated using atmospheric transport models [ l o ] . The impact of two large smelters in t h e Kola a r e a in Russia on t h e pollution climate was observed in north eastern Finland [lo, 201. The deposition on deciduous forests was roughly the same a s on the bulk collectors. However, on conifer stands it was up to twice a s much 1111. The deposition of Ca was substantially higher in south eastern Finland than in other parts of the country. This was largely because of the use of oil shale for energy production in Estonia about 100 kilometres south across the Baltic sea 1121, and was from 0.5 to 0.8 g/(rn2a) 191. It thereby had the potential of buffering 50 to 70 per cent of the acidifying impact of sulphur deposition, assuming no limitations in the chemical contact and reaction kinetics.

3.2. Aquatic effects A synoptic survey of lake water chemistry was carried out during t h e autumn overturn in 1987, following an objective sampling scheme [131. The measured sulphate concentration in 70 % of t h e lakes was higher t h a n 50 peq/l. However, in the northern subregion of the country, covering about one third of the land area, measured sulphate concentrations were above 50 pgA in only 22 % of the lakes [141. Many Finnish lakes a r e humic, "brown water" lakes. High concentrations of organic carbon were found in lakes throughout the country, but especially in lakes in southern and central Finland where the median total organic carbon was 14 mg/l. The organic anions, with a median concentration of 89 peq/l slightly dominated over sulphate, t h e second-most important anion. The median concentration of sulphate was 71 peq/l. I n acidic lakes which have lost their acid neutralisation capacity (ANCeO) organic anions dominated (with a median 88 peq/l) over sulphate (median 54 peqll). The contribution of nitrate in

434

acidic lakes was only 0.8 peq/l [14]. The lakes dominated by organic anions were mostly located in central Finland, on t h e west coast, and in eastern Finland. In the south the lakes were mainly dominated by sulphate and in the north by bicarbonate. The concentrations of organic anions and sulphate were slightly lower in acidic t h a n in other lakes. This indicates t h a t acidity occurs in oligotrophic lakes and is associated with the scarcity of bases rather than with an exceptionally high contribution of acids. I t was estimated that 1300 to 3100 lakes in Finland are acid because of anthropogenic sulphur deposition [ 141. Lake acidity affects biological organisms directly through the low pH and the associated increase of mobile inorganic aluminium. I t can also have a n indirect effect through responses in the food web. Biotic effects were studied in detail within HAPRO in 140 survey lakes, mostly acid lakes located in southern Finland. The species composition of macrophytes, phytoplankton, zooplankton, diatoms and benthic invertebrates were correlated with lake acidity [151. Acidification therefore affected the whole ecosystem. Fish species were ranked in terms of sensitivity t o acid conditions [16].Direct toxicity was observed, especially in the reproductive phase [171. A slow decline of fish populations rather than an abrupt fish kill was the observed response to acidification both in demographic [161 and in experimental 1171 investigations. The synoptic lake survey and diatom investigations on acidification history 1181 were used t o generalize fish investigation results. On this basis, it was estimated that 300-700 lakes have lost one or more fish populations. In addition, acidification has had a significant impact on the structure of fish populations in 800-1600 lakes, i.e. 1.4 to 2.8 % of the total of 56,000 lakes.

3.3. Forest effecte The sampling grid of forest research investigations was linked with t h a t of t h e forest inventory, a national system for monitoring forest resources. The synoptic survey grid was used t o measure soil, vegetation and tree characteristics. Experimental and modelling research was carried out in addition t o the survey approach. The growing stock and t h e growth in Finnish forests have increased substantially in recent decades [4]. Changes in the structure and age distribution of forests have been the primary reasons for this. Acidifying deposition or other forms of air pollutants have not had a significant negative effect on wood production. Early warning signals of potential future damage were investigated in terms of changes in sensitive vegetation and in soils.

A widespread decline of sensitive epiphytic lichens was observed in southern Finland, where the deposition and concentration of air pollutants is highest. Long term monitoring of canopy litter indicated t h a t there has been a gradual decline since the mid 1960s [19]. Elevated concentrations of heavy metals were observed in lichen vegetation in rural forest areas over large regions [20]. Similar patterns were observed in a Nordic study on heavy metal concentration in mosses [21]. Hypotheses of the mechanismk) of the lichen decline were developed but remained untested. Sulphur and heavy metal pollution can

435 induce toxic effects. Nitrogen loading can alter t h e species competition relationships. Despite the lack of direct experimental evidence it is apparent t h a t the change in the chemical composition of the atmosphere has been the major cause of the observed change in lichen vegetation. A link was established between diameter growth and defoliation of individual trees [22]. However, no spatial correlation was observed between defoliation and pollution load on a regional scale [231. The time series of systematic defoliation observations were too short to allow trend analysis and will remain so for a t least five more years. Hence there were no firm observations of trees or tree populations which, on a large regional scale, could be interpreted a s early warning signs of a productivity decline. In experimental research, links were established between the gas exchange of trees and the levels of nutrition and inorganic aluminium in soils [241. Scots pine trees on extreme oligotrophic, dry sites showed adverse nutrition effects in terms of discoloration and growth decline [251. Tree damage was also observed on ordinary soils in the vicinity of heavy metal [261 or ammonium [27] sources. Hence there is evidence of air pollution damage of trees on a restricted geographic scale. The most widespread effects were those caused by ammonium emissions in f u r farming regions 1271. Model calculations indicated leaching of base cation nutrients from the top soil [281, a phenomenon which was also firmly established on the basis of lake survey results 1141 and ground water observations 1291. Air pollutants thus affected tree nutrition over large regions, although marginally, since t h e r e were no major responses in terms of productivity or canopy characteristics. Unfortunately, present models of stand development and growth are insufficient for predicting possible future responses. 3.4. Other individual findings

Agricultural soils, sampled in 1974 and resampled i n 1987, were analysed for their chemical characteristics. Both soil pH and the concentrations of most macro nutrients were higher in 1987 than in 1974 because of liming and the application of fertilizers. The deposition from the atmosphere had a n effect on t h e cadmium concentration in soils [301. Anthropogenic deposition from the air was the main source of heavy metals (Cd, Hg and Pb) in t h e sediments of forest lakes [311. The waste problem of coal fired power plants was studied a s a separate issue. Desulphurization, which will be implemented on a large scale to combat acidic deposition, produces gypsum waste. The amount of wastes will increase significantly in Finland by the year 2000. The largest utilization potential is within geotechnical construction, although there a r e environmental concerns 1321. The potential impacts of forest decline on the international market of forest products was also a subject a r e a within HAPRO. Model scenarios indicated t h a t t h e market mechanisms a r e fairly robust. Only extreme decline, e.g. in central Europe, would have effects on trade flows and prices of Finnish products [33].

436

4. INTERCOMPARISONS AND CONCLUSIONS

H A P R O was a n interdisciplinary programme and yielded particularly valuable results because many research fields were involved. For example, the results of the lake survey were used in order to draw conclusions about the status of forests. The smooth and consistent gradient of sulphate concentrations in lake water followed the spatial distribution of sulphur deposition, a s estimated with the atmospheric transport model (Fig. 1).This is a n important finding for the following reasons. Firstly, it gives credibility both to the lake survey and t o the atmospheric deposition model. Secondly, it indicates the dominance of anthropogenic deposition, since there are no natural sulphur sources that could possibly generate the observed spatial pattern of sulphur concentration in lakes. Thirdly, forests are probably subject to the same spatial pattern of sulphur load because in the Finnish landscape most of the water entering the lakes has percolated through the forest. Finally, as the deposition is described with a model, it is feasible to compute deposition scenarios in order t o evaluate and assess various emission reduction measures. Sulphur deposition in 1987 g/(m' a) Sulphate in lake water

1.2

'

Figure 1. Sulphur deposition (contour lines) and lake water aulphate concentration in 1987 (dot symbols). Each dot represents five lakes.

431

The Finnish research programme was, a s far a s we know, the only national programme where a consistent spatial correlation was established between the model estimated s u l p h u r deposition and t h e spatial pattern of sulphur in ecosystems. This was feasible because of t h e clear gradient of sulphur from south t o north. Moreover, t h e smooth topography of t h e country is a n asset when constructing and testing atmospheric transport models. In addition, the uniform patterns of bedrock and land use facilitate interpretation of synoptic inventories such as the lake chemistry survey (Fig. 2 ) .

Figure 2. Forests, lakes and typical Finnish topography. Photo: Jyrki Luukkonen. The change in lichen vegetation in forests followed the same geographic pattern [19]. I t is unclear whether sulphur deposition has directly affected t h e lichen vegetation. Nitrogen loading and the atmospheric deposition of most heavy metals followed similar geographic patterns. The causal relationships, therefore, remained unclear and should be studied by using an experimental approach. However, i t is apparent t h a t , one way or the other, a i r pollutants have been the cause of a substantial change in lichen vegetation. Measurements on deposition were compared with those on surface water chemistry in order to estimate the role of the different chemical compounds in

438

surface water acidification.The contribution of nitrogen deposition was estimated to be insignificant. The impact of sulphur deposition was quantified against that of the natural sources of surface water acidity. It was established t h a t sulphur deposition affects aquatic biota in various ways especially in southern Finland. One of the specified research objectives was "to study whether the harmful effects of air pollutants are on the increase". For lakes, it was established, firstly, t h a t fish populations in acidified lakes continue t o deteriorate. Secondly, new lakes are being acidified at the current rate of (sulphur) deposition. However, the rate of increase in the number of acidified lakes was estimated a s low. Less than 100 lakes were estimated to bypass zero alkalinity over the next ten years. Mass balance calculation was the method used to estimate responses in forest soils. Comparing input (deposition) with output (leaching), it was shown that K, Ca and Mg are being depleted from forest soils [361. Nitrogen, on the other hand, accumulates in vegetation. Given the role of the different elements in limiting growth, it is possible t h a t the present deposition mix tends to increase rather than decrease forest growth. Silviculture, climatic variability, changes in the structure of forests, and other factors, however, obscured the potential general impact of air pollutants on forest growth. In average conditions in Finland, the change in plant nutrition is a threat t o forests not in the short but in the long term (> 60 years). I t was established t h a t the main body of natural resources have remained intact both in lakes and in forests. Forest growth and tree survival have not been affected. Large rivers and lakes are strongly buffered and have been only marginally acidified. It was concluded t h a t there i s no reason for concern about existing, large scale ecological damage but one should be aware of the possibility of future damage. Pollution abatement thus aims to prevent damage rather than t o restore ecosystems. In the process of programme assessment, critical loads were estimated [34]and integrated assessment models were used in order to compare options of pollution abatement [351.

Loss of genetic diversity has taken place in lakes and in forest with respect to certain species and populations. Pollution abatement cannot bring back those genotypes. This can be judged to be a serious, irreversible consequence of air pollutants. Changes in soils and in the frequency of biotic organisms are reversible and will respond to pollution abatement. Nitrogen deposition has not so far caused significant adverse effects. Because of the increasing trend of nitrogen deposition, as demonstrated in [ 101, detailed evaluation of whether the situation is changing is necessary. 5. POLICY IMPLICATIONS

5.1. Utilization of natural resources

Renewable natural resources, especially forests, are vital to the national economy. It was an important finding that trees in Finland have not shown signs of decline on a large geographical scale. On the other hand, there are signs indicating ecosystem responses in terms of changes in soils and in sen-

439

sitive vegetation. Stand rotation on certain sites in Finland can be as long a s 160 years. Observations of gradual changes in the ecosystems therefore create genuine concern about the future of the resource. Sustainable forestry has been the "trade mark" of Finnish forestry policy. The industrial products are exported mainly to central Europe and t o the United Kingdom. Consumers in those countries are environmentally aware, creating an additional incentive for the Finnish forest sector to maintain and develop acceptable forestry practices. This has brought the sector into a new situation. Sustainability, both in the narrow sense referring t o wood production and in the broad sense referring t o biodiversity, is a vital goal even from the viewpoint of direct business interests. The standing stock of Finnish forests, about 1900 million cubic metres, is equivalent to harvest for over 30 years a t the present rates. The annual growth a t present is 1.3 times higher than the harvest. In the worst case imaginable, a horror scenario, growth would be radically reduced and trees would start dying. Even then it would be logical to maintain forest industries until the standing stock has been exhausted. This kind of exploitation is business-as-usual in the utilization of oil and natural gas, but is ruled out a s an acceptable practice in forestry. Fortunately, the probability for such a scenario in Finnish forests is practically zero. Given the amount of wood t h a t already exists in forests, acidic deposition can not affect the availability of timber within the next 20 to 30 years. Concern about the environment is viewed a s a long term potential for the Finnish economy, in particular for the forest sector. The largest environmental problems, like acidic deposition and the threat of the greenhouse effect, are consequences mainly of the utilization of non-renewable resources. The forest sector draws upon a renewable resource base. Productivity can be maintained without decreasing the potential for future generations to enjoy the benefits from the resource. 62. Air pollution abatement and liming

Sulphur emissions have been successfully reduced in most European countries. If this development can continue for another 10 to 20 years, sulphur deposition will not threaten Finnish ecosystems apart from exceptional and insignificant cases. The remaining impacts can be taken care of by liming and other related measures. This optimistic scenario has been the basis for recommendations about large scale liming. To ensure the potential for future action, HAPRO urged a continuation of liming experiments. If, however, emission reduction can essentially solve the problem, there is no need for large scale liming. Liming, after all, adds to the material fluxes from the industrial system t o the environment. The guiding principle of environmental protection is t o decrease those fluxes. In addition, liming is costly and has negative side effects L371. The total annual costs of reducing sulphur emissions in the near term in Finland were estimated a t 1.3 billion FIM (about 300 million U.S. dollars). If 1.9 billion FIM are invested annually in Russia and in Estonia, and if the emissions elsewhere in Europe are reduced by 60 % from the level of 1980, sul-

440 phur deposition in Finland would be less than 0.5 g/(m2a) [7].This would solve the major part of the problems caused by sulphur deposition. It remains to be seen whether the expenditure is economically and politically acceptable in these countries. Nitrogen deposition has so far been rather insignificant in Finland in terms of acidification, aquatic effects, and adverse responses of vegetation. Nitrogen deposition is a potential future threat to ecosystems, for example to the Baltic sea.

6.3. Research management HAPRO was a unique exercise and many lessons were learned about research management. I t proved effective to have t h e programme secretariat as a n independent unit located a t t h e ministry, i.e. a t a n administrative level above research institutes and universities. This made it possible to make direct contacts to research groups without complicating interventions by the administration of different institutes. The rapid beginning of t h e programme meant t h a t there was a shortage of time for screening t h e projects. The topic was politically so interesting that postponing research was not an option. It was possible to some extent to adjust the programme during the second and the third year. In retrospect, the advantages of the quick s t a r t outweighed the disadvantages. The budget framework and the time schedule of the programme were met. I t is understandable, given the size of the programme that pressure developed towards the end of it. Voices were raised in favor of continuation. However, it was necessary to keep to the original plan. Firstly, this was the first programme of this size in the country and it was important to maintain the credibility of its management. Secondly, having a fixed time frame was i m p o r t a n t f o r creating t h e momentum: I t would otherwise have been difficult to organize joint reporting by more than a hundred authors. Thirdly, it was estimated that the best way of promoting fut u r e projects was to r u n t h e existing project according to the predetermined schedule. This strategy was partly successful, although a well-justified proposal for a continuation programme on critical loads was rejected. HAPRO gave new insight in the process of environmental research probably to all participants. My initial perception of the role of research in environment a l protection was a s depicted in Fig. 3a. New information according to this model emerges from individual research projects. Review papers organize and evaluate t h e information and serve a s t h e basis for environmental assessments. Alternative policy options are prepared. The options are internally cons i s t e n t a n d u n d e r s t a n d a b l e to policy m a k e r s a n d t o t h e g e n e r a l public. Politicians then choose between t h e options, modify them, and transform them into legistlation and statutes. This chain of activity existed in HAPRO but was not the only important channel for preparing decisions. Frequently, a case was met a s depicted in Fig. 3b. The media picks u p a n individual scientist describing individual research findings, often direct measurements with concrete interpretation. The public responds to the media report, starts t o discuss and make contacts with politicians. Ministers and other high ranking politicians react and, first, a s k t h e best experts to prepare a statement on the issue. Policy decisions are then taken to achieve an improvement.

44 1

a Primary

measurements * Report

Scientific

Policy

Political * Improvement in the environment

* review * Assessment *options * choise

b

Primary measurements

1

Assessment

4t

Press Public * Political Polrtical * Improvement release * concern reaction * decision in the environment

Figure 3. Linkage between research and improvements in t h e environment: (a) through review and assessment; (b) through media. The latter chain of events has the disadvantage of being ineffective. I t can focus on random problems a s i t circumvents t h e review and assessment processes. I t tends to omit t h e deep understanding of causal linkages. However, this "mediacratic" line of events strengthens democracy over the alternative, potentially bureaucratic procedure. Recognition of t h e mediacratic line of events (Fig. 3b) may have implications for research management. I t is important in this situation that a large number of scientists is available for making initiatives and giving statements. In this way environmental issues can be illuminated from many different angles by small, concrete pieces of information. A balance between the two approaches might serve society best. Anyway, the support of t h e general public is the only solid basis both for research and for the management and protection of environmental resources.

Reprints: Reprints of individual HAPRO publications, a s given in t h e reference list, can be ordered from the author of this chapter. 6. REFERENCES

1

2 3 4

5

J. Merilainen, Ann. Bot. Fenn. 4 (1967) 51. S. Huttunen, In M. Treshow (ed.) Air Pollution and Plant Life, Wiley, New York, 1983. H. Arovaara, P. Hari and K. Kuusela, Commun. Inst. For. Fenn. 122 (1984) 1. E. Tomppo and M. Siitonen, Paper and Timber 73 (1991) 2. Finnish Research Project on Acidification (HAPRO), General Plan, Ministry of the Environment, Ministry of Agriculture and Forestry, Helsinki 1985.

442

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 34 35 36 37

P. Kauppi, P. Anttila and K. Kenttamies (eds.), Acidification in Finland, Springer-Verlag, Berlin, Heidelberg, New York, 1990. P. Kauppi, P. Anttila, L. Karjalainen-Balk, K. Kenttamies, J. Kamari, I. Savolainen, Forsurningen i Finland, HAPROs slutrapport, (available in Swedish and in Finnish), Ministry of the Environment, Report 90,1990. S.M.Joffre, T.Laurila, H.Hakola, V.Lindfors, S.Konttinen and P. Taalas, in ref. 6, p. 43. 0.Jarvinen and T.Vanni, in ref. 6, p. 151. J.-P.Tuovinen, L.Kangas and G.Nordlund, in ref. 6, p. 167. A.Hyvarinen, in ref. 6, p. 199. P.Anttila, in ref. 6, p. 111. M.Forsius, V.Malin, I.Makinen, J.Mannio, J.Kamari, P.Kortelainen and M.Verta, Environmetrics l(1990) 73. M.Forsius, J.Kamari, P.Kortelainen, J.Mannio, M.Verta and K.Kinnunen, in ref.6, p.759; J.Kamari, M.Forsius, P.Kortelainen, J.Mannio and M.Verta, AMBIO 20 (1991) 23. L.Heitto, in ref. 6, p. 963; P.Kippo-Edlund and A.Heitto, in ref. 6, p. 973; P.Eloranta, in ref. 6, p. 985; P.Huttunen and J.Turkia, in ref. 6, p. 995; J.Sarvala and S.Halsinaho, in ref. 6, p. 1009; J.J.Merilainen and J.Hynynen, in ref. 6, p. 1029. M.Rask and P.Tuunainen, in ref. 6, p. 911. P.J.Vuorinen, M.Vuorinen a n s S.Peuranen, in ref. 6, p. 941. P.Huttunen, K.Kenttamies, A.Liehu, M.Liukkonen, T.Nuotio, 0.Sandman and J.Turkia, in ref. 6, p. 1071. M.Kuusinen, K.Mikkola and E.-L.Jukola-Sulonen, in ref. 6, p. 397. E.Kubin, in ref. 6, p. 421. A.Riihling, L.Rasmussen, K.Pilegaard, A.Makinen and E.Steinnes, NORD 1987:21. P.Nojd, in ref. 6, p. 507. E.-L.Jukola-Sulonen, K.Mikkola and M.Salemaa, in ref. 6, p. 523. H.Arovaara and H.Ilvesniemi, in ref. 6, p. 715. H.Raitio, Acta Universitatis Ouluensis, Ser. A 216 (1990) 1. K.Heliovaara and R.Vaisanen, in ref. 6, p. 447. A.Ferm, J.Hytonen, P.Lahdesmaki, P.Pietilainen and A. Patila, in ref. 6, p. 635. M.Johansson and I.Savolainen, in ref. 6, p. 253. J.Soveri and T.Ahlberg, in ref. 6, p. 865. R.Ervio, R.Makela-Kurtto and JSippola, in ref. 6, p. 217. M.Verta, J.Mannio, P.Iivonen, J.-P.Hirvi, 0.Jarvinen and S.Piepponen, in ref. 6, p. 883. J.Ranta, in ref. 6, p. 1209. H.Seppala, RSeppala and M.Kallio, in ref. 6, p. 1217. A.Henriksen, J.Kamari, M.Posch, G.Lovblad, M.Forsius and A.Wilander, NORD 1990:124. M.Johansson, J.Klmari, R.Pipatti, ISavolainen, J.-P.-Tuovinen and M.Tahtinen, in ref. 6, p. 1171. K.Kallio and L.Kauppi, in ref. 6, p. 811. J.Derome and A.Patila, in ref. 6, p. 1093.

T Schneider (Editor). Acidification Research Evaluation and Policy Applications 1992 Elsevier Science Publishers B V

443

STATUS OF ACIDIFICATION RESEARCH IN CZECHOSLOVAKIAAND ITS RELATIONSHIPTO POLITICSAND ECONOMICS IN EUROPE T.Paces Czech Geological Survey, Czechoslovakia Acidification of water and soil is related to political and economic systems. Data from 1987 (Anonymous, 1990a, 1991a) indicate that the socialist countries with centrally planned economies show large differences with the democratic countries with market economies. The socialist countries a s a group have similar values for environmental and economic parameters that are inferior to those of the other group. This is documented by statistic data on acidic emissions, consumption of energy and gross national products of selected European countries, USA, Japan and China (figures 1 and 2). Czechoslovakia, after 42 year with a planned economy, is approximately in the centre of the cluster and represents a country with very serious acidification problems in relation to its economic status. High consumption of energy per gross national product causes the country to use local soft coal with a high sulphur content. The existing power plants do not have any desulphurisation devices and emit large quantities of SO2 into the atmosphere. This is the major cause of environmental acidification (Paces, 1985). This acidification is most severe in the western part of the country - Bohemia. This part is drained by the Elbe river. In this river the acidification is partly responsible for the increased concentration of strong acid anions and a decrease in concentration of weak acid anions represented by bicarbonates (table 1).Also soils are acidified (table 1) a s well as lakes (table 2). Acidification is held responsible for the fast increase in the damage and dieback of forests (figure 3) and the changes in the rates of weathering and erosion (Paces, 1991). Acidification, in spite of all the evidence, was virtually ignored by the socialist government and environmental information was either classified o r its dissemination was suppressed. Within the Geological Survey of Prague (present title is Czech Geological Survey), it was claimed that acidification is a geochemical process and we started to investigate this phenomenon in 1978 by monitoring a small catchments in the northern Bohemia near Chomutov. This is a n area hardest hit by industrial emissions from power plants and the local chemical industry. A comparison with monitoring results of the mass balance of sulphur, nitrogen, chloride, base cations and hydrogen ions in a catchment located in less polluted region of the Bohemian - Moravian Highland near Pacov indicate how serious the biogeochemical cycles of elements are affected by acidification (Paces, 1985). Later, data from studies of other catchments and lakes were published in a volume of the International Workshop on Geochemistry and Monitoring in Representative Basins (Moldan

444

and Paces, eds. 1987) organized by the Geological Survey in Prague. After the political changes in 1989, environmental issues became priorities to be solved by the present federal and republic governments. Present research of acidification includes: (1)monitoring of input and output in 6 small representative catchments and lysimetric measurements in the soil of the severely acidified Krusne hory mts., less acidified Slavkovsky lea mete. (Kram and Hruska, 1991), and the least acidified Bohemian-Moravian Highland. A project GEOMON will start in 1992. This project, headed by the Czech Geological Survey will include monitoring of input and output of chemical elements in 14 catchments in the Czech and Slovak Federal Republic; (2) evaluation of hydrochemical and sedimentary records in acidified lakes of the Sumava mts.; (3) mapping of critical loads (Hettelingh et al., 1991) and integrated monitoring of small catchments (Anonymous, 1990b, 1991b) within the framework of the Convention on Long-Range Transboundary Air Pollution (UN-ECE). Air pollution emission and acidification by industrial sourcea is monitored and made public through daily announcement of critical concentrations of SO2 and NO, by the Czech and Slovak Hydrometeorological Institutes. Ecological impacts of acidifying emissions are studied by the Czechoslovak Academy of Sciences and the Department of Forestry at the T.G.Masaryk University a t Brno. Local forest authorities are responsible for reforestation of damaged areas after the dieback of the spruce. Czechoslovakia will not be able to reduce acidification rapidly because it will require fundamental changes i n energy consumption, installation of expensive desulphurisation equipment and modernisation of the chemical industry. These changes will come with the privatisation of state-owned industries and with the political changes that take place in central and eastern Europe.

445

Table 1 Changes in the acidification status of Elbe river and the soil of the Orlicke hory mountains, data by Paces, 1982 and Pelisek, 1984

Elberiver

1982

1976 mmo1.m-?yr-1

36

NOSHCOr

170 40 252

5.6

330 Czech soil

1953

1981 pH H20

A0 Hh

B

3.6 3.8 4.9

4.2 4.7 5.2

Table 2 Acidification of Cerne (Black) lake in Sumava mountains in Czechoslovakia, data by Vesely, 1987

Depth of lake m

Acidity* p o l H+.rl

1936 0 5

50-70

15

90-135

25

30

160 - 180

m 215 210 1% #I5

215

* titration with NaOH with phenophtalein indicator

446

40

-

DDR

30-

P'

a

5

R'

&

g

20-

g 10 -

0

USSR' ' 0 DDR BRD s 0 0 8 j

A.

'USA us$/c

441 60-

-

f

-o-

Czech republic

-m-

Slovak republic

40-

1

-5

10

-

0 1960

1970

1980

1990

time in years

Figure 3. Development of forest damage in Czechoslovakia, data from "blue book" of Czech Ministry of Environment, 1990 REFERENCES

Anonymous (1990a1, World Resources 1990 - 1991, Oxford University Press, Oxford Anonymous (1990b1, Pilot Programme on Integrated Monitoring, 1 Annual Synoptic Report 1990, Environmental Data Centre, National Board of Waters and the Environment, Helsinki, 88 p. Anonymous (1991a1, Data by PlanEcon, USA, published in Lidove Noviny, March 14,1991, Praha Anonymous (1991b), Pilot Programme on Integrated Monitoring, 2 Annual Synoptic Report 1991, Environmental Data Centre, National Board of Waters and the Environment, Helsinki, 200 p. Hettelingh J.P., Downing R.J. and De Smet P.A.M. (eds.1 (19911, Mapping Critical Loads for Europe, CEE Technical Report no. 1, RIVM report no. 25910001, National Institute of Public Health and Environmental Protection, Bilthoven, 79 p. Kram P. and Hruska J. (19911, Hydrogeochemical balance of acidic catchment Lysina (Slavkovky lea mnts.1 with extremely high content of aluminium in discharge, Casopis pro mineralogii a geologii, v. 36, No. 4, Prague Moldan B. and Paces T. (eds.1 (19871, Extended Abstracts, GEOMON, International Workshop on Geochemistry and Monitoring in Representative Basins, Geological Survey, Prague Paces T. (19821, Natural and Anthropogenic Flux of Major Elements from Central Europe, Ambio, vol. 11, pp. 206 - 208

448

Paces T. (19851,Sources of Acidification in Central Europe Estimated from Elemental Budgets in Small Basins, Nature, vol. 315,pp. 31 - 36 Paces T. (19911,Changes in rates of weathering and erosion induced by acid emissions and agriculture in central Europe, In: Land Use Changes in Europe, F.M.Brouwer et al. (eds.1, Chapter 14,pp. 317 - 323, Kluwer Academic Publishers Paces T. and Pistora Z. (1979),Antropogenni Ovlivneni Chemickeho Slozeni Labske Vody, Vodni hospodarstvi, B., vol. 11,pp. 305 - 307, Praha Pelisek J. (19841,Changes in Acidity of Forest Soils of the Orlicke Mts.caused by Acid Rains, Lesnictvi vol. 30,pp. 955 - 962,Praha Vesely J. (19871,The development of acidification of lakes in Bohemia, In: GEOMON, Extended Abstracts, pp. 80 - 82,Geological Survey, Prague

T Schneider (Editor). Acidification Research Evaluation and Policy Applications All rights reserved

0 1992 Elsevier Science Publishers B V

449

S w i s s National Research Program "Forest Damage and A i r Pollution" (NFP 14+)

The

Frank Haemmerlil, Norbert Krauchil, Martin Stark2 1 Swiss Federal Institute for Forest, Snow and Landscape Research, CH-8903 Birmensdorf.

Switzerland 2 Program Management NFP14+, Sigmaplan, Zihringerstrasse 61. CH-3012 Bern. Switzerland

Abstract The ohjective of this paper is to present a review of the NFP 14+. The studies of this multidisciplinary research program were carried out hetween 1985 and 1989. They were concentrated on three forest sites at different altitudes, one in the densely populated Mittelland, one in a protection forest in the Prealps, and one in the Alps proper. All sites are dominated hy Norway spruce (Picea ahies). The main goal of the case studies was a comprehensive characterization of the sites, giving the opportunity to evaluate the role of air pollution in forest health. The program included various ecological investigations as well as measurements of meteorological parameters, gaseous pollutants and atmospheric deposition. The results indicate neither a temporal nor a spatial dependence of crown defoliation on air pollution for the study sites. Nevertheless, there is experimental evidence that the presentday ozone levels in Switzerland have to he considered as a risk factor for more sensitive tree species.

1 MAIN OBJECTIVE AND CONTENTS OF THE RESEARCH PROGRAM ~

In 1980 the Swiss National Science Foundation got the order from the Executive Federal Council to conduct a National Research Program about 'Air Cycle and Air Pollution in Switzerland' (NFP 14). Because of the general impression that the crown defoliation of trees had increased hetween 1982 and 1984, the program was extended hy a supplementary program called 'Forest Damage and Air Pollution' (NFP 1 4 t ) . The projects of this multdisciplinary research program were conducted hetween 1985 and 1989. They were concentrated on one case study in the densely populated Mittelland, and two in protection forests, one in the Prealps, one in the Alps proper (Figure 1). The main goal of the program was a comprehensive characterization of these three sites, giving the opportunity to evaluate the role of air pollution in forest health. More than 20 different scientific groups from several research institutes were integrated in the research program, which included various forest ecological studies. Four towers were installed at different altitudes on the Laegeren ridge, another tower was erected in the Alptal, and another in Davos to measure the air chemistry and meteorology of the sites. Table 1 gives a general account of the focal points of this research program. The program was completed this year with the publication of six partial synthesis reports - three of them concerning forest relationships. A comprehensive report will he puhlished in 1992.

450

Figure 1 The geographical situation of the study sites

Table 1 Focal points of the NFP 14 + tree and stand vituliry

- crown condition (terrestrial and aerial inventory)

- growth - nutrient supply of spruce needles

- mycology of spruce needles

meteorology and uir chemistry - climate and weather conditions

- air pollution situation (gaseous pollutants, atmospheric deposition above and below the canopy of spruce and beech, stemflow of heech) - fog chemistry

physiology und biochemistry of trees under the influence of guseous pollutants

- photosynthesis and stomata hehaviour of spruce needles

- condition of the wax layer of spruce needles - fumigation experiments in the laboratory and in the stands

soil ecology - water balance - nutrient and heavy metal contents

- condition of root mycorrhizas

- litter input and litter decomposition

45 1

2. ECOLOGICAL CHARACTERISTICS OF THE STUDY SITES

Topographic, climatic and soil specific requirements for plant growth are extraordinarily varied in a mountainous country like Switzerland. Considering these circumstances the three study sites were chosen in three distinct regions differing greatly in altitude. Since Norway spruce is the most widely spread species in Swiss forest, the main analyses were made for spruce. The study site in the Mittelland is situated at 685 m a.s.1. on the southern slope of the Laegeren; it belongs ecologically to the heech forest area (Galio odoratio-Fagetum typicum & Pulmonario-Fagetum typicum). Today, it is a mixed spruce-fir stand with beech. The dominant spruce is between 100 and 150 years old. The geological subsoil is either moraine and sandy molass or, rarely, clay-rich molass marl. The acidity of the surface layer varies from acid over moderately acid to slightly alkaline since different soil types occur: cambisol. luvisol and pararendzina (Tahle 2). The soil is rich in clay and silt. The study site in the Prealps is situated at 1185 m a.s.1. on the western slope of the Alptal. The spruce-fir forest is based on natural regeneration (ecologically: Veronico urticifoliae-Piceetum to Sphagno-Piceetum typicum). It mainly consists of 120- to 150- year-old spruces. The soil is rich in clay, often saturated and partially moving. The surface layer of this gleysol is of moderate to high acidity. Finally, the study site in the Central Alps is situated at 1660 m a.s.1. in the high valley of Davos. A characteristic of the natural spruce forest (Larici-Piceetum)is the wide age range of the trees which varies between 120 and 370 years. An acid iron-humus-podsol has developed on the gneiss crystalline subsoil. Climatologically the sites Laegeren and Alptal are intluenced hy a maritime climate. Davos, on the other hand, is already continentally intluenced resulting in a rougher climate and a high radiation intensity (Table 2). Typical for the Alptal is a high annual level of precipitation due to blocked air masses on the northern side of the Alps. Special, climatically induced stress situations occur from time to time in Davos and in the Alptal due to the Fiihn, an alpine wind which causes abrupt air temperature changes. In relation to the air quality situation none of the study sites was close to major emission source. The sites may therefore be considered as representative for pollution situations covering large areas. Table 2 Climatic and soil-specific characteristics of study sites (from Stark et al. 199I , Liischer 1991) Characteristics (1987188189) temperature [“C] precipitation [mml

Laegeren

Alptal

Davos

7.2 18.3 18.5

5.7 16.7 17.1

2.9 13.8 14.2

1084 I1017 I866

2600 12300 I2150

986 I809 I671

gleysols

humus form

cambisol, luvisol pararendzina mull

iron-humuspodsol raw-humus

rooting depth permeability acidity (surface layer) nutrient supply

60 - 90 cm slighlty reduced pH 4.5-7.5 normal

SQU

soil type

raw-humus, anmoor 30 60cm poorly permeable pH 3.6-5.4 normal

-

-

10 30 em

excessive pH 3.5-4.5 normal

452

3. VITALITY OF TREES ON THE STUDY SITES The condition of the tree crowns on the three sites changed only minimally between

1986 and 1988. This was demonstrated by a low needle loss level on the Laegeren, and

essentially higher needle loss values in the Alptal and in Davos. Figure 2 shows further that the extent of crown defoliation on the sites Alptal and Davos is distinctly higher than the comparative value of the large scale forest damage inventory for the regions Prealps and Alps. According to the internationally approved definition, trees with more than 25 per cent foliage loss are considered to be damaged. About the half of the trees in the stand Alptal and about one third in the stand Davos fell into this category. Therefore, it might be expected that this high needle loss level would be reflected in the growth behaviour of spruce. As Figure 3 shows, this is not the case. The average development of radial growth does not exhibit an unusual decrease in the eighties on both sites. It is therefore difficult to denote the observed level of spruce defoliation as new and abnormal. On the other hand, we notice on the Laegeren a growth depression of spruce since the sixties, even though the actual needle loss level is relatively small. These apparent contradictions show that a clear picture of stand vitality can only be obtained through the combined observation of crown defoliation and growth. Despite all its drawbacks, crown defoliation still remains a measure of vitality; significant correlations have been found between needle loss and radial growth of spruce (Keller L Stark 1991). Table 3 shows that the nutrient supply of young spruce needles on all sites varies between sufficient and optimal with a few exceptions. Using the threshold values of Bergmann (1988) and Anonymus (1987) no distinct deficiencies have been noted. The spruce collectives on all sites did not show any relation between the observed crown transparency and the nutrient supply of needles. Taking the growth depression of spruce on the Laegeren into consideration, it may be that on1 these trees are in a critical situation. Growth depression may be related to the barely suficient content of magnesium in the needles. On acid soils in the Fichtelgebirge (Germany) and in the Vosges, needle yellowing has been detected as symptom of magnesium deficiency. No such yellowing was observed on the Laegeren site. As the soil here is well supplied with magnesium, there is the hypothesis that root uptake is hindered by the high level of calcium.

4. GASEOUS POLLUTANTS AND THEIR INFLUENCE ON TREE CROWNS

In order to evaluate the air quality situation on a specific site the Swiss Clean Air Act (LRV 1985) provides threshold values for air pollutants. These lawful limits, which should guarantee an overall protection for man, plants, animals and environment, are based upon a great number of scientific studies and mostly concur with the recommendations of recognized technical organizations (WHO,VDI, UN-ECE). As shown in Table 4, the levels on the sites due to primary pollutants such as sulfur dioxide and nitrogen dioxide are usually small, especially at Alptal and Davos, where the daily average limits of the Swiss Clean Air Act were never exceeded. On the other hand the secondary pollutant ozone has to be considered as a serious stress factor. It is striking to see how many times the highest allowed hourly average of 120 pg/m3 ozone was exceeded each year on the Laegeren and in the Alptal.

453 Percentage o f trees with defoliat. 1 2 5 % 100

Lageren-Midland

Alptal-Prealps

Davos-Alps

80 60

858687888990

858687888990

=

NFP14*

858687888990

Region

Figure 2 Crown defoliation in the stands of NFP 14+ and in the comparative regions ofthe Swiss forest damage inventory (ohserved trees: Laegeren 214, Alptal 240, Davos 519) (from Keller &stark 1991)

4

rniII Imet re I

3

2

A

L D

1

I

Davos (D) 01

t

1 - L L I . J

---I&

IIU

LLILUJ,

1

LA

+

,LL+-LLdLl

1890 1900 1910 1920 1930 1940 1950 1960 1970 1980 1990

Figure 3 Average development of radial growth of dominant and co-dominant spruces hetween 1890 and 1989 (Observed trees: Laegeren 16, Alptal 20, Davos 17). These collectives are slightly more defoliated than the whole stands (from Joos 1991).

454

Table 3 Nutrient supply in 6 month old needles of older spruces (from Stark 1991)

Element N

Laegeren

Alptal

Davos

+

0

0-

P

t

0 -

t+

Mg Ca K Zn B Fe

0

+ ++ ++

++

t+

++ ++ 0-

+t

0

t+

++ ++

++ +

++

0

Table 4 Annual mean concentrations (AM, pg/m3) of S 0 2 , NO2 and 0 and number of exceedences (NE) of the short-time threshold values of the Swiss Clean Air Act ( V 1985) [from Star in pressl. Maximum I-day average concentration for SO : 100 pg/m , tor N02: 80 pg/m , maximum I-hour average for 0 3 : 120 pg/m3. The Act &Crees that these limits may only be exceeded once a year.

Y

5

455

The influence of gaseous pollutants on tree crowns was examined in three projects. The first project dealt with the question of whether significant alteration in gas exchange of spruce needles under the influence of ambient air pollutants could be measured (Hlsler 1991). To this end, the gas exchange of the two youngest needle age-groups of older spruces was measured at 5-minute intervals for 3 years on the southern slope of the Laegeren and at Davos. Measurements were conducted with branch chambers made by Walz (Germany). Considering the available results, the hypothesis of a direct and immediate influence of air pollutants on the gas exchange of needles cannot be completely excluded, but it is certain that meteorological conditions were the determining factors for net photosynthesis and stomatal behaviour of young needles during the study period. It was further established that photosynthetic activity, as measured some 30 years ago, does not differ significantly from toda 's measurements. Nothing can be said about the question as to whether photosynthesis in order needles, which have been exposed for several years to today's air quality conditions, has k n chronically affected. If so, it would not be too detrimental for the photosynthetic performance of the whole tree, since 80 per cent of the photosynthesis occurs in the two youngest age-groups of needles. The wax layer of needles builds a barrier between the plant and the atmosphere. The goal of a second project was to figure out whether the wax layer changes under the influence of air pollution (Giinthard-Goerg 1991). To this end, 1-, 3- and 5- year old needles of 6 mature spruces from Laegeren and Davos, as well as 6-month-old needles of 4-year-old spruce seedlings were used for the analyses. The seedlings were exposed to different doses of ozone, sulfur dioxide and ambient air. The results show that exposure to ambient air did not lead to any deterioration in wax structure. The condition of the wax layer was always mainly determined by the expos' 'on to weather conditions. Only an experimental exposure to a high level of ozone (300pg/m ) led to retardation in the development of the wax layer. The third project dealt with the question as to what extent ozone concentrations such as occur in Switzerland may negatively effect forest plants (Landolt & Luthy-Krause 1991). To determine this, various I- to 5-year-old forest plants were fumigated in the closed-top chambers of the Swiss Federal Institute for Forest, Snow and Laryiscape Research in Birmensdorf for a maximum of 20 weeks with 0, 100 and 200 pglm ozone. Supplementary fumigations with ambient air were conducted in open-top chambers on the Laegeren and in Davos for a maximum of 2 years. As a measure for the evidence of pollutant effects either visible symptoms or biochemical parameters in needles/leaves were used. As shown in Table 5 , Scots pine (Pinus silvestris) was most sensitive to ozone in the classic gas exposure experiment. Needle yellowing occurred after only three weeks of 100 pg/m3 ozone exposure. Within the concentration range of 200 pg/m3 ozone European beech (Fagus silvatica) reacted after 6 weeks with brown necrotic spots on leaves. In the same concentration range Norway spruce (Picea abies) showed visible symptoms too, but much later or even several weeks after the fumigation had ceased. The symptoms were also hardly or not at all reproducible for spruce. Finally, Silver fir (Abies aha) never reacted with visible symptoms. Regarding the biochemical changes induced by ozone, pine was established as the most sensitive species of those tested. During fumigation with ambient air on the Laegeren and in Davos, only the well known bio-indicators red clover (Trifolium pratense var.'lucrum') and a poplar hybrid (Populus euramericana var.'Dorskamp') exhibited any symptoms. In the study on pine and spruce in ambient air compared with controls in filtered air, no obvious reactions were observed. In conclusion, these experiments demonstrate the sensitivity of various plants to ozone exposure is very different.

P

456

experimental fumigation with ozone

100 d m 3 Pinus silvestris Fagus silvatica Picea abies Abies alba

I

zoo pg/m3

BC

VS

+

+

+

-

.

-

(-)

VS

-

I

BC

experimentalfumigation with ambient air Lawren

VS

+

-

+

+

nl

-

+

-

I

BC

-

Davos

VS

I

BC

nt

nt

nt

nt

nt

-

?

.

?

nt

nt

nt

nt

5. ATMOSPHERIC DEPOSITION

Results about atmospheric deposition are available for the wet deposition above and below the canopy of spruce and beech (Table 6). Figure 4 shows the annual load of protons, nitrate and ammonium above and below spruce canopy for all three observation sites. The load of protons above the canopy of the Laegeren and Alptal is comparable to the average value measured in Germany (032 kg/ha*year, Fuhrer et al. 1988). At the alpine site Davos it falls well below that value. For the Laegeren and Alptal the load below spruce is higher than that in the open field. In the beech crown cover instead it seems that protons are buffered, since the acid deposition onto the soil below beech is very small (Table 6). The wet deposition of nitrate and ammonium in the open field is for the two sites Laegeren and Alptal on a similar level to the mean values found in Germany (6,7 kg nitrate/ha* ear; 9.0 kg ammonium/ha*year ; Fuhrer et al. 1988). On the alpine site Davos it falls signif;Ycantlybelow these values as already for the proton load. Figure 4 shows further that the nitrogen deposition below spruce is notably high on the Laegeren. For this site it is clearly shown that measuring precipitation in the open field comprises only a part of nitrogen deposition onto the soil in spruce stands. An essential part is probably due to dry deposition of nitrogen compounds to the tree crowns during p o d s without rain. Regarding the role of the observed deposition for the investigated sites, there is currently no evidence for existing problems for tree vitality. It does not Seem that the nutrient supply is limited on any site by the proton load. The two sites Laegeren and Alptal, which are more affected by acid deposition, show a sufficient supply of calcium, magnesium, potassium and manganese in the soil solution. A critical situation for the fine roots cannot be deduced from the calcium/aluminium-ratio in the soil solution. Regarding the actual load of atmospheric nitrogen on the Laegeren and in the Alptal, neither site exhibits signs of a fertilization overdose. The nitrogen content in spruce needles is on all sites below the optimal range. The atmospheric deposition of such metals as lead, cadmium and zinc seems to he harmless for all three observation sites. There is no evidence of disturbance of litter decomposition or of the release of nutrients in the soils. Regarding the Swiss Clean Air Act (LRV 1985), the deposition of these metals measured by Bergerhoff method does not exceed the threshold values.

451

Table 6 Wet deposition of selected chemical constituents (from Klijti et al. 1991, Keller & Klijti 1988).

nitrate

ammonium

20 15 10 5

0

A d -

LA AL

DA

~~

LA AL

above canopy

DA

L A AL

DA

below canopy

Figure 4 Wet deposition above and below the canopy of spruce (observation period: 1986/87; LA: Laegeren, AL: Alptal, DA: Davos) [from Kloti 1991, Keller & Kloti 19881.

45 8

6. SUMMARY AND CONCLUSIONS The National Research Program NFP14+ was based on three case studies in different parts of Switzerland. The observation period was rather short. Therefore, the results cannot be unreservedly generalized. Nevertheless, some valuable conclusions can be drawn. Assessment of crown condition alone is insufficient for the estimation of stand vitality. The condition of the mountain forest in Davos and in the Alptal is considered normal based on several indices (e.g. growth. nutrient supply), even though the actual needle loss level is high. About a third to a half of the trees on these sites have more than 25 per cent crown defoliation. The internationally used damage limit would consider all these trees as damaged. This general damage definition for all ecological zones of Europe and for all tree species should therefore be critically examined. Nevertheless, the crown defoliation remains a measure of vitality. The investigations indicate neither a temporal nor a spatial dependence of crown condition on air pollution. It can be concluded for the study sites that air pollutants have hardly affected the crown condition of spruce within the observation period. In a spatial point of view, the pollution on the alpine site Davos is generally low, while the needle loss level is high. On the other hand the pollution on the Laegeren, in the densely populated lowland, is relatively high whereas the extent of crown transparency is small. On all sites the actual levels of gaseous pollutants such as sulfur dioxide and nitrogen dioxide as well as the loads of metals such as lead, cadmium and zinc do not exceed the lawful threshold values in the Swiss Clean Air Act (LRV 1985). This is, however, not the case for ozone. There is experimental evidence that the current ozone levels have to be considered as a risk factor for more sensitive tree species. The margin existing between the present-day ozone concentrations and those which have been experimentelly shown to produce damage to Scots Pine is on a toxicological scale small. On the other hand, the assumption that ozone is damaging to Norway spruce and Silver fir should be critically questioned, as sensitivity to ozone varies greatly between species. Acid and nitrogen deposition onto forest soils is distinctly higher on the Laegeren and in the Alptal than on the alpine site in Davos. Nevertheless there is currently no evidence of an existing problem for the tree vitality due to this deposition on all sites. The long term risk of this factor for the different forest types in Switzerland is, however, unknown. Another long-term risk can be the potential climate change by the emission of greenhouse gases. Even if, or perhaps just because, these potential risks due to our modern civilisation are so difticult to assess, it is more than sensible to pursue efforts to reduce air pollution.

7. BIBLIOGRAPHY

1 Anonymus: Grundsatze f i r die Dungung im Wald. Bayerisches Staatsministerium fiir Erntihrung, Landwirtschaft und Forsten, Munchen (1987) 2 BergmaM, W. : Erniihrungsstorungen bei Kulturpflanzen. 2. Auflage. Fischer Verlag, Stuttgart (1988) 3 Fuhrer, H.W.; Brechtel, H.M.; Ernstberger, H.; Erpenheck, C.: Ergebnisse von neuen Depositionsmessungen in der Bundesrepublik Deutschland und im benachbarten Ausland. BOM, Mitt.Dt.Verb.Wasserwirt und Kulturbau, 14 (1988) 4 Gunthard-Goerg, M.: Die Einwirkung von Luftschadstoffen und Klimafaktoren auf die Wachschicht von Fichtennadeln. In: Stark, M. (4s.): Luftschadstoffe und Wald (1991) 5 Haemmerli, F.; Schlaepfer, R.: Forest Decline in Switzerland. In: Huettl, R.F.: Forest Decline in Atlantic and Pacific Region, Springer Verlag. Berlin, New York (paper submitted: June 199 I 1.

459 6 Hasler, R.: Vergleich der Gaswechselmessungen der drei Jahre (Juli 1986 - Juni 1989). In: Stark, M. (eds.): Luftschadstoffe und Wald (1991) 7 Joos, K. : Jahrringanalysen auf den Beoabchtungsflachen Davos, Alptal und LIgeren. In: Stark, M.(eds.): Luftschadstoffe und Wald (1991). 8 Jutzi, W. (eds.): Luftschadstoffe und ihre Erfassung. Ergehnisse aus dem Nationalen Forschungsprogramm 14 "Lufthaushalt, Luftverschmutzung und Waldschiden. Teil 1 . Verlag der Fachvereine an den schweizerischen Hochschulen und Techniken, Zurich ( 1 991). 9 Keller, H. M.,Kliiti, P.: Teilprojekt Bestandesniederschlag, Schlusshericht zu Handen der Programmleitung NFP14+. Eidg. Forschungsanstalt fLir Wald, Schnee und Landschaft, Birmensdorf (1988). 10 Keller, W.;Stark, M.:Wachstum und Kronenverlichtung auf den Beohachtungsflachen Ligeren, Alpthal und Davos. In: Stark, M . (eds.): Luftschadstoffe und Wald (1991). 1 I KWi, P.; Gehrig, R.; Portmann, W.: Depositionen. In: Schuphach, E. (eds.): Meteorologie und Luftchemie in Waldhestinden ( I 99 I ) 12 Landolt, W. ; Luthy-Krause, B. : Wirkungen umweltrelevanter Ozon-Konzentrationen auf verschiedene Pflanzen. In: Stark, M. (eds.): Luftschadstoffe und Wald (1991) 13 LRV: Luftreinhalte-Verordnung des Schweizerischen Bundesrates, EDMZ, Bern (1 985). 14 Luscher, P. : Gesamtschweizerische Einordnung der drei Beohachtungsflachen aus physiographisch-hodenkundlicher Sicht. In: Pdnkow, W. (eds.): Belastung von Waldhoden (1991). 15 Pankow, W. (eds.): Belastung von Waldhiiden. Ergehnisse aus dem Nationalen Forschungsprogramrn 14 "Lufthaushalt, Luftverschmutzung und Waldschiden, Teil 6 . Verlag der Fachvereine an den schweizerischen Hochschulen und Techniken, Zurich (1991). 16 Schlaepfer, R.; Haemmerli, F.: Das "Waldsterhen" in der Schweiz aus heutiger Sicht. Schweiz. Z. Forstwes. 141 (3): 163-188 (1990). 17 Schupbach, E. (eds.): Meteorologie und Luftchemie in Waldbestinden. Ergehnisse aus dem Nationalen Forschungsprogramm 14 "Lufthaushalt, Luftverschmutzung und Waldschaden, Teil4. Verlag der Fachvereine an den schweizerischen Hochschulen und Techniken, Zurich (1991). 18 Schiiphach, E.; Wanner, H. ( 4 s ) : Luftschadstoffe und Lufthaushalt in der Schweiz. Ergehnisse aus dem Nationalen Forschungsprogramm 14 "Lufthaushalt. Luftverschmutzung und Waldschaden, Teil 2. Verlag der Fachvereine an den schweizerischen Hochschulen und Techniken, Zurich (1991). 19 Stark, M. (eds.): Luftschadstoffe und Wald. Ergehnisse aus dem Nationalen Forschungsprogramm 14 "Lufthaushalt, Luftverschmutzung und Waldschaden, Teil 5. Verlag der Fachvereine an den schweizerischen Hochschulen und Techniken, Zurich (1991). 20 Stark, M.; Primault, B.; Schupbach, E.: Die Beohachtungsflachen an der Ggeren, im AIptal und Davos. In: Stark, M. (eds.): Luftschadstoffe und Wald (1991)

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T. Schneider (Editor), Acidification Research. Evaluation and Policy Applications 1992 Elsevier Science Publishers E.V

463

A comparison of some national assessments J a n Nilssona and Ellis Cowlingb aswedish State Power Board, S-162 87 Vallingby, Sweden bNorth Carolina State University, Raleigh, North Carolina 27695 USA

Abstract During the past two years, several countries in Europe and North America completed major scientific research and assessment programs on acidification and related air-pollution problems. Most of these programs culminated in publication of detailed documents which summarize the results obtained in each national program. In this paper we have tried to compare these national documents for similarities and differences in six specific features: 1) natural, cultural, and economic resources a t risk; 2) pollutant emissions of concern; 3) research and assessment approaches used; 4) scientific findings obtained; 5) policy options considered; and 6) use of research and assessment findings in making environmental decisions within each country. Based on these comparisons, we have drawn a few generalizations about the impacts of acidification and air pollution on soils, crops, forests, surface waters, fish, wildlife, engineering materials, cultural resources, public health, and visibility. We have also described a few lessons learned in various countries about the interface between science and environmental decision making. 1. INTRODUCTION This paper was prepared as a summary contribution for a n International Conference entitled Acidification Research: Evaluation and Policy Applications. This Conference was held 14-17 October 1991 a t the Maastricht Exposition and Congress Center (MECC) in The Netherlands. It was sponsored by The Netherlands' Ministry of Housing, Physical Planning and Environment and its National Institute of Public Health and Environmental Protection. The Conference was organized in four sections: 1) New Research Results on the Acidification Problem; 2) Results from National Research Programmes; 3) State-of-the-Art of Acidification Research; and 4) Acidification Policy. This summary paper was designed t o compare and contrast several of the recently completed national research and assessment documents with regard to six specific features: 1)natural, cultural, and economic resources a t risk; 2) pollutant emissions of concern; 3) research and assessment approaches used;

464 4) scientific findings obtained; 5 ) policy options considered; and 6) use of research and assessment findings in making environmental management decisions within each country.

Based on these comparisons, we also have tried to draw a few generalizations about both the scientific facts and also about public perceptions in various countries about impacts of acidification and air pollution on soils, crops, forests, surface waters, fish, wildlife, engineering materials, cultural resources, public health, and visibility. Finally, we have described a few of the lessons learned in various countries about the interface between science and environmental decision making. The national documents used a s the primary focus for comparison in the present paper include the following: The Netherlands -- Acidification Research i n The Netherlands (Heij and Schneider, 1991); Sweden -- Air Pollution '90 -- Action Programme for Air Pollution and Acidification ( S N V , 1990) and Mercury i n the Environment: Problems and Remedial Measures ( S N V , 1991); Finland -- Acidification in Finland (Kauppi, Antilla, and Kenttamies, 1990); United Kingdom -- Review Group Reports on: Effects of Acid Deposition on the Terrestrial Environment, (DOE, 1988); Effects of Acid Deposition o n Buildings and Building Materials (DOE, 1989a); Acidity in United Kingdom Fresh Waters (DOE, 1989b);Acid Deposition in the UK 1986-1988 (DOE, 1990a); Oxides of Nitrogen in the UK (DOE, 1990b); Acidity in Scottish Rivers. A Chemical and Biological Survey (Doughty, 1990); Critical and Target Loads Maps for the United Kingdom (DOE, 1991); Canada -- The 1990 Canadian Long-Range Transport of Air Pollutants and Acid Deposition Assessment Report including its Executive Summary and its separate components on: Emissions and Controls, A t m o s p h e r i c Sciences, Aquatic Effects, Terrestrial Effects, Human Health Effects, Socio-Economic Effects, and Quality Assurance Studies (FPRMCC, 1990); and United States -- State of Science and State of Technology Reports including four volumes: Emissions, Atmospheric Processes and Deposition; A q u a t i c Processes and Effects; Terrestrial, Materials, Health, and Visibility Effects; and Control Technologies, Future Emissions, and Effects Valuation (NAPAP 1991a). The national documents for the USA also include: the 1990 NAPAP Integrated Assessment Report (NAPAP,199lb); and a report by a n Oversight Review Board entitled The Experience and Legacy of NAPAP (NAPAP, 1991~). As a stimulus for further thought, we also have made some comparisons between the major scientific and assessment findings in these recent national

465

documents with: 1) some related observations in one east European country -Czechoslovakia (Thomas Paces, Personal Communications); and 2) the conclusions drawn more than ten years ago in the final report of the Norwegian Joint Research Project (SNSF project 1972-1980) -- A c i d Precipitation -- Effects on Forest and Fish (Overrein et al, 1981). Readers with an interest in other aspects of the history of acid deposition research may wish to examine an earlier review paper by Cowling (1982). In this paper, we have used the term national document to include both research documents (reports dealing with scientific findings) and assessment d o c u m e n t s (reports dealing with policy options o r management recommendations). This was done whether o r not these different types of reports were published separately (as in the cases of the United Kingdom and the United States) o r published in the same document (as in the cases of the national documents for The Netherlands, Sweden, Finland, Canada, and Norway). To save space in the paper we have often used the following abbreviations: NL = The Netherlands, S = Sweden, SF = Finland, UK = United Kingdom, CS = Czechoslovakia, CAN = Canada, and USA = United States. In comparing the several national documents we have used two general approaches -- a rough numerical approach, and a more detailed narrative approach. The rough numerical comparisons take the form of tables in which we have used numbers to indicate our judgement of relative importance within a given national document: 3 = major matter of concern; 2 = moderate matter of concern; 1 = minor matter of concern; 0 = matter not dealt with so far as we could tell. The narrative comparisons are simply that -- a few sentences (sometimes quotations from the documents themselves) describing our judgement about the features of the national documents being compared. Before beginning the comparisons, a few acknowledgements and disclaimers may be in order: We are grateful for advice and counsel received from many colleagues prior to, during, and following the Conference in Maastricht. We have tried our best to study the national documents thoroughly and objectively. In spite of our best efforts, however, we are sure t o have overlooked o r misunderstood some important aspects of these sometimes very long documents. For this reason, we look forward eagerly to both personal communications and published reactions to our efforts from colleagues in all countries who are willing to help improve our collective understanding of these important environmental problems.

466 2. COMPARISON

OF NATIONAL RESEARCH AND ASSESSMENT

DOCUMENTS 2.1. Who Prepared the Documents? What Methods Were Used?

Table 1 contains a very brief overview of the duration of research efforts, management approaches, and types of documentation developed in each national program. In most cases, some type of international peer review was used in addition to the within-country efforts to insure that high standards of scientific quality and integrity were maintained in developing the national documents.

Discussion In The Netherlands and Sweden, the assessment documents were prepared by a relatively few persons within o r with close connections to the federal government. Scientists who played important roles in writing research documents also played important roles as advisors in evaluating alternative policy options and occasionally in formulating national environmental goals. In the USA, a very large number of scientists and policy analysts from different disciplines and geographically distant regions were involved. This was also but somewhat less true in Canada, where most assessment team members were from Environment Canada or other federal and provincial bodies. These differences seem to reflect various "cultures": In The Netherlands and Sweden, a high degree of consensus existed between the government and the scientific community. Thus scientists had a relatively strong influence on decision making because the process called for policy analysts and scientists to work closely together. In the USA, greater separation was maintained between the persons involved in the research program and those involved in developing the assessment document. Also, many different federal agencies were involved. This sometimes led to compromises in interpreting scientific results. To some extent these differences in ways of conducting national research and assessment programs can be explained by differences in starting points. In most of Europe and Canada, there was a higher degree of public and scientific consensus about the seriousness of impacts and the need for international cooperation in limiting emissions than was true in the UK and the USA. The smaller scientific and policy-analysis communities in The Netherlands, Sweden, Finland, Norway, and Canada also made i t more necessary for scientists involved in research also to be involved in policy analysis.

Table 1 Management of the assessments in various countries

The Netherlands Duration of Research Program Management of research and assessment program

EuroDe Sweden

Finland

United Kingdom

First phase (198588) Second phase (1988-90) Small assessment team worked closely with research scientists

Three phases (1976-88) Fourth phase (1988-93) Main work done by staff a t Swedish EPA & collaborating scien-

Book. Part 1: Results and conclusions relevant for policymaking.

Three basic Book on primary reports on: scientific Consequences findings in of deposiEnglish. tion of sulphur and Assessment nitrogen, Air pollureport in Finnish and tion in Swedish urban areas, Mercury problems and environmental goals

(1985-90) Small assessment team worked closely with research scientists

Part 2: Thematic reports with summary of scientific results.

North America United States

(198590) Scientific review groups covenng various fields

tiStS

Documentation

Canada

Reports on Acid Deposition, Acid waters, Terrestrial effects, Nitrogen deposition, Photochemical oxididants, Buildings, Critical loads

Eight assessment teams mainly from Environment Canada and provincial orpanizations Eight separate reports: Executive summary; Emissions and control; Atmospheric sciences; Aquatic effects; Terrestrial effects; Human health effects socioeconomic studies

First phase (1980-90) Second phase (1991-) Ten task groups prepared 27 State of Science/ Technology &DO&

Integrated assessment document prepared in policy questiodanswer format: Effects of concern? Related to acid deposition? Sensitivity to change? Possible control scenarios?

P 4

m

468 2.2. PollutantEmissions of Concern

Significant differences in the pollutant emissions of concern were evident among the various national documents:

EuroDe nts of Concern

SO, emissions NOx role in acidification NO, rolein 03 formation NHx emissions

NL

S

SF

UK

CS

3 3 2 3

3 3 2 2

3 2 2 1

3 1 3 2

3 1 1 1

North h a a k a CAN USA

3 1 3 0

3 1 3 1(0F

a In the USA, NHx emissions were dealt with only in the NAPAP State of Science reports, not in the Integrated Assessment. In addition to this rough numerical comparison we found i t useful also to examine the amounts of emissions in each country in the several ways shown in Table 2. Table 2 Emissions of sulfur dioxide-S, nitrogen oxide-N, and ammonia-N in some countries (In Canada and the USA, emissions are shown only for the eastern parts of each country, roughly east of the 103rd meridian).

ElXOQ€!

Eastern North America CAN USA

S

SF

UK

CS

Emissions of sulfur dioxide-8: 125 loo Total, ktondyr Density, kg/ha/yr 35 3 Per capita, kg/yr 8 1 2

165 5 33

1900 74 32

1200 93

120 3 14

76 2 15

750 29 l3

250 19 16

298 0.7

60

35 1 7

450

160 12 10

66 0.2 4

NL

Total, ktondyr Density, kg/ha/yr Per capita, kg/yr

165 46 11

Fmissions of ammonia-N: Total, ktondyr 1% Density, kghafyr 55 Per capita, kg/yr 13

2

7

18 8

77

1915 3 94

20

m 17 52 4702 9 25 984 2

5

469

Discussion Globally, natural emissions of sulfur and nitrogen oxides are significant. In Europe and North America, however, natural emissions are small (10% or less) compared to anthropogenic emissions. Emissions densities in the various countries varied over a very wide range -by a factor of 35 in the case of sulfur oxide emissions, by a factor of 70 in the case of nitrogen oxide emissions, and by a factor of more than 200 in the case of ammonia emissions. The large differences i n human population density between most of central Europe and the UK compared to the Scandinavian countries are reflected in the much larger emission densities for The Netherlands and the UK compared to those for Sweden and Finland. Similar differences in population density are apparent between Canada and the USA. In The Netherlands and Sweden, per capita emissions of sulfur were very low (8 and 12 kg/yr) in comparison with eastern Canada (94 kg/yr), Czechoslovakia (77 kg/yr), and the USA (52 kg/yr). Finland and the UK showed intermediate figures (33 and 32 kg/yr). Per capita emissions of nitrogen oxides were much more similar in all countries with only a factor of 2.5 separating the values for the lowest country (The Netherlands) and the highest country (USA). The ratios of SOx/NOx (on an elemental weight basis) were less than one in the Netherlands (0.76)and Sweden (0.831, but vaned from two to six in other countries (2.17 in Finland, 2.53 in the UK, 4.80 in Czechoslovakia, 6.33 in eastern Canada, 2.04 in the eastern USA). These figures are reflections of differences in rates of energy use, dominant types of industries, and the strength of control measures used in various countries. Emissions of sulfur dioxide, and to a lesser extent, nitrogen oxides, have been the dominant pollutants of concern with regard to acidification problems in both Europe and North America. In The Netherlands, however, emissions of ammonia are now recognized as the emission of concern -- with larger total emissions, a larger emissions density, and even a larger per capita emissions rate for ammonia than for either sulfate sulfur o r nitrate nitrogen. The importance of total nitrogen deposition also has risen in important as nitrogen saturation and eutrophication as well a s acidification of ecosystems has become more common in Europe. These developments also have led to some important changes in the language of air pollution discussions, especially in Europe, where the terms acidification, acidifying deposition, total potential acidity, total nitrogen deposition, nitrogen saturation, and eutrophication all are growing in frequency of use and concern in both scientific debates and public discussions. In North America, both nitrogen oxide emissions and ammonia emissions were not considered to be a s important as they were in Europe. Although emissions inventories for ammonia were included in the scientific reports for

410

the USA,the major regions of ammonia emissions were considered too distant to be important in areas where significant effects of acidification had been observed. We could find no mention of ammonia emissions in the assessment document for the USA.

As discussed in a later section of this paper, we found the emissions data in Table 2 to be important also in understanding some of the conclusions drawn and the management strategies suggested in the national documents. 2.3. Domestic and InternationalPollutantSources of Concern

In all countries, only a fraction of the total deposition within the country originated from emissions sources within the country itself. Table 3 shows the distribution between domestic and foreign sources to the extent that this is discernible from the national documents. Table 3 Origin of the deposition from sources within each country itself, % of total deposition

EuroDe NL S ~

-N

SF

UK

CS

CAN

USA

10 25 4 0 1 0 1 5

70 60 70

50 XI 40

50 35

80 96

~

Origin of SO, Origin of NO, Origin of NH,

30

804030

--

--

The extent of transboundary exchange of pollutants is a very complex function of the atmospheric half-life of the pollutant, the relative size of the country, location of emission sources within the country, the relative magnitude of domestic sources and sources outside the country, the strength and direction of dominant winds, and many other factors. 2.4. Depositionof Acidification-Relevant Substanw

Rates of deposition varied greatly within all countries. Within Sweden, for example, the areal deposition of sulfur in the north is only about 10% of that in the south. Also, within a given region, differences in type of vegetation, topography, and other landscape and climatic factors influence the rate of deposition. The deposition to a coniferous forest edge is sometimes five times larger than that t o the interior of a forest or to a typical agronomic crop. Because of these large variations within countries, we have presented in Table 4 the estimated mean regional deposition in the most affected regions within various countries. Still higher deposition values were reported in close

47 1

proximity to large point sources, in cloud-impacted forests, and other areas with special conditions. Table 4 Deposition of some acidification-relevant compounds. The values shown are estimates of regional mean values in acidification-affected areas within each country, k g k d y r

NL

S

Euro~ea SF UK

CS

North Arne* CAN USA

..

osition of sulfur oxides-& 24 25 Total, kg S k d y r 8 I2 Wet, kg S k d y r

16 10

30 10

160 100

10

Denosition of n itroven oxides-N: Total, kg N h d y r 19 10 Wet, kg N k d y r 5 5

6 4

10 4

18 10

10 5

f ammonium -N and ammonia-N: 6 30 Total, kg N/ha/yr 60 10 Wet, kg N k d y r 14 5 4 5

-_

_-

16

5

--

14

4

5

..

ition of Ca Wet, k g k d y r

+ MP:

9

3

5

5

13

17 10

l5 8

--

~~

a In the national document for Norway (SNSF report), wet deposition was estimated to account for about 70% of the total s u l h r deposition.

Discussion The data in Table 4 show several intriguing similarities and differences in areal rates of deposition in the most-affected areas within each country: As expected, total (wet plus dry) deposition is greater than wet-only deposition. In Europe, however, the tota1:wet deposition ratio usually was greater than in North America. In the most-affected areas within Europe, for example, total (wet plus dry) sulfate deposition was usually 2-3 times greater than that for wet-only sulfate deposition -- 3.0:l in The Netherlands, 2.1:l in Sweden, 1.6:l in Finland, 3.0:l in the UK, and 1.6:l in Czechoslovakia. By contrast, in North America, the tota1:wet ratio for sulfate deposition is only 1.3:l in eastern Canada and 1.7:l in the eastern USA. Regional estimates in Canada indicate that total sulfate deposition is only about 15%greater than

472

wet sulfate deposition, except for areas near large point sources of sulfur dioxide emissions. The distinctions among the most-affected regions of Europe and North America is less striking with regard to nitrate deposition. The tota1:wet ratio for nitrate deposition was -- 3.8:l in The Netherlands, 2:l in Sweden, 1.5:l in Finland, 2.51 in the UK, 1.8:l in Czechoslovakia, 2.0:l in Canada, and 1.9:l in the USA. The deposition data in Table 4 for the UK refer primarily to Scotland and Wales which contain the most-affected regions. In some parts of southern England, however, the total deposition of sulfur and nitrogen is much greater than in the more acid-sensitive regions of Scotland and Wales (as much as 60 kg of SOX-sulfurhdyr and 70 kg of total nitrogenihdyr (NOxN plus NHx-N). The tota1:wet ratio for deposition of ammonia was sometimes higher than for either sulfate or nitrate deposition -- 4.2:l in The Netherlands, 2.0:l i n Sweden, 1.5:l in Finland, and 6 : l in the UK. Neither total nor wet deposition data for ammonia were found in the national documents for Canada and the USA. In the national documents for Canada and the USA, expected rates of acidification usually were evaluated in terms of wet sulfate depositionihdyr rather than in terms of total (wet plus dry) sulfate deposition, total sulfur deposition, or total sulfate plus nitrate depositionihdyr. I n Europe, total deposition of sulfurlhalyr o r total deposition of sulfur plus nitrogedhalyr (sulfate ion + sulfur dioxide + nitrate ion + ammonia + ammonium-ion/ha/yr) were more commonly used. In the national document for The Netherlands, these ideas were extended further in two closely related concepts called potential acid deposition and actual acid deposition which are defined as follows: "Various substances are considered to contribute to acidification: - the deposition of oxidized sulfur compounds (SOX) includes dry deposition of S02, and sulfate aerosol (SOe), a s well as wet deposition of sulphate (SO4).1mol SOx can lead to the production of 2 mol H+; - the deposition of oxidized nitrogen compounds (NOy) includes the dry deposition of NO, NO2, HN02, HNO3, and nitrate aerosol (NOg), a s well as wet deposition in the form of NO3. 1mol NOVcan lead to production of 1mole H+; - the deposition of reduced nitrogen compounds (NHx) includes the dry deposition Of NH3 and NH4 aerosol, as well a s dry deposition of NH4.- 1 mol NHx can lead to the production of 1mol H+. It should be mentioned t h a t these figures indicate the maximum (potential) contribution. The actual acid load (H+) depends on what happens to the compounds [the chemical and biological processes taking place in ecosystems1 after deposition." [In many ecosystems, especially

473

those deficient in nitrogen, the difference between potential and actual acid deposition can be quite large].

As also shown in Table 4, regional average wet deposition of base cations (Ca

+ Mg) in the most-affected regions does not vary much from country to country.

Nevertheless, base cations can influence the acidity of air and precipitation and sometimes even its acidifying effects in ecosystems. In both Europe and North America, the main sources of Ca and Mg are dust from arable land. In southeastern Finland and some areas of central Europe, however, the amounts of basic cations in dust from some combustion sources in eastern Europe sometimes are so high that they effectively neutralize the sulfate and nitrate in wet and dry deposition. In the most-affected regions of the northeastern USA, wet deposition of nitrate nitrogen seems to be larger than in the most-affected regions of some other countries. As shown in Table 4, the total deposition of nitrate nitrogen is somewhat larger in The Netherlands and the USA than in Sweden, Finland, the UK, and Canada. It is striking that gaseous ammonia and ammonium-ion deposition are considered major air-pollution problems in the national documents for most European countries. But these pollutants are dealt with only very briefly in the national documents for Canada and the USA. Why is this so? We speculate a s follows: The importance of these pollutants derives from three closely related concepts which have gained general scientific acceptance only recently (last 5-8 years): 1)the concept of acidifying deposition (as opposed to acidic deposition). This concept includes the notion of gaseous ammonia and ammonium ion a s acidifying nutrients rather than j u s t a s acidneutralizing constituents in wet and dry deposition; 2) the concepts of nitrogen saturation and optimum nutrition of ecosystems (as opposed to wide-spread nitrogen deficiency in ecosystems); and 3) the concept of nitrogen-induced (as opposed to phosphorus-induced) eutrophication of large lakes, estuaries, and even ocean waters. Much more evidence for these concepts was accumulated in Europe, and especially in The Netherlands and Sweden, than in North America. Manures from large populations of domestic livestock have been regarded a s the largest source of airborne ammonia and ammonium ion in Europe. In central Europe, large domestic animal populations exist in close proximity to areas sensitive to acidifying deposition. In the assessment documents for Canada and the USA, however, these distances apparently were regarded as too long for ammonia to be contributing significantly t o acidification impacts. 2.5. Effects on Surface Waters

Much of the original impetus for development of national programs of research and assessment on the acid-deposition problem derived from

414

discovery during the late 1960s and early 1970s of linkages between: 1) pollution-induced changes in the chemistry of precipitation; 2) concomitant changes in the chemistry of lakes and streams; and 3) changes in fish populations. The original focus of concern was on emissions of sulfur dioxide which led to changes in the sulfate content and therefore the sulfuric-acid content of rain and snow. Later, concerns were broadened to include: Dry deposition of acidic aerosols and gases; Emissions of nitrogen oxides which led to formation of nitric a s well a s sulfuric acid; The concept of acidifying as well a s acidic deposition and thus the inclusion of ammonia and ammonium ion a s part of the several causes of acidification of soil and surface waters; The discovery of episodes of acidic stream water a t times of spring snow melt and after prolonged summer droughts; The acidification of soils and eventually ground water; and The effects of the changing acidity and alkalinity of surface waters on the productivity, health, and reproductive capacity of fish, amphibians, aquatic insects, benthic invertebrates, and other aquatic organisms. Through this series of discoveries relating to the effects of acid deposition on surface waters and aquatic ecosystems came much of our present understanding of the acid deposition issue. It is interesting t o compare the recent national assessment documents in the context of this series of unfolding discoveries and concepts:

Eurom Concern about effects on surface waters

NL

S

Sl?

UK

1

3

3

3

North Amen= CAN USA 3

3

The Netherlands The national document for The Netherlands contains no significant discussion of acidification effects on surface waters and aquatic biota. There are two major reasons for this: 1) there are very few natural lakes or streams in the country. Although a high fraction of the total surface area is surface water, almost all such waters are either highly polluted rivers o r man-made canals with a n extremely high water table; and 2) the subject of surface water effects (mainly focused on small ponds) was discussed in the assessment document completed earlier in 1988 by the Dutch Priority Program on Acidification.

Since Swedish research scientists and their colleagues in Norway have been (and remain) in the forefront of studies of the effects of acid deposition on

475

aquatic ecosystems since the mid-l960s, it was not surprising that the current national document for Sweden contained significant reports of progress in understanding air pollution and acid deposition effects on aquatic ecosystems. "About 16,000 of Sweden's total of around 85,000 lakes are so seriously affected by acidification that sensitive species had greatly declined in number or disappeared completely. Areas where more than 25 per cent of all lakes and watercourses are seriously affected [now cover more than half the total land area of southern Sweden]. The situation of the flora and fauna in 6,000 of these lakes has been markedly improved by means of liming. At least a quarter of the total length of watercourses would be seriously harmed by acidification if i t were not for liming." "Water chemistry studies indicate that the acidification situation for surface waters in Gotaland (southern Sweden), Svealand (central Sweden) and coastal areas of Norrland (northern Sweden) in general has not changed to any great extent since the middle of the 1970s. On the other hand, acidification in the mountains and adjacent areas in northern Sweden has continued to worsen over the last 10-15 years. Deleterious biological effects due to acidification of minor watercourses have increased considerably in the 1980s. Iron, aluminum and manganese are dissolved out of the ground and then precipitated in the water. Fish and many bottom-dwelling animals have decreased in number."

United Kingdom Many studies have shown that streams draining from forest plantations are more acidic and contain higher concentrations of aluminum than streams draining from grassland and moorland. The major factors appear to be the increase in strong acid anions in the drainage water caused by the "airfiltering" effects of the tree canopies and the production of organic compounds in the soil as a result of drying. Sediment core studies have shown that lakes with pH less than 5.0 became common only during recent decades and only in areas of high acid deposition. The downward trend in pH generally began around 1850 and typically involved a decline of 0.5 to 1.5 pH units. All studies point to atmospheric deposition of acidic and acidifying substances as the major causal factor. The ecological effects of acidification are of fundamental importance in the

UK. Populations of salmon and brown trout show evidence of decline, as do several species of frogs. The dipper, an insectivorous riparian bird, has been shown to decrease in abundance along streams in both Scotland and Wales.

A report on Acidity in Scottish Rivers summarizes effects on surface waters a s follows: "In recent years, evidence has been accumulating of the acidification of rivers and lochs in parts of Scotland. In the mid-l970s, a number of lochs in southwestern Scotland with granite catchments were found to be acidified.

476

Several of these lochs were fishless and others supported only sparse fish populations. Subsequent studies in Central Scotland showed that streams draining base-poor catchments with large areas of mature coniferous forest were particularly susceptible to acidification and had impoverished fish and invertebrate communities. More recent work has provided further evidence of acidification and its effects on fish populations, particularly in Galloway. Paleoecological studies have shown that lochs here and in other parts of Scotland (Arran, Rannoch Moor, and the Cairngorms) are acidified." Canada

"Forty-three percent of Canada's land area is sensitive to acidic deposition. Sensitive terrain generally has non-carbonate bedrock and coarsely textured, shallow surface deposits ...The coincidence of sensitive terrain and acidic deposition defines the damage area of concern for aquatic effects. In eastern Canada these conditions occur in an area east of the Manitoba-Ontario border and roughly south of James Bay." "An inventory of lakes with an area slightly smaller than the region of concern described above...found almost 800,000 water bodies greater than 0.18 hectares in area. A special data base has been compiled that contains physical and chemical information for 8,505 lakes across eastern Canada." "The regional acidification of eastern Canadian lakes is primarily due to the deposition of atmospheric sulphate rather than to nitrogen deposition or to natural acidifying agents, such as organic acids or sulphide minerals in the bedrock. Acidification arising from land use changes is not considered to be important." "Apart from the long-term acidification of waters, temporary episodes of acidification may also produce conditions that are lethal to aquatic biota. Storage of acids within the winter snowpack can lead to the release of exceptionally high concentrations of sulphate during early melt stages that may cause short-term acidification of surface waters. Acidification episodes can also be caused by temporary storage of sulphate in the catchment, particularly in wetlands, during dry seasons. Release of sulphate at the onset of the next period of wet weather causes short term acidification of the runoff water."

USA "Within acid-sensitive regions of the United States, 4% of the lakes and 8% of the streams are chronically acidic. Florida has the highest percentage of acidic surface waters (23% of the lakes and 39% of the streams). In the MidAtlantic highlands, the mid-Atlantic Coastal Plain, and the Adirondack Mountains, 6%-14% of the lakes and streams are chronically acidic; about three times that many become temporarily (days to weeks) acidic during storms and snowmelt conditions in these regions. Virtually no (~1%) chronically acidic surface waters are located in the Southeastern Highlands or the mountainous West."

411

"Acidic deposition is the dominant source of acid anions, excluding chloride, in about 75% of the acidic lakes and 50% of the acidic streams in the National Surface Water Survey (NSWS). Most of these have probably become more acidic (declined in pH) because of acidic deposition. These acidic surface waters are found primarily in the mid-Atlantic Highlands, the Adirondacks, New England, the mid-Atlantic Coastal Plain, Florida, and the eastern portion of the upper Midwest." "Natural organic acids are the dominant source of acid anions excluding chloride, in about 25% of the acidic lakes and streams in the NSWS, and these are found in Florida, the mid-Atlantic Coastal Plains, the upper Midwest, and New England." "Acid mine drainage is the major source of acid anions in about 25% of the acidic streams in the NSWS. These acidic streams are located in the midAtlantic and Southeastern Highlands."

26.Ef€ectson Forests The assessment documents for all European and North American countries contained significant sections on the effects of air pollutants and acid deposition on forests: EuroDe Concern about effects on forests

NL

S

SF

UK

3

3

3

2

North Amerka CAN USA 2

2

In the national documents for all six countries, several general conclusions were drawn: Airborne gases and aerosols as well as precipitation, cloud water, and fog were recognized a s significant sources of several of the 16 essential elements needed for plant growth and development, and, most notably, as significant sources of nutrient sulfur and nitrogen; The cumulative atmospheric deposition of acidic and acidifying substances (including sulfate, nitrate, and ammonium ions a s well a s gaseous ammonia, amines, and other nitrogen compounds) were recognized as potential causes of increased acidity and decreased fertility of forest and other wildland soils; Forest trees and other natural vegetation were recognized to be subject to a wide variety of natural and human stress factors. These include natural competitive stresses, natural climatic stresses such as drought, frost, and wind, natural nutrient stresses such as nutrient deficiencies and excesses, natural biotic stresses -- especially insects and fungal diseases, human disturbance stresses, and air pollutant stresses. Thus, it was

47 8

always difficult to reliably distinguish between the individual and combined effects of airborne chemical stresses and other natural or human stress factors acting singly or in various combinations; Ozone was recognized as a significant stress factor for forest trees and other natural vegetation in many parts of Europe and North America. The emphasis given to these several generalizations varied considerably in the national documents from each country in Europe and North America.

The Netherlands The major conclusions regarding effects on forests in The Netherlands include the following: "Increasing nitrogen deposition over a period of several decades has led, first of all, to a removal of nitrogen deficiencies and increased growth, and secondly, to nutrient imbalances as a result of magnesium, potassium, and phosphorus deficiencies. More and more forests are moving from a situation of nitrogen deficiency to a situation of nitrogen saturation. At the moment [1990]about 15% of the Dutch forest soil is saturated. At the same time, the nitrogen input, together with SOX deposition, is causing considerable soil acidification. The present contribution of nitrogen to actual soil acidification is about 35%, and that of sulfur about 65%. The combined action of increased nitrogen availability and soil acidification has led to a decline in cation availability (nutrient deficiency). This is resulting in a greater risk of damage to forests by pests and plagues, frost and drought." "In The Netherlands, soil acidification is the greatest risk factor. The research carried out in the context of the second phase of the Dutch Priority Program on Acidification has confirmed the hypothesis that Dutch forest soils are degrading (radical physico-chemical changes in the long run) by deposition of acidifying substances. This hypothesis was proposed about five years ago on the basis of various measurements and data from other countries. Confirmation of this hypothesis was obtained from input-output budgets and model analyses, and through nation-wide monitoring of soil solution chemistry. A major concern are virtually irreversible changes in the soil caused by depletion of the aluminum buffer, and the consequences of a decline in pH (to between 2.8 and 2.9) associated with aluminum depletion, which in the event of unchanged deposition) is the expectation for Dutch forest soils. In any case, large changes in the soil and thus in the conditions of forests stand locations, are to be expected. This could lead to changes i n vegetation and soil fauna."

Sweden The assessment document for Sweden is less worrisome with regard to effects on forests in large part because of lower rates and cumulative amounts of deposition of sulfur and nitrogen compounds, and lower ozone concentrations

479

than in The Netherlands. Nevertheless, the document contains the following statements regarding effects on forests:

"A considerable portion of Swedish forest is suffering from impaired vitality. Approximately 20 per cent of spruce trees and 14 per cent of pine have abnormal needle loss (more than 20 per cent of needles). Deciduous trees such as beech, oak, and birch display more extensive damage than spruce o r pine. Such damage is manifested in the form of thinning and a n altered growth pattern in the crown. It is not possible to distinguish a clear trend in the damage caused to spruce and pine over the last six years [198419901". "Increasing soil acidification and the impoverishment of forest land thereby caused appear to be the greatest long-term threat to Swedish forests". "Uncultivated [mostly forest] land in the south of Sweden has been acidified to a considerable extent over the last few decades. Acidification of soils in Skdne and Halland, as well as in southern Smdland (all provinces situated in Gotaland) has penetrated to a depth of several meters, where i t affects the superficial ground-water. This acidification has meant that the easily available store of nutrients such as calcium and magnesium in the soil has diminished between 30 and 70 percent in southernmost Sweden. A t the same time metals such a s aluminum are liberated. Such metals are poisonous to plants and animals." "There are strong indications that forest soils in southern Sweden are approaching nitrogen saturation, with resultant acidification, nitrogen leaching and nutrient imbalance in the trees. Large areas of southern Sweden may be suffering from nitrogen saturation within 10-20 years unless nitrogen deposition can be reduced."

United Kingdom Surveys of the health of forests in the UK suggest that significant changes in crown density have taken place in sitka spruce and scots pine. Beech forests are not in good health in some parts of the UK. Air pollution cannot be excluded as a contributing stress factor. There is no direct proof of pollution-related decline of forest trees in the UK. But some forests are subjected t o pollution climates which probably cause stress. There is good evidence that air pollutants decrease the frost resistance and winter hardiness of forest trees. Damage by some fungal diseases and insects seem to be increased by exposure to air pollutants. Canads

"In Canada, most of the forests exposed to high levels of acid deposition and ozone are those close to populated areas. These forests are some of the most productive in the country and the most heavily utilized. Typical uses include

480

recreation, tourism, wildlife habitat, aesthetics and forest products activities (woodlots, maple syrup, quality softwood and hardwood lumber)." "There is general agreement that the current episode of maple decline in Canada is more severe and more extensive than those which have occurred in the past. Although no current evidence currently exists for a direct (foliar/airborne) o r single component causal role for acidic deposition o r ozone in any of the current tree declines i n eastern Canada, studies conducted in Ontario and Quebec indicate that acidic deposition and other long range transported air pollutants may be indirect (soilhutrient) o r contributing (pre-disposing) factors in sugar maple decline. The role and importance of these additional stresses on the trees already under natural stresses such as climatic extremes (drought, temperature), or attack by insects and disease, remains unknown but must be viewed with concern." "Symptoms of maple decline and tree mortality have continued to increase in severity and extent. Other species in the same stands have also been affected. In many areas subject to acidic deposition and experiencing tree decline, soils have become deficient in several essential nutrients. Nutrition research has indicated that soil chemistry is a n important factor in the decline syndrome and that in the short term, forest fertilization can provide a n effective ameliorative treatment." "Dendrochronology data from Ontario maple decline studies also indicate that significant growth reductions have occurred in declining a s well as outwardly healthy trees since the mid-1940s to mid-1950s in regions experiencing moderate to high levels of acidic deposition and ozone." "Research a t New Brunswick has circumstantially linked the white birch deterioration along the Bay of Fundy with exposure to acidic marine fog."

USA "Ozone is important in a decline of pines in southern and central California and is the pollutant of greatest concern with respect t o possible regionalscale impacts in North American forests." "Within the Los Angeles Basin and the southern and central Sierra Nevada Mountains, ozone has been documented to contribute to decline in the health of mixed conifer forests in the San Bernardino Mountains. There is evidence for alteration of productivity, ecosystem dynamics, and physiological processes of ponderosa and Jeffrey pines by ambient levels of photochemical oxidants. Field studies support the occurrence of these alterations in time and space with patterns of oxidant occurrences. "Controlled exposure studies with a wide variety of forest species have shown either no effect, mixed results, o r negative effects on growth a t ambient levels. As a result, quantitative, reproducible exposure-response functions are not as available for tree species as they are for crop species."

48 1

"There is no conclusive evidence of widespread forest damage from current ambient levels (pH 4.0-5.0) of acidic deposition in the United States". "Although crop production rates are high and most forests appear healthy, acidic deposition and associated air pollutants affect some terrestrial ecosystems." "There are indications that acidic deposition and associated pollutants have contributed to growth reductions and mortality of red spruce a t high elevations in the northern Appalachians and to growth reductions in red spruce in the high elevations of the southern Appalachians." "Long-term changes in the chemistry of some sensitive soils are expected, but it is uncertain whether these will result in reduced forest health, how the effect would be manifest, how much of the forest resource would be affected, o r how long it would take for such effects to occur." 2.7. Effects on Agriculture

-

Significant differences in concern about effects of acidification and air pollutants on agricultural crops were evident among the various national documents:

Concern about effects on agricultural crops

NL

S

SF

UK

0

2

1

1

North A& CAN USA

3

3

The national documents for the European countries, include only very brief statements about the effects of gaseous air pollutants and acid deposition on agricultural crops and none about effects on domestic livestock. By contrast, the assessment documents for Canada and the United States, deal both explicitly and a t length with effects of gaseous air pollutants and acid deposition on agricultural crops. But they too contain no information about effects of ozone or acid deposition on animal agriculture. Sweden

"Ozone levels in Sweden during the summer months are near or above the levels which, in the short or long term, harm crops. In western Sweden the present ozone levels reduce the crop of spring wheat by around 10 percent, while barley seems to be less susceptible. Crops such a s meadow grass, oats, and potatoes are also affected".

United Kingdom Ambient concentrations of ozone are known to decrease yields of some sensitive crops in some areas of the UK. Interactions between pollutant

482

stresses and other stress factors such as diseases and insects also have been demonstrated in various locations. Current concentrations of sulfur dioxide and nitrogen oxides in the UK usually are too low to cause significant harm by either pollutant acting alone. Increased damage by ozone has been observed when ozone and sulfur dioxide o r ozone and nitrogen oxides occur concurrently. Canada

The Canadian assessment document on Terrestrial Effects includes a chapter on effects of acid deposition and other air pollutants on agriculture crops. Ozone is the major air pollutant of concern to agriculture in Canada and critical levels for foliar injury are included in the assessment document. A significant reanalysis was made of the adequacy of Ontario's present ambient air quality standard for ozone (80 ppb for one hour) to protect agricultural crops and ornamental plants from harm. The conclusion was that attainment of this standard would provide economic benefits to agriculture in Ontario alone of between $170 million and $1.9 billion per year. Effects of sulfur dioxide on crops is not discussed but effects of peroxyacetyl nitrate (PAN) and nitrogen oxides are mentioned as potential causes of direct effects and of interactive effects with ozone. Field experiments in Canada and the United States have shown no significant direct effects of acid deposition on yields of crop plants (with the exception of a single variety of soybean (Amsoy). Also, no important interactions between acid deposition and ozone were reported.

USA The United States assessment document also includes a section dealing with effects of acid deposition and other air pollutants on agriculture. Although sulfur dioxide is reported to cause occasional damage to crops in the vicinity of major point sources of pollutants under unusual meteorological conditions, ozone is the major air pollutant of concern to regional crop production. A major study called the National Crop Loss Assessment Network (NCLAN) provides a very thorough analysis of ozone impacts on the yield of agricultural crops. The important crops found to be damaged by ozone include alfalfa, corn, cotton, soybeans, sorghum, forage, rice, and both spring and winter wheat. These nine crops account for 75% of the value of total crop production in the United States. The NAPAP assessment document contains estimates of the potential changes in wheat, corn, soybean, and sorghum production in each of the 10 major agricultural production areas of the United States. The document contains two major conclusions about agricultural effects: "Ozone represents a significant stress factor in agricultural crop production in the United States. Depending on species, location, and exposure, yield reductions in crops have been estimated to range from 2% to 56% a t ambient ozone concentrations"; and

483

"Acidic deposition a t ambient levels is not responsible for regional crop yield reduction." NCLAN provided estimates of the range of economic impacts of ozone on crop production in the United States ($1 to $5 billion per year). The NAPAP assessment document concluded that: "Substantial increases in tropospheric ozone levels above background level [approximately 30 ppbl, can be expected t o cause biological damages or crop yield reductions. However, an evaluation of economic effects requires a n allowance for the action of market forces..." These market forces were not considered in developing the NCLAN estimates of economic impacts. 2.8. Effects on Natural Terrestrial Flora and Fauna

Relatively few studies have been made of the effects of air pollutants on terrestrial fauna and flora (non-tree species). This topic has played a minor but recently growing role in most national research programs. It has rather strong implications for the air pollution policy especially in some European countries:

EuroDe N

L

S

S F U K

North America CAN USA

Overall impact'concern

3

2

2

2

1

0

Vascular plants Lichens, mosses Birds Heavy metals in moose Land snails

3 2 2 0 0

2 2 2

2 2 0 0 0

2

1 2

0 1 1 0 0

3 2

2

2 0 0

2

3 0

The Netherlands Increasing atmospheric deposition of total nitrogen strongly stimulates growth of nitrogen-loving grasses (e.g. Molinia caerulea and Deschampsia flexuosa) and herbaceous plants (e.g., Urtica sp. and Rubus idaeus) over large forest areas. Heathlands in The Netherlands are being rapidly transformed into grassland, mainly due to acidification and excessive inputs of nitrogen. About one third of the heathland is still healthy, about one third contains large amounts of grass and will probably change into grassland within 3-5 years, and one third has already changed into grassland. Some rare heathland species have almost disappeared, probably due to direct effects of gaseous sulfur dioxide coupled with soil acidification. The critical

484

level for sulfur dioxide impacts on heathland in The Netherlands is estimated to be about 8 pg sulfur dioxidelcubic meter -- a n amount which is frequently exceeded. The main cause of these vegetational changes is excessive deposition of nitrogen. The changes are o r major interest i n policy-making. Separate critical loads for nitrogen have been established based on the criterion of “vegetation changes” in both heathlands and well-drained sandy forest soils. Changes in lichen populations also have been observed in The Netherlands -decreases early in this century, probably caused by sulfur dioxide, followed by increases in recent decades probably caused by ammonia which now neutralizes the harmful effects of sulfur dioxide.

Sweden Vascular flora, particularly in deciduous forests in southern Sweden, have changed over the last 15-35years. Species which prefer only weakly acid soils have declined while species with a preference for nitrogen-rich environments have increased. The same tendency may be seen for fungi. Even larger changes in fungal flora are expected by the year 2000. Extensive changes in epiphytic mosses and lichens have been observed in southern Sweden. Approximately 130 species are affected. Many species have disappeared or become threatened during the past 40 years. These changes are believed t o be caused by changes in both nitrogen and acid deposition. Effects on birds are seen especially among species (such as osprey) that feed on aquatic organisms. There also are signs that populations of overwintering coniferous forest birds have decreased in areas of damaged forests in southern Sweden. Acidification causes decreases in available soil calcium. These changes have had severe effects on populations of land snails i n several parts of southern Sweden. Strong accumulation of cadmium in kidneys and livers of moose has been established in a gradient from south to north. These organs have been declared unfit for human consumption in affected parts of the country. United Kingdom

In the northern and western UK, semi-natural vegetation covers a large proportion of the total land surface. Few data exist on the impact of air pollutants on natural vegetation other than some bog plants that are known to be affected by acidifying deposition. Little is known about effects of increased deposition of nitrogen compounds, including ammonia, on a wide array of plants species. Nitrogen deposition has caused some changes in vegetation that affect grouse habitat.

485

Many species of birds live near lakes and streams, with some totally dependent on fresh waters for food. Osprey and other fish-eating birds have been adversely affected. Effects on dipper populations have been studied in some detail. Both in Wales and Scotland, dippers are scarce on streams with pH less than 5.5 to 6.0. On a Welsh river, abundances of breeding dippers declined by 80 to 90% between the 1950s and 1980s a t the same time that the pH of the river water decreased.

Food chain effects through lichens and some other forage organisms, have resulted in accumulation of cadmium in kidneys and livers of moose and deer living in poorly buffered areas. These organs have been declared unfit for human consumption in several provinces. Defoliation in declining maple stands has been shown to decrease populations of some birds. Increased acidity in deposition results in decreased growth, size, and ground cover by some cryptograms, including several species of Cladonia and the feather moss, Pleurozium Schreberii. Rainfall events of pH 3.5 or lower pose a threat t o some species of herbs and lichens. This might have serious wildlife implications, since lichens are an important winter food source for ungulates. Wetland birds, e.g., the common loon and osprey, are affected, probably through changes in quantity and quality of their food supply in acidified lakes.

OthercoUntries In contrast to the situation in several countries in Europe, impacts of air pollution on terrestrial fauna and flora were not identified a s a n effect of concern in the 1980 Norwegian assessment document o r the 1990 assessment document for the USA. In many European countries, impacts on natural flora and fauna now play a n important role in establishing environmental goals. These differences with Norway and the USA might be due to: Significant differences in the magnitude of effects, perhaps because nitrogen deposition impacts are greater in some parts of Europe. Possible differences in societal and government attitudes between 1980 and 1990 or from one country to another. Changes in natural fauna and flora and disappearance of species are now matters of major concern in public discussions of the environment in some countries. 29. Effects on Soils

Results Acidification of soils can be expressed in various ways. It is often defined as a decrease in acid neutralizing capacity (ANC) or sulfate absorption capacity

486

(SAC). The soil solution characteristics are of key importance in evaluating the effects of acidifying deposition. Effects on forest soils were of great concern i n many of the national documents, especially those for European countries:

EuroDe N Overall impact/concern

3

L

S 3

S F U K 2

3

North America CAN USA 1

1

The Netherlands In The Netherlands, acidification of forest soils by airborne sulfur and nitrogen compounds has led to rapid depletion of base cations and increasing soil-solution concentrations of aluminum in the root zone of trees. A1uminum:calcium ratios are very high and aluminum concentrations frequently exceed the critical value of 2 lg/l for damage to tree roots. Mobilized aluminum is currently the main buffer system i n sandy soils. Mathematical models of these soils indicate th a t current rates of acid deposition will cause the exhaustion of the aluminum buffer system within 10 to 100 years. These acidifying effects are attributed primarily to total sulfur deposition (65%) and total nitrogen (35%)deposition. Sweden

Acidification of forest soils has increased dramatically during the past 50 years. The extent of these effects is most pronounced in southern Sweden and decreases progressively to the north. No acidification of soils has been detected in Norrland. Decreases in pH of 0.5 to 1 pH unit (and sometimes even 2 pH units) have been observed in large parts of southern Sweden. At present, the total land area with a pH of mineral soil below 4.4 is about 650,000 hectares. In this area, soil acidification has penetrated to a depth of several meters. Also, the easily-available base-cation content of these soils has decreased by 30-70% since 1950. About half of the acidification of superficial layers of forest soils results from natural processes, especially growth of forests and timber harvesting. Acidification of deeper layers of forest soils cannot be explained by natural processes. These deep-soil effects can only be explained by the cumulative deposition of acidifying substances from the atmosphere. About 90% of current changes in soil acidity are attributed to sulfur deposition, the remainder to nitrogen. Finland

Time-series data on soil acidification are not available from Finland. But mass balance calculations indicate similar extent and distribution of soil

487

acidification a s in Sweden. In southeastern Finland base cations often neutralize the strong acid anions in precipitation and thus minimize soil acidification.

United Kingdom As in other European countries, atmospheric deposition of sulfur from anthropogenic sources superimposed on already substantial deposition of sulfur from marine sources are believed to cause changes in the chemistry of terrestrial ecosystems and soils in the UK. Acidification of some sensitive soils also has been demonstrated in several areas of the UK. The extent to which acidity of upland peats has increased is not clear. Large inputs of nitrogen, especially ammonia, may have significant impacts in some sensitive soils. Long-term changes in soil acidity are suspected to decrease rates of litter decomposition and subsequent release of nutrients and to affect mycorrhizae on forest trees. The Critical Loads Advisory Group for the UK has recently published a series of maps showing the geographical distribution of critical loads for acidity and sulfur deposition into soils. These maps were then combined with maps for non-marine sulfur, nitrogen, calcium, and magnesium deposition in order to identify regions where critical loads for deposition of acidity and sulfur are exceeded. These maps show that large areas of west-central England and of west-central and northern Scotland are receiving total loads substantially in excess of the critical loads for both sulfur and acidity.

Canada Impacts of acid deposition on forest soils in Canada have not been well documented. But nutrient deficiencies have increased in recent years and decreases in soil pH have been detected. These changes in properties are thought to be contributing factors in the current changes in condition of maple and birch forests in southern Ontario and Quebec.

A recent analysis based on known soil properties and bedrock geology in eastern Canada indicates that 46% (400 million hectares) of the land surface mapped is considered sensitive, 21% moderately sensitive, and 23% less sensitive to acidification. The proportion of total land area considered sensitive was 82% for Quebec and 34% for Ontario. It is considered unlikely that current rates of nitrogen deposition to forests in Canada will produce nutrient imbalances in the short-term. Thus, sulfur deposition is considered the dominant cause of soil acidification in Canada.

USA Most soils in the eastern USA, where emissions and deposition of sulfur and nitrogen, and emissions of volatile organic compounds are concentrated, are inherently less sensitive to acidification than those in large parts of eastern Canada. Model simulations suggest that most soils in the eastern USA will not

488

experience large changes in chemical properties a t current rates of sulfur deposition. Even if sulfur deposition were increased by 30%, base saturation of most forest soils is expected to decrease by only 2-4% in a 50 year period. If sulfur deposition were decreased by 50%, base saturation is estimated to increase by 1-2% in 50 years. In eastern spruce-fir forests, especially at high elevations above cloud base, however, some changes in soil chemical properties and significant changes in rates of growth, susceptibility to winter injury, and rates of mortality of the trees have been detected. Regardless of region or elevation, soils within sprucefir forests are by nature strongly acidic (pH 3.5-4.5) and have mineral horizons with low base saturation (3-15%). Peak soil aluminum concentrations and aluminudcalcium ratios a t several sites in the southern Appalachians approach o r exceed the toxicity threshold for red spruce seedlings. Values in the northeastern USA typically have remained below this threshold. The total nitrogen content of spruce-fir soils is especially high in the southern Appalachians. Under current rates of sulfur and nitrogen deposition, the combination of large soil nitrogen pools and low carbodnitrogen ratios are expected to result in increased nitrate and aluminum concentrations in the soil solution. Discussion

The national documents for several countries suggest that biological processes and intensive forest management are of substantial importance in the acidification of humus and the upper-most portions of the mineral soil. I n deeper horizons, however, it appears that atmospheric deposition is the principal cause of acidification. Soil acidification is a matter of major concern in most of the national documents for European countries. Dramatic changes have been demonstrated in soil acidity, buffering status, ion exchange capacity, sulfate absorption capacity, weathering rates, and most importantly, nutrient supply and availability. These processes are thought to pose a threat to the long-term productivity and both the natural function and structure of terrestrial ecosystems in Europe. Impacts on soil chemistry do not seem to be very obvious in North America. In some areas, especially in eastern Canada, however, it appears that soil aluminum concentrations exceed critical levels. Also, essential nutrients in some soils are near or below suggested critical levels. These differences between Europe and North America might be explained by: Real differences in intensity and duration of land use. The landscape of Europe has been under intensive management by humans for many centuries. As a result, essential nutrient reserves have been depleted through erosion, intensive cropping and grazing, litter removal, treecutting, burning, and other land-use practices.

489

More intensive study of soil processes in Europe. Studies of the impact of airborne chemicals on soils were given very high priority i n many European countries. Detailed studies of soils early in this century also made it possible to remeasure soil properties on the same plots after time periods of 10 to 70 years. There are few places in North America, where reliable remeasurements of soil properties can be made over these time scales. In part this reflects differences in approaches in earlier forestry research. Detailed characterization of soils on fixed plots is more common in Europe than in North America. The 1981 final report of the SNSF project in Norway, included the following statement about acidification of soils: "Acid precipitation may result in changes in the properties of the soil. At present it is difficult to draw any definite conclusions in the time required. Severe effects can hardly be expected to be observed. Some land use changes (spreading of coniferous forests) have for a long time been known to affect acidity of the soil. Acidification may affect micro-organisms and invertebrates. The ecological importance of this is uncertain." 2.10. Effects on Groundwater

Acidification and other changes in groundwater and drinking-water quality were of some concern in most national documents: N Overall impactlconcern Major changes of concern

L

2

pH NO2 Al-

Euro~e S S F U K

th A m e d CAN USA

3

1

1

1

1

pH Cu Cd

pH

pH

PH

PH

Pb cu A1

Pb Pb

"he Netherlands

In a study a t 150 locations, the nitrate content of groundwater exceeded the drinking water standard (50 mg/l) in 30% of the coniferous forest sites and 13% of deciduous forest sites. The aluminum content of shallow groundwater exceeded the drinking water standard (7 pmoV1) in 90% of the coniferous forest sites and 70% of deciduous forest sites investigated.

490

Sweden

Groundwater accounts for about 50% of the municipal water supply i n Sweden. Approximately 1.1 million permanent residences and 2 million temporary residences (summer houses) use groundwater from their own wells. Acidified groundwater is found in most areas where soils and lakes are acidified; this is especially true of shallow groundwater. Acidification during the past 40 years has been the primary cause of changes in groundwater quality. Alkalinity in groundwater has decreased over large areas. Acid well water sometimes results in corrosion of household plumbing systems with resulting costly damage and dissolution of heavy metals, especially cadmium and copper. Very high concentrations of copper have been found in drinking water from many private wells. In a study of 300 households in western Sweden more than 25% had copper concentrations greater than 3 pgA.

Finland Acidification of groundwater has been documented and this has enhanced public understanding of the acidification issues in Finland. Groundwater is a significant source of drinking water especially in the countryside. The general public and policymakers, however, appear less concerned about human health effects than about ecological effects of acidification.

United Kingdom Groundwater acidification is not dealt with in the national documents for the

UK. Canada

Groundwater sources make up a significant proportion of all drinking water used in Canada. Twenty-six percent of Canadians use groundwater from domestic wells and approximately 38% of municipalities rely entirely or partly on groundwater. The overall rate of groundwater acidification in Canada is unknown. Short term acidification of shallow groundwater has been reported in association with large acid-loading events such as spring snow melt. If acidic deposition continues, the frequency and duration of near-surface pH depressions in the groundwater are expected to increase.

USA Acidification of groundwater is not specifically addressed i n the NAPAP Integrated Assessment report. Lead and methyl mercury are the only substances that are considered likely to pose increased health risks due to acid deposition. Discussion

Acidification of groundwater is generally regarded a s a slow process that results from the cumulative effects of acidifying deposition over decades of

49 I

time. Short-term acidification of shallow groundwater has also been reported, usually in association with acid-release episodes such as spring snow melt. Acidification of groundwater has been observed in large parts of Sweden. Here, groundwater is a n important source of both municipal and household water supplies. Acid well water results in corrosion in household plumbing systems, with resultant damage caused by leaks. Sharply increased concentrations of copper pose potential health risks. In The Netherlands, groundwater is strongly acidified and contains high concentrations of nitrates. Since groundwater from individual wells is not so widely used, however, these effects are of only moderate concern. Parts of southeastern Canada and the northeastern USA may experience groundwater acidification in the future. The 1980 Norwegian assessment document does not specifically address groundwater acidification. 2.1 1. Effects on Human Health

Results Air pollution can affect human health both directly and indirectly -- directly when people inhale toxic gases o r aerosols, and indirectly when acidification causes mobilization and/or dissolution of toxic metals that are then ingested with drinking water or foods. Mobilization of lead from soldered joints by acidified drinking water and accumulation of mercury in fish from acidified lakes are examples of indirect effects. Considerable variation in concern about these two types of effects on human health were evident in the various national documents: NL Direct health effects Indirect health effects

0 0

EuroDe S SF

UK

0 1

0 1

1 2

North A m x i ~ a CAN USA

2 1

3 1

The Netherlands Human health effects are part of another research program and hence not evaluated in Acidification Research in the Netherlands. Effects of long-term exposure to ozone are of concern. Acid aerosols are under suspicion. All drinking water is treated and meets EEC-standards. Sweden

The ambient air quality standard for ozone is exceeded in large parts of Sweden. Acidification of ground water is a matter of great concern because of increased risk to human health due to increased intake of heavy metals. Several studies have shown a connection between copper in drinking water and diarrhea in infants. A possible connection between copper toxicity and

492

cirrhosis of the liver is currently being debated. Although a causal linkage between aluminum and cadmium in drinking water and human health has not been established, their presence in drinking water from private wells is a matter of concern. Accumulation of methyl mercury in fish from acidified lakes has been demonstrated repeatedly. The guideline of 1 p g k g in pike is exceeded in thousands of lakes. The livers and kidneys from moose have been declared unfit for human consumption in many acidified regions of Sweden.

United Kingdom Effects of acidification and air pollution on human health is not dealt with specifically in the review group documents for the UK. The potential risk to humans from acidification is discussed briefly in the report on fresh water. Most of the surface water used for domestic water supplies receive treatments that do not include adjustment in pH or other measures to decrease metal solvency or corrosivity. Thus, soft acid waters may dissolve copper and lead from plumbing systems.

Canada Although a great deal remains to be learned about the health effects of air pollutants, there is now incontrovertible evidence that ozone and other pollutants are contributing to observed effects on human health i n Canada. In many Canadian cities, current peak ambient concentrations of ozone a r e sufficient to cause transient respiratory effects in healthy people. There is strong evidence that human health also is impaired by exposure to acid aerosols. The combination of ozone and acid aerosols is suspected to increase individual susceptibility to respiratory disease. An estimated 300,000 people in acid-sensitive areas obtain their drinking water from unregulated sources that may be affected by acidic and acidifying deposition. More monitoring of such water supplies for the presence of toxic metals is required before the risk to public health can be evaluated.

In some acid-sensitive areas in Canada, moose, deer, and caribou accumulate such high concentrations of cadmium in liver and kidneys that these organs have been declared unfit for human consumption.

USA Ozone and other photochemical oxidants are the most significant air-pollution threats to public health in the United States. About 145 million people live in areas that exceed the national air quality standard for ozone -- 0.12 ppm for one hour. The annual ambient air quality standard for sulfur dioxide (0.03 ppm) is attained virtually everywhere in the USA. A similar situation exists for nitrogen oxides. A t present, only the Los Angeles area exceeds the 0.053 ppm annual average air quality standard for nitrogen oxide. There is an emerging

493

database on the effects of acute (1-2 hour) exposures to nitrogen oxides on the lungs of some sensitive asthmatics. The databases on health effects and exposures to acidic aerosols have major uncertainties that preclude quantitative risk assessment. However, the information available justifies concern over the potential for health effects a t high ambient concentration of acid aerosols, especially for acid sulfate particles. Preliminary sensitivity analysis suggests that continuing acid deposition could significantly increase drinking water lead exposures. In the eastern USA, the population potentially a t risk because they obtain drinking water from shallow wells includes about 18,000 children and 7,000 women. Methyl mercury exposure due to acid deposition is not thought to be a significant problem, except for individuals who consume substantial amounts of fish from acidified lakes and streams. People who consume typical amounts of wild game are not likely to have significantly increased exposure to toxic metals due to acidic deposition.

Discussion Human health effects of air pollution and acidification were not discussed in the national documents for The Netherlands, Norway, or Finland. The direct health effects of ozone and acidic aerosols were of major concern in the national documents for Canada and the USA. By contrast, indirect health effects of acidification were not well quantified in the documents for either country. In Sweden, indirect health effects of acidification via drinking water and food are judged to be a substantial future health risk. Increased methyl mercury in fish is a large problem today. The widely varying attention given to health effects of air pollution and acidification in the national documents from different countries may be explained in part by the following: Real differences in human exposures. Almost all drinking water in The Netherlands is chemically treated. In Sweden and Finland, however, ground water from wells is commonly used. Soils and ground waters in Sweden are more strongly acidified than in North America and thus cause more obvious indirect effects. The high population density in urban areas in North America leads to high concentrations of ozone and other oxidants as well as acid aerosols. Different research priorities. In Europe, indirect health effects generally have had high priority, while direct health effects are of greater concern in North America. Similar differences in priority also are evident in the case of research on terrestrial effects.

494 2.12. Effects on Visibility

-

The most striking of all differences in concern about effects of air pollutants among the various national documents was the complete absence of discussion about pollutant-induced changes in visual range in Europe and its substantial emphasis in the documents for the USA:

Concern about effects on visibility

NL

S

SF

UK

CAN

0

0

0

0

1

America USA 3

The national document for the USA included substantial emphasis on both the fundamental physics of haziness in the atmosphere and the quantitative anaIysis of relationships between particulate matter in air and the quality of scenic vistas in parks and rural areas. A unique computer-assisted method for analysis of visibility effects is presented in the State of Science document on visibility. We infer from the complete absence of discussion of visibility impacts of air pollutants in the national documents for European countries that Europeans generally must have come to accept loss of visual range in scenic vistas and in rural and urban areas. They apparently regard it a s a n unfortunate (but unavoidable) consequence of the high population density and style of modern life in Europe. Apparently, both the public and federal and state government agencies in the USA continue to regard scenic vistas a s a n important part of the natural and cultural heritage of the nation. 2.13. Effects on Engineering Materials and Cultural Resources Significant differences were evident among the various national documents i n degree of concern about effects of acid deposition and air pollutants on engineering materials and on cultural resources such a s monuments, statuary, and historical buildings: EuroDe Concern about effects on engineering materials and cultural resources

NL

S

SF

1

2

1

UK 3

North America CAN USA 2

3

The Netherlands

In the national document for The Netherlands, we could find no discussion about effects of acidification on engineering materials or cultural resources. We are aware that effects on buildings and water works are,of concern within the country, however.

495

The national document for Sweden emphasizes the effects of acidification on underground installations and constructions: "One effect of acidification is that installations in contact with soil and water corrode more quickly than they otherwise would. ...Waterpipes, road culverts and similar constructions are affected both via the ground and via acidified water. 'I

"Zinc is the metal most sensitive to acidification, and corrosion of zinc constructions increases with falling pH and alkalinity in water. Lead is also considered to be sensitive to acidification, as are copper, cast-iron and carbon steel." "There is a certain correlation between the occurrence of corrosion damage in water pipes in Sweden and acidified lakes. The Swedish Water and Waste Water Works Association has estimated the total cost of corrosion damage to the water supply network throughout Sweden amounts to approximately 1 billion kroner per year. The Corrosion Institute carried out a study on the situation a t the beginning of the 1980s, in which i t was estimated that approximately one third of indoor corrosion damage has been caused by acidification."

United Kingdom Laboratory studies have shown clearly that air pollutants can damage most building materials. There is increasing evidence t h a t pollutants act synergistically. Commonly used damage functions usually do not take account the complex interactions among environmental variables t h a t affect the stability of building materials (e.g., the chloride content of air, synergism in the effects of nitrogen and sulfur oxides, the role of moisture in both vapor and liquid forms). There is no unequivocal evidence that present rates of weathering of stone and most metals in the structure of historic buildings are significantly different from those in the recent past. Different materials show different sensitivities to air pollutants. Various types of stone and metals interact strongly with sulfur and nitrogen oxides. Concrete shows reaction to carbon dioxide, while paints, plastics, and organic materials show greatest sensitivity to photochemical oxidants. It has so far not been possible to determine with confidence, relationships between pollutant exposure and rates of weathering or damage including the effects of present exposures and future changes in exposures which might result from changes in pollutant emissions. Thus, i t is not possible to develop reliable estimates of the costs of damage to buildings in the UK.

496

Canada In the national document for Canada, we could find no discussion of effects of acidification on engineering materials and cultural resources. But we are aware of concern within Canada about accelerated weathering a n d deterioration of buildings in urban areas -- especially buildings of special historical value.

USA The national documents for the USA include three separate State of Science and Technology reports dealing with materials: Effects of Acidic Deposition on Materials, Processes of Deposition to Structures, and Distribution of Materials Potentially at Risk from Acidic Deposition. There is also a n extensive discussion about materials damage in the 1990 NAPAP Integrated Assessment. Emphasis is given to three broad categories of materials mainly in urban environments: 1) functional construction, 2) structures with unique artistic qualities, and 3)structures with historical value. A few excerpts and conclusions drawn in these reports are instructive:

In functional construction, the following materials may be listed in order of decreasing sensitivity to acidic deposition: galvanized steel > marble > painted steel and limestone > painted wood > bricwmortar > copper > aluminum concrete. "The key concern for construction materials is additional public and private expenditures to maintain an acceptable level of fbnctionality and appearance for the duration of the structure's economic life. This can involve using more durable materials or stricter design standards as well a s maintenance actions on existing structures...Aesthetic and safety concerns may exist if maintenance does not always occur promptly in response to physical damage." "Marble stone statuary ranks high in terms of sensitivity, lack of protective options, and irreplaceability but is not exceptionally prevalent. Other statuary (e.g., bronze) is less sensitive and is possible to protect, reducing the importance of damages that otherwise would be irreversible." "Important historical buildings in the United States are listed in the National Register of Historic Places. This source indicates that there are over 275,000 historic buildings in an 18-state region of the Northeast, about 35,000 of which are constructed partially o r wholly of stone. All but 3% of the properties are located in urban areas. Five percent of the registered historic structures are of special value [and are] listed a s National Historic Landmarks." "The value of cultural resources derives from their symbolic a s well a s from their practical functions in society. ...Since the early 1980'9, annual expenditures for rehabilitation of existing structures have equaled or

497

exceeded investment in new construction (in the range of hundreds of billions of dollars), a n unknown portion of which was performed on historical structures. ...It is important to bear in mind, however, that not all of these expenditures are due to air pollution. ...Future incremental damages due to sulfur-related acidity will be smaller, even without further controls on SO2 emissions." 3.0. THE CRITICAL LOADS APPROACH -- IN WHAT COUNTRIES IS IT

USED?

Results Many nations are looking for more effective and scientific means for assessing the expected effects and comparing the benefits and costs of various types of emissions controls. Many would prefer an alternative to mandating arbitrarily defined percentage decreases or use of best available technology (BAT). The critical loads approach is an alternative concept that has some promise both nationally and internationally. Despite many differences in definitions and approaches, critical load calculations in different countries have so far yielded similar answers. This has led to increased confidence in the critical load concept as a method for developing pollution control policy. The critical load approach is used a s the basic tool in the current negotiations to reach international agreements in Europe on control measures to decrease emissions of sulfur and nitrogen under the Convention on Long-Range Transport of Air Pollutants within the UN-ECE. Readers with an interest in the origins of the critical loads approach may find i t useful to review the volumes edited by Nilsson and Grennfelt (1988) and Malanchuk and Nilsson (1989). The extent to which the critical loads approach has been used in the national documents for various countries is illustrated below: Euro~e Pollution issue of concern Acidification Nitrogen deposition Ozone

North America

NL

S

SF

UK

3 3 3

3 3 3

3 1 1

3 1 3

CAN

USA

2 0 3

0 0 3

The Netherlands The critical loads approach has been used extensively in both scientific discussions and in formulating environmental policy goals for The Netherlands. A set of concrete examples are given in Tables 5-8. The critical loads for various effects of total acid deposition on forests and surface waters are presented in Table 5. Similarly, the critical loads for

498

various effects of total nitrogen deposition on forests and heathlands are presented in Table 6. Table 5 Average critical loads of total acidifying deposition on terrestrial ecosystems on well drained sandy soils and surface waters in the Netherlands, mol/ha/yr Effects

Coniferous forests

Deciduous forests

Root damage Aluminum depletion Aluminum leaching to groundwater Decline of fish populations

1100-1400'

1200

1400-17OOb 1500

500

300

Surface waters

400

aThe first value is related to a critical aluminum concentration of 0.2 moVm3 and the second value to a critical aluminudcalcium ratio of 1.0. bThe first value is related to a critical aluminudcalcium mol ratio of 1.0 and the second value to a critical aluminum concentration of 0.2 moVm3.

Table 6 Average critical nitrogen loads for terrestrial ecosystems on well drained sandy soils in The Netherlands, molkdyr

Vegetation changes Elimination by grasses Frost damagelfungal diseases Nutrient imbalances Nitrate leaching to groundwater

Coniferous forests

Deciduous forests

400-1400

600-1400

Heathlands

700.1100 1500-3000 800-1250a 900-1500

1700-2900

aThe worst case would be total inhibition of nitrification.

2000-3600

499

These two sets of scientifically determined critical loads were then used by the Parliament of The Netherlands to establish target loads for both total acidifj6ng deposition and total nitrogen as described in Table 7. These politically determined target loads for various years were then translated into specific emissions-reduction targets for the years 1994 and 2000 a s outlined in Table 8. By fulfilling these targets leading up to the year 2010, The Netherlands will do its part (together with other nations of Europe) in attaining the objective of keeping atmospheric deposition below scientifically determined critical loads and thus avoid o r minimize harm to the forests, heathlands, and surface water of The Netherlands.

As indicated by the percentage changes in emissions outlined in Table 8, achievement of these environmental quality objectives will require rather drastic measures which will require that abatement strategies be developed for various categories of sources including agriculture, industry, households, etc. Table 7 Comparison of scientifically determined critical loads and politically decided target loads for total acidifying deposition and total nitrogen deposition in The Netherlands, molshdyr Scientific estimate gf critical load

Politically decided tarFet loads Year 2000 Year 2010

Total acidifvinv deDositiou: Forests and soils

1200 - 1700

2400

1400

1600

loo0

Total n i t r a n deDosition: Forests Heathlands

400 - 1500 700 - 1100

Table 8 Emissions reduction targets for The Netherlands, percentage change relative to 1980 emissions

Ammonia emissions Nitrogen oxide emissions Sulfur dioxide emissions

Year 1994

Year 2000

Year 2010

30

70

80-90

20

50 80

80-90 80-90

60

500

Sweden As shown in Table 9, similar use has been made of the critical loads approach in setting environmental goals for Sweden.

Table 9 Critical loads, target loads, current (1990) deposition, and percentage decrease in emissions of total sulfur and total nitrogen in various parts of Sweden a s specified in Parliament-approved environmental goals for Sweden, k g h d y r Critical load

Target load

Deposition in 1990

Reduction required (%I

5 5 3

10-20 5-10 3-5

75 50 40

10 8 6

10-20 5-10 25

50 20 0

Total sulfur deDosition: Whole country Gotaland Svealand Norrland

3-8

Total nitroTen denosition: Whole country Gotaland Svealand Norrland

5-15

The critical load for sulfur in the major provincial areas of Sweden vary from 3 kg sulfurlhdyr in the north to 8 k g h d y r in a few parts of southern Sweden. Thus, the target load for sulphur deposition is set a t 5 kg sulphudhdyr in Gotaland and Svealand (southern and central Sweden) and a t 3 k g h d y r i n Norrland. The target load value of 3 kg sulfurhdyr in Norrland is justified by the fact that a great deal of acid accumulated in the snow pack during winter is released during spring snow melt. Similarly, the critical load for nitrogen varies from 5 kg nitrogen/ha/yr in lowproductive ecosystems mainly in the north to 15 kg nitrogenlhdyr in highly productive ecosystems in the south. Thus, the target load for nitrogen deposition is set a t 10 kg nitrogenhalyr in Gotaland, 8 k g h d y r in Svealand, and 6 kghalyr in Norrland, since growth and accompanying uptake and utilization of nitrogen is smaller in the north.

50 I

These target loads for sulfur and nitrogen have now been accepted by the Parliament as environmental goals for Sweden. Progress toward these target loads is planned by the years 1995 and 2000 as outlined in Table 10. Table 10 Emission reduction targets for sulfur, nitrogen compounds, and volatile organic compounds in Sweden by the years 1995 and 2000

Pollutant

Reference year

Total sulfur Nitrogen oxides Ammonia Volatile organic compounds

1980 1980 1990 1988

Emission reduction target pe reduct&: Year 1995 Year 2000

65 30 25b

80 50a 50b 50

aParliament has ordered a further evaluation of this target load. bThese reductions in emissions will apply mainly in southern and southwestern Sweden. Alternative strategies for achieving these reductions are currently being evaluated.

Finland The first critical loads maps for acidity, sulfur, and nitrogen and their exceedences were published in the national document for Finland. The Finish Council of State decided in 1991 to decrease sulfur emissions by 80% by about 2000. According to model estimates, the critical load for sulfur will be achieved if Finland and other countries decrease their emissions by this amount. If so, only very small areas in Finland will still be threatened by acidification. The possible continuing effects of nitrate and ammonium deposition are causing some uncertainty about this optimistic scenario.

United Kingdom Drawing on the work of various review groups in the UK, critical load maps have been prepared for acid deposition effects on soils. Similar maps also have been prepared for acidification of fresh waters, so far only for Scotland. National surveys are currently in progress throughout the UK. Generally, critical loads for soils are lower than critical loads for fresh waters. Critical load exceedence maps also have been prepared for much of the UK. There is generally good agreement between the areas where exceedences are shown on these maps and areas where acidification impacts have been reported.

502

Provisional maps of target loads for sulfur have been prepared by the government. Abatement actions are being taken t o meet the requirements of the European Community's emissions standards. Even though sulfur deposition will have been reduced by the year 2005, some of the more sensitive ecosystems in the UK will still be sustaining damage. These still-unprotected areas with regard to soil acidification constitute about 8% of the area of the UK. Catlilda

In order to protect surface waters from acidification, the critical loads approach was used in establishing the agreement among the provinces that sulfur dioxide emissions in Canada should be decreased by 50% by 1994. The critical load for sulfur in all of eastern Canada was estimated to be 20 kg sulfatehdyr (7 kg sulfurhdyr). Integrated atmospheric, water quality, and aquatic biota models were used to establish emissions reduction targets and to identify the least-cost means by which these targets could be achieved. A variety of federal, provincial, international political, engineering, and social factors were considered in reaching the final emission control decisions shown in Table 11. Table 11 Reductions in sulfur emissions in various Canadian provinces intended to meet the target of a n overall 50% decrease in sulfur emissions in eastern Canada by 1994, % reduction in each province

Province Manitoba Ontario Quebec New Brunswick Nova Scotia Newfoundland Prince Edward Island

Based on 1980 regulated emissions 25 83 45 14 7 2A 17

Based on 1980 actual emissions 0

50 45

16 0

20 0

The critical loads approach has not been used for other pollutants and/or resources a t risk in Canada although some effort has been made with respect to sulfur-induced changes in soil chemistry.

USA The concepts of critical loads and target loads have not been addressed in the national document for the USA.

503 4.0. USE OF EMISSIONS-REDUCTIONSCENARIOS

-

Many of the national documents used emissions-reduction scenarios a s a technique of assessment, but the extent of use varied from country to country:

Use of scenarios in assessment

NL

S

3

1

S

F

U

3

K

1

North A m e m CAN USA 2

3

The Netherlands In the national document for The Netherlands, long-term potentials for management of acidifying deposition were explored using the Dutch Acidification Systems Model (DAS). To illustrate this approach, the results of three illustrative scenarios are presented in Table 12. The deposition targets presented in these scenarios are derived from explorations of possible emissions decreases in The Netherlands and other countries in Europe. Four features of these three scenarios are important: Changes in amounts of deposition. In all three scenarios, a 54% decrease in potential acid deposition is proposed by the year 2000. This probably would be achieved mainly by decreasing emissions of ammonia. In scenario 2, further decreases totaling 71% and 75% are proposed In scenario 3 further decreases totaling 75% and 85% are proposed by 2010 and 2050. Expected changes in forest soils. Significant decreases in amounts of acidifying deposition generally lead to rapid improvement in the soilsolution chemistry of forests. Change from a n average annual deposition of 4800 to 2200 mol H+/hafyr are expected t o decrease the area exceeding the critical aluminum concentration and the aluminudcalcium ratio from about 75% of the forest soil area to 40%. In scenario 2 the exceedence is expected to be negligible by the year 2050 and thus depletion of the aluminium buffer system probably would be prevented. Expected changes in forest trees. The most significant positive effects on forests are expected to be lower nitrogen content of leaves and decreased leaching of nitrate, both of which probably will make the trees less susceptible to frost, insect pests, and fungal pathogens. Increased growth of trees (compared to unchanged deposition) is also expected in the southeastern part of The Netherlands. Decreasing deposition before the year 2000 is expected to have larger beneficial effects than decreases after the year 2000.

504

Expected changes in heathlands. According to the DAS model, scenario 1 offers no prospect for the continued existence of typical dry heathland vegetation, while scenario 3 does. Wet heathland vegetation is expected to be successful under scenarios 2 and 3 if sod-cutting also is applied. Table 12 Three possible scenarios of change in management of acidifying deposition and expected effects on forest soils in The Netherlands. Part 4:Potential acid deposition targets set by the government. Part 8: Changes in percentage of forest land area where critical soil-solution concentrations probably would be exceeded, 1990-2050. Current vear 1990

h€tA Scenario 1 Scenario 2 Scenario 3

Scenario 1 Scenario 2 Scenario 3

Future Years

m

2010

2050

(Target potential acid deposition)a 4800 2200 2200 2200 4800 1400 1200 2200 4800 2200 1200 700 (Percentage of forest land area where critical soil-solution concentrations would be exceeded$ 75 40 40 25 75 40 a0 0 75 40 ? 0

apotential acid deposition includes SOX + NOx + NHx (mols H+/ha/yr). bThe critical soil-solution concentrations selected for this analysis included aluminum concentration, calciudaluminum ratio, ammoniudpotassium ratio, and nitrate concentration. These estimates were based on calculations using the Dutch Acidification Systems Model (DAS).

Sweden The national document for Sweden contains no systematic analysis of alternative emission or deposition scenarios. Environmental goals for the nation have been established mainly through consensus-forming processes by scientists and policy experts assembled by the Swedish Environment Protection Board. Mathematical models have been used for analysis of effects on surface waters. These models indicate that a 50% decrease in acid loading beginning in the year 2000 and continuing to 2030 would be required to bring about half the

505

acidified lakes in southern and central Sweden to pH 6.0 or higher. In order to achieve the same effect in the most sensitive parts of southern Sweden, a decrease in acid loading of at least 80% would be required beginning in the 1990s. These forecasts mean that liming of acidified lakes must continue for several decades. If present rates of sulfur and nitrogen deposition are not decreased, acidification and associated depletion of base cations in forest soils are expected to continue for several decades. This is projected to cause long-term decreases in productivity of large areas of forest in southern and central Sweden. In addition to decreases in deposition, both liming and nutrient supplementation are expected to be necessary to avoid impoverishment of forest soils. Also, a t present rates of deposition, substantial areas of southern Sweden are projected to suffer from nitrogen saturation within 10-20 years. This would result in continuing soil acidification, leaching of nitrate, nutrient imbalances in forest trees, and major changes in species composition of natural flora and fauna. Continuing acidification of soils and surface and ground waters also is expected to pose a continuing threat to human health. But much more research is needed to improve present understanding of these possibilities.

Finland Finland developed a n integrated acidification assessment model called HAKOMA. The model was then used to evaluate alternative energy-use and specific emissions-control scenarios. Both basic scenarios for Finland are adaptations of the present energy system. The first involves only the current emissions-reduction plans. The second involves maximum feasible reductions in emissions. Specific methods by which to decrease emissions of sulfur dioxide, nitrogen oxides, and ammonia also have been explored. Increasing use of natural gas and enhanced energy conservation are among the options studied. The RAINS model also has been used to analyze the effects on Finland of similar emissions-reduction options in the rest of Europe. More detailed analyses have been used to estimate changes in deposition within Finland that might result from different energy-production options in the Russian Republic and the other Baltic States including Estonia, Latvia, Lithuania, and Poland. Analyses of relative cost-effectiveness also have been made for various methods of decreasing emissions. The following conclusions were drawn from these scenario analyses in Finland: The critical loads for forest soils and lakes are presently exceeded in large parts of Finland; Maximum feasible reductions in emissions both in Finland and elsewhere in Europe will be necessary t o substantially decrease total acid deposition in Finland;

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Even maximum feasible decreases in emissions of sulfur in Finland would not be sufficient to stop acidification of forest soils in Finland; The critical load for forest soils could be achieved in most areas of Finland if maximum feasible reductions were applied both in Finland and abroad; Uncertainty about NHx emissions in Finland and abroad are the principal causes of uncertainty in estimates of future acidifying deposition in Finland.

United Kingdom There is no comprehensive analysis of future emissions-reduction scenarios for the UK. Modeling of future trends is an important part of the Review Group Report on Acidity in Fresh Waters, however. These models indicate that: Restoration of most surface waters to their pristine state probably would require a decrease in sulfur deposition of about 90%. Within a few years, many waters would be expected t o respond positively to such a drastic change. A few waters, however, are almost certainly irreversibly acidified. A decrease in emissions of about 50% would be required to prevent further biological changes in many moorland surface waters in Wales.

Because of the "air-filtering" effect of forest canopies, forest cover increases the total transfer of acidifying deposition from the atmosphere by about 30%. For this reason, the effect of planting new forests in a n upland moorland catchment is a sensitive region is equivalent to a 30% increase in acidifying deposition. Canada

In the national documents for Canada, four scenarios were compared. We describe two of
507

large parts of Ontario and Quebec, but there will be little change in wet deposition of sulfate in the most acid-sensitive regions of the Atlantic Provinces. Similarly, as shown in Table 13, aquatic-chemistry and aquatic-biota models predict a substantial improvement in both lake chemistry and lake biota in large parts of Ontario and Quebec. Much less substantial improvement is expected in certain aggregations of tertiary watersheds (AG) in parts of Newfoundland and Labrador. It is not clear whether the effects of these projected decreases in acid loadings could be documented in terrestrial ecosystems. Table 13 Predicted percentage of lakes with a pH <6.0 in certain aggregations of tertiary watersheds (AG) in eastern Canada under the two emissions-reduction scenarios

Scenario 1 Scenario 4

Ontario AG 19

Quebec AG 14

Newfoundland AG 7

Labrador

38 10

31 10

47 47

22 16

AG 10

USA Detailed comparisons of several illustrative emissions scenarios were included in the 1990 NAPAP Integrated Assessment for the USA. Expected results from the two main illustrative scenarios (S1 and S4) are discussed below and illustrated in Table 14: Scenario S1-- No new emissions regulations beyond those in effect in 1985; Scenario S4 -- 48% (10 million-ton) decrease in emissions of sulfur dioxide by the year 2000.

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Table 14 Total annual sulfur deposition and expected changes in lake water quality in selected regions of the USA under two illustrative emissions scenarios (S1 and s4) Sulfur der, .osition. k g m Eastern USA

Appalachian Mountaiu

Year

s1

s4

s1

SQ

1980 1985

11

11

12

14 9 9

7

2010 2030

11

7 7

14 16 14

2050

7

6

7

m

Adirondack lakes (% with acidneutralizing w t v
s1

s4

14

14

22 19

10 9 6

8

The following conclusions were drawn from these scenario analyses for the USA: Under scenario 4, by the year 2000 annual deposition of sulfur is expected to decrease by 36% compared to 1985. Also, the cumulative deposition of sulfur from 2000 t o 2030 is expected to be 33% less under scenario 54 than under scenario S1. By the year 2030, only small differences in annual rates of sulfur deposition are expected between S4 and S1. Under S4, however, lake water chemistry is expected to improve about 20 years sooner than under S1. Among all regions of the USA, the most substantial improvements in lake water chemistry are expected in the Adirondack region. With regard to terrestrial effects, the health of red spruce trees affected by winter injury is expected to improve. Also, a decrease in sulfur deposition of 50% over 10 years is expected to cause a slight (2-4%) increase in base saturation in some shallow forest soils, but most soils will not be affected greatly. With regard to human health effects, under scenario 54 the lead content of surface water and ground water is expected to decrease. Acidic sulfate aerosols in urban areas are projected to decrease by 30-50% Reduced health risks due t o exposure to gaseous sulfur dioxide would occur only for human populations close to large point sources.

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Under scenario S4, visual range in the eastern USA is estimated to improve by 20-30%. The atmospheric process, water chemistry, aquatic biota, soils, and visibility models used to make these predictions depend on many assumptions and embody many uncertainties.

5. LESSONS LEARNED FROM THIS COMPARISON OF NATIONAL AssEssMENTs Lesson 1. Importance ofMultidisciplinaryand Policy-FocusedAppmachee Complex environmental problems such a s acidification and air pollution require for their solution both: A multi-disciplinary scientific approach; and An integrated and policy-focused research and assessment program.

Leason 2. Benefits of International Cooperation International cooperation often enhances several features of integrated research and assessment programs. These features include: Scientific quality and integrity, Policy relevancy of scientific findings, Public understanding and acceptance of assessment findings, and Personal satisfactions that scientists, policy analysts, and decision makers in industry and government, as well as communications specialists, find by becoming full participants in their respective parts of the whole research and assessment program.

Leason 3. Importanceof Public Pemptions of Values The national documents sometimes differed from each other because of differences in public perceptions about societal values within various countries.

Example 1. Visibility impairment is included in the national document for the USA but is barely touched upon in the national documents for The Netherlands, Sweden, the UK, and Canada. From this it appears that citizens of the USA place a higher value on scenic vistas than citizens of some other countries. Example 2 . Sulfate aerosols are more thoroughly discussed in the national documents for the USA than in those for other countries. Medical authorities in the USA apparently consider respirable particulate matter to be a more significant hazard to public health than similar experts in other countries.

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Example 3. Effects of acidifying deposition on sport fishing is more fully developed in the national documents for Sweden, Finland, the UK, Canada, and the USA than for The Netherlands. Because of severe pollution of its major river and canal systems, sport fishing is less common, and for that reason, probably less highly valued by citizens in The Netherlands than in many other countries. Lesson 4. Importance ofTransboundaryFlows of Pollutants The national documents for various countries sometimes differed from each other because of differences in the fractional part of atmospheric deposition that derives from sources outside the nation itself. Examples: Transboundary aspects of pollution climate are discussed more fully in the national documents for The Netherlands, Sweden, Finland, and Canada, than in those for the UK and the USA.

Leason 5. Capitalizing on Competence in Science and Policy M y & Differences in the distribution of competence in science, engineering, and policy-analysis among federal, state/provincial, university, a n d industrial specialists, and the degree of engagement of this competence with acidifying deposition and other air-pollution problems were evident in the national documents for various countries.

Leason 6. Achieving SignificantInput to Policy Discussions The effectiveness of national assessment documents i n facilitating development of public policy is determined in part by the following: Degree of public and political consensus about the seriousness of the problem and the solutions that are being considered; Quality and continuity of scientific and policy leadership; Freedom from political interference in development of scientific and assessment findings and development of policy recommendations; Extent and quality of inter-agency, multi-disciplinary, and inter-institutional cooperation and communication; Fiscal and managerial authority that are commensurate with responsibility; and Skill in articulation and timely communication of scientific and assessment findings.

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Lesson 7. Conceptual Importance of the Words we Use "The quality of our scientific understanding is revealed by the words we commonly use." This generalization is demonstrated by the frequency with which some specific terms and concepts appear in the various national assessment documents. Example 1 . Acidic deposition

vs

Acidifying deposition

Example 2. Acid deposition

vs

Acidification

Example 3. Acidic deposition

vs

Atmospheric deposition

Example 4. Annual average deposition vs

Cumulative deposition

Example 5. Chemical meteorology

vs

Chemical climatology

Example 6. Buffering capacity

vs

Acid neutralizing capacity

Example 7. Forest health

vs

Forest productivity

Example 8. Tree health

vs

Ecosystem Health

Example 9. Forest condition

VS

Genetic diversity

Example 10. Nitrogen nutrition

vs

Nitrogen saturation

Example 11. Waldsterben

VS

Neuartige Waldschiiden (New types of forest damages)

Example 12. Critical loads approach

vs

Illustrative scenario approach

Example 13. Air-quality related values

vs

Air pollution problems

Lesson 8. Conclusions from Forestry Research In the national assessment documents from all countries, damage to forests was attributed to an unexpected convergence of: natural stresses, and/or human disturbance stresses, and/or air pollution stresses, acting simultaneously o r sequentially. Damage to forests by air pollutants usually occurred because the pollutants predisposed the tree (increased their susceptibility) to: biological stresses such as fungi or insects, and/or climatic stresses such as drought or frost.

512

Detrimental changes in the visible condition of some trees in a forest sometimes occurred in stands where the productivity of the whole forest remained unchanged (or even increased!).

Lessons. Importance of EfFects of Concern The major "effects of concern" varied greatly from one country to another. Example 1 . Effects of acidifying deposition on surface waters appeared to be much more important in Sweden, Finland, the UK, Canada, and the USA than in The Netherlands. Example 2 . Effects of acidifying deposition on heathland appeared to be much more important in The Netherlands and the UK than in Sweden, Canada, or the USA. Example 3 . Acidification of forest soils has been found in large parts of Sweden and Finland and in certain parts of The Netherlands and the UK. It also has been detected in a few locations in Canada and the USA. For these reasons, effects of acidifying deposition on forest soils is considered to be the most significant long-term threat to the health and productivity of forests in The Netherlands, Sweden, Finland, and the UK and is considered to be a potential long-term threat to forests in Canada and the USA. Example 4 . Effects of acidifying deposition on agricultural crops were assumed to be unimportant in The Netherlands, Sweden, and the UK and were proven to be unimportant in the USA and Canada. Example 5 . Ozone has been proven to be a significant cause of economic losses in agricultural crops in the USA and Canada. Example 6. Ozone is considered to be the most significant current threat to forests in North America, while it is considered only a contributing stress to forests in most countries in Europe.

Lesson 10. PhilomphicalAspects of Assessment Processes Some philosophical differences were found i n the national assessment documents. Example 1 . The assessment documents for The Netherlands, Sweden, the UK, and Canada did not provide an explicit definition of criteria for evaluation of cause-and-effect relationships. By contrast, the assessment document for the USA defined three criteria (patterns of linkage) for evaluation of cause-and-effect relationships:

513

"The criterion of consistency requires that a temporal and/or spatial correlation be found between field measurements of the suspected cause (or causes) and field measurements of observed effects." "The criterion of responsiveness requires that a quantitative relationship be found in controlled experiments between exposure to the suspected cause (or causes) and the amount or extent of measured effects. Also the effects observed in controlled tests must be similar to those observed in the field." "The criterion of mechanism requires that there be a physical, chemical, and/or biological process (or series of such processes) by which the observed effects could reasonably be induced by the suspected cause (or causes)." In reviewing the national documents from various countries, we believe that explicit identification of criteria for analysis of cause-and-effect is highly desirable in principle. In practice, however, we recognize that the rigorous use of these (or other) criteria also has certain disadvantages. For example, controlled experiments frequently are relatively short in duration and therefore make it difficult to discover cumulative and longterm effects (e.g., on soils). Also, controlled experiments frequently are conducted with unrealistically high doses and thus may lead to conclusions about possible causes that may o r may not hold under more realistic conditions. Finally, controlled experiments with whole ecosystems or with very large, mature trees sometimes a r e nearly impossible to perform reliably.

Example 2. The assessment document for the USA also included a system for explicit treatment of certainty and uncertainty in assessment findings: "A system of rank-ordered stars is used to communicate the degree of scientific consensus associated with summary statements and competing hypotheses. This "Star System" provides a format for communicating extensive, technically complex information about uncertainty. Definitions of Star Rankings for answers to the Integrated Assessment Questions: o - Unsupported hypothesis * - Best judgment - low confidence ** - Best judgment - limited confidence *** - Reasonable confidence **** - High confidence." We believe that this (or some other) ranking system for rough quantification of uncertainty should be used more widely i n future programs of policy relevant research and assessment. We also call attention to the Guidelines for the Formulation of Scientific Findings to be Used for Policy Purposes that are contained in the report of the NAPAP Oversight Review Board (NAPAP, 1991~).

5 14

Lesson 11. Use of the CriticalLoads Approach Large differences in use of, knowledge about, and/or acceptance of the concepts of critical loads, critical levels, and target loads were discernable among the several national assessment documents.

Example 1 . The assessment document for The Netherlands included a chapter dealing specifically with the concepts of critical loads and critical levels: "The critical load concept constitutes, together with a technology-based approach, the two track approach that is the basis for the development and adoption of the Netherlands' Acidification Abatement Plan." The assessment document for The Netherlands also included critical loads for acids and sulfur and nitrogen compounds a s well a s critical levels for ozone with regard to effects on forests, heathlands, and groundwater. Example 2 . Similarly, the assessment document for Sweden included critical loads for sulfur and nitrogen compounds as well as critical levels for ozone with respect to effects on surface waters, forest and crop lands, soils, and public health. The environmental goals for Sweden use these critical loads and levels as the foundation for recommendations of target loads for emissions of sulfur oxides and nitrogen oxides both for Sweden and Europe as a whole. Example 3. The assessment documents for the UK also dealt explicitly with the concept of critical loads with regard to effects on soils and surface waters. They also include maps showing areas within the UK where critical loads for acidity and sulfur are exceeded. The Netherlands and the UK now appear to be providing significant leadership in developing and applying the critical loads approach for both national and international decision making. Example 4 . The assessment document for Canada applied the concept of critical loads and target loads to the major "effect of concern" in Canada -acidification of surface waters resulting from atmospheric deposition of sulfur compounds. These concepts were used a s the foundation for recommending target loads that vary from c8 kgfhdyr to 20 k g h d y r of wet sulfate deposition for various parts of eastern Canada including the Atlantic Provinces, Labrador, Quebec, and Ontario. Example 5 . The assessment document for the USA contains no significant discussion of the concepts of critical loads, critical levels, or target loads. Instead it contains a series of illustrative emissions scenarios mainly dealing with sulfur emissions and secondarily with nitrogen and volatile organic substances. These illustrative scenarios are suggested a s the foundation for future governmental and industrial decisions about emissions targets.

515

Lesson 1 2 Understandingthe Critical Loads Approach Critical loads and critical levels are matters of scientific judgement: They describe quantitatively "What the nature can stand". Target loads, on the other hand, are matters of social, economic, and political judgement : They describe quantitatively "What society prefers" o r "What society believes it should do". Bearing in mind these matters of scientific judgement on one hand and matters of social, economic, and political judgement on the other, three factors should be considered in establishing target loads: The critical loads and/or critical levels for each of the resources a t risk and the values held dear by that society; Various technical means and managerial mechanisms by which a wide range of possible target loads might be achieved (alternative scenarios); and, finally, A searching evaluation of these loads and/or levels, means and mechanisms, and values held dear by society, in the light of relevant social, economic, ethical, technological, and political benefits and costs to all sectors of society.

6. SUMMARY

This paper has been prepared to encourage a continuing process of learning about the interface between science and environmental decision making in various countries. Governments and industries in many countries of Europe and North America have invested substantial sums of money, human resources, imagination, and creativity in national programs of research and assessment on the issue of pollutant-induced acidification of our environment. These programs have substantially advanced human understanding of this phenomenon and its effects in various parts of the world. It has been a pleasure for us to compare and contrast these several national programs. We hope that our initial efforts will encourage others to join in this enterprise by: drawing additional inferences about the phenomena and effects of acidification; and learning additional lessons about how science can be made more useful in making decisions about the management of nature and the human institutions that seek to use it wisely. As indicated earlier in this paper, we look forward eagerly to both personal communications and published reactions to our efforts from colleagues in all countries who are willing to help improve our collective understanding of these environmental problems and possibilities for their solution.

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7.l3EFmmvcm Cowling, E. B. Acid precipitation in historical perspective. Env. Sci. & Tech. 16:llOA-123A. DOE. 1988. The Effects of Acid Deposition on the Terrestrial Environment in the UK. Department of the Environment, United Kingdom, London. DOE. 1989a. The Effects of Acid Deposition on Buildings and Building Materials. Department of the Environment, United Kingdom, London. DOE. 1989b. Acidity in United Kingdom Fresh Waters. Department of the Environment, United Kingdom, London. DOE. 1990a. Acid Deposition in the UK 1986-1988. Department of the Environment, United Kingdom, London. DOE. 1990b. Oxides of Nitrogen in the UK. Department of the Environment, United Kingdom, London. DOE. 1991. Critical and Target Loads Maps for the UK. Department of the Environment, United Kingdom, London. Doughty, C. R. 1990. Acidity in Scottish Rivers. A Chemical and Biological Survey. Edited version of a report prepared for the Department of the Environment. Scottish River Purification Boards. Edinburgh. FPRMCC. 1990. The 1990 Long-Range Transport of Air Pollutants and Acid Deposition Assessment Report. Part 1 Executive Summary. Part 2 Emissions and Controls. Part 3 Atmospheric Sciences. Part 4 Aquatic Effects. Part 5 Terrestrial Effects. Part 6 Human Health Effects. Part 7 Socio-EconomicStudies. Part 8 Quality Assurance studies. Federal/Provincial Research and Monitoring Coordination Committee, Environment Canada, Ottawa. Heij, G . J., and T. Schneider. 1991. Acidification Research i n The Netherlands. Final Report of the Dutch Priority Programme on Acidification. Studies in Environmental Science 46. Elsevier Science Publishers, Amsterdam. Kauppi, P., P. Anttila, and K. Kenttamies, Eds. 1990. Acidification in Finland. Springer-Verlag, Berlin.

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Malanchuk, J. L., and J. Nilsson, Eds. 1989. The Role of Nitrogen in the Acidification of Soils and Surface Waters. Nordic Council of Ministers, Copenhagen. Nilsson, J., and P. Grennfelt, Eds. 1988. Critical Loads for Sulphur and Nitrogen. Reprint of the workshop held a t Skokloster, Sweden, 1988. Nordic Council of Ministers, Copenhagen. NAPAP. 1991a. Acid Deposition: State of Science and Technology. Volume I Emissions, Atmospheric Processes and Deposition. Volume I1 Aquatic Processes and Effects. Volume I11 Terrestrial, Materials, Health, and Visibility Effects. Volume IV Control Technologies, Future Emissions, and Effects Valuation. National Acid Precipitation Assessment Program, Washington DC. NAPAP. 1991b. 1990 Integrated Assessment Report. Precipitation Assessment Program, Washington DC.

National Acid

NAPAP. 1991c. The Experience and Legacy of NAPAP. Report of the Oversight Review Board to the Joint Chairs Council of the Interagency Task Force on Acidic Precipitation. National Acid Precipitation Assessment Program, Washington DC. Overrein, L. N., H. M. Seip, and A Tollen, Eds. 1981. Acid Precipitation -Effects on Forest and Fish. Final Report of the SNSF-Project 1972-1980. Ministry of the Environment, Oslo.

SNV. 1990. Air Pollution '90 -- Action Programme for Air Pollution and Acidification. Swedish Environmental Protection Agency, Stockholm. S N V . 1991. Mercury in the Environment: Problems and Remedial Measures. Swedish Environmental Protection Agency, Stockholm.

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T Schneider (Editor). Acidification Research Evaluation and Policy Appllcations 0 1992 Elsevier Science Publishers B V All rlghts reserved

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The Dutch Acidification Systems (DAS) model: The emissions and air transport modules. K.F. de Boer, J.W. Erisman, F.A.A.M. de Leeuw, T.N. Olsthoom and R. Thomas National Institute of Public Health and Environmental Protection (RIVM), P.O. Box 1, 3720 BA Bilthoven, The Netherlands

Abstract This poster gives an outline of two modules of the DAS model: the emissions module and the air transport module. The poster presents some graphic examples, which are not repeated here.

1. THE EMISSIONS MODULE OF DAS The emissions module of the DAS model serves as the head of the modelled causality chain (from emissions up to ecological effects). The module uses national total Dutch and foreign emissions of from several economic sectors as input. The main sources of data are: For past and present emissions (1950 - 1990): National Bureau of Statistics (CBS) RIVM/RIS@ (Dutch and Danish national research institutes) ECE/EMEP (European database of emissions) National statistics of all countries considered For future emissions ( 1 990 - 2050): Dutch National Environmental Policy Plan ( N W ) [ 11 ECEEMEP [2] The air transport module of DAS requires emissions from 20 Dutch and 19 European so-called acidification areas [3]. The Dutch acidification areas are about 60x60 km in size, using administrative borders, the European areas are larger the less they contribute to the deposition of acidifying compounds on the Netherlands. The total Dutch emissions are reallocated to emissions per area based on the TNO Emission Inventory System [4].Reallocation includes the effective height of emissions in one of the three classes "low", "medium" and "high". Emission height classes are associated with the economic sectors. Powerplants and refineries are associated with high emissions, while agriculture, traffic and households are considered as low emittors. The other sectors fall into the medium category.

522

2. THE AIR TRANSPORT MODULE OF DAS. The air transport and transformation module is focussed on SO,, NO, and NH, and their derivatives. In the transport module the averaged deposition for each of the 20 acidification areas in the Netherlands is calculated, using the emission data from the 39 source areas. The emissions and depositions are related to each other through transfer matrices. These matrices have been generated by means of the long-range transport model TREND [5, 61. This statistical model includes transport, dispersion, chemical transfomation and deposition of pollutants. For distances close to a source, the effective source height of emissions strongly determines the contribution to the deposition. Different sets of transfer matrices are derived for the three emissions height classes "low", "medium" and "high". Using the transfer matrices, depositions are calculated for an averaged Dutch landscape. The average dry deposition on forests, heathland or heathland pools is different due to different roughness lengths. To compensate for this effect, the calculated dry depositions are multiplied by empirically derived, compound and region specific correction factors to estimate the dry deposition flux on these systems [7].

REFERENCES 1. 2. 3.

4. 5.

6.

7.

Ministry of Housing, Physical Planning and Environment. 'To choose or to lose; the National Environmental Policy Plan', Second Chamber, session 1988-1989, 21137, nos 1-2 Economic Commission for Europe. 'National strategies and policies for air pollution abatement', United Nations, Geneva/New York, 1987 A.H. Bakema, K.F.de Boer, G.W. Bultman, J.J.M. van Grinsven, C. van Heerden, R.M. Kok, J. Kros, J.G. van Minnen, G.M.J. Mohren, T.N. Olsthoorn, W. de Vries and F.G. Wortelboer. 'Dutch Acidification Systems Model - Specifications', Dutch Priority Programme on Acidification, report no. 114.1-01, 1990 TNO. 'Gegevens Emissieregistratie le en 2e ronde' (Data from the Emission Inventory System, first and second inventory), TNO, Delft, 1985 J.A. van Jaarsveld and D. Onderdelinden. 'TREND; and analytical long-term deposition model for multi-scale purposes', report no. 228603009, National Institute of Public Health and Environmental Protection, 1990 W.A.H. Asman and J.A. van Jaarsveld. 'A variable-resolution statistical transport model for ammonia and ammonium', report no. 228471007, National Institute of Public Health and Environmental Protection, 1990 J.W. Erisman. 'Acid deposition in the Netherlands', report no 723001002, National Institute of Public Health and Environmental Protection. 1991

T Schneider (Editor). Acidification Research Evaluation and Policy Applications 0 1992 Elsevier Science Publishers B V All rights resewed

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The relationship between research and policy on acidification impacts in the nature conservation agencies of Great Britain Andrew M Farmer English Nature, Northminster House, Peterborough, PEI IUA, UK.

1 Nature Conservation Agencies in Great Britain Until April 1991 Britain had one government nature conservation agency - the Nature Conservancy Council (NCC). In April this was split into three agencies covering England, Wales and Scotland - English Nature, Countryside Council for Wales and the Nature Conservancy Council (Scotland). English Nature advises the UK Government on nature conservation in England. It promotes the conservation of England’s wildlife and natural features in a UK and international context. It selects, establishes and manages National Nature Reserves and notifies Sites of Special Scientific Interest (SSSIs), i.e. areas of particular conservation interest. English Nature provides advice and information about conservation and supports and conducts research relevant to these functions. 2 Acidification Research by NCC and English Nature

In the early 1980s the extent of damage to natural ecosystems by acid rain was very poorly understood. NCC highlighted this form of pollution as a potentially major threat to the nature conservation resource of Britain. A number of early studies were initiated which were aimed at identifying whether acid rain was causing damage. This was done in conjunction with other bodies funding research, so that NCC focused on particular habitats and species of nature conservation importance. Once a relationship between pollutant and damage was established, it became necessary to identify the geographical extent of these effects. A series of surveys or monitoring programmes was undertaken. Again this has focused on the nature conservation resource base. Thus the current surveying of freshwaters to identify sensitive and potentially acidified water bodies is limited to those which occur on SSSIs. Areas of research and results are given in Woodin and Farmer (1991j.

3 Policy Implications of Research Findings The results gathered from the research funded by the conservation agencies and other bodies is brought together for a number of purposes. The way that information is gathered and used by English Nature on pollution research is shown below.

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1. Information Gathering - Research, Survey and Monitoring: Commissioned Research. Field Work by English Nature Staff. Research by Outside Bodies. 2. Information Processing in English Nature: Interaction between Science and Policy Directorates. 3. Primary Output: Policy Decisions. Advice. Publications.

4.Targets: Central Government. Local Authorities and Regulatory Bodies. Land Managers. Advice to English Nature Regional Staff. Industry. Other Conservation Organisations. Other Scientists. Educationalists. General Public. The main target objectives for the use of this information are: to influence opinion and to contribute to discussion in the policy making community; to encourage good practice with land managers, industry and regulatory authorities; to foster the inclusion of nature conservation concerns in the agenda of others; to fulfil legal duties. Advice to government on pollution damage can used to seek that it initiates pollution reduction policies that safeguard the nature conservation resource base. Thus critical loads and exceedence maps reveal that even after implementation of the EC LCP directive in 2003, a large area of upland Britain will continue to be acidified. Much NCC research focused on the uplands and many SSSIs occur here. Current emission reduction targets will not be sufficient to safeguard a significant part of the nature conservation resource base of this part of Britain. Also needed is advice to our regional staff responsible for the day to day management of sites. The required advice is often different from that needed by government and may not relate to areas of pollutant research that most researchers currently focus on, for example, the siting of new industrial plants, etc., requires advice especially on the local effects SO2 and particulate deposition on a wide range of species and habitats. However, much of this information is lacking and there is little current research in these areas. 4 Conclusions English Nature pollution research is mostly influenced by two of its target audiences. This is primarily from government with the need to provide information on major effects, extent of damage, etc, with less aimed at responding to regional staff. A review is needed of how well research priorities reflect the needs of the target audiences of English Nature. 5 Reference Woodin, S.J. and Farmer, A.M. (Eds) (1991) The effects of acid deposition on nature conservation in Great Britain. Nature Conservancy Council, Peterborough.

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The Dutch Acidification Systems (DAS) model: The forest Module SOILVEG J.J.M. van Grinsven, C. van Heerden and J.G. van Minnen National Institute of Public Health and Environmental Protection, P.O. Box 1, 3720 BA, Bilthoven, The Netherlands

Abstract Within the framework of the Dutch Priority Programme on Acidification, the SOILVEG model has been developed. The SOILVEG model predicts growth and phenology of Douglas fir in relation to availability of N, Ca, Mg and K in the root zone, ambient concentrations of NH,, SO,, O,, and acid atmospheric deposition. Douglas fir is a representative tree for production forest. The effects are integrated for the tree using the concept of nutrient and carbon limited growth. The behaviour of leaves and roots with respect to gas and nutrient exchange is strongly aggregated. The effects of SO,, 0, on photosynthesis and of pH and Al on root uptake, were described by means of dose-effect relationships. The calculation of allocation and growth from the availability of carbon and nutrients and the actual nutrient status and phenology is a dynamic procedure, which mimics the various hierarchical decision processes in the tree, aimed a t optimally using available resources. The model was applied to evaluate the effects of different emission reduction strategies on growth of Douglas fir for the major sandy soil types in the Netherlands. The present regionalization is still very preliminary as only deposition and air quality were regionalized. Simulations show that differences of tree growth and nutrient status between soils, scenarios and regions are most strongly related to differences of external N-inputs. N plays a key role in direct impact of air pollution via leaf uptake of NH,, and in indirect impact on the soil acidification status via proton production by nitrification and NH, uptake. The role of N for tree growth in the Netherlands is ambiguous. Low emission scenarios lead to N-shortage and growth reduction, eg. in coastal regions, while high emission scenarios cause Nsurplus, growth reduction and enhanced soil acidification in the south-eastern regions. Simulations indicate that reduction of photosynthesis by direct impact of SO, and 0, will not exceed 5%. However, reduction of Mg uptake by roots, due to low pH can amount to 50% and reduction due to high Al levels in soil solution, can amount to 20%. But compared to reductions at the start of simulations in 1950, the maximum additional reduction is only 10%. In the present version of SOILVEG forest damage due to drought, frost and plagues are not included. This omission leads to an underestimation of forest damage. Moreover, the sensitivity of forest to these "traditional" factors tends to increase with increased leaf N contents and reduction of the fine root biomass, which both are typical effects of the "new" stress factors. Therefore, traditional and new stress mechanisms are now combined in SOILVEG.

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A modelled assessment of critical load exceedences for sulphur over the United Kingdom G W Campbell and J G Irwin Warren Spring Laboratory, Department of Trade and Industry, Gunnels Wood Road, Stevenage SG1 2BX, United Kingdom Abstract Inputs of sulphur through dry and wet deposition to the United Kingdom have been modelled for the years 1955, 1987 and for a future scenario representative of the period 2000-2005. The results have been compared with estimated critical loads for soils.

1. THE MODEL The model is a statistical one which allows for oxidation of SO2 to sulphate aerosol, dry deposition of SO2,and wet removal of sulphur. Short-range dispersion ( < 50km)is described by a Gaussian plume model using meteorological data from the nearest of 50 sites. Over longer distances vertical dispersion is represented by a height-dependent eddy diffusivity. Emissions from large point sources are treated individually while area sources are based on a lOxlOkm grid. Contributions from other European sources and unattributable background inputs in westerly air masses are considered separately'. 2. EMISSIONS

From 1955 to 1987 UK emissions of S q decreased from 2.6 to 1.8 Mtonnes (as S). New larger power stations and the closure of smaller units in urban areas, together with reduced domestic and industrial emissions, led to decreased emissions over most of the country and an increase in the average height of emission. In estimating future emissions a 'middle economic growth' scenario was selected. Power station emissions were calculated taking account of the European Community Large Combustion Plant Directive, the anticipated lifetime of present stations and an increasing use of natural gas and low sulphur coal. This gave a total emission of 1.0 Mt (as S) with high-level sources accounting for about 60%. The contribution from the rest of Europe was calculated for 1987 by running the EMEP model without UK emissions; the 1955 contribution was taken to be the same as in 1987. For the future scenario transfer functions were used, assuming implementation of the LCP Directive in EC countries but no decrease in emissions in Eastern Europe. A background concentration of lOueql-' was used in all three cases based on trajectory analysis of precipitation composition at remote sites in the west of the United Kingdom2.

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3. EXCEEDENCE MAPS The overall change over the c50 year period is shown in Figure 1. In 1955, the critical loading for sulphur was exceeded by >0.2 keq ha-'over 77% of the land area; by 1987 this figure had decreased to 49%. The future scenario predicts a further decrease to 17% by early next century. It should be noted that this map has not been produced on the same basis as the published United Kingdom target loads map. Differences in assumptions about future emissions, atmospheric transport and, importantly, background deposition mean that the two cannot be compared directly. Clearly, there are very considerable uncertainties in these calculations. For example, orographic enhancement of wet deposition is not included in the model and may lead to exceedences in mountainous areas. Similarily, changes in background inputs over time and with latitude require investigation.

Figure 1. Areas where critical loads for sulphur are exceeded by >0.2 keq ha-'. 4. REFERENCES

1 Heyes CA, Irwin JG, Pemn DA (1991) Modelling sulphur concentrations and depositions in the UK in 1987 and 1955. Stevenage: Warren Spring Laboratory, LR 772. 2 Stedman J (1991) Measurements of background sulphur and scavenging ratios at a site in the west of Northern Ireland. Amos Environ, 25A, 699-708. 5. ACKNOWLEDGEMENT

The funding of this work by the United Kingdom Department of the Environment within their air pollution research programme (contract no PECDI7I 10/29) is gratefully acknowledged.

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DRY DEPOSITION OVER GRASSLAND: SEASONAL INFLUENCES, CHEMICAL EQUILIBRIA AND SURFACE WETNESS M.A.H.G.Plantaz, A.T.Vermeulen, P.J.de Wild, G.P.Wyers and J.Slanina Netherlands Energy Research Foundation ECN, The Netherlands

Dry deposition of NH3, NH4NO3 and HNO3 over flag, homogeneous grassland will be measured under different atmospheric and surface conditions. Special attention will be paid to periods with dew occurrence. The measurements will cover a period of one year (four seasons) and will be continuous and automated up to a high degree. The aim is to provide the modelling of dry deposition with respect to influences: - of meteorological and seasonal conditions; - of the presence of water layers on the surface; - of chemical equilibria. Deposition fluxes of the target components will be derived from vertical concentration profiles (1 and 5 m measured by thermo denuder systems), using both the gradient method and a modified Bowen-ratio technique. Therefore, profiles of temperature, windspeed and moisture are also measured, besides radiation (netto and global), soil heat flux and soil heat storage. From these, fluxes of sensible and latent heat and of momentum are calculated, and by combination with the concentration profiles-, the component fluxes. Under dew conditions, dew samples are also taken and analyzed, and dew volumes per square meter are measured. As profile-related methods fail under such stable conditions, fluxes of target components are estimated by supposing an analogous exchange behaviour with H20(,); therefore fluxes of H20g) (and of sensible heat and momentum) are also measured directly using the eddy-correlation technique. This combination of atmospheric flux estimates and dew water analysis will yield information on the effects of water layers on deposition of HN3, NH4N03 and HNO3 and on the equilibria that are involved in different stages of the deposition process. The allseason coverage and the combination of methods will improve knowledge and modelling of meteorological (seasonal) influences. The experiment is planned for 1992; until then, the set up will be tested at Schagerbrug, near ECN. Preliminary data, obtained during the testing period, will be presented.

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The Dutch Acidification systems (DAS) model A. Tiktak, A.H. Bakema, K.F. de Boer, R.M. Kok and T.N. Olsthoorn National Institute of Public Health and Environmental Protection, PO Box 1, 3720 BA BILTHOVEN, The Netherlands

Abstract The Dutch Acidification Systems (DAS) model is presented. The model aims a t evaluating the long-term effectiveness of acidification abatement strategies on soils, forests, heathlands, aquatic systems, agricultural production, monuments and construction materials in a regionalized way. Some results for the soil, forest and heathland modules are presented. Within the framework of the Dutch Priority Programme on Acidification, the Dutch Acidification Systems (DAS) Model has been developed. The model aims a t evaluating the long-term effectiveness of acidification abatement strategies on a number of receptor systems in the Netherlands and describes the entire causality chain from emissions to effects in a regionalized way. The time-span for the model is one hundred years, starting in 1950. DAS contains modules for the calculation of emissions, air transport and effects on soils, forests, heathlands, aquatic systems, agriculture, monuments and construction materials. Each module can be run independently. The individual modules are described in detail in accompanying posters. The air pollution components included are the acidification agents NH,, NO, and SO,. As some of the effect modules use additional components, such as ozone, the base cations (Ca'', M e and K') and the anion Cl-, these components have also been included. Results are presented using emission data for the past (1950-1990) and the near future (1990-2000). Future emission data are based on acidification abatement strategies as announced in the National Dutch Environmental Policy Plan Plus (NEPP'). No information about emissions is available for the period after the year 2000. Instead, three deposition targets, which aim a t an acid deposition of 2200 (scenario l), 1230 (scenario 2) and 700 (scenario 3) mol,.ha-'.yr-' have been formulated for the year 2050. The air transport and transformation module SRM calculates the deposition and concentration of the acidifying components by means of transfer matrices. The matrices include a linear relationship between emissions and concentrations or deposition of a compound. Different matrices are available for compounds from low (<50 meters), medium (50-100 meters) and high (>lo0 meters) sources. The module generates depositions for an average Dutch landscape. However, the dry deposition to forests is about 20% higher than the average deposition. To account for this, the average dry deposition is multiplied

532

by compound and region specific correction factors which are derived empirically. Using the emission and air transport modules of DAS, the average potential acid deposition to the Netherlands will be 2200 mol,.ha".yr-' in the year 2000, which is a reduction of more than 50% compared to the deposition in 1989 (which was 4800 rnol.ha".yr-'). The strong deposition reduction up to 2000 is caused by a dramatic reduction of the NH, emission in the Netherlands, but considerable reductions are also expected for NO, and SO,. The soil module of DAS is the process orientated REgional Soil Acidification (RESAM) model. RESAM describes changes in soil chemistry, both in the solid phase (minerals, Al-hydroxides and adsorption complex) and in the liquid phase. Processes included are weathering of primary minerals and Al-hydroxides, cation exchange reactions, nitrogen transformations and nutrient cycling by the vegetation. The model includes the major components in A13+,Ca", M P , K+, Na+, NH,+, NO,', SO:., Cl, HCO; and forest soils, i.e. H+, RCOO-. Deposition reduction leads to a fast improvement in soil solution chemistry, i.e. a n increase in the pH value and a reduction in A13+and NO, concentrations and the A13+/Ca2+and NH,+/K+ratios. The exceedance of the critical A13+ concentration from about 75% of the forest soil area (the current situation) reduces to 40% in the year 2000, the exceedance of the critical A13+/Ca2+ ratio reduces from about 65% to 40%. Deposition reductions according to scenario 2 and 3 cause a decrease in the exceedance of these parameters to negligible values for the year 2050. In scenario 1, the critical A13+concentration and A13+/Ca2+ ratio will still be exceeded in a considerable area of forest (25 and 10%) respectively) a t that time. The forest module of DAS is the model SOILVEG. This model predicts growth and phenology of Douglas fir in relation to the availability of carbon from photosynthesis and N, Ca, Mg and K from atmospheric deposition and soil processes. Indirect and direct effects of acid deposition are included. The model has been parameterized for Douglas fir stands on four sandy soil types. The scenario analysis for these forests indicates that the nitrogen content of needles is reduced significantly. As a result, the sensitivity of Douglas for plagues, frost and drought stress will reduce. In the South-Eastern part of the Netherlands the increase of nitrogen and total acid deposition between 1970 and 1990 leads to a net decrease of needle mass and wood production. In relatively clean areas, however, there is a positive effect of enhanced nitrogen deposition on needle mass and wood production. The reduction of photosynthesis by direct impact of SO, and 0, will not exceed 5%. However, additional reduction of M e uptake by roots, due to low pH can amount to 15% and reduction of M e uptake due to high aluminum concentrations can amount to 20%. The heathland module CALLUNA simulates the nitrogen cycle and the competition between Calluna vulgaris and Deschampsia flexuosa for nitrogen. Simulations indicate that scenario 1 does not offer any prospects for the continued existence of dry heathland vegetations, whereas scenario 3 offers good prospects for such vegetations.

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LONG-TERM IMPACT OF THREE DEPOSITION SCENARIOS ON DUTCH FOREST SOILS W. de Vries, J. Kros, C. van der Salm and J.C.H. Voogd DLO The Winand Staring Centre for Integrated Land, Soil and Water Research (SC-DLO), P.O. Box 125, 6700 AC Wageningen, The Netherlands

The long-term impact of three deposition scenarios on Dutch forest soils has been evaluated using the model RESAM (Regional Soil Acidification Model). RESAM describes changes in the chemical composition of the soil solid phase (minerals and adsorption complex) and soil solution due to natural and maninduced processes. The model is part of the overall Dutch Acidification Systems model (DAS). Deposition scenarios for total acidity (SO,, NO, and NH,) have been generated by the DAS air transport model for the period 1990 to 2050 based on expected emissions in the near future (1990-2000) and deposition targets (2000-2050) as shown in Table 1. Table 1 Average deposition trends of total acidity for the Netherlands for three scenarios during the period 1990-2050. Scenario

1 2 3

Acid deposition (mol, ha-' yr-') 1990

2000

2010

2050

4420 4420 4420

2240 2240 2240

2240 1400 1230

2240 1230 700

The regional application of RESAM has been restricted t o seven tree species and fourteen non-calcareous sandy soil covering about 65% of the Dutch forest area. The Netherlands has been subdivided in twenty deposition areas. Presentation of the model results is restricted t o pH, A1 concentration, AVCa ratio and NH,/K ratio in the topsoil (top 20 to 30 cm) and to pH, Al, NO, and SO, concentration in the subsoil (at the bottom of the rootzone). The parameters in the topsoil are important indicators of forest stress, whereas the parameters in the subsoil are important indicators of potential groundwater pollution. In order to gain insight in the reliability of the model predictions, model results of the soil solution chemistry were compared with soil solution

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measurements in 150 forest stands during the period March to May 1990. The model results and field data agreed well for pH and SO,; they were fair for the A1 concentration (topsoil), AWCa ratio and NO, concentration and unfavourable for the NH4/K ratio and A1 concentration (subsoil). Future trends in soil solution parameters in response t o three emissiondeposition scenarios for the period 1990-2050, showed that deposition reductions generally lead to a fast improvement of the soil solution chemistry. It leads to an increase in pH and a decrease in A1 and NO, concentration and AUCa and NH,/K ratio. However, for the NO, concentration and NHJK ratio there was a clear time lag between deposition reduction and concentration reduction which is mainly due t o N mobilization from the litter layer. The effect of deposition reductions according to scenario 2 in some relatively densily forested deposition areas is shown in Table 2. Table 2 Decrease in areas exceeding a critical A1 concentration (0.2 mol, m-,) and a critical Al/Ca ratio (1.0 mol mar') in response to scenario 2 during the period. Deposition area

Area exceeding critical [All

Area exceeding critical AYCa

(%I

(%I

1990 2000 2010 2050

1990 2000 2010 2050

Drenthe Noord Overijssel Twente Veluwe Achterhoek Eindhoven Venray

98 62 86 69 87 98 61

50 40 64 21 70 70 40

15 17 46 1 60 27

1 5 23 0 30 14 12

84 56 90 65 88 79 59

Netherlands

80

45

23

8

70

32

-_ 51 62 38 81 69 35

4 19 48 6 62 33 23

0 0 3 0 4 0

51

22

1

33

3

Table 2 shows that a deposition reduction to 2240 mol, ha-' yr" in the year 2000, will reduce the exceedance of a critical A1 concentration and AVCa ratio from 80% and 70% of the forest soil area (current situation) t o 51% and 45% respectively. Reductions up to 1400 mol, ha-' yr-' in 2010 will reduce t h e exceedance of a critical A1 concentration and Al/Ca ratio in the topsoil to less than 25%. However, there are large differences in the various areas. In 2050 t h e exceedance of these criteria will be almost negligible.

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SOIL AND SOIL SOLUTION COMPOSITION OF 150 FOREST STANDS IN THE NETHERLANDS IN 1990

W. de Vries, E.E.J.M. Leeters, C.M. Hendriks, W. Balkema, M.M.T. Meulenbrugge, R. Zwijnen and J.C.H. Voogd DLO The Winand Staring Centre for Integrated Land, Soil and Water Research (SC-DLO), P.O. Box 125, 6700 AC Wageningen, The Netherlands

During the period March to May 1990, the chemical composition of the humus layer, the mineral topsoil (0-30 cm) and the mineral subsoil (60-100 cm) has been determined for 150 forest stands. All stands were part of the national forest vitality inventory. They were all located on non-calcareous sandy soils. Tree species included were Scotch Pine, Black Pine, Douglas Fir, Norway Spruce, Japanese Larch, Oak and Beech. Measurements for the humus layer and mineral topsoil included total contents of C, N and P and exchangeable contents of H, Al, Fe, Ca, Mg, K, Na and NH,. Soil solution measurements in the mineral soil included the cations mentioned before and NO,, SO,, C1, HCO, and DOC. An important aim of the research was to assess the effect of deposition levels, tree species and site characteristics on the level of A1 mobilization (acidification) and N accumulation (eutrophication) in the soil. Important characteristics of the chemical composition of the humus layer and mineral topsoil are shown in Table 1. Table 1 Median values for the N contents in organic matter and exchangeable base cation and A1 contents in the humus layer and mineral topsoil of 150 forest stands Tree species

Scotch Pine Black Pine Douglas Fir Norway Spruce Japanese Larch Oak Beech All species

N contents (%) _______ humus topsoil layer

Base saturation (%)

A1 saturation (%)

humus layer

humus layer

2.0 1.9 2.4 2.2 2.1 2.6 2.4 2.2

34 36 38 38 42 46 34 38

1.9 2.0 2.2

1.9 1.8 2.4 2.0 2.0

-__

topsoil

topsoil

6 5

69

7

67 63 68 43 64 66

6

5 17 6 6

70

536

The N contents in the organic matter of the humus layer were somewhat higher than in the mineral topsoil. Values in the humus layer were much higher than in relatively unpolluted areas, where the N content was generally less than 1.5%.This suggests N immobilization caused by high N deposition. In the humus layer, base saturation was high (about 30-70%)and A1 saturation was low (about 3 t o 13%) compared t o the mineral topsoil. This is due to base cation mineralization and a negligible A1 mobilization in this layer. In the topsoil, base saturation was low (often below 20%) whereas A1 saturation was high (about 30 t o 80%). This implies that acid input is hardly buffered by exchange with base cations. Except for the Oak, the variation between tree species appeared to be small. Important characteristics of the soil solution composition of the mineral topsoil are shown in Table 2. Table 2 Median values of important soil solution parameters in the mineral topsoil of 150 forest stands __-__

Tree species

Scotch Pine Black Pine Douglas Fir Norway Spruce Japanese Larch Oak Beech All species

___

Concentrations ( m ~ l , . m ' ~ )

Ratios (mol.mo1-l)

__

pH

A1

Ca

SO,

NO,

NH,/NO,

NH,/K

AVCa

3.60 3.67 3.41 3.43 3.55 3.70 3.67 3.59

0.81 0.62 1.43 1.16 0.81 0.40 0.52 0.69

0.39 0.36 0.90 0.58 0.47 0.51 0.31 0.45

1.09 0.80 2.56 2.04 1.02 0.83 0.65 1.03

0.56 0.27 0.92 0.50 0.71 0.54 0.33 0.55

0.35 0.75 0.70 1.18 0.46 0.25 0.60 0.45

1.oo 1.19 2.56 2.36 1.67 0.54 0.60 1.12

1.22 1.27 1.42 1.38 1.15 0.63 1.09 1.13

Low pH values and high concentrations in NO,, SO, and A1 occured below all tree species, especially below Douglas Fir and Norway Spruce. Most probably this is due to a high dry deposition on - and a high evapotranspiration of these tree species. The ratio of NH, t o NO, was generally below 1.0, indicating a reasonable rate of nitrification. Consequently, the ratio of NH, to K remained below a critical value of 5. The ratio of A1 to Ca was generally above 1.0, which is considered a critical value. The median NO,/SO, ratio was about 0.6. In the Netherlands, the N/S ratio in the input is a t least 1.0. Assuming that SO, behaves a s a tracer, this implies that a considerable amount of nitrogen is fixed or disappears as a result of uptake, immobilization and/or denitrification.

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AUTOMATED DENUDER SYSTEMS FOR DRY DEPOSITION STUDIES OF ACIDIFYING COMPOUNDS G.P.Wyers, A.T.Vermeulen, R.P.Otjes, A.Wayers, J.J.Mols and J.Slanina Netherlands Energy Research Foundation ECN, The Netherlands

Sampling of atmospheric trace gases using conventional denuders is a laborious task. Since it involves many manipulations in the field and the laboratory, the precision of this technique is limited. Automation of the technique increases its precision and offers the possibility to perform measurements over a long period which allows a study of seasonal and daily influences on concentrations and fluxes. Initially, emphasis was placed on the development of thermodenuders. Thermodenuders for ammonia, in which N H 3 is collected on a V205 coating and, partly, converted to NO, by heating and detected by an NO, monitor, were used to study dry deposition over a forest (Speulderbos) for a period of five months. The relative precision of these instruments is 5% for concentrations of 1-20 g/m3. Another line of development involved the so-called wet denuders, in which trace gases are collected in a solution film on the walls of an annular, rotating denuder. The denuder is automatically emptied and refilled. The samples are stored by a fraction collector and analyzed off-line for N H 3 , HNO3, HN02, HCl, H202 and S02. One of these instruments was used to measure the concentrations of these trace gases for a full year. A major advantage of wet denuders is that they feature quantitative collection of trace gases over a large concentration range. Recently a wet denuder has been constructed in which the absorption solution is continuously refreshed and on-line analyzed for ammonia. With the continuous-flow rotating denuder a very high precision can be obtained (better than 1% for 20-minute measurements) which makes it well-suited for measurement of, usually very small, concentration gradients for dry deposition studies. In a recent joint field experiment at a heather field (Leenderheide), organised within the framework of the EUROTRAC-subproject BIATEX, three of these instruments were used to measure profiles of ammonia above the field. Preliminary results will be presented.

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AQUACID: Modelling the acidification of shallow heathland lakes in The Netherlands. The aquatic systems module of DAS. F.G. Wortelboer Laboratory for Water and Drinking Water, National Institute of Public Health and Environmental Protection, P.O. Box 1,3720 BA Bilthoven, The Netherlands. The model AQUACID describes the changes in the ecosystems of shallow heathland lakes in The Netherlands as induced by high levels of atmospheric deposition. The characteristics of the lakes themselves led to the formulation of a new model, instead of using one of the more widely known models. Most important factor was the complete hydrological isolation of these lakes from surface water and groundwater, the lakes receiving atmospheric input only. Furthermore, changes in the macrophyte vegetation and accumulation of organic matter in the sediment, along with acidification, were recognized as important aspects of the lakes and should be described by the model. For this, the inclusion of various biological and biochemical processes into the model was needed. The heathland lakes are found in sandy soils with low calcium contents. Low amounts of nutrients and organic material used to be the normal situation. They are therefore weakly buffered and acid-sensitive. The surface area ranges from 1 to 100 ha, with a n average depth of 1m. The heathland lakes formerly had a very low production of biomass. Deposition only brought minor amounts of nutrients. Vegetation was dominated by submersed plants, such as Littorella uniflora,Lobelia dortrnanna, Isoetes lacustris. Due to acidification, the pH ofthe water dropped from about 6.5 to less than 4, the lakes showing an increase in concentrations of carbon dioxide and the nutrients nitrate and ammonium, and an accumulation of organic matter in the sediment. The vegetation has become dominated by Juncus bulbosus and Sphagnum spp. Acidification is seen as a balance (whether or not disturbed) between acidity and alkalinity producing processes. These processes are regarded the most important with respect to the ability of these aquatic systems to recover from acidification. The model simulates the competition between the characteristic submerged macrophyte species Littorella uniflora and Juncus bulbosus. The modelling is based upon the ecophysiology of the species themselves, which differ in their productivity, the site of CO, and nutrient uptake (sediment vs. water), the source of nitrogen (nitrate vs. ammonium; Wortelboer, 1990). Because of differences in the degradability of the organic matter produced by the macrophytes species, four organic matter fractions were distinguished: of each species, both a labile and a refractory fraction. Other processes included in the model are aerobic mineralization, denitrification, sulphate reduction, methanogenesis, oxidation of ammonium, sulphide and methane. Interaction with the lake surroundings takes place through exchange with the atmosphere, wet and dry deposition, and the input of allochtonous organic matter. A schematic representation of the model is given in Figure 1.(See also Wortelboer, 1992.)

540

Figure 1.Scheme representing the main characteristics of the model AQUACID. The model, consisting of a set of non-linear ordinary differential equations, was implemented using the interactive modelling environment FAME (Wortelboer & Aldenberg, 1992), in which the chemical equilibria calculations were incorporated into the integration algorithm. Calculations were performed over simulation periods of 100 years. The input of allochtonous organic matter is necessary to refrain the system to become carbon deficient, as the carbon dioxide is continuously depleted by exchange with the atmosphere. The presence of macrophyte vegetation depends on carbon dioxide levels in both water and sediment. The speciation of nitrogen in the system (and thus in the deposition) is important in determining the species composition of the vegetation. Furthermore, the results indicate the importance of the mineralization processes in neutralizing the acidifying effect of the primary production by the macrophytes. Results of scenario analyses to establish the critical deposition levels will be published elsewhere.

References Wortelboer, F.G., 1990. A model on the competition between two macrophyte species in acidifying shallow soft-water lakes in The Netherlands. Hydrobiol. Bull. 24:91-107. Wortelboer, F.G., 1992. AQUACID: Acidification model of shallow soft-water lakes in The Netherlands. National Institute of Public Health and Environmental Protection, Bilthoven, in prep. Wortelboer, F.G. & T. Aldenberg, 1992. FAME: Friendly Applied Modelling Environment. Version 3.0 User Manual, National Institute of Public Health and Environmental Protection, Bilthoven, report no. 779200001.

T. Schneider (Editor). Acidification Research. Evaluation and Policy Applications

0 1992 Elsevier Science Publishers B.V. All rights reserved

AN INTERNATIONAL RESEARCH PROGRAM ON ACID RAIN AND EMISSIONS IN ASIA Wesley K. Foell Resource Management Associates of Madison, 520 University Avenue, Suite 300, Madison, Wisconsin, 53703. U.S.A. FAX: (608) 283-2881. Abstract Strong economic and population growth will accelerate the rapid development of fossil fuel energy systems throughout Asia. If the present trends continue, by early next century Asian emissions of SO, will exceed the present emission levels in North America and Europe combined. In response to the concern that these emissions have the potential to cause significant damage in Asia, a group of international specialists has established a project on Acid Rain and Emissions in Asia. In the initial phase of this project, work is underway to develop an integrated assessment model to assist policy makers in evaluating options to reduce precursor emissions and to catalyze the process of international policy dialogue on acid rain in Asia. 1. INTRODUCTION

Fossil fuels are now used in large quantities throughout Asia. These present systems, however, are dwarfed by the substantial fossil fuel systems planned in many Asian nations over the next two decades. The accelerated development of these energy systems is propelled by the very high economic and population growth rates of the AsiaPacific region. The vast expansion of these energy systems combined with plans for a major fuel shift to coal in a number of predominant countries, will result in greatly increased atmospheric emissions of a wide range of substances of concern, including most prominently, acidifying compounds and greenhouse gases. In addition to their global implications, the long-term and regional/local implications of these atmospheric emissions touch not only the natural environment, but also have far-reaching implications for important commercial and cultural activities such as forestry, agriculture, and tourism. In response to the growing concern about acid rain and emissions of other atmospheric pollutants, a group of international specialists convened a workshop on this topic in Bangkok in 1989. In November 1990 in Bangkok, an expanded group of prominent experts attended a second workshop under the broader sponsorship of the World Bank, the United Nations Environment Program (UNEP), the Economic and Social Commission for Asia and the Pacific (ESCAP), and the US. Dept. of Energy (DOE). The participants had the objective of reviewing the problem in more depth, and if

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appropriate, to lay out a detailed program and plan of action'. As a result of the current and future situation in Asia, the Workshop concluded that acidifying compounds and other related emissions from energy facilities have the potential to cause (and in some locations are already causing) significant damage in Asia. In China for example, there already appears to be considerable evidence of the effects (Zhao et al., 1988). The participants were greatly concerned by the implications of the planned initiatives for accelerated development of massive fossil-fuel energy systems in many Asian nations, which could entail greatly increased emissions of acid deposition precursors. Because these would likely be emitted from large energy faculties with high stacks, they could increase the transport of these acidifymg substances to sensitive geographical areas. Developing and implementing approaches to reduce anthropogenic emissions will require the efforts of many organizations, both public and private. Currently, there is little regional coordination. However, as national regulators and the development community become increasingly aware of this issue and the consequences of no action become clear, sound programs and projects could be developed in the region. In order to reduce the emissions of acid precursors, the Asian countries will have to shift their energy policy through: Expanded use of cleaner fuels and renewable energy sources 0 0 Upgrading of older inefficient facilities 0 Instituting energy efficiency measures Industrial restructuring for regional energy efficiency 0 Installation of pollution control devices 0 Strategies to reduce acidifying emissions will also result in reduction of other pollutants such as greenhouse gases. There are two international initiatives under way to address greenhouse gas emissions: the Intergovernmental Panel on Climate Change (IPCC), under the auspices of UNEP and World Meteorological Organization (WMO); and the Global Environmental Facility (GEF), administered by the World Bank in association with UNEP and UNDP. Similar initiatives for acid rain in Asia may be needed because of the regional and global implications and to obtain regional agreement. Based on this conclusion, the Workshop recommended the development of a specific research project and plan of action. In addition the Workshop strongly recommended that an integrative program be developed for the purpose of formulating and assessing long-term strategies to confront the emerging problems of acid rain and the anthropogenic atmospheric emissions of a wide range of substances of concern at both the national and Asia-wide levels. 2. NEED FOR AN ACID RAIN PROGRAM IN ASIA

Our concern about the acid rain problem in Asia is heavily influenced by

'The Third Annual Workshop on Acid Rain and Emissions in Asia was held in Bangkok at the Asian Institute of Technology from November 18-22, 1991. Although the detailed output from that workshop has not been incorporated into this report, a few of the more general programmatic ideas regarding the future direction of the project's collaborative efforts have been noted in this paper.

543 consideration of Asia's population, its growing economy, and the associated systems of energy consumption and production. Growth of both population and economy are powerful driving forces behind the rapidly growing use of fossil fuels whose combustion is ultimately responsible for the emissions leading to acid deposition. Both the size and the growth rates of Asia's population and economy intensify this concern. Asia currently accounts for more than 55% of the world population and by the year 2010, according to United Nation scenarios of population growth rates, could be close to four billion people. The economic growth rate of Asia is far higher than for any other world region. Japan is currently responsible for more than half of the region's income. However, as is well-known, the economic growth rates of several of the industrializing countries of the region are extremely high and it is quite likely that several countries will establish developed industrialized economies early in the next century if not before. If current trends continue through the end of the century, the Western Pacific countries would have an economy comparable in size to those of North America or Western Europe. The acid rain impact in Asia may already be significant (Rodhe et ul., 1988; Bhatti et aL, 1990). The expansion of these energy systems, propelled by the very high economic and population growth rates of the region, combined with a major fuel shift to indigenous coal, will result in increased atmospheric emissions of acidifymg compounds and greenhouse gases. The total projected SO, emissions for the Asian countries in year 2010 of 76 million tons exceed the current emissions of North America and Eastern/Western Europe combined. Table 1

Emissions of SO2 in million tons per year Region/Country 1985* Europe 53 U.S.A. 21 Asia 28 China 19 India 3 * 1986 estimates for Asia

1990 50 24 35

2000 39 25 53 34 5

2010 39 21 76 49 9

Europe: Based on RAINS "currentreduction plan" scenario U.S.A.: Based on NAPAF' Integrated Assesrnent, Draft Sept. 1990 Asia: Base case scenario, Foe11 and Green, 1990.

An overview of key elements of the acid rain situation in Asia has been briefly sketched in Figure 1 (Bhatti et ul., 1990). Relative sulfur dioxide emissions in 1985 by country or region are indicated by vertical bars. Emissions are significant in China, India, Japan, and the Koreas, and are small but very rapidly growing in several other

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countries of south and southeast Asia. Deposition patterns of emissions are largely determined by wind flow patterns, shown by arrows in the figure. In winter (January) the flow is generally from the land mass to the ocean, while in summer (July) the reverse is true. Typical monitored values of pH are shown in the circles. Low pH values (4.5) occur in Japan and southern China where emissions are large; elsewhere the pH values are 6 to 7. Because of the paucity of monitoring sites, we cannot discount the possibility of low-pH at other unsuspected locations. In Asian countries where the recognition of this phenomenon is just beginning, there is very limited information available on the possible consequences of acid deposition. Some of the ecosystems of the Asian region are very similar to those found in areas of North America and Europe where the majority of the research on acid deposition impacts has been conducted. Thus, in some cases it should be possible to extrapolate the potential effects found in these western nations to the corresponding ecosystems of Asia. However, many other environments in Asia are very different from those for which acid deposition impacts have been studied, and research will have to be conducted to analyze and assess the relative vulnerability of these various anthropogenic and natural environments to acidic inputs. Nevertheless, based on knowledge of the various components (i.e., soils, flora, fauna, climate, etc.) of the ecosystems in question, their relative vulnerability to acidic inputs and their distribution in the Asian region, it is possible to attempt to predict the potential effects of acidic deposition (Bhatti et al., 1990). For a given area to be at high risk from acidification, a number of conditions must be present simultaneously. First, the area must be receiving or at risk of receiving high levels of deposition. Second, the soils of the area must be sensitive to acidification, and the area must have vegetation, fauna, aquatic organisms, man-made materials and/or large human populations vulnerable to increased inputs of acidity. In light of the present and future situation in Asia, the time has come to develop an integrated program of assessment and policy analysis for the purpose of analyzing long-term strategies for acid rain problems at national and at Asia-wide levels. A key component of such a program entails the application of the analytic and policy development experience gained in the West to this emerging issue in Asia. A predictive tool could be built to help decision-makers project future trends in emissions, estimate the regional consequences for acid deposition levels, evaluate the vulnerability of natural and artificial systems, and determine the costs and effectiveness of alternative mitigative actions that might be taken. Such a policy analysis exercise can start to raise environmental awareness in the region and begin a dialogue that could help ameliorate (or prevent the worsening of) an environmental problem in the early stages. 3. ENERGY AND EMISSIONS 3.1 Current Situation

The current energy system which has evolved to fuel Asia’s economic development is striking in many respects. First, it is almost completely dependent on fossil fuels (more than 95% of its commercial energy). Second, approximately 60% of the consumed fossil fuel is in the form of coal, with most of the remainder being fuel oil. Both of these fuels contribute significantly to the emissions of acid rain precursors

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in Asia. Equally important from an environmental perspective are the extremely high growth rates of Asian energy systems. During the past decade these annual rates have greatly exceeded (by factors of 2 to 4) those in Western industrialized countries, which are now relatively stable at a few percent or less. There is little reason to expect that per capita energy use in the developing countries of Asia will stabilize anywhere near its present level of approximately 0.5 tons of oil equivalent per year, which is approximately ten percent of the per capita use in North America and Europe. If the Western experience can be used as a guide, even with strong conservation measures we can expect continued very high energy growth in Asia for many years. Of particular concern is the continued growth of coal use. Coal use in the region in 1987 was approximately 1360 million metric tons with the three largest users being China, India, and Japan. Preliminary data and estimates presented at the 1989 Workshop indicated that regional coal use in the year 2000 could be as high as 2,300 million metric tons, with very large growth increments in China and India, and significant growth rates in several other countries, including Indonesia, Thailand and the Koreas. In addition to the concern about the level of total emissions early in the next century, we anticipate increasingly serious problems associated with the high density of emissions sources from a number of localized areas. Among those areas of concern are eastern and central India, northeast and southwest China, the Korean peninsula and Japan, northern and possibly southern Thailand, and western Java. The majority of these emissions would likely result from uncontrolled coal combustion and/or motor vehicle use. 3.2 Framework for Emissions Analysis Described briefly below, is a framework developed by Foe11 and Green to calculate a range of emissions scenarios for 13 developing Asian countries for a range of energy scenarios based on a variety of socio-economic and technical assumptions. From these energy scenarios, national values for sulfur emissions from fossil-fuel combustion were calculated, based on fuel characteristics, combustion technology, and emission control assumptions. The objective of the analysis was to calculate SO, emissions resulting from fossil fuel consumption in a base year (1986) and two future years, 2000 and 2010. Energy use in the scenarios is disaggregated into 5 sectors (industry, residential/commercial/agricultural, transportation, energy sector energy use, and other) and three fuel types (coal, oil, and natural gas) for each country. A uniform emission factor for SO, is applied to each fuel type in each sector. Such methodology is often adopted in national studies to account for the variations in sulfur, ash, and energy content within fuel types and differences in combustion technology. Additionally, the emission factor used will be influenced by the control strategy adopted for each sector. The resultant emissions are then summed to give the total SO, emissions for each country. The energy consumption scenarios which were developed for this analysis were in general based upon scenarios taken directly from the literature, e.g., from a countryreport developed by an energy agency in the country of interest. This approach was used as much as possible for the so-called base case scenarios. Wherever possible, both the assumptions and the resulting energy consumption patterns were compared with

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available related studies. In essentially every case, the "alternative scenarios" required a considerable amount of supplementary analysis because of the need to introduce different assumptions on a sectoral or fuel-by-fuel basis. The general scenario assumptions are described below. (For a more detailed description, see Foell and Green, 1990). 3.3 Scenario Assumptions Four main scenarios were developed. s1: Bme Case Scenario. s2: EfFciency Case: Energy improvements are introduced by sector into the residential, industrial, and transportation sectors as well as the electricity distribution network. Significant improvements over the base case values were introduced progressively in 2000 and 2010. These improvements were in general on the order of 15 to 30 percent depending on the country and sector involved and were assumed to be in part price-induced and in part through other interventions. A further assumption, probably reasonable in comparison with the gross assumptions made for such a long time horizon, is that these efficiency improvements occur equally across all industrial fuels, i.e., they are fuel independent. s3: Low Carbon Emission Care: This scenario combines the energy efficiency improvements of S2 and shifts to fuels with low carbon and acid-precursor emissions. The study examined the availability of natural gas, and where judged appropriate, assumed substitution for a fraction of the coal or oil used for electricity generation in the year 2010. In addition, new and renewable electric generating sources which are free from all precursor emissions were introduced in that year to a small degree. s4: Control Scenario: Sulfur controls were assumed to be implemented on electric power plants in a phased approach. In this approach, 50 percent of all new power generation between 1986 and 2000 was assumed to have 90 percent sulfur removal; all new power generation between 2000 and 2010 was assumed to have 90 percent sulfur removal. The controls were assumed to be applied to the Base Case energy and emission figures. 3.4 Scenario Results Of particular interest in the results of this analysis are the continued high growth rate of primary energy consumption and the continued dependance on fossil fuel energy sources, particularly coal and oil. In aggregate, commercial energy consumption for the region as a whole in the Base Case scenario is expected to grow at above 4.5 percent per year from 1986 through 2000 and only slightly less than 4.5 percent per year from 2000-2010. As a result, primary commercial energy is expected to reach approximately 2000 million tons of oil equivalent (MTOE) in 2000 and 3000 MTOE in 2010. Even when considering significant improvements in energy efficiency, energy growth in the region remains at relatively high levels. Efficiency improvements assumed in S2, the Efficiency Scenario, result in a net savings of about 250 MTOE in 2000 and 600 MTOE in 2010 from the base case. However energy growth rates for both periods remain at about 3.5 percent per year, still far higher than those of industrialized western countries.

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Also of interest is the continued dominance of China’s energy use in the regional totals. In 1986, China was responsible for nearly 55 percent of the region’s total primary energy, Despite the assumption of a significant decrease in China’s overall growth rate for energy (from 5.2 percent in 1980-1987 to 3.9 percent in 2000-2010), China is expected to comprise 48 percent of the region’s total energy use in 2010. Coal has supplied about 75 percent of China’s commercial energy needs during the 1980’s, and this situation is not expected to change during the 1990’s or into the next century.

ASIA REGION SO2 EMISSIONS 8o 70

8

Ease Case S1

Effie. Case 52 Low Carbon 53

Control S4

1

l1 o09 8 5

1990

1995

2000

2005

2010

Figure 2 Projected SO, emissions for the Asia Region The regional sulfur dioxide emissions are shown in Figure 2 for the Base Case and for the three alternative scenarios. The total emissions for 1986 are approximately 28 million tons of SO,. For comparison, the estimated U.S. sulfur dioxide emissions in 1985 were 21 million tons. Notable in 1986 were the emissions of China (19 million tons), India (3.2 million tons), and the Republic of Korea (1.2 million tons). By far the largest portion of the 1986 emissions is from coal-based facilities. The base scenario shows very high growth rates of the emissions, as might be expected from the scenario assumptions described above, and in particular the continued or growing reliance on coal in several of the major countries. Coal is responsible for the largest part of the emissions and, in both the base year and in 2010, accounts for approximately two-thirds of the total, followed by petroleum which accounts for around thirty percent. The contribution of natural gas is very small. By the years 2000 and 2010, total emissions for the 13 countries have grown to 53 and 76 millions tons respectively. Annual growth rates of SO, are 4.6 and 3.7 percent in the periods 1986-2000 and 2000-2010, respectively. Of this total, coal is responsible for 35 million tons in the year 2000 and 50 million tons in the year 2010; nearly 65 percent of

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the total emissions by 2010. The sulfur control strategy, S4, leads to a reduction of 3 million and 12 million tons in 2000 and 2010, respectively. Interestingly, this pollution control strategy, although quite ambitious in its rate of implementation, does not ultimately lead to the magnitude of reduction achieved by efficiency improvements and fuel shifts. Of course, the economics of these alternatives need to be investigated in some detail. If efficiency improvements, fuel shifts, and emissions controls were all implemented, it appears that the emissions could possibly be stabilized between 2000 and 2010. However, reduction in the absolute level would require more drastic measures. 4. DESIGN OF A RESEARCH PROGRAM

Environmental problems that have emerged in the last two decades can be characterized by two important factors: they are triggered by rapid socioeconomic development; and the spatial resolution of their impacts is usually well beyond the local and often even the national scale. These two important factors strongly influence the basis for building an analysis and policy framework for problems such as acidification and global climate change. In addressing the first factor, socioeconomic development, attention should be paid not only to the costs of abating but also to the costs of not abating. Especially in the developing countries of Asia, it is important to point out that benefits of abatement do exist. Concurrent with the explicit recognition of these benefits has been the emergence of the concept of sustainability, as introduced by the World Commission on Environment and Development. Development policies, which in the long run are not sustainable because of their increasingly negative impacts on the productivity of human and natural systems, are proving extremely costly for industrialized countries. The second factor, namely, the increasing tendency of environmental impacts to extend their domain to points at greater distances from the pollution source, has been strongly demonstrated by the nature of the acidification problems in Europe and North America. In North America, the acid precursor emissions of the midwestern industrial heartland of the U.S. inflict their damage on the states of the northeastern part of the U.S. or on Canada, hundreds or even a few thousand kilometers distant. In Europe some 30 countries influence each others environment to such a degree that national solutions are not adequate. An environmental decision in one country can greatly influence the situations of its neighbors. Because of the importance of the above factors, any attempt to develop approaches to avoid the emerging problems of acid rain and emissions in Asia need be comprehensive in scope. That is, the analysis should take into account the entire cause-and-effect chain of events which leads to the damage. This would include the energy supply/utilization system, its consequent emissions of precursors, emission control measures, transport and transformation of pollutants, acidic deposition, and the spectrum of impacts. These factors are shown schematically in a sequential systems diagram in Figure 3. As shown by the feedback loop, the analysis should also include the effects of various strategies and interventions to address the problem, including the economic consequences. For example, an analysis would examine the entire chain of

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consequences of a planned action such as expansion of the coal-fired electricity generation system over a long-term time horizon, as well as the consequences of alternative measures such as emission control, alternative fuel choices, and technology choices.

'igure 3 The Cause-Effect Chain for Acid Rain, Including Feedback to Policy Decisions

Approaches should also be comprehensive in the geographical sense. With increasing socioeconomic and industrial development, we are observing in Asia, a shift from small-scale (domestic) uses of fossil fuels to large-scale, tall-stack emissions associated with power plants and industrial facilities. This is increasing the propensity for long-range transport of pollution. Transboundary transport and deposition are inevitable, with, increasingly, damage to sensitive ecosystems in regions or countries far removed from the sources. Thus, mitigation strategies involving regional compacts or international agreements may be necessary. In recent years, integrated assessment models have been utilized for international negotiations on acid rain. The purpose of these models is to provide negotiators or regulators with a full regional picture of the problems associated with the entire causal process from energy systems and emissions through to the ultimate impact on natural and man-made systems. The model user can analyze the regional and national implications of various scenarios, which include options for energy use, control strategies, and mitigation policies. Such an effort for the acid rain problem in Europe has led to the Regional Acidification, INformation and Simulation (RAINS-EUROPE) model (IIASA, 1990). The RAINS model has been developed as a tool to assist policy advisors in

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evaluating options to reduce acid rain in Europe. The model can be used for both scenario analysis and optimization analysis. The optimization analysis can assist with questions such as: maximum SO, reduction for a given budget; reductions needed for a specific deposition target; or minimum cost to achieve specified deposition target. RAINS-EUROPE consists of sub-models (modules), each describing various aspects of the acid rain system. The user is guided through the model via a set of menus and options. A first set of sub-models deals with pollution generation and cost of abatement. For scenario analysis, users can select one of several preset energy use scenarios or can create their own, based on expectations of fuel use. The technical constraints (such as power plant efficiencies) are shown, and the user is warned when these constraints are violated. The user can then pick an abatement strategy for each country. The second set of sub-models deals with atmospheric transport and deposition of acidifying compounds. Calculations are based on a sub-model that consists of a transfer matrix for long range transport of pollutants in Europe, developed under the cooperative program for the Monitoring and Evaluation of Long-Range Transmission of Air Pollutants in Europe (EMEP). The outputs from the first set of sub-models are taken as inputs and transformed into deposition patterns for Europe. Comparisons between scenarios and abatement strategies can also be carried out. The third set of sub-models includes regionalized environmental impact models for forest soil acidification, lake acidification, susceptibility of groundwater, and direct effect of SO, on forests. All of these models are driven by the deposition patterns calculated by the second set of sub-models. The outputs from this module allows the user to compare scenarios and abatement strategies. In Europe, the Critical Loads Approach is being used in the negotiations for a new Sulfur Protocol. For this approach, knowledge of the sensitivity of ecosystems is necessary to derive national emission reductions. This knowledge is available in the form of critical loads maps for Europe (Hettelingh et al., 1991) which can be connected to RAINS-EUROPE to calculate the difference between deposition patterns and critical loads. The international collaborative Project on Acid Rain and Emissions in Asia has established as its initial objective the adaptation of the RAINS-EUROPE model to the Asia region. RAINS-ASIA will be developed by integration of the following modules applied to the Asian situation: (1) Energy and Emission Module Atmospheric Transport and Deposition Module (2) (3) Ecosystem Sensitivity Module Initially, RAINS-ASIA will be used to conduct a preliminary analysis of different energy and emission scenarios. The impacts of these energy and emission scenarios consist of (a) acid deposition and (b) comparison of acid deposition to the best estimates for sensitivity of ecosystems in Asia based on the European and North American data. The subsequent phases of this project will allow for refinement of the impact data based on specific work on critical Asian ecosystems, more detailed scenarios analysis, and inclusion of other pollutants. A specific sub-objective of this initial project phase is to involve scientists and policy makers from Asian countries and development institutions in all aspects of the project, to increase their awareness, and to transfer the model RAINS-ASIA and

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associated technologies to participating Asian governments and organizations.

5. SUMMARY AND CONCLUSIONS As a result of the present and future socio-economic and energy situation in Asia, emissions of acid precursors are likely to reach levels that could have far reaching implications for human and environmental systems throughout the region. To address this emerging problem, a group of specialists from Asia, North America, and Europe has established a project on Acid Rain and Emissions in Asia. The broad goals of this international collaborative project are to: develop approaches that assist in deriving regional and sub-regional 0 policies to avoid the rapidly emerging problems of atmospheric emissions in Asia; provide preliminary basic assessments needed to catalyze the process of 0 inter-governmental policy dialogue on acid rain precursor emission control in the Asia Region; and, develop basic strategies for policy advice, institution building, and 0 investment initiatives in countries of the Asia Region to deal with the problem of acid rain precursor emission control. Although the initial focus of the project will be limited to acidification, the overlap with several related issues of concern (e.g., emissions of greenhouse gases and other pollutants) is recognized. Later phases of this project must therefore address some of these related issues as well. 6. REFERENCES 1.

2. 3. 4. 5. 6.

7. 8.

Zhao Dianwu and Xiong Jiling: 1988.”Acidificationin Southwestern China,” In: Rodhe, H.and Herrera, R. (Eds) Acidification in Tropical Countries. SCOPE Report 36, pp. 317-376,John Wiley and Sons, Chichester, U.K. Rodhe, H. and Herrera, R. (Eds): 1988.Acidification in Tropical Countries. SCOPE Report 36, John Wdey and Sons, Chichester, U.K. Bhatti, N., Streets, D.G., and Foell, W.K.: 1990 “Acid Rain in Asia,” In: Foell, W.K. and Sharma, D. (Eds) Proceeding of the Second Annual Workshop on Acid Rain and Emissions in Asia. pp. 533,Asian Institute of Technology, Bangkok, Thailand. Foell,W.K., and Green, C.W.: 1990. “Acid Rain in Asia: An Economic, Energy and Emissions Overview,” In: Foell,W.K. and Sharma, D. (Eds) Proceeding of the Second Annual Workshop on Acid Rain and Emissions in Asia. pp. 123-145,Asian Institute of Technology, Bangkok, Thailand. International Institute for Applied Systems Analysis: 1990. The Rains Model of AcidiJication: Science and Strategies in Europe. Alcamo, J., Shaw, R., and Hordijk, L. (Eds). Kluwer Academic Publishers, Dordrecht, the Netherlands. Hetteligh, J.-P., Downing, R.J., de Smet, PAM., (Eds): 1991. Mapping CriticalLoadsfor Europe. Coordination Center for Effects Technical Report No. 1,RIVM Report No. 259101001,Bilthoven, the Netherlands. Workshop on Acid Rain in Asia: November 13-17,1989.“Report of the Workshop on Acid Rain in Asia,” Asian Institute of Technology, Bangkok, Thailand. Second Workshop on Acid Rain in Asia: November 19-22,1990.“Report of the Second Workshop on Acid Rain in Asia,” Asian Institute of Technology, Bangkok, Thailand.

T Schneider (Editor). Acidification Research Evaluationand Policy Applications 1992 Elsevier Science Publishers B V

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DEPOSITION OF ACIDJFYING COMPOUNDS D. Fowlercr, J.N. C a w , M A Suttona, R Mourn@ K.J. Hargreavesa, J.H. Duyzerb, M.W. GallagheF

eInstituteofTemsttd * Eoology Bush Estate, Pdcuik Midlothian,EEI26 OQB, United Kingdom bTN0

PO Box6011,2600 JA Delft The Netherlands cunivemity ofManchester Institute of Science& Technology Sackville,Manchester U60 lQD, United Kingdom Abstract Inputs of acidifying compounds to terrestrial ecosystems include deposition of the gases NO2, NO, HN02, HNO3, PAN,NH3, SO2 and the ions NO3-, NH4+, SO$-and H+ in precipitation, cloud droplets and particles. Recent research has identified particular ecosystems and regions in which terrestrial effects are closely linked with specific deposition processes. The air pollution related forest health problems at high elevation in the Appalachian mountains for example are primarily due to the cloud water deposition while the health of Dutch heathlands appear to be linked primarily with inputs of ammonia. Areas of large wet deposition in northern Europe show large spatial gradients in inputs and peak values up to a factor of 2 in excess of those provided by long range transport models as a consequence of seeder-feeder scavenging. Aerosol fluxes t o shoot vegetation a r e small relative t o gaseous deposition and seldom provide deposition velocities larger than 2 mms-1. There is also evidence of aerosol production in the few metres of the boundary layer close to the surface. Dry deposition rates for gases onto vegetation are controlled by atmospheric processes for the very reactive gases HNO3 and HC1. For NO2 uptake has been shown to be entirely stornatal with deposition increasing from zero a t night to 6 mms-1 over rapidly growing vegetation during the day. Rates of SO2 deposition are also regulated by stomatal uptake in the absence of surface water or ammonia. In the presence of NH3 (rc 0-200 sm-1) rates of deposition may reach the upper limit set by turbulent exchange processes (Vg = (r, + %)-I but the chemistry of NH3 and SO2 interactions on natural surfaces and its dependence on surface condition has not been quantified. Ammonia exchange over natural plant communities has been shown to

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be bidirectional, with NH3 emission over polluted Dutch heathland in dry conditions and deposition in wet conditions and is very similar to agricultural surfaces. For the cooler, wetter, less polluted regions such ecosystems represent a very efficient sink for NH3. Mechanistic models of deposition processes to extrapolate to regional scales have now been developed for S02, NO2, cloud droplet and wet deposition processes, but they do not satisfactorily incorporate NH3, SO2 co-deposition and are not yet adequate to extrapolate NH3 fluxes to the landscape scale. INTRODUCTION The atmospheric trace gases and particles which contribute t o acidification of ecosystems include the primary emitted pollutants SO2 and HCI,the secondary pollutant gases N02, HNO3 (and possibly HN02) and the SO$-, N o s - and H+ ions in aerosols and precipitation. Some of these components are not acidic directly and only lead to acidification as a consequence of chemical or biochemical processes within plants and soil (van Breeman and van Dijk, 1988). A notable addition to the list of acidifying compounds in recent years has been NH3, emitted largely from agricultural sources (Apsimon et al. 1987) and present close to sources in gaseous form and as the NH4+ ion in aerosols, cloud droplets and precipitation after a few hours in the atmosphere. The acidification resulting from NH3 or NH4+ deposition results from the microbial nitrification of NH4+ to NOS-. The physical and biological fate of deposited NO2, NOS-,NH3 and N H 4 + is not fixed or unique for each compound. For example an ammonia molecule may dry deposit t o a leaf surface into a water film and be washed by rain into a soil where it is nitrified and leached as NOS- from the soil profile. OR the same NH3 molecule may following deposition be taken up by a plant directly or the root system as NH4+ following its transfer to the soil. Such complications mainly with deposited nitrogen compounds lead to considerable difficulty in quantifying the actual acidification resulting from k n o w n deposition rates. For the purposes of this review the biological or chemical processing of deposited gases or ions will not be considered beyond the initial sites of uptake. For the sites of uptake it is clear that initial chemical processing may influence the deposition process directly and will be considered further especially under the subject of dry deposition of gases. This paper is not a review in the conventional sense as it does not provide a uniformly detailed consideration of the entire subject. It does however identify areas of the subject in which important developments have occurred during the last 5 years and an attempt is made to show which aspects of the subject are most important for the policy makers (i.e. what do we most need to get right). This necessarily requires us to identify particular components of the pollution climate which cause the largest acidification or impact and as such components are now always the dominant feature it is necessary to identify specific ecosystems and pollutants which are considered to be particularly sensitive.

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Wet deposition The chemical composition of precipitation is extensively monitored and there are networks on regional scales for Europe and North America (EMEP, NADP) and global background stations (BAPMON) Iversen et al. 1991 (Fig. 1). Such networks are not uniform and within the regions there are areas with intensive detailed networks e.g. the Netherlands and areas in which monitoring is not adequate to define the inputs to sensitive ecosystems, e.g. in many of the acid sensitive high altitude regions of Europe and North America. While interest and field measurements in wet deposition is growing in countries with rapid industrial development it is likely that the detailed networks to establish the scale and intensity of acidic input will follow the identification of damage rather than precede it. 1.5

Figure 1

Non-marine sulphate concentrations (mgl-1)in precipitation in North America and Europe (1990)

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The process of wet deposition scavenging of pollutants is reasonably well understood and has been recently reviewed as a part of the National Acidic precipitation assessment programme of the USA (NAPAP, 1990). This understanding is sufficient to explain observed wet deposition of aerosols and gases and the concentrations of major ions SO$-, Nos-, NH4+ and H+ at sites where the concentration of aerosol and gaseous precursors are known. The major limitation in the modelling of wet deposition processes on regional scales lies in the other processes; in quantifying the oxidation rates of sulphur and nitrogen oxides, in dry deposition and in describing the local variability in wet deposition scavenging due to orographic effects. The latter process is very important to the current discussion because many areas experiencing acidification effects from wet deposition lie in uplands with complex terrain on which the inputs show a very considerable spatial variability. The importance lies in identifying sensitive areas which receive wet deposition smcient to induce biological effects or in excess of a critical load yet which may be smoothed out of regional deposition estimates as a result of being sub-grid scale. The effect results from two different (though related) processes, the seeder-feeder enhancement of wet deposition with altitude (Fowler et al. 1988)and the turbulent deposition of cloud droplets (occult deposition) Dollard et al. 1983. The seeder-feeder process has been shown to increase wet deposition at an upland site by a factor of 4 relative to an adjacent lowland site. As rainfall amount was increased by approximately a factor of 2 any model which used only the spatial rainfall field would underestimate inputs at such sites by a factor of 2. The process of seeder-feeder scavenging illustrated in Fig. 2 results from the incorporation of aerosol SO$-, NOS-, NH4+ and H+ into orographic cloud above hills and the scavenging of the hill cloud droplets by rain droplets falling from a higher level cloud. FRONTAL CLOUD ( SEEDER)

TYPICAL CONCENTRATIONS (ueq 1-1 )

1

Cap cloud

Rain 200m 800m

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(FEEC

-

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200m

Figure 2

The seeder-feeder scavenging process by which wet deposition on elevated topography is enhanced by scavenging of polluted orographic cloud

557

The process necessarily leads to large spatial gradients in wet deposition and in complex terrain the measurement of such gradients is a difficult task and requires a large array of collectors (Dore et al. 1992). Even with such detailed measurements the physical difficulties in operating high altitude precipitation collectors precludes long term large area measurements of wet deposition using conventional techniques. Alternative techniques have been applied to estimate the inputs in complex terrain. The measurements of 210 Pb deposition from soil inventories by Graustein and Turekian, 1986 show the effect of altitude on wet deposition. These methods provide a longterm (- 1OOy) estimate of the scavenging of submicron aerosol by precipitation and dry deposition and have been further developed by Mourne et al. (1992). He has shown that the lead 210 Pb inventory averaged over the Pennine ridge a t Great Dun Fell increases from the valley to the ridge by a factor of 4 and that the point of maximum deposition lies l k m downwind of the summit a s predicted from the more detailed event studies (Fig. 3). Such methods applied to other sites in more complex terrain have quantitatively demonstrated the scale of this enhancement at a range of sites. Clearly the detailed monitoring of a n entire region o r country is prohibited by cost and practicality. With a n understanding of the seederfeeder process it has been possible to develop a simple model of the process and to generalize over regional scales. For the UK such a model has provided a revised wet deposition map and in the high rainfall areas of northern and western Britain the inputs have been increased by up to 70% (Dore et al. 1992). The same exercise should be extended to other regions. In Scandinavia, for example, the coastal mountains of Norway in particular are subject to very similar meteorological conditions, and close to the coast the seeder-feeder scavenging process contributes significantly to the total input.

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Altitude (rn) A.S. L.

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Figure 3

Lead 210 inventories on a transect a t Great Dun Fell showing the point of maximum deposition and the degree of orographic enhancement (from Mourne et al. 1992)

558

For other upland areas of Europe and in E. North America the above generalizations do not apply quantitatively because a much larger proportion of the wet input occurs through convective storms rather than frontal rain. In these circumstances, and with the much greater variability in topography and altitudes there is a requirement for more field study. The modification of wet deposition processes by snow at much greater altitude have been studied by (Ronseaux and Delmas, 1988) but the effect on annual inputs at sites where most wet deposition falls as snow remain poorly understood.

~udandfbg~~tdepoeition Work in Europe and N. America during the last six years has provided a mechanistic understanding of the cloud droplet deposition process and has quantified the dependence of this process on the physical properties of the underlying surface and atmospheric conditions. The main emphasis in this work has been on the 'interception' of cloud or fog droplets by vegetation which for high altitude or coastal vegetation may represent a large input which is not detected by precipitation collectors (and hence was the origin of the term occult (Kerfoot, 1968). Equally important are the chemical characteristics of hill cloud and fogs. Droplets are activated on aerosols, and in a polluted atmosphere the ionic composition of aerosol is dominated by SO$-, NOS-,NH4+ and H+. Following activation the droplets may grow as the air cools, in the case of hill cloud by lifting the air further up a hillside, or in the case of fog, following radiative cooling at the surface. The concentrations of the above ions in hill cloud and fog lie in the range 20 to 2000 peq 1- (Saxena et al. 1991) and generally exceed the concentrations in rain (Fig. 2). The largest concentrations to which hill vegetation is exposed (10002000 peq 1-1)are very much larger than those in rain and have been identified as an important mechanism by which high elevation forests may be damaged by pollutants (Johnson and Siccama, 1983, Leith et al. 1989, Fowler et al. 1989). Hill cloud and lowland fog droplets generally have diameters between 3 and 20 pm, and it is the dependence on the droplet size of deposition rates on a range of vegetation that has been obtained from process studies. The major limitation on estimating inputs by the cloud droplet deposition process is the lack of monitored cloud water composition. The rates of turbulent deposition vary with windspeed, droplet size and the aerodynamic roughness of the vegetation. The maximum rate of droplet deposition is provided by the rate of transfer of momentum (vm)so that the ratio of deposition velocity of cloud droplets V&m provides a good illustration of the efficiency of the deposition process. The results from measurements by Gallagher et al. (1988) (Fig. 4) show that over the most common drop size range for hill cloud droplets (radius 4-12 pm) the rate of deposition is approximately equal to that for momentum. In these conditions the forest or moorland surfaces are efficient sinks for the droplets and rates of deposition can readily be calculated for a range of surface and atmospheric conditions. To model the inputs satisfactorily it is necessary to know the land use, windspeed, concentration of major ions in cloudwater and cloud liquid water content. A substantial collection of data for Eastern North America has been provided by the mountain cloud chemistry programme (Weathers et al. 1988) and 'campaign' data exist for 5 sites in the UK and 6 sites in Germany. These

559 10'

1

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Moorland

lo-* 0

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10 12

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16

Droplet radius (urn )

Figure 4

The variation with particle size of the ratio of turbulent deposition rate for cloud droplet deposition Vt to that for momentum (from Gallagher et al. 1990)

consistently show the large concentration of ions in cloudwater to which vegetation is exposed but are not adequate for precise estimates of inputs. In the absence of such monitoring data it is possible to provide a n estimate of the lower limit of cloudwater deposition by assuming that the concentrations of major ions in cloud are on average proportional to their regional average concentrations in rain, and using field data to establish the constant of proportionality. From precipitation and cloud composition studies a t Great Dun Fell, Fowler et al. (1988) estimated the mean ratio of concentrations of major ions in cloud and rain during precipitation. Such data provide much smaller concentrations than the median for cloudwater from all events and are therefore likely to be an underestimate. Using these data cloudwater deposition has been modelled for the UK incorporating observed cloud frequency land use and wind speed data (Fowler et al. 1992). Alternative methods for modelling cloud droplet deposition have been adapted for individual sites in N. America by Lovett et al. 1983 using a n empirical model and a much more detailed mechanistic approach over areas of complex terrain has been used by Gallagher et al. (1992). Although differing in detail, similar conclusions are drawn by all authors on the importance of the deposition process for upland forests. For high elevation forests in Europe and N. America cloud water deposition has been estimated to contribute up to 75% of the total atmospheric inputs of S and N. Even a t modest altitudes of lOOOm in N. Britain the input by this process may be as much as 40% of total wet deposition. Fog water deposition at low elevation sites has only been shown to contribute significantly to the hydrological input at coastal sites (e.g. California and Japan). However, the concentrations of major ions during

-

560

winter months in radiation fogs in parts of Europe (e.g. the Po Valley, Fuzzi, 1988) are large enough to represent a potential threat to some sensitive tree species and merit further investigation. If the total regional amounts of acidic pollutants deposited in cloud or fog droplets were expressed as a fraction of the total deposition this deposition mechanism would be dismissed as irrelevant. However the importance of this process for inputs on hills, the implications for land use and critical loads, and the mechanism for exposure of vegetation to damaging concentrations of pollutants, highlight the need t o know more about the composition of hill and mountain cloud.

Dryde~tionofparticles The subject of particle deposition has received little attention during the last 5 years but remains an area of substantial controversy. On the one side, wind tunnel and field micrometeorological methods have consistently reported small rates of aerosol deposition on vegetation, with deposition velocities in the range 0.1 to 1 mm s-1. Such small values for the deposition velocity are consistent with the gradual accumulation of aerosol during dry weather and with the known particle size dependence of impaction efficiency for particles in the size range 0.1 to 1.0 pm (radius), (Chamberlain 1975). However there have also been measurements using surrogate surfaces and a range of leaf washing techniques (Lindberg et al. 1990) which imply rather larger rates of particle deposition, especially for coarse particles (radius >> 1 pm). Further, the modelling studies of Wiman (1988) also suggest larger rates of particle deposition close to a forest edge. Thus the forest present in areas of southern Sweden for example, with large edge perimeter of forest areas. may be much more important as a sink for aerosols than the wind tunnel studies suggest. Recent measurements of particle deposition rates have provided the particle size dependence of aerosol deposition using micrometeorological techniques (Fig. 5). These results, from measurements by the eddy covariance method using a sonic anemometer and an active scattering aerosol spectrometer probe (ASASP) above a heather dominated heathland in the Netherlands, show apparent emission of particles smaller than 0.1 pm and deposition of particles in the size range 0.1 to 0.5 pm increasing to 5mm 5-1 at 1pm. The process responsible for the apparent upwind fluxes are not known but such fluxes of very small particles are not consistent with the literature on resuspension of particles from foliage (Slinn 1983). At this early stage an analysis and interpretation of the field data no quantitative explanation of the results has been provided and the only plausible mechanism advanced has been the possibility of rapid gas to particle conversion. Such rates of deposition to short vegetation make the aerosol dry deposition pathway a small component of the annual budget. For forests however with much larger leaf area and smaller characteristic dimensions in the case of conifers, together with larger windspeeds, the question is still open and detailed measurements of the kind made by using ASASP methods over forests are necessary to address this question.

561

26th April 1991

30

tt

-30

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ptle radius (pm) Figure 5

The size dependence of aerosol particle deposition onto heathland during the Leende experiment in 1991

DRY DEPOSlTIONOFGASES

H N a and HCl These very reactive gases are deposited on vegetation as rapidly as they can be transported to the foliar surfaces by turbulent transfer. Rates of deposition are therefore large and are determined by windspeed and the aerodynamic roughness of the surface. Recent measurements, reported by Dollard et al. 1987 and early measurements by Hubert 1983 consistently reported canopy resistances (re) close to zero. The only measurements in which non-zero surface resistances have been detected a r e those of Johannson and Granat (1986) for measurements of HNO3 deposition to snow a t temperatures below -543. NQa Stomata represent a n important sink for pollutants and in many circumstances are the major sink. For Nos, field data over rapidly growing cereal crops show rates of NO2 deposition very close to those expected for stomata1 uptake, and no leaf surface uptake, a s shown in Figure 6. For slow growing moorland vegetation a t low ambient NO2 concentrations, very small rates of NO2 deposition have been observed (1-2 mm s-1). Similarly in cuvette measurements on Scots pine (Pinus syluestris) shoots, Johannson, 1987 has shown the presence of a large surface resistance to NO2 uptake at low ambient NO2 concentrations (< 10 ppb). An important complication in the measurement of NO2 exchange by gradient micrometeorological techniques is that of air chemistry in the lower layers of the atmosphere between NO, NO2

562

v)

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v,

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LI 300 600 900 1200 1500 1800 2100 2400

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Figure 6

Canopy resistances for NO2 and 0 3 over a wheat crop showing the close link between these gases

and 03. In practice the analysis of vertical profiles in the concentrations of these three species to provide fluxes between the atmosphere and the ground requires more information than the gradients in concentration and windspeed. In particular, the photolysis rate coefficient JN02 is required. The analysis of NO2, NO and 0 3 flwdgradient data by Duyzer (1991) and by Kramm et al. (1991) has shown the magnitude of these effects in field conditions. Grassland on the Halvergate marshes was shown to be a significant sink for NO2 and stomata were the primary sink. The field data for NO2 are at last therefore beginning to show a more consistent pattern of uptake for agricultural vegetation,in which stomatal uptake dominates, and which can therefore be simulated using the big-leaf resistance analogy, Hicks and Matt, 1988. The regional concentrations of NO2 may therefore be used together with a stomatal resistance model and land use data to compute NO2 deposition. For hill land and forests the inputs are less certain,as a consequence of poor understanding of the factors which control the magnitude of the internal or surface resistance, but it is at least possible now to compute regional estimates of NO2 dry deposition fluxes.

HNoa

The presence of gaseous nitrous acid at concentrations in the range 0.1 to 5 ppbV has been detected at night in measurements in Germany (Lammel et ul. 1989 and in the UK Kitto and Harrison, 1992). This gas is rapidly photolysed during the day to form the OH radical and nitric oxide and is therefore an important nocturnal reservoir for the OH radical. Measurements by Kitto and Harrison (1992) indicate that the HNO2 present close to the ground at night may be the product of heterogeneous reaction of

563

NO2 on moist vegetation and may therefore be closely linked with NO2 deposition processes. Production rates of HNO2 at the ground or on aerosol surfaces, have also been estimated by Lammel et al. 1989. Such hypotheses require further testing but are an indication of the growing tendency t o incorporate atmospheric chemistry field studies with those of trace gas flux measurements.

soa This gas is absorbed readily by stomata and in many laboratory and field studies the primary sink has been shown to be stomata (Fowler 1978,Black & Unsworth 1979, Garland et al. 1978). In the absence of any mesophyll resistance to SO2 uptake by stomata, and using a constant and small rate of uptake on external plant surfaces, it has been possible to model SO2 deposition to the landscape using land use, meteorological and monitored SO2 concentration data (Fig. 7). These modelling approaches have been applied in North America and in parts of Europe and it is now time to extend the application of these models to permit comparisons between the large scale, long range transport model estimates of dry deposition fluxes and the process based resistance analogy models working largely from monitoring data. The most pressing current limitation to our understanding of SO2 uptake by vegetation is the leaf surface uptake, its variability and the chemical processes taking place on surfaces. In particular, it is known that leaf surface uptake is not constant; the presence of liquid water on leaf surfaces has been shown to change the surface resistance. Fowler and Unsworth (1979)showed that initially dew reduced surface resistances to zero but with time, as dewfall rates decreased, a surface resistance reappeared. These changes were interpreted as effects of the equilibrium chemistry of SO2 but have not been confirmed by detailed measurements of the reactants and products in the field.

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Figure 7

The seasonal dependence of SO2 deposition velocity onto a spruce forest simulated using a big leaf model

564

More recently it has been suggested that the presence of gaseous ammonia may influence SO2 deposition rates on leaf surfaces (Adema 1986, van Hove et al. 1989) and some field data have been obtained which are consistent with these suggestions (Fowler et al. 1991). The laboratory work at large NH3 and SO2 concentrations is certainly indicative of the co-deposition hypotheses but do not show that the process occurs at the concentrations experienced in field conditions. The field observations of small canopy resistances in the presence of concentrations of NH3 and SO2 adequate for their reaction at the surface to produce (NH412SO4 are also consistent with the co-deposition hypothesis and the data (Figure 8) provide an example of this phenomenon. The detailed behaviour of surface resistance for SO2 and the apparent surface concentration for NH3 (at zo) show the presence of a significant surface resistance to SO2 deposition with concentrations of NH3 still adequate for the complete reaction of S02. The determination of rate coefficients from data of this nature will continue to be problematical because of the difficulty in controlling or measuring the wide range of variables which all influence the chemistry. It is clear however that there is an effect of rates of NH3 on SO2 exchange. Also consistent with these data are the earlier field data for SO2 deposition in the UK and the USA which found large and fairlyconstant canopy resistances for non-stomata1uptake, typically of 400 s m1 or greater. These measurements were generally made at ambient SO2 concentrations in the range 10 to 30 ppb, between 1and 2 orders of magnitude greater than the likely NH3 concentrations! Simulation of the surface Chemistry of these gases within the models of net exchange is a relatively straightforward matter when the mechanisms of reaction are understood and shown to be consistent with field data. m

S

The subject of ammonia exchange over natural surfaces has become central to the current debate on acidification. First, the microbial nitrification of deposited NH4+ leads to acidification; second, over substantial areas of the Netherlands, Germany, Denmark and the UK, deposited NH4+and NH3 (wet and dry) dominate the deposition budget for deposited nitrogen, and third, the presence of atmospheric NH3 has been shown to influence rates of deposition of SO2 as described above. "he processes of surface-atmosphere NH3 exchange are similar to those for NO2 or SO2 but the strongly bidirection nature of the fluxes of this gas lead to some agricultural fields representing net sources while adjacent fields act as sinks for the gas and the spatial heterogeneity in fluxes, at least in agricultural districts, is a major feature of the behaviour of NH3. Over heathland, moorland and forests the rates of NH3 deposition have been shown to be large, and most studies show that the gas is deposited at rates close to the upper limit set by rates of turbulent transfer (Duyzer et al. 1987, Sutton, 1990). These data led to a rather simplified approach t o estimating the likely annual inputs of NH3 for natural vegetation (Sutton, 1990). These generalizations appear to be valid for measurements made at sites in N. Britain with small ambient NH3 concentrations (0.1 to 5 ppb) and

565

Ammonia and sulphur dioxide exchange at Halvergate 100

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Figure 8

Fluxes and resistances in the exchange of NH3 and SO2 over the Halvergate grassland (U.K.)

during periods of high humidity. However, during periods of low humidity and especially at sites with large ambient NH3 concentrations (as during the Dutch Heathland experiment at Leende) in warm dry conditions, large surface resistances to NH3 deposition have been observed. At this site even upward fluxes were recorded. Figure 9 shows an example of the bidirectional exchange of NH3 over the Dutch heathland during this study, during which the change from emission to deposition occurs at the transition from dry to wet surfaces of the vegetation late in the day as radiative cooling at the ground led to dew formation on the vegetation. "he flux is bi-directional on both seasonal and daily scales. A pulse of ammonia emission occurs following fertilizer application, which may result from fertilizer residues in the soil as well as an increase in the equilibrium NH3 concentration, or 'compensation point', in plant tissues. A second peak in emission may occur following hay cutting or during senescence of annual crops (Dabney and Bouldin 1990). On a daily scale, emission is favoured during warm dry conditions, whereas deposition generally occurs when the surface is wet of frozen (Sutton et al. 1992). At present, the integration of these fluxes over annual scales for crop vegetation must be regarded as very approximate.

566

Ammonia exchange at Leende Heide '-

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Figure 9

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humid

Fluxes of N H 3 over Dutch heathland at Leende showing the bidirection nature of N H 3 exchange at this site

k u g b f a l land stamfbw measurements

Many research groups throughout the last three decades of acid rain research have made measurements of the change in precipitation chemistry as the water and solutes pass through plant canopies. The work has mainly (though not exclusively) been applied to forest vegetation (e.g. Miller et al. 1984, Cape et al. 1987a). The measurements have provided a valuable contribution to the general field of bio-geochemical cycling of a range of elements and the objectives of the measurement have been varied. One objective of such measurements has been to estimate the input of a range of elements from the atmosphere. Many reviews of the different techniques applied and their success have been published (Lindberg et al. 1991). It is not the purpose of this review to repeat the conclusions of the earlier reviews of the throughfall and stemflow measurements. Rather it is to identify scientific issues which have not been solved by the work so far and which require a solution before any such techniques are applied as a monitoring instrument to obtain areal estimates of inputs for critical loads or other mapping purposes. Although in some studies, notably the Integrated Forest Study

567

(Lindberg et al., 1991) there has been apparent agreement for sulphur between modelled deposition from the atmosphere using actual concentration and meteorological data, and measured deposition in throughfall and stemflow, the errors involved are potentially large. Quoted error estimates for deposition modelling to forests are f 50%, or more where occult deposition is important. Error estimates for deposition in throughfall and stemflow are typically 1530%. The apparent agreement between modelled deposition and throughfall and stemflow data only holds for long-term averaging (one or more full years) and does not apply on weekly or monthly time scales, despite the use of hourly averaged data for modelling, and event sampling for deposition. There are large uncertainties in the short-term processes which control storage and release of sulphate in forest canopies. Moreover at some sites (e.g. Cape et al. 1987b) measured annual deposition below canopy was shown to be so far in excess of modelled deposition from the atmosphere that significant additional sources of sulphate were required to make up the deficit, suggesting a role for the internal cycling of sulphur. If internal cycling of sulphate is a major contributor to sulphate measured below canopy then throughfall and stemflow measurements cannot easily be used to quantify net deposition. Attempts to quantify sulphur cycling have been made using radioisotopes. Garten (1988,1990) injected trees with 35so4 and followed the partitioning of the isotope in foliage and throughfall, and deduced that leaching (internal cycling) contributed only a few percent to below-canopy deposition. His studies, on isolated or edge trees, probably underestimated the contribution in a closed forest stand, because below-canopy deposition a t the forest edge would have been much greater than in a uniform block of forest, as described above. Cape et al. (1992) applied 36so4 to the forest floor and followed the appearance of the isotope in foliage and throughfall. Apparent leaching was initially almost loo%, falling to c.5% of below-canopy deposition after 6 weeks. These data illustrate the uncertainties in using isotopes, where assumptions about equilibrium within the canopy do not in general apply. The role of internal cycling remains unclear, but given the relative uncertainties involved, throughfall and stemflow measurements may provide an adequate upper estimate of sulphur deposition for critical load assessment if appropriate measures are taken to avoid the sampling bias introduced by the large spatial heterogeneity below forest canopies. For nitrogen compounds (ammonium or nitrate) existing methods of sampling throughfall and stemflow are generally inadequate, given the rapid microbial degradation of water samples. Interpretation of the results is even more difficult than for sulphur, as nitrogen may be taken up by foliage, by stems, by epiphytic lichen or other microflora, or may be leached (as amino acids) from the canopy. Once measurement techniques have improved, deposition to the forest floor may be quantifiable, but canopy interactions will still dominate the pathway between atmospheric deposition and the ground. CONCLUSIONS Two important conclusions may be drawn from the above review for science and application in the policy field: firstly, that current uncertainties in estimates of S and N inputs by dry deposition are likely to lead to continual changes in estimated regional inputs of S or N of typically 50%, as current

568

understanding is refined, and that this uncertainty must be incorporated in the critical load calculations to avoid failure to protect given ecosystems when emission controls are fully implemented. Secondly, and this applies to the regions dominated by wet deposition, that the spatial resolution of total inputs in complex terrain must be improved to match the current scales of information on landscape sensitivity to acidic inputs . The initial conclusion follows from the inability of the research community to unravel the scientific understanding at the pace and timetable agreed by policy makers. It follows from these two statements and the high cost of emission controls that a cost-benefit analysis would show that a much larger investment in research would be entirely justified on financial grounds simply to reduce the uncertainty in inputs and therefore the magnitude of emission controls necessary. An alternative and more practical appeal would be to encourage more widespread application of current understanding by modellers. Modellers have become a n effective channel of communication between the process orientated scientists and policy makers. This would therefore assist in using the best available information to direct policy.

ACKNOWLEDGEMENTS The authors gratefully acknowledge financial support for this work from the Air Quality Division of the U.K. Department of the Environment and the Commission of the European Communities.

REFERENCm Adema, E.H. (1986). On the dry deposition of NO2, SO2 and NH3 on wet surfaces in a small scale wind tunnel. Proceedings of the 7th World clean air congress, Sydney, p. 1.18. Apsimon, H.M., Kruse, M. and Bell, J.N.B. (1987). Ammonia emissions and their role in acid deposition. Atmos. Enuiron. 21, 1939-1946. Black, V.J. and Unsworth, M.H. (1979). Resistance analysis of SO2 fluxes to Vicia faba. Nature,282,68-69. Cape, J.N., Fowler, D., Kinnaird, J.W., Nicholson, I.A. and Paterson, I.S. (1987a). Modification of rainfall chemistry by a forest canopy. In Pollutant Transport and Fate in Ecosystems. eds, P.J. Coughtrey, M.H. Martin and M.H. Unsworth. Blackwell, Oxford. pp. 155-169. Cape, J.N., Fowler, D., Leith, I.D. and Paterson, I.S. (198713). Measurements of sulphate in throughfall, and dry deposition of SO2 to a forest. In Direct Effects of Dry and Wet Deposition on Forest Ecosystems - in Particular Canopy Interactions. ed. P. Mathy. Air Pollution Research Report 4(EUR 112641, CEC, Brussels, pp. 86-91.

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Cape, J.N., Sheppard, L.J., Fowler, D., Harrison, A.F., Parkinson, J.A., Dao, P. and Paterson, I S . (1992).Contribution of canopy leaching to sulphate deposition in a Scots pine forest. Environ. Pollut. 75,229-236. Chamberlain, A.C. (1975).The movement of particles in plant communities. In Vegetation and the Atmosphere, Vol. 1, pp. 115-201,ed, J.L. Monteith. London: Academic Press. Duyzer, J.H., Boumann, A.M.H., Diederen, H.S.M.A and Van Aalst, R.M. (1987).Measurement of dry deposition velocities of N H 3 and NH4+ over natural terrains. Report R 87/273 Netherlands Organisation for Applied Scientific Research (TNO). Dabney, S.M. and Bouldin, D.R. (1985).Fluxes of ammonia over a n alfalfa field. Agron. J. 77,572-578. Dollard, G.J., Unsworth, M.H. and Harvey, M.J. (1983).Pollutant transfer in upland regions by occult deposition. Nature, 302,241-247. Dore, A.J., Choularton, T.W., Fowler, D. and Storeton-West, R.L. (1992).Field measurements of wet deposition in a n extended region of complex topography. Q. J. Roy. Met. Soc.(in press). Dore, A.J., Choularton, T.W. Fowler, D. and Crossley, A. (1992).Orographic enhancement of snowfall. Environ. Pollut. 75, 175-180. Fowler, D. (1978).Dry deposition of SO2 on agricultural crops. Atmos. Environ. 12,369-373. Fowler, D. and Unsworth, M.H. (1979). Turbulent transport of sulphur dioxide to a wheat crop. Quart. J. Roy. Met. SOC.105,767-783. Fowler, D., Cape, J.N., Leith, I.D., Choularton, T.W., Gay, M.J. and Jones, A. (1988).The influence of altitude on rainfall composition a t Great Dun Fell. Atmos. Environ. 22, 1355-1362. Fowler, D., Cape, J.N., Deans, J.D., Leith, I.D., Murray, M.B., Smith, R.I., Sheppard, L.J. and Unsworth, M.H. (1989).Effects of acid mist on the frost hardiness of red spruce seedlings. New Phytol. 113, 321-335. Fowler, D., Duyzer, J.H. and Baldocchi, D.D. (1991).Inputs of trace gases, particles and cloud droplets to terrestrial surfaces. Proc. Roy. SOC. Edinburgh m,35-59. Fuzzi, S. (1988).Fog chemistry and deposition in the Po Valley. In Acid Deposition at High Elevation Sites, ed. M.H. Unsworth and D. Fowler. Kluwer, Dordrecht, pp. 443-452.

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Gallagher, M.W., Choularton, T.W., Morse, A.P. and Fowler, D. (1988). Measurements of the size dependence of cloud droplet deposition a t a hill site. Q.J. Roy. Meteorol. Soc. 114,1291-1303. Gallagher, M.W., Beswick, K., Choularton, T.W., Coe, H., Fowler, D. and Hargeaves, K.J. (1992). Measurements and modelling of cloudwater deposition to moorlands and forests. Environ. Pollut. 75,97-107. Garten, C.T.(1988). Fate and distribution of sulphur-35 in yellow poplar and red maple trees. Oecologia, 76,43-50. Graustein, W.C.and Turekian, K. (1986). 2loPb and 137Cs in air and soils measure the rate and vertical profile of aerosol scavenging. J. Geophys. Res. 91,14355-14366. Hicks, B.B. and Matt, D.R. (1988). Combining biology, chemistry and meteorology in modelling and measuring dry deposition. J. Atmos. Chem. 6, 117-131. Huebert, B.J. (1983). Measurements of the dry deposition flux of nitric acid vapour to grasslands and forests. In Precipitation Scavenging, Dry Deposition and Resuspension, pp. 785-794, eds, Pruppacher, H.R., Semonin, R.G. and Slinn, W.G.N. New York: Elsevier. Iversen, T., Halvorsen, N.E., Mylona, S. and Sandnes, H. (1991). Calculated budgets for airborne acidifying components in Europe, 1985,1987,1988,1989 and 1990. Det Norske Meteorologiske Institutt, Technical Report no. 91. August 1991. Johannson, C. and Granat, L. (1986). An experimental study of the dry deposition of gaseous nitric acid to snow. Atmos. Environ. 20,1165-1170. Johannson, C. (1987). Pine forest: a negligible sink for atmospheric NOx in rural Sweden. Tellus, 39B, 426-438. Johnson, A.H. and Siccama, T.G. (1983). Acid deposition and forest decline. Environ. Sci. Technol. 17, 294a-305a. Kerfoot, 0. (1968).Precipitation on Vegetation. Forestry Abstracts 29,8-20. Kitto, A-M.N., and Harrison, R.M. (1992). Nitrous and nitric acid measurements a t sites in south-east England. Atmos. Environ. (in press). Lammel, G., Perner, D. and Warneck, P. (1989). Nitrous acid at Mainz: observation and implication for its formation mechanism. I n PhysicoChemical Behaviour of Atmospheric Pollutants. eds., G. Restelli and G. Angeletti, Kluwer, Dordrecht. pp. 469-476.

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Leith, I.D., Murray, M.B., Sheppard, L.J., Cape, J.N., Deans, J.D., Smith, R.I. and Fowler, D. (1989).Visible foliar injury of red spruce seedlings subjected to simulated acid mist. New Phytol. 113,313-320. Lindberg, S.E., Garten, C.T., Cape, J.N. and Ivens, W. (1991).Can sulphate fluxes in forest canopy throughfall be used to estimate atmospheric sulphur deposition? - A summary of recent results. In Precipitation Scavenging and AirISurface Exchange Processes, ed, W.G.N. Slinn. Washington: Hemisphere. (in press). Lovett, G.M. and Reiners, W.A. (1983).Cloud water: a n important vector of atmospheric deposition. In Precipitation Scavenging, Dry Deposition and Resuspension eds, H.R. Pruppacher, R.G. Semonin and W.G.N. Slinn, pp. 84106.Kluwer. Dordrecht. Miller, H.G. (1984).Deposition-plant-soil interactions. Phil. Trans. Roy. SOC.

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LIST OF PARTICIPANTS

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575 Adema, E.H. Agricultural University 6700 EV WAGENINGEN The Netherlands tel.: 8370-82100 fax: 8370-84457

Al, G.J.A. Ministry of Housing, Physical Planning and Environment P.O.Box 450 2260MB LEIDSCHENDAM The Netherlands tel.: 70-3 174155 fax: 70-3174448

Amann, M. USA Schlossplatz 1 A-2361 LAXENBURG Austria tel.: 2236-71521 fax: 2236-71 3 13

Bakema, A.H. National Institute of Public Health and Environmental Protection P.O.Box 1 3720 BA BILTHOVEN The Netherlands tel.: 30-743704 fax: 30-250740

Barth, H. European Economic Commission, DG XII 200, Rue de la Loi B-1049 BRUSSELS Belgium tel.: 32-2-2358160 fax: 32-2-2363024

Bekker, M. Ministry of Agriculture, Nature and Fisheries P.O.Box 20401 2500 EK DEN HAAG The Netherlands tel.: 70-3792250 fax: 70-3825752

Bertills, U. Swedish Environmental Protection Agency S-17185 SOLNA Sweden tel.: 46-8-7991000 fax: 46-8-283008

Bleuten. W. Dept.of Physical Geography, Univ. of Utrecht P.O.Box 801 15 3508TC UTRECHT The Netherlands tel.: 30-540604 fax: 30-532780

Bouma, W.J. CSIRO, Division of Atmospheric Research Private Bag no. 1 3195 ASPENDALE, Victoria Australia tel.: 61-3-5867666 fax: 61-3-5867600

Bresser, A.H.M. National Institute of Public Health and Environmental Protection P.O.Box 1 3720 BA BILTHOVEN The Netherlands tel.: 30-743043 fax: 30-250740

Briill, N.J.H.C. The Netherlands America Institute P.O.Box 2225 6202HA MAASTRICHT The Netherlands tel.: 43-897602 fax:

Buijsman, E. National Institute of Public Health and Environmental Protection P.O.Box 1 3720 BA BILTHOVEN The Netherlands tel.: 30-7423 18 fax: 30-28753 1

Cowling, E.B. College of Forest Resources, North Carolina State University RALEIGH, NC USA tel.: 919-515-7564 fax: 919-515-7231

Cuypers, C.M. Catholic University Leuven E.van Evenstraat 2b 3000 LEUVEN Belgium tel.: 32-16-2831 14 fax: 32-16-283253

576 Dalziel, R.T.K. Powergen, Ratcliffe Technology Centre RATCLIFFE-ON-SOAR, Nottingham NG11 OEE United Kingdom tel.: 602-832257 fax: 602-83271 1

Denvent, R.G. Department of the Environment B 358 Romney House 43 Marsham Street LONDON SWlP 3PY United Kingdom tel.: 71-2768881 fax: 7 1-2768299

Downing, R.J. National Institute of Public Health and Environmental Protection P.O.Box 1 3720 BA BILTHOVEN The Netherlands tel.: 30-743532 fax: 30-250740

Draaijers, G. University of Utrecht P.O.Box 80115 3508TC UTRECHT The Netherlands tel.: 30-534014 fax: 30-540604

Driscoll, C.T. Syracuse University 220 Hinds Hall SYRACUSE NY 13244-1190 USA tel.: 315-4433434 fax: 315-4431243

Dumas, M.T. Ministry of Housing, Physical Planning and Environment P.O.Box 450 2260 MB LEIDSCHENDAM The Netherlands tel.: 70-3174422 fax: 70-3174448

Ek, R.van Dept.of Physical Geography, Univ.of Utrecht P.O.Box 80115 3508TC UTRECHT The Netherlands tel.: 30-534014 fax: 30-540604

Elshout, A.J. KEMA P.O.Box 9035 6800ET ARNHEM The Netherlands tel.: 85-562381 fax: 85-515022

Farmer, A.M. English Nature NGhminster House PE11UA PETERBOROUGH United Kingdom tel.: 733-340345 fax: 733-68834

Forsberg, E. Ministry of Environment S-10333 STOCKHOLM Sweden tel.: 46-8-7632041 fax: 46-8-7231 160

Fowler, D. Institute of Terrestrial Ecology Bush Estate PENICUICK MIDDLOTHIAN EG25 OQB United Kingdom tel.: 44-31-4454343 fax: 44-31-4453943

Gaasbeek, P. Shell Nederland B.V. P.O.Box 1222 3000BE ROTTERDAM The Netherlands tel.: 10-4696035 fax:

Giilli, Mrs.B. Feded Office of Environment CH-3003 BERN Switzerland tel.: 44-31616857 fax: 44-31618057

Gregor, H.D. Umweltbundesamt Bismarckplatz 1 D-1000 BERLIN 33 Germany tel.: 49-30-89032130 fax: 49-30-89032285

511 Grennfelt, P. Swedish Environmental Research Institute P.O.Box 47086 402 58 GOTEBORG Sweden tel.: 3 1-460080 fax: 31-482180

Grinsven, J.J.M.van National Institute of Public Health and EnvironmentalProtection P.O.Box 1 3720 BA BILTHOVEN TheNetherlands tel.: 30-743397 fax:

Ham, J.van SCMO-TNO P.O.Box 186 2600AD DELFT The Netherlands El.: 15-696877 fax: 15-613186

Hannessen, H. Ministry of Housing, Physical Planning and Environment P.O.Box 450 2260 MB LEIDSCHENDAM The Netherlands tel.: 70-3174427 fax: 70-3174448

Heammerli, F. Institut de Recherches sur la Foret, la Neige - et le Paysage CH-8903 BIRMENSDORF ZH Switzerland tel.: 41-17-392111 fax: 41-17-392215

Heij, G.J. National Institute of Public Health and EnvironmentalProtection P.O.Box 1 3720 BA BILTHOVEN The Netherlands tel.: 30-743108 fax: 30-250740

Hertz, J. Swiss Federal Institute for Forest, Snow and Landscape Research (WSL) Ziircher&rasse 111 BIRMENSDORF 8903 Switzerland tel.: 1-7392464 fax: 1-7392488

Hettelingh, J.P. National Institute of Public Health and EnvironmentalProtection P.O.Box 1 3720 BA BTLTHOVEN The Netherlands tel.: 30-743048 fax: 30-250740

Holland, M.R. ETSU, B149 Hanvell Laboratory OXFORDSHIREOX11 ORA United Kingdom tel.: 44-235-432522 fax: 44-235-432662

Hordijk, L. Centre for Environment Studies AgriculturalUniversity P.O.Box 9101 6700 HB WAGEMNGEN The Netherlands tel.: 31-8370-84919 fax: 31-8370-84919

Hultberg, H. Swedish Environmental Research Institute P.O.Box 47086 402 58 GOTEBORG Sweden tel.: 3 1-460080 fax: 31-482180

Innes, J.L. Forestry Commission Alice Holt Lodge, Wrecclesham GUlO 4LH FARNHAM Surrey United Kingdom tel.: 420-22255 fax: 420-23653

Irving, P.M. NAPAP 722 Jackson Place NW WASHINGDON DC 20503 USA tel.: 202-2961002 fax: 202-2961009

Irwin, J.G. Warren Spring Laboratory Gunnels Wood Road STEVENAGE SG12BX United Kingdom tel.: 44-438-741122 fax: 44-438-360858

578

Jattke, A. Institut for Industrial Production Hertzstrasse 16 D-7500 KARLSRUHE 21 Germany tel.: 49-721-6084551 fax: 49-721-758909

Johansson, M.P. Technical Research Centre of Finland P.O.Box 208 ESPOO 02151 Finland tel.: 358-0-4565053 fax: 358-0-4565000

Jongebreur, A.A. Institute of Agricultural Engineering Mansholtlaan 10 - 12 6700 AA WAGENINGEN The Netherlands tel.: 8370-76346 fax: 8370-25670

Kauppi, P.E. Finnish Forest Research Institute Unioninkatv 40a SF-00170 HELSINKI Finland tel.: 358-0-857051 fax: 358-0-625308

Keizer, V. Ministry of Housing, Physical Planning and Environment P.O.Box 450 2260 MI3 LEIDSCHENDAM The Netherlands tel.: 70-3174350 fax: 70-3174449

Kooij, J.van der N.V. SEP P.O.Box 575 6800AN ARNHEM The Netherlands

Kops, J.A.M.M. KEMA P.O.Box 9035 6800ET ARNHEM The Netherlands tel.: fax:

Koster, L. Shell Nederland B.V. P.O.Box 1222 3000BE RO'ITERDAM The Netherlands el.: 10-4696035 fax:

Koviics, E. Ministry for Environment and Regional Policy Institute for Environment Protection Aga 4-4 H- 1113 BUDAPEST Huncarv tel.: fax:

Kros. J. The Winand Staring Centre for integrated land, soil and water research P.O.Box 125 6700 AC WAGENINGEN The Netherlandq tel.: 8370-74366 fax: 8370-24812

Landmann, G. Programme DEFORPA, Centre de Recherches Forestiers - INRA 54280 CHAMPENOUX France tel.: 33-83394072 fax: 33-83317160

Leaf. D.A. US EPA 401 M Street S.W.(ANR-445) WASHINGDON DC 20460 USA tel.: 202-2609306 fax: 202-2600892

Lee, J. University of Manchester, Department of Environmental Biology

Lekkerkerk, Mrs.L.J.A. IKC-V&M P.O.Box 42 6710BA EDE The Netherlands tel.: 8380-71461 fax: 8380-24737

Oxford Road MANCHESTER M13 9PL United Kinedom tel.: 44-61-2753888 fax:

tel.: 85-430858 fax: 85-430858

579 Loch, J. Swiss Federal Institute for Forest. Snow and Landscape Research WSL) Zurcherstrasse 1 1 1 . BIRMENSDOFW 8903 Switzerland tel.: 1-7392486 fax: 1-7392488

Lubbers, F. N.V. SEP Utrechtseweg 310 6812 AR ARNHEM The Netherlands tel.: 85-721 111 fax: 85-430858

Lumme, E. Ministry of the Environment, Air Pollution Control and Noise Abatement Division P.O.Box 399 SF HELSINKI Finland tel.: fax:

Marseille, Mrs.H. Ministry of Housing, Physical Planning and Environment P.O.Box 450 2260 MB LEIDSCHENDAM The Netherlands tel.: 70-3174402 fax: 70-3174449

Matzner, E. Lehrstuhl fur Bodenkunde, BITOK, Univ.Bayreuth Postfach 101251 8580 BAYREUTH FRG tel.: 921-553508 fax:

Mill, W.A. ul.Kosutha 6 41-8 KATOWICE Poland tel.: 1540005 fax:

Minnen, J.G.van National Institute of Public Health and Environmental Protection P.O.Box 1 3720 BA BILTHOVEN The Netherlands tel.: 30-743704 fax: 30-250740

Mohren, G.M.J. De Dorschkamp, Institute for Forestry and Urban Ecology P.O.Box 23 6700 AA WAGENINGEN The Netherlands tel.: 8370-95322 fax: 8370-24988

Nagel, H. Institut fiir Oekosystemforschung Magdalenstr. 17-19 0- 1 130 BERLIN Germany tel.: fax: 37-2-23722034

Nilsson, J. Vattenfall 16287 VALLINGBY Sweden tel.: 8-7395000 fax:

Novo, A. ENEL-CRTN V.Rubattino 54 20 134 MILAN0 Italie tel.: fax:

Paces, T. Czech Geological Survey Malostranske Nam. 19 PRAHA 11821 Czechoslovakia tel.: 2-590282 fax: 2-7980965

Pasma, B.R. Province of Limburg P.O.Box 1094 6202 MA MAASTRICHT The Netherlands tel.: 43-897565 fax: 43-618712

Plantaz, M.A.H.G. Netherlands Energy Research Foundation, ECN P.O.Box 1 1755 ZG PETTEN The Netherlands tel.: 2246-4006 fax: 2246-3488

580

Puckett, K.J. Environment Canada, Air Quality Processes, Research Div. 4905 Dufferin Street DOWNSVIEW Ontario Canada M3H 5T4 tel.: 1-416-7394836 fax: 1-416-7395708

Pul, W.A.J.van National Institute of Public Health and Environmental Protection P.O.Box 1 3720 BA BILTHOVEN The Netherland% tel.: 30-7428 18 fax: 30-287531

Reijnders, L. Netherlands Society for Nature and Environment Donkerstraat 17 351 1 KB UTRECHT The Netherlands tel.: 30-33 1328 fax:

Reinds, G.J. Winand Staring Center P.O.Box 125 6700 AA WAGENINGEN The Netherlands tel.: 8370-74364 fax: 8370-24812

Reuther, M. GSF-Forschungszentrum Ingolstadter Landstrasse 1 8042 NEUHERBERG FRG tel.: 89-31873086 fax: 89-31873365

Salm, C.van der Winand Staring Centre P.O.Box 125 6700 AC WAGENINGEN The Netherlands tel.: 8370-74326 fax: 8370-24812

Schiirer, B. Umweltbundesamt Bismarckplatz 1 D-100 BERLIN33 Germany tel.: fax:

Schliipfer, R. Swiss Federal Institute for Forest, Snow and Landscape Research CH-8903BIRMENSDORF Switzerland tel.: 41- 1-7392225 fax: 41 -1-7392215

Schneider, Mrs.M. Nassauplantsoen 7 3761 BH SOEST The Netherlands tel.: fax:

Schneider, T. National Institute of Public Health and Environmental Protection P.O.Box 1 3720 BA BILTHOVEN The Netherland% tel.: 30-742970 fax: 30-250740

Schutter, M.A.A. National Institute of Public Health and Environmental Protection P.O.Box 1 3720 BA BILTHOVEN the Netherlands tel.: 30-742938 fax: 30-287531

Slanina, J. Netherlands Energy Research Foundation, ECN P.O.Box 1 1755ZG PETTEN The Netherlands tel.: 2246-4236 fax: 2246-3488

Slobbe, P.T.M. Ministry of Economy P.O.Box 20101 2500EC DENHAAG The Netherlands tel.: 70-3796420 fax: 70-3796210

Smet. P.A.M.de National Institute of Public Health and Environmental Protection P.O.Box 1 3720 BA BILTHOVEN The Netherlands tel.: 30-743528 fax: 30-250740

58 1 Smeulders, S.H. Ministry of Housing, Physical Planning and Environment P.O.Box 450 2260 MB LEIDSCHENDAM The Netherlands tel.: 70-3174395 fax: 70-3174449

Smits, T.F.C. IKC-NBLF P.O.Box 20023 3502LA UTRECHT The Netherlands tel.: 30-852330 f a : 30-891864

Soveri, J. Hydrological Office P.O.Box 436 00101 HELSINKI Finland tel.: 358-0-7314289 fax:

Staaf, H. Swedish Environmental Protection Agency Research Dept. S-17185 SOLNA Sweden tel.: 46-8-79914.42 fax: 46-8-283008

Steenis, Mrs.O.van National Institute of Public Health and Environmental Protection P.O.Box 1 3720 BA BILTHOVEN The Netherlands tel.: 30-742970 fax: 30-250740

Suaaten, A.van Ministry of Amiculture, Nature and Fisheries P.O.B&x 20461 2500 EK DEN HAAG TheNetherlands tel.: 70-3793280 fax: 70-382572

Teunenbroek, T.van AMECO Hooghiemstraplein 96 3514 AX UTRECHT The Netherlands tel.: 30-769292 fax: 30-720909

Thomelof, E. Swedish Environmental Protection Agency 5-17185 SOLNA Sweden tel.: 46-8-7991280 fax: 46-8-283008

Tiktak, A. National Institute of Public Health and Environmental Protection P.O.Box 1 3720 BA BILTHOVEN The Netherlands tel.: 30-743367 fax: 30-74297 1

Venis-Pols, Mrs.P.W.A.M. SCMO-TNO P.O.Box 186 2600AD DELFT The Netherlands tel.: 15-696885 fax: 15-613186

Verhoeff, K. Ministrv of Amiculture. Nature and Fisheries P.O.Bo;( 20461 2500 EK DEN HAAG The Netherlands tel.: 70-37921 10 fax:

Vermeulen, A.T. Netherlands Energy Research Foundation, ECN P.O.Box 1 1755ZG PE'ITEN The-Netherlands tel.: 2246-4194 fax: 2246-3488

Vonk, Mrs.M.B. National Institute of Public Health and Environmental Protection P.O.Box 1 3720 BA BILTHOVEN The Netherlands tel.: 30-74321 1 fax: 30-250740

Vonkeman, G.H. Committee for Long Term Environmental Policy P.O.Box 90740 2509LS DENHAAG The Netherlands tel.: 70-3472351 fax: 70-3857901

I _

582 Voorburg, J.H. Mansholtlaan 10-12 6708 PA WAGENINGEN The Netherlands tel.: 8370-76595 fax: 8370-25670

vos, w. IBG De Dorschkamu P.O.Box 23 6700 AA WAGENINGEN The Netherlands tel.: 8370-95412 fax: 8370-24988

Vries, W.de The Winand Staring Centre for integrated land, soil and water research P.O.Box 125 6700 AC WAGENINGEN The Netherlands tel.: 8370-74477 fax: 8370-24812

Waters, D. Institute of Hydrology WALLINGFORDOX10 8BB United Kingdom el.: 491-3880 fax: 491-32256

Webb, J.D. ADM Nobel House, 17 Smith Square LONDON SWlP 3JR United Kingdom tel.: 71-238-5698 fax: 71-238-5732

Weiss, D.I. Umweltbundesamt Spittelaverlande5 1090 W E N Austria tel.: 22-31304316 fax: 22-31304900

Werner, L. Institut fiir Oekosystemforschung Magdalenstrs. 17-19 0-1 130 BERLIN Germany tel.: fax: 37-2-23722034

Whitehead, P.G. Institute of Hvdrologv Crowmarsch Gifforr WALLINGFORD, OXON OX10 8BB United Kingdom tel.: 491-38800 fax: 491-38097

Wijers, G.P. Netherlands Energy Research Foundation, ECN P.O.Box 1 1755ZG PE'ITEN The Netherlands tel.: 2246-4155 fax: 2246-3488

Wijnheijmer, E.H.M. Departmentof AgriculturalResearch P.O.Box 59 6700 AB WAGENINGEN The Netherlands tel.: 8370-74340 fax: 8370-24060

Wild, P.J. Netherlands Energy Research Foundation, ECN P.O.Box 1 1755ZG PETTEN The Netherlands tel.: 2246-4006 fax: 2246-3488

Wilson, R.B. Department of the Environment Room B 352, Romney House 43 Marsham Street LONDON SWlP 3PY United Kinedom tel.: 71-2768316 fax: 7 1-2768299

Wolters, G.J.R. Ministry of Housing, Physical Planning and Environment P.O.Box 450 2260 MB LEIDSCHENDAM The Netherlands tel.: 70-3174366 fax: 70-3 174449

Wortelboer, F.G. National Institute of Public Health and EnvironmentalProtection P.O.Box 1 3720 BA BILTHOVEN The Netherland3 tel.: 31-30-743128 fax: 31-30-250740

583 Wright, R.D. National Power, NP TEC LEATHERHEAD KT22 7SE United Kingdom tel.: 372-375063 fax: 372-375515

Wiister, H. UN/ECE Palais des Nations CH 1211 GENEVA 10 Switzerland tel.: 41-22-734601 1 fax: 41-22-7349825

Zijlstra, W.M.

Zuidema, F.C. Ministry of Agriculture, Nature and Fisheries P.O.Box 20401 2500EK DENHAAG The Netherlands tel.: 70-3793459 fax: 70-3478167

VNO P.OTBox 93093 2509 AB DEN HAAG The Netherlands tel.: 70-349748 1 fax: 70-3850707

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