Acid Rain: Report number 14

The views expressed in this Report are those of the authors of the papers and contributors to the discussion individually and not necessarily those of their institutions or companies or of The Watt Committee on Energy Ltd. Published by: The Watt Committee on Energy Ltd 18 Adam Street London WC2N 6AH Telephone: 01-930 7637 This edition published in the Taylor & Francis e-Library, 2005. “To purchase your own copy of this or any of Taylor & Francis or Routledge’s collection of thousands of eBooks please go to www.eBookstore.tandf.co.uk.” © The Watt Committee on Energy Ltd 1984 ISBN 0-203-21031-X Master e-book ISBN

ISBN 0-203-26813-X (Adobe eReader Format) ISSN 0141-9676

THE WATT COMMITTEE ON ENERGY REPORT NUMBER 14

ACID RAIN

Papers presented at the Fifteenth Consultative Council meeting of the Watt Committee on Energy, London, 1 December 1983

The Watt Committee on Energy Ltd A Company limited by guarantee: Reg. in England No. 1350046 Charity Commissioners Registration No. 279087

AUGUST 1984

Contents

page Members of Acid Rain Working Group

iv

Foreword

vi

Introduction Section 1

The fate of airborne pollution

viii 1

1.1

Introduction

2

1.2

Emissions

5

1.3

Transformations

7

1.4

Contribution of motor vehicles

10

1.6

Loss processes to above the boundary layer

12

1.5

Deposition

10

1.7

Networks and measurements

13

1.8

Trends in acid precipitation

15

1.9

Modelling

16

1.10

Episodes

18

1.11

Some important questions

19

1.12

Recommendations

22

Vegetation and soils

24

2.1

Historical aspects

25

2.2

Effects of ambient air pollution on metabolism and growth of plants

29

2.3

Effects of acid precipitation on plants and soils

37

2.4

The Ulrich Hypothesis on effect of acid depositions on forest ecosystems

42

2.5

Lichens as indicators of atmospheric sulphur pollution

45

2.6

Central European forest die-back and its relevance to Britain

47

Section 2

iv

Section 3

Freshwater

51

3.1

Introduction

52

3.2

Techniques for assessing acidification

53

3.3

Effects of acidification on flora and fauna

54

3.4

Evidence for acidification outside U.K.

55

3.5

Evidence for acidification in U.K.

57

3.6

Extent and distribution of acid waters in Great Britain

59

3.7

Conclusions

63

Section 4

Remedial strategies

65

4.1

Technologies for reducing emissions

66

4.2

Impact of SO2 control on electricity prices

73

4.3

Existing and prospective legislation governing British emissions

76

4.4

Scientific support for an acid rain strategy

79

4.5

Summary

83

Fifteenth Consultative Council meeting of the Watt Committee on Energy

85

Abbreviations

86

Appendices 1. 2

THE WATT COMMITTEE ON ENERGY Member institutions

88

Policy

90

Members of Executive, June 1984 Recent Watt Committee Reports 57

90

Members of Acid Rain Working Group Chairman Professor K.Mellanby CBE Institute of Biology Dr Helen ApSimon Dr R.W.Battarbee Dr J.N.B.Bell Dr W.O.Binns T.R.Carrick Dr M.J.Chadwick A.J.Clarke

Imperial College of Science & Technology, London University College London Imperial College, London Institute of Chartered Foresters Freshwater Biological Association University of York Central Electricity Generating Board

v

Dr A.T.Cocks D.H.Crawshaw C.J.Davies Dr B.E.A.Fisher J.A.Garland D.Hammerton R.Harriman N.H.Highton A.V.Holden A.S.Kallend Dr A.W.C.Keddie Dr J.A.Lee Dr P.S.Maitland Professor T.A.Mansfield Dr A.R.Marsh C.Martin Dr H.G.Miller G.S.Parkinson Dr M.A.Plint Dr P.Roberts Dr R.A.Skeffington Dr F.B.Smith Dr J.H.Stoner Dr M.H.Unsworth (observer) Secretary J.G.Mordue

Royal Society of Chemistry North West Water Authority Operational Research Society Institute of Mathematics and its Applications Royal Meteorological Society Institution of Water Engineers and Scientists Dept of Agriculture & Fisheries for Scotland Beijer Institute, Stockholm Pitlochry Central Electricity Generating Board Warren Spring Laboratory Victoria University of Manchester Institute of Terrestial Ecology University of Lancaster Central Electricity Research Laboratories Institution of Public Health Engineers Institute of Chartered Foresters Institute of Petroleum Royal Geographical Society International Flame Research Foundation Central Electricity Research Laboratories Royal Meteorological Society Institution of Water Engineers and Scientists Institute of Terrestial Ecology

Note The Working Group appointed four sub-groups; the membership of each is listed on the first page of the relevant section of this Report.

Foreword

What is loosely described as ‘acid rain’ is not a new phenomenon. The burning of coal and other fossil fuels must have always resulted in the production of sulphur dioxide, and, where the combustion temperatures are high, of oxides of nitrogen. These may be present in various stages of oxidation and are often referred to as simply SOx and NOx. The Clean Air Act 1956 with its limitations on the burning of raw coal in urban areas has virtually eliminated ‘smog’ in British cities but has not directly reduced the SOx emissions. It is only during the last decade or so that Acid Rain has become a topic of discussion vying with nuclear energy in its emotive power. Initially attention was mainly concerned with the alleged effect of these gases and the acids formed therefrom on lakes and rivers in Scandinavia. This concern was soon followed by reports of serious damage to, for instance, the Black Forest, and, more locally, to lakes in the Galloway area and damage in other parts of Scotland. In the case of these and many other examples, suggestions, still to be verified, have been made about the probable origin of the pollutants. During a general discusion about the future programme of the Watt Committee, Dr A.A.L.Challis, then Chief Scientist at the Department of Energy, told me that acid rain was a field in which we could be of real help. Much of the large amount of data available was self-contradictory, and though Dr Challis did not imagine that we could provide a solution to the problem, we might, he thought, be able to clarify it and give a lead as to the work needed to give a truer picture. A wide range of experts—not just physicists, chemists and mathematicians, but experts in combustion chemistry, meteorology, aerodynamics, biology, water engineering and forestry—was called for, all of whom could be brought together through the existing Watt Committee organisation. A number of our member institutions had, indeed, indicated their support for the idea of a Watt Committee study of acid rain. Knowing of Professor Kenneth Mellanby’s wide knowledge and enthusiasm in this field, I asked him if he would be prepared to lead such a team. After a short hesitation he agreed, and said that he would aim at producing the first draft of a report in about three months, leading to a series of papers for our next Consultative Council meeting on 1st December 1983. This he did, and the general reaction was that the work, having made so much progress in such a short time, must be continued. The first task was to get the present report into print: this would enable us to decide where further work was most needed. In spite of the many hours that he and the thirty or so people involved in the present report had given, Professor Mellanby agreed to continue the study, selecting his team in line with the questions to be answered. As Chairman of the Watt Committee I should like to thank them all for the splendid job that they have done and for their willingness to continue. Finally, I would emphasise again that we are still clarifying the

vii

problem, and can only be any help because of the far deeper studies and more intensive research done by other workers in Europe and North America. April 1984 J.H.Chesters Chairman, The Watt Committee on Energy

Introduction

When I was asked to be Chairman of the Watt Committe’s working group on acid rain, I said that we had three questions to answer. These questions are: (1) Is there an acid rain problem? (2) If so, what is the cause of the problem? (3) How can this problem be solved? The papers here produced show how far we have gone to answer these questions. It will be clear that, though considerable progress has been made, simple and comprehensive answers cannot, as yet, be given. First, is there a problem? There is no doubt that environmental damage—to crops, to trees, to buildings and to freshwater—does occur, and that, at least in some cases, this is caused by acid emissions from the use of energy derived from fossil fuels. However, it is not easy to quantify this damage, or to be sure that ‘acid rain’ is always its main cause. This brings us to the second question. We must be careful not to be misled by spurious correlations between damage and the emission of substances which may, or many not, cause the damage. Here it must be pointed out that it is clearly necessary to consider all the substances emitted when fossil fuels are burned and when the use of energy by industry affects the environment. It is often gaseous emissions that are involved, and the rain, as such, may not always be as important. We must also realise that many different substances—oxides of nitrogen and ozone in particular—may be involved in addition to sulphur dioxide. It is obvious that, if we cannot as yet give clear answers to the first two questions, we cannot answer the third with any precision. Thus we must be very careful not to advocate expensive ‘solutions’ which may, in the end, be found to have little effect. The impression given by the media, that acid rain is a new problem, is quite false. As long ago as 1872 R.A. Smith, the first British Alkali Inspector, published a book on chemical meteorology, in which he described rain from industrial areas as acid as, or even more acidic than, anything found today. For a hundred and fifty years we have recognised the harmful effects of air pollution, a serious local problem in our cities and industrial areas. In fact, of course, many of the harmful effects of air pollution have been greatly reduced in recent years. The pea-soup fogs described by Dickens have disappeared from London and our main cities. The damage to plants and buildings from gaseous pollution, mainly sulphur dioxide, has been greatly reduced. In fact, we now find that roses in city parks are once more attacked by black spot; the fungus that causes this disease was previously killed by sulphur dioxide to which the rose bushes were more resistant. Urban air pollution in Britain has been largely conquered by the policy of ‘dilute and disperse’. By discharging toxic flue gases from high chimneys, ground-level concentrations of sulphur dioxide no longer

ix

cause acute phytotoxicity. Further from the source of pollution, even greater dilution reduces the sulphur dioxide to apparently harmless levels. Until recently we believed that we had by these means completely conquered this type of pollution. We did not alway recognise that harmful effects may be produced by these very dilute gases at great distances from their sources, probably because we did not fully realise that chemical changes may occur to the substances discharged, producing compounds that have quite different properties. We need to distinguish between primary pollutants, mainly oxides of sulphur and nitrogen, which are emitted as such in industrial areas, and secondary pollutants, the most important of which are sulphates and sulphuric acid derived from SO2 and nitrates and nitric acid derived from oxides of nitrogen. The transformations from primary to secondary pollutants, particularly of sulphur, may be slow, and may take several days during which the contaminated air may be blown a great distance—perhaps from Britain to Norway, or from the Ruhr to Sweden. I have found it useful to describe precipitation containing mainly primary pollutants as ‘primary acid rain’, and that containing secondary pollutants as ‘secondary acid rain’. Not all my colleagues agree with this distinction, which may not always be clear-cut. Rain falling at intermediate distances from pollution sources may contain both primary and secondary pollutants. Nevertheless, I think that the general concept may be useful. Primary acid rain is a local phenomenon, and is often accompanied by high levels of gaseous pollution. Where damage to plants and buildings occurs, it is generally the gases, and not the wet deposition, which are mainly to blame. It is primary acid rain that was described over a hundred years ago by R.A.Smith, and it is its analysis that often hits the headlines today. Though primary acid rain must have some effect on soils and poorly buffered fresh waters, its effects on buildings and vegetation seem generally to be minimal, and so it can sometimes be somewhat of a ‘red herring’. Our working group has therefore devoted its main attention to the important and difficult problem of secondary acid rain, to a study of its production and of its effects. Secondary acid rain is very dilute, so it seldom causes direct phytotoxicity. It usually falls where the air is clean, with low levels of oxides of sulphur and nitrogen. However, in high-rainfall regions soils that contain little calcium and poorly buffered fresh waters may be seriously affected by the great volume of dilute acid falling on them. I may appear to have given greater importance to problems of gaseous pollution, and to have underestimated the importance of ‘acid rain’. This is certainly not my intention. However, I think it is essential that we distinguish between the various processes and reactions involved. We know a good deal about the control of gaseous pollution, very little about how to reduce or eliminate the effects of wet pollution. We still do not fully understand the causes of fishless lakes and dying trees, bu we can be sure that many of the simplistic explanations often circulated do not give us a true picture of the situation. We are still very far from being able to give firm and simple solutions to all the problems concerning acid emissions. We can control, and have controlled, the gross damage caused by gaseous pollution in many areas. We know how, at a cost, emissions may be reduced. We still do not know how the reduction of emissions would reduce the damage from secondary acid rain. We have to unravel the problem, identifying the causes, processes and effects We have to investigate all the methods for reducing emissions, including the improved combustion of fuels, to reduce the amounts of toxic substances in the flue gases. We need to assess the value of remedial methods such as the liming of acid lakes, which may alleviate the situation while a more fundamental solution is being discovered. However, one other important method is clearly available. This is to be more economical in the use of energy and of fossil fuels. The Watt Committee in a previous report* showed that this was possible without reducing our productivity or our standard of living. Also the substitution of renewable sources of energy—

x

solar power, hydroelectric generation, wind power—could make a moderate but significant contribution. The greater use of nuclear power could also be important. So there are many ways of tackling the problem, but first and most important we need to understand just what the problem really is. Hill Farm, Wennington, Huntingdon. March 1984 KENNETH MELLANBY

* The rational use of energy. Watt Committee on Energy Report No. 3, London, 1978.

THE WATT COMMITTEE ON ENERGY REPORT NUMBER 14

Section 1 The fate of airborne pollution F.B.Smith

This paper presents the work of Sub-group 1 (Production, Transformation, Transport and Deposition) of the Watt Committee working group on Acid Rain. Membership of Sub-group 1

Dr F.B.Smith (Chairman) Dr H.ApSimon Dr A.T.Cocks Dr B.Fisher Dr J.Garland Dr A.S.Kallend Dr A.W.C.Keddie Dr A.R.Marsh Professor K.Mellanby G.S.Parkinson Dr M.A.Plint Dr M.H.Unsworth (observer)

The fate of airborne pollution

1.1 Introduction Gaseous and particulate pollution injected into the atmosphere can be subjected to many physical and chemical processes. The main processes, shown schematically in Figure 1.1, are outlined only briefly in this introduction, but are described more fully later. In general, many different pollutant species are found in the atmosphere. Those of principal importance for acid deposition are sulphur dioxide, nitric oxide, ammonia, non-methane hydrocarbons and their derivatives. These are derived from different sources, some natural, others resulting from man’s activities. In Europe, the latter predominate for most of the important species. These ‘anthropogenic’ sources have a spatial distribution that requires careful determination.* In principle, it is important to know the elevation of each source above ground, any inherent plume rise due to buoyancy, the average magnitude of the emission and any trends and cycles that may be evident. Emissions often exhibit diurnal, weekly and annual cycles in response to man’s needs and working patterns, as well as more unpredictable variations in response to weather conditions, for example. There is thus a level of basic uncertainty in emission values from any area at any time, even in those areas of Europe where detailed analyses have been made. Having entered the atmosphere, the pollutants are advected away by the wind. The principal property of the wind in this respect is its direction. Both the wind direction and the wind speed are determined by

Figure 1.1 Processes involved in deposition of atmospheric pollutants.

THE FATE OF AIRBORNE POLLUTION

3

objectively analysing winds and pressures measured at a rather sparse network of meteorological observing stations (typically positioned 50–100 km apart over land) at certain fixed hours of the day to form continuously changing wind fields. Clearly the interpolated wind field is not exact, particularly in the case of complex topography, and this causes errors in pollution trajectories that grow exponentially with distance. There has been only limited testing of our ability to predict the trajectory in the atmosphere of pollution emitted from a single point source, and then usually in relatively simple meteorological situations and out to distances of only a few hundred kilometres. However, most important sources in reality have broad geographical distribution and our interest lies in depositions over some period of time: both these factors tend to reduce the effective errors to within acceptable limits when ecological consequences are assessed. In addition to advection, the motion of the air is usually turbulent. This turbulence causes dispersion, that is, mixing of the pollutant plume with the surrounding atmosphere and with plumes from other sources. In consequence, different species may be brought together, resulting in chemical reactions; in addition, they may be dispersed down to the surface, where they may be partially deposited, and to the top of the mixing layer, where some may be temporarily ‘lost’, often in a sporadic manner, to the less turbulent atmosphere above. Chemical reactions may take place in cloud, often during the uptake into growing raindrops. These reactions are sometimes rapid, implying that what comes out in the rain is physically and chemically very different from that which entered the cloud. Other transformations take place out of cloud in what are called gas-phase reactions. On the whole, these changes take place more slowly, transforming the primary emitted pollutants (e.g. sulphur dioxide (S02) and nitrogen oxide (NO))* to secondary pollutants (e.g. sulphate and nitric acid) at a typical rate of a few per cent per hour. Thus, in very general terms, primary pollutants are important at short range near to the main source regions whereas secondary pollutants become relatively more important at long distances including the more remote parts of Europe. Owing to the transformations involved, the deposition Table 1.1 European and U.K. emissions since 1900, million tonnes per year 1900 SO2 Europe 16 SO2 U.K. 2.8 NOx U.K. 0.68 NOx expressed as equivalent NO2.

1910

1920

1930

1940

1950

1960

1970

22 3.2 0.69

22 3.2 0.72

25 3.2 0.72

25 3.6 0.82

25 4.6 0.99

32 5.6 1.35

52 6.0 1.64

Table 1.2 U.K. emissions of SO2 and NOx since 1971, million tonnes per year SO2 Power Station s All source s

*

1971

1972

1973

1974

1975

1976

1977

1978

1979

1980

1981

1982

2.80

2.87

3.02

2.78

2.82

2.69

2.74

2.81

3.10

2.87

2.71

2.60

5.83

5.64

5.80

5.35

5.13

4.98

4.98

5.02

5.34

4.67

4.23

4.00

Anthropogenesis is the study of the origin of man. The word has sometimes been wrongly used elsewhere, but in this Report the expressions ‘man-made’ and ‘man-induced’ are generally preferred, where appropriate. *See Appendix 2.

4

INTRODUCTION

1971

1972

1973

1974

1975

1976

1977

1978

1979

1980

1981

1982

NO2 — 0.73 0.81 0.72 0.76 0.79 0.79 0.81 0.88 0.85 0.82 0.77 Power Station s Vehicl — 0.42 0.45 0.44 0.47 0.45 0.46 0.48 0.49 0.49 0.48 0.49 es* All — 1.73 1.85 1.72 1.70 1.74 1.77 1.80 1.89 1.79 1.71 1.67 source s CO — 7.86 8.30 8.07 7.80 8.06 8.27 8.62 8.78 8.85 8.62 8.83 All source s* ‡ Total — 2.40 2.60 2.73 2.77 2.93 3.03 3.16 3.35 3.37 3.36 3.29 hydroc arbons (metha ne equiva lent weight ) * These are the best estimates currently in a consistent format, but recent W.S.L. work suggests that vehicles emit somewhat more NOx ( 40% of total) and significantly less CO. ‡ Principally vehicles. NOx represented as equivalent NO2.

properties and the ecological effects may be quite different at the two ranges. Different species are absorbed, sedimented and deposited by impaction on to the underlying surface at different rates according to their chemical and physical character and the nature of the surface itself. The whole process is called ‘dry deposition’. Typically, about half the sulphur dioxide emitted into the turbulent mixing layer of the atmosphere is dry-deposited (ignoring its partial conversion to sulphate) in about 30 hours during which time it may have travelled some 1000 km. Other chemically reactive gases tend to be deposited at much the same rate, whereas aerosols, like sulphate, tend to be dry-deposited very much more slowly. As Figure 1.1 shows, the other main deposition process is due to precipitation (rain or snow) and is called ‘wet deposition’. The main route for wet deposition is through the involvement of pollution in developing cloud droplets within the cloud itself. This process, termed ‘rain-out’, tends to be a very efficient removal process, although the exact removal rate is imprecisely known and almost certainly varies with the amount and character of the pollution, the microphysics of the cloud itself and the average size and temperature of the raindrops or snowflakes involved. Less important in general is the take-up of pollution into falling precipitation below cloud. Because precipitation is an occasional event, so is wet deposition. Wet deposition fields over a short period of time, up to a month or two, are therefore intrinsically very patchy and very difficult to assess from conventional meteorological measurements. A very heavy shower occurring between meteorological stations can sometimes affect a heavily polluted airmass and result in an undetected large local deposition (perhaps

THE FATE OF AIRBORNE POLLUTION

5

Figure 1.2 Sulphur dioxide emissions in Great Britain in 1982 in thousands of tonnes; National Grid squares 20 km×20 km. Data by courtesy of Warren Spring Laboratory.

10% or more of the annual total). Situations like this, in which significant deposition occurs within a relatively short period of time, are called ‘episodes’. A third process exists which may be rather important in some mountainous areas. This is the process of fog (or occult) deposition. Fog or low cloud droplets can contain high concentrations of pollution and, when blown by the wind over hills and mountains covered by trees, heather and other scrub, can be deposited on to the vegetation very efficiently. The contribution this makes to the long-term deposition in such areas is not very well known, and it is certainly a contribution that would be very difficult to model except in a statistical way. In general, dry deposition accounts for some two-thirds of the total sulphur deposit in Europe. Wet deposition and fog deposition account for the remainder. These last two are particularly important in areas of high rainfall, especially in mountainous areas to the east or northeast of significant source areas. 1.2 Emissions Table 1.1 gives estimates of sulphur dioxide emission values for the United Kingdom and for Europe since the beginning of the century and corresponding values for nitrogen oxides for the U.K. alone. As would be expected, they all show a very marked increase up to 1970. Table 1.2 gives the U.K. situation in more detail between 1971/72 and 1982. It shows a fairly persistent downward trend in sulphur dioxide emissions from 5830000 tonnes in 1971 to an estimated 4000000 tonnes in 1982, although significant year-to-year variations about this trend clearly exist. Power-station emissions have declined only slightly over this time, and in

6

INTRODUCTION

consequence they constitute now roughly two-thirds of all emissions compared with less than a half in 1971. Figure 1.2 shows how the total SO2 emissions were distributed spatially across the U.K. in 1982. The nitrogen oxide emissions (given as equivalent NO2) are also given in Table 1.2. In reality the emissions include Table 1.3 Annual emissions of sulphur dioxide, oxides of nitrogen and non-methane hydrocarbons for European countries about 1 978 SO2

NOx

NMHC

Albania 100 10 10 Austria 430 275 280 Belgium 760 410 390 Bulgaria 1000 240 240 Czechoslovkia 3000 600 600 Denmark 456 240 220 Finland 540 200 200 France 3600 1650 2000 G.D.R. 4000 680 680 F.R.G. 3630 3350 2450 Greece 704 500 260 Hungary 1500 220 220 Iceland 12 10 15 Ireland 174 90 105 Italy 4400 1550 1750 Luxembourg 48 50 30 Holland 480 700 600 Norway 150 110 170 Poland 3000 1000 1000 Portugal 168 110 200 Romania 2000 460 460 Spain 2000 850 1050 Sweden 550 260 380 Switzerland 116 160 260 Turkey 483 600 600 Western U.S.S.R. 16000 5000 5000 U.K. 5020 1800 1158 Yugoslavia 2950 210 210 Estimated values in thousands of tonnes. NOx expressed as equivalent NO2. NMHC values represented by total mass.

both NO and NO2, the former tending to predominate. The totals show little overall trend, although from 1975 the levels first increased systematically until a maximum was reached in 1979, since which time they have decreased, mainly as a result of reductions from industry and from power stations. The contribution from vehicles has shown a small increase over the period—of the order of 15% — and now constitutes

THE FATE OF AIRBORNE POLLUTION

7

roughly 30% of the total emissions, compared with about 45% from power stations (but see the footnote to Table 1.2). Emissions of carbon monoxide have increased by roughly 10% and those of total hydrocarbons by about 30% over the same period. Of the latter, it seems that the relevant non-methane hydrocarbons (NMHC) constitute roughly one-third of the total when expressed as an equivalent weight of methane. The remaining methane is chemically very stable and is therefore fairly unimportant as far as the formation of acidic species in the atmosphere is concerned. At present it is not known for certain how much detail is required concerning the composition of NMHC, although some, but not all, modellers consider it likely that they can be ‘lumped’ together as one species beyond a few tens of kilometres from significant sources. It is possible that control of NMHC emissions could be a way of controlling the formation of acidic species. European emissions are given by country for a period around 1978 in Table 1.3. The uncertainty in the SO2 emissions is estimated to be 10–15% at best, and considerably larger for many countries, especially those in eastern Europe. The NOx and NMHC emissions are even more uncertain. Considered as average emissions per head of population, in Western Europe, values for SO2 all range below 0.1 t per year, whereas in Eastern Europe values range up to 0.25 t per year, largely reflecting the nature of the fuel used. Emissions of NOx are much more uniform, however, at around 0.03–0.04 t per year, with only the F.R.G.* standing out as exceptional with 0.055 tonnes. Similarly, for NMHC the U.K. emissions at 0.02 t per year per head contrast with other countries’ values of around 0.04 t per year, for reasons that are not entirely clear, although different assumptions about sources and emission factors are a probable reason. The tables presented here do not contain the contribution of natural emissions. These are estimated to be some 10% of the total sulphur emissions for Europe. Background air coming into Europe from the Atlantic carries some further sulphate (typically at a concentration of 0.5–2 µg m−3). The origins of this sulphate are still being debated; it is possible some of it comes from North America and other very distant man-induced sources. This sulphur comprises a small fraction of the total airborne amount available for deposition to the ground surface of Europe (typically 5%), although in more remote areas, like northern Scandinavia, the background contribution can be dominant. Natural emissions of NOx are not thoroughly quantified. Sometimes air concentrations at remote sites are surprisingly large and are not readily explained. Lightning is thought to contribute some 20% of the natural generation over the world as a whole, but its incidence obviously varies considerably in time and space. Figure 1.3 shows the sulphur dioxide emissions for Europe in about 1978 in 150-km grid squares prepared in connection with the EMEP Programme (see below) for studying the long-range transport of pollution.* 1.3 Transformations The major acid precursors emitted as a result of man’s activities are sulphur dioxide (SO2) and nitric oxide (NO). These materials undergo oxidation to sulphate aerosol and gaseous nitric acid respectively. Molecular oxygen, although abundant, reacts relatively slowly with SO2 and nitrogen oxides and, hence, more reactive oxidising species present at much lower concentrations in the troposphere are responsible for most of the production of acidic species. *

Throughout this Report, the Federal Republic of Germany (‘West Germany’) and the German Democratic Republic (‘East Germany’) are referred to, in abbreviated form, as F.R.G. and G.D.R. respectively.

8

INTRODUCTION

Figure 1.3 EMEP emission data in the 150 km×150 km grid network: thousands of tonnes of sulphur per year.

The major homogeneous gas-phase oxidation route for SO2 is through a reaction with the hydroxyl radical, OH: OH is produced by a complex chemical mechanism involving ozone, nitrogen oxides, carbon monoxide, hydrocarbons, water vapour and sunlight. Reaction with SO2 is not a major loss process for OH and it is possible that some OH is reformed during the aerosol production step. Hence, the rate of the homogeneous production of acidic aerosol from SO2 is probably approximately proportional to SO2 concentration. Under typical conditions in the tropospheric boundary layer, maximum average homogeneous oxidation rate constants of around 1 % per hour are predicted. SO2 is soluble in water and the dissolved species can undergo oxidation in cloud and rain drops:

*See

list of abbreviations, Appendix 2 (page 54).

THE FATE OF AIRBORNE POLLUTION

9

SO2 solubility is inversely related to H+ concentration, and, as dissolution of SO2 produces an increase in H +, SO uptake in droplets is self-inhibiting. 2 Oxidation leads to a further increase in acidity: Although catalysed oxidation by molecular oxygen may in certain special circumstances produce efficient aqueous oxidation of dissolved SO2, ozone and hydrogen peroxide are probably the most important stable oxidants. Some ozone in the troposphere is the result of stratospheric incursion and both ozone and hydrogen peroxide are produced by complex chemical processes similar to those involved in OH production. Ozone is generally more abundant than SO2 in the boundary layer and oxidises it rapidly in aqueous solution under neutral and slightly acid conditions, but the rate is inversely related to H + concentration and is relatively low at the pH values of acid rain and clouds. The rate of reaction of hydrogen peroxide with dissolved SO2 is proportional to H+ concentration, however, and this cancels the inverse H+ concentration effect on SO2 solubility when sufficient hydrogen peroxide is available; thus a pH-independent rate of sulphate production is produced over the range of likely atmospheric circumstances. However, hydrogen peroxide is probably appreciably less abundant than SO2 under most conditions in the boundary layer and, hence, as its formation rate is low, the oxidation is limited by peroxide availability and the amount of sulphate formed may be virtually independent of SO2 in many circumstances. Model calculations indicate that aqueous oxidation in non-precipitating clouds involving ozone or hydrogen peroxide is unlikely to produce an overall rate constant for the oxidation of SO2 in the boundary layer that would be substantially greater than 1 % per hour. It has been postulated recently that aqueous droplets may scavenge oxidising radicals from the gas phase. Models, involving many unvalidated assumptions, indicate that such processes could lead to rapid oxidation of dissolved SO2. The effect of SO2 concentration on oxidant concentrations in such systems is not well defined at present. NO is initially oxidised to nitrogen dioxide in the troposphere, mainly by reaction with ozone: NO2 is oxidised homogeneously to nitric acid in the daytime by reaction with OH: The rate constant for this reaction is about 9 times that for the given corresponding reaction for SO2. As nitrogen oxides play a crucial role in determining OH concentrations, nitric acid production is unlikely to be simply related to nitrogen oxide concentrations. Heterogeneous processes in aqueous droplets are unlikely to provide a substantial pathway for nitric acid production. Overnight, NO2 can be further oxidised in the gas phase to nitrogen trioxide: This in turn can react with NO2 and water vapour to produce nitric acid: This process may be the major production route for nitric acid in some circumstances and the dependence of the reaction rate on nitrogen oxide concentrations is unlikely to be simple. Concentrations of oxidants within an expanding combustion-generated plume from a point source are generally lower than those in the ambient atmosphere because of the higher nitric oxide concentrations. Thus, rate constants for acid production would be lower although the produced acid concentrations within the plume boundaries could be higher because of the higher concentrations of primary species.

10

INTRODUCTION

1.4 Contribution of motor vehicles Motor vehicles can contribute to the acidification of rain in two ways. First, roughly a third of the NOx emissions in the U.K. are derived from motor vehicle exhausts. As stated above, nitric oxide is oxidised in the atmosphere to nitrogen dioxide and thence to nitric acid, the latter being fairly efficiently scavenged in rain systems and observed as the nitrate ion in collected precipitation. Second, motor vehicles contribute both NOx and hydrocarbons to the ambient air on a widely distributed basis. These are the essential precursors to oxidants formed photochemically in the atmosphere. The oxidation rates of both SO2 and NO diluting into the ambient air from point sources are strongly dependent upon the concentrations of these precursors. Ambient air quality is similarly thought to influence the progress of oxidation in clouds because the oxidant precursors are again hydrocarbons and NOx in the ambient air. Because of these complex interactions between pollutants from different sources, it is not possible to make a realistic quantitative assessment of the contribution of motor vehicles to acidic deposition or to the acidification of rain. 1.5 Deposition Deposition processes clean the atmosphere of pollutants and deposit them on to land or sea. Wet deposition —the removal by rain and snow—has long been studied, using funnels to collect the precipitation for chemical analysis. However, removal by rain is not the only significant deposition process. There is also dry deposition, which means dust fall and various processes for direct absorption of the pollutant by the vegetation, soil, water or buildings that form the surface. Thirdly there is the collection of wind-blown fog or cloud water on vegetation and other surfaces. These three processes ultimately remove all the acidifying materials from the atmosphere. Provided sufficient care is taken to expose the collector correctly, to avoid contamination and to analyse the samples by suitable techniques, wet deposition can be measured reliably by use of rain collectors. The contamination of the collector by dry deposition may have a small effect on the result but this effect may be largely avoided by using ‘wet-only’ collectors which place a lid over the funnel whenever no precipitation is falling. Networks of rain gauges have been used to measure the variation of wet deposition across Europe. There must be quite large uncertainties in the deposition at individual locations between the sampling points, but to some extent these average out and the total annual deposition over the area is probably known to within about 30%. Although the networks give a direct measure of the amount of deposit in precipitation, they do not tell us anything about the mechanism for removal. Sulphate and nitrate in rain could result from incorporation of particles containing these compounds or from absorption and reaction of the gaseous sulphur and nitrogen oxides. Several mechanisms have been proposed for the incorporation of the particles, and one of these, involving condensation of water on to the particles during cloud formation, seems particularly significant and capable of accounting for a large proportion of the material in rain. There is also evidence from field data that a large fraction of the gaseous compounds is incorporated into rain but the mechanisms are difficult to quantify. If sulphur dioxide is to be removed in rain, it requires an oxidation process that can operate sufficiently rapidly in the cloud droplets; although hydrogen peroxide and ozone and some free radicals are possibly capable of doing this, the factors that control the amount of sulphate formed in this fashion are not known. Uncertainty in the mechanism for removal by rain makes it difficult to be certain of the detailed effects of a reduction of emissions from a particular source area on rain chemistry.

THE FATE OF AIRBORNE POLLUTION

11

Dry deposition processes are different for particles and gases. Very large particles* (diameter d>10µm) fall slowly, like fine rain; intermediate particles (1
*

Quoted diameter ranges are to be taken as indicating orders of magnitude.

12

INTRODUCTION

Figure 1.4 Total sulphur depositions per unit area (g S m−2 per year) from the EMEP model for a two-year period (1978 and 1979).

1.6 Loss processes to above the boundary layer Like all fluids moving over a solid surface, the atmosphere has a layer called the boundary layer (or mixing layer) which is strongly affected by the ground—by its roughness and temperature. It is characterised by turbulence which diffuses pollution rather rapidly throughout the layer. It has a depth which is constantly varying in response to the changing surface beneath, but whose magnitude typically lies between a few tens of metres and about 2 km. Most emissions are into the boundary layer, although at night, when the layer is often very shallow, the major emissions may be above it and are only taken into the layer when it deepens in response to increased buoyancy forces reflecting the radiational warming of the ground when day returns. When the layer deepens, turbulence ensures that pollution is mixed through its whole depth and that air and potential oxidants, like ozone entrained from above, mix with it. When the layer shrinks, however, some of the pollution is left behind in a relatively low-turbulence airstream in which mixing is depressed. This complex behaviour significantly affects chemical transformation processes which depend on continued mixing of plumes with ambient air. It can also isolate, for a while at least, some of the pollution from dry deposition processes at the ground. Other atmospheric processes can also result in the upward transport of pollution which may take it above the range of the boundary layer. These processes include the effects of orographie uplift over mountains, slow synoptically-induced upwelling, complex flows near warm and cold fronts, breaking gravity waves and deep convective cloud. Without going into further details of these processes, it is understandable that some may be involved with the formation of rain when much of the pollution carried by the air will be washed out. However, some will not be, and the pollution may be carried within the upper troposphere over very long distances, perhaps several thousand kilometres. So far, an adequate quantification of this effect has

THE FATE OF AIRBORNE POLLUTION

13

Figure 1.5 Total sulphur depositions (g S m−2 per year) implied by interpolation between measurements made at the EMEP network of monitoring stations over a two-year period (1978 and 1979).

not been made. As far as Europe is concerned, it is not thought to be a very large item in the total budget, especially since there is a tendency for the outflow by these means to be partially balanced by an inflow from other regions. 1.7 Networks and Measurements 1.7.1 European atmospheric chemistry network (EACN) This network was started in 1955, from an earlier Swedish network, by Egner, Rossby and Eriksson whose interests were respectively agricultural, meteorological and geochemical. The U.K. established ten stations as its contribution to the International Geophysical Year 1957–58: they were reduced in 1966 to three, which have continued to the present. One of these stations, Eskdalemuir, also reports to the Background Atmospheric Pollution Monitoring Network (BAPMoN) which is a worldwide operation of more recent vintage coordinated by the World Meteorological Organisation (WMO). All the EACN stations take monthlong samples of air and rain composition.

14

INTRODUCTION

1.7.2 The OECD programme (1972–76) Following proposals submitted by the Scandinavian Council for Applied Research (NORDFORSK) to the OECD (Organisation for Economic Cooperation and Development) in May 1970, ten member countries in western Europe, including the U.K., agreed in April 1972 to participate in a cooperative technical programme to measure the long-range transport of air pollution (LRATP). The initially-stated objective was to determine the relative importance of local and distant sources of sulphur to regional air pollution and sulphur depositions; the actual work went far beyond this single objective. A network of 56 stations monitoring pH values and concentrations of sulphur dioxide and sulphate in air and precipitation on a daily basis was established, coordinated by the Norwegian Institute for Air Research (NILU). Estimates of sulphur emissions were made and models simulating the atmospheric fate of sulphur were developed and run operationally. The U.K. played a full role in this work, and mounted the first air-sampling programme, using the Meteorological Office aircraft, which helped to establish some of the basic transformation and deposition parameters required by the models. The programme established, in broad terms, the contribution that each European country makes through its emissions to the long-term depositions in every other country, and pointed to the importance of so-called ‘background’ contributions originating from outside Europe. 1.7.3 The EMEP programme (1977–) The second programme, starting in 1977, brought in many of the eastern European countries, including the USSR, recognising the truly international character of the problem, the need to extend and improve the data input and the need to involve all the emitters if reductions in emissions were to be achieved. The programme is mounted jointly under the auspices of UNECE, UNEP and WMO, and called EMEP, the European Monitoring and Evaluation Programme. The objectives are similar to those of LRTAP but also include the monitoring of some of the other major constituents in acid rain as well as the development of an understanding of their role and the development of suitable models which additionally incorporate significant air chemistry. U.K. work relevant to this programme includes a Central Electricity Research Laboratories/Meteorological Office study of the chemical and physical behaviour of the plume and its pollutants from Eggborough power station in South Yorkshire. The plume is identified by specially injected inert tracers which can be recognised in real time by the highly-instrumented Meteorological Research Flight Hercules aircraft at distances out to at least 600 km. Search for the plume is assisted by forecasting ahead with a numerical weather-prediction program. The total sulphur depositions over a two-year period according to the EMEP model are shown in Figures 1.4 and 1.5. 1.7.4 American/Canadian programmes Considerable effort is now going on in North America. Monitoring and modelling programmes, similar to those in Europe, have been under way since about 1977, but with increased vigour since 1980 following the Memorandum of Intent on Transboundary Air Pollution agreed between the U.S.A. and Canada. Many different models are being developed and tested, some highly complex, and a very ambitious and costly programme of monitoring, using ground stations and aircraft, is being established.

THE FATE OF AIRBORNE POLLUTION

15

1.8 Trends in Acid Precipitation Short-term fluctuations in precipitation composition are such that monitoring over a number of years is necessary to reveal trends. The EACN network provides the longest data set for trend analysis and the data from this have been the subject of several studies. A statistical examination of data from 120 sites, each of which has five or more years’ data, showed that 29 showed a statistically significant increase in acidity with time and five indicated a significant decrease. In Scandinavia, 12 sites with at least 18 years’ data for the 20year period 1956–75 showed increases in hydrogen-ion concentration averaging 7% per year, compared to a year-to-year scatter of 50–100% of the mean values. The quality of some of the data has been severely questioned by some workers, and changes in techniques of analysis have led to consistent but uncertain effects that bring into question some of the trends evident in these long-term records. Similarly, sulphate that was not of marine origin showed a positive significant trend at 23 sites and one negative trend out of the 120 stations. The trend in sulphate at the 12 Scandinavian sites averaged +2.5% per year. However, more stations (55) showed a significant increase in nitrate concentration with none showing a significant negative trend. The increase in nitrate concentration at the 12 Scandinavian sites was 6% per year. A detailed analysis does not bear out the expectation of a steady increase in acidity over the period 1955– 75 paralleling, for example, the increase in sulphur emissions. Rather, the monthly data show that the increased average levels of hydrogen ion, where such increases occur, arise from an increased frequency of intermittent high monthly values occurring from about 1965 onwards. At several stations this appears to have occurred quite suddenly, around 1965, leading to an apparent step in average acidity which may be associated with changes in sampling and analytical technique. However, so far it has not been possible to isolate from the data any clear anomaly that might be associated with such changes. The large monthly variation similarly masks any seasonal trend, but in Scandinavia it appears that winter deposition exceeds that in summer, having larger means and standard deviations. At one site, Kise, the amplitude of the hydrogen-ion mean concentration fluctuation corresponds to 0.15 pH units. There are few long-running rainfall analysis data sets for individual sites within the U.K., some of which show no significant trends. The longest set is from Rothamsted, where the data recorded over 120 years indicate a twofold increase in sulphate since 1930 and a fivefold increase in nitrate over the entire period, with a doubling in nitrate since 1900—increases that compare with the emissions given in Table 1.1. Measurements of pH values are available only from 1930 and also show an increase in acidity. However, there have been changes in site and operational procedures that demand that such trend data should be treated with caution. Recently, extensive rainfall composition data have been made available for Scotland, part of Lincolnshire and a few other sites in England and Wales in the Report of the U.K. Acid Deposition Group (published by Warren Spring Laboratory), but these networks have not been established long enough for trend analysis. It might be expected that the temporal and spatial variation in sulphate and nitrate and hence acidity would follow the emission pattern in both domains. However, the limitations of data availability and quantity together with the large variation of composition at any one site are such that no detailed relationship with emissions can be identified.

16

INTRODUCTION

1.9 Modelling Acid rain is concerned with the transport of material typically over distances of 1000–2000 km and travel times of a few days. These large time- and space-scales complicate the description of transport considerably. However, modelling can be used successfully as a tool to interpret measurements, primarily to link sources with observations of airborne concentrations and wet deposition, and to improve our understanding of processes. Because of the necessity of simplifying many complex processes referred to in the Introduction by some degree of parametrisation, it is not realistic at present to attempt accurate simulations of all transport and removal processes during any given single event. However, models used in a statistical sense to build up long-term average fields of concentration and deposition do appear to have a measure of success, giving values that are within a factor of two of observed values. A common approach is to describe transport in a simplified way by assuming that large air parcels (say 100 km×100 km in area and 1000 m high) move with an average velocity over the region considered, depositing and taking up pollution as they proceed. Table 1.4 Budget of annual total depositions Emitting country Receiv Area, Czech G.D. ing km2× . R. countr 103 y

Belg.

F.R.G Polan Neth. U.K. . d

Franc U.S.S Norw Other Unde e .R. ay s c.

Total

Czec 128 4.5 1.8 0.1 1.0 0.9 0.1 0.3 0.4 0.1 0 2.1 0.8 12.2 h. G.D. 108 0.6 5.5 0.1 0.9 0.3 0.1 0.2 0.2 0 0 0.4 0.3 8.6 R. Belgi 30.5 0 0.1 2.6 0.9 0 0.2 0.7 1.1 0 0 0.3 0.4 6.3 um F.R.G 250 0.2 0.6 0.2 2.7 0.1 0.1 0.3 0.5 0 0 0.4 0.4 5.6 . Polan 313 0.5 0.8 0 0.1 2.2 0 0.1 0.1 0.1 0 0.8 0.3 5.1 d Neth. 41 0.1 0.2 0.5 1.3 0 1.2 0.8 0.4 0 0 0.3 0.3 5.1 U.K. 244 0 0.1 0 0.1 0 0 3.3 0.1 0 0 0.1 0.4 4.2 Franc 544 0 0 0.1 0.2 0 0 0.2 1.4 0 0 0.3 0.4 2.7 e U.S.S 3363 0.1 0.1 0 0.1 0.1 0 0 0 1.3 0 0.3 0.4 2.5 .R. Norw 324 0.03 0.08 0.01 0.07 0.04 0.01 0.15 0.04 0.03 0.07 0.14 0.27 0.94 ay Czech., Czechoslovakia; Belg., Belgium; Neth., Netherlands; Undec., Undecided. Depositions in grams of sulphur per square metre per year. Note The budget is given by the EMEP Meteorological Synthesising Centre-West’s Lagrangian trajectory model when applied to 1978–79. It gives the estimated contributions to the average deposition in one country arising from other countries. Only a selection of countries is given and these are ranked in order of deposition magnitude. It is interesting to note that Czechoslovakia receives over 12 times as much sulphur deposition as Norway.

THE FATE OF AIRBORNE POLLUTION

17

Emitting country Receiv Area, Czech G.D. Belg. F.R.G Polan Neth. U.K. Franc U.S.S Norw Other Unde Total ing km2× . R. . d e .R. ay s c. countr 103 y Note that the figures quoted for the U.S.S.R. refer only to the area of that nation within the analysis area of the model—very roughly, that part within Europe. Reproduced by courtesy of the Meteorological Synthesising Centre—West, Oslo, Norway.

Modelling studies have indicated that an efficient removal of SO2 and sulphate in precipitation is necessary to explain routine precipitation-chemistry measurements. A high efficiency of removal would imply that the overall system was quasi-linear, at least in a statistical sense. If other evidence should point to a lower efficiency, the deposition measurements could perhaps only be explained by increasing the contribution from very long-range transport of sulphur which cannot be directly attributable to sources within Europe. Three types of model may be mentioned. 1.9.1 Statistical models Statistical models use statistics of wind, boundary-layer depth and precipitation. The models are simple and very economical to run. Resulting deposition fields may have validity, but only over long times—of at least one year— when they show surprisingly good agreement with smoothed fields based on observations. Best agreement is obtained when the sporadic nature of rainfall is recognised and incorporated, using simple statistical techniques. 1.9.2 Operational models Operational models use real hour-by-hour meteorological data, but, owing to the limitations of these data and the rather sparse observing network, they still require the use of much parametrisation and have validity over rather long times—of the order of a few months. The models used in EMEP are of this type. Table 1.4 gives EMEP’s main results. 1.9.3 Complex models Complex models are being developed in several places, especially in North America. They are essentially mesoscale models which in a sense supplement the actual observed data by generating physically plausible data through the three-dimensional equations of continuity (and motion), taking into account the effects of the known local topography. In some cases they must rely on the output of numerical weather prediction models; in others they require considerable non-standard meteorological data. All need a highly sophisticated computer facility. Their intended uses range from aiding field experiments, testing the validity of simpler models and gaining a greater understanding of the nature and importance of various complex processes that are simulated to establishing various statistics of episodes of high concentration and deposition.

18

INTRODUCTION

Figure 1.6 Daily measurements of amount and composition of precipitation made at Goonhilly by Warren Spring Laboratory (Irwin, 1983). Rainfall, sulphate, nitrate and hydrogen ions are progessively more episodic.

1.10 Episodes Two kinds of episodes, sometimes related, can be identified: episodes in terms of deposition and episodes in terms of concentration within precipitation. The importance of these kinds of episodes is a matter of continuing research and debate. So far, no important ecological damage has been identified as being dependent on a single rain event. As noted later, effects on soils and freshwaters appear to depend on longer-term average depositions. However, the longterm average depositions are often made up of contributions from relatively few large episodic depositions, which suggests that it may be important to understand and to be able to model episodes. Considering deposition episodes first, episode-days have been defined as those days with the highest wet depositions which, when summed, make up 30% of the annual wet deposition total. Figure 1.6 shows the episodic character of Goonhilly in Cornwall in 1980. In terms of sulphate, nitrate and hydrogen ions the station is highly episodic, 30% of the total deposited H+ ions being deposited in only about 5 or 6 days per year. Episodes of this type sometimes occur in a slow-moving active frontal system when convergence of moist air at low levels, sometimes enhanced by the effect of topography, produces heavy and prolonged rainfall.

THE FATE OF AIRBORNE POLLUTION

19

Episodes in concentration (that is, concentration in precipitation) may be defined as days when the concentration exceeds 3 times the long-term mean. These days occur in a very limited number of meteorological situations, principally for deposition in Northern Europe when an active front draws into itself heavily polluted air from a stagnant anticyclone covering the industrialised areas of central Europe. 1.11 Some Important Questions 1.11.1 Long-range transport of pollution In urban areas, local pollution dominates air concentrations and dry deposition. Wet deposition values, which are relatively much smaller than dry, may depend more on sources some distance away, since local pollution will generally have had insufficient time to diffuse upwards into the clouds and get taken up into the droplets of precipitation. In rural areas, remote sources normally make a larger relative contribution. The fraction of the total deposition that originates from outside the country depends very much on the size of the country and on the magnitude and location of its emissions. As is to be expected, small countries tend to receive more from other countries than do large countries. In 16 out of 28 EMEP countries, foreign sulphur depositions exceed depositions from sources within the country. The U.K. is not one of these, since few sources lie upwind in the prevailing westerly wind direction. Roughly 75% of our sulphur depositions are of national origin. Nevertheless, some of the most acidic U.K. deposition episodes are related to European sources in easterly winds. In remote areas, like western Scotland and northern Scandinavia, especially when they receive high amounts of precipitation, there seems to be a significant contribution to the sulphur deposition from sources outside Europe. Large reductions in European emissions might not therefore generate correspondingly large reductions in depositions in these areas. 1.11.2 Sources of western Britain’s acid rain Acid rain is experienced in Wales and western Scotland, for example, in areas that could be ecologically sensitive. Measurements made in 1975–76 in the OECD programme were analysed on a sector basis of incoming air trajectories. Table 1.5 and Figure 1.7 summarise the findings. They show that at Eskdalemuir in southern Scotland the highest concentrations of sulphur in air and rain are in southeasterly airflows, whereas the most acidic rain comes from the northeast. However, by far the most sulphur and acidity Table 1.5 Results from analysis of trajectories and measurements of air and precipitation concentrations at Eskdalemuir Sectors

1

2

3

4

5

6

Days, %

29

14

3

6

18

23

Concentrations In air SO2 SO4 In rain SO4

5.1 2.9 2.4

20.1 10.2 5.0

29.2 10.1 10.1

21.8 6.6 4.9

10.0 2.6 2.3

3.2 1.9 1.7

20

INTRODUCTION

Figure 1.7 Sectors at Eskdalemuir: trajectories and measurements of air and precipitation concentrations (see results tabulated in Table 1.5). Sectors Days, % H+

1

2

3

4

5

6

29

14

3

6

18

23

35

58

55

70

48

17

Depositions S 390 190 20 20 50 120 Acid 8.2 4.5 0.1 0.4 1.1 2.0 Note See also Figure 1.7. Measurements made in 1975 and 1976 during OECD Programme. Units of air concentration, µg m−3; units of S deposition, mg m−2 per year; units of acid deposition, m eq m−2 per year.

comes in frequent rains—in less concentrated amounts— in southwesterly flows off the Atlantic. The origin of this pollution may be naturally-occurring oceanic sources or it may be very distant sources in North America or elsewhere (including Europe, from which the pollution may have travelled out over the Atlantic and then returned). 1.11.3 Effect of single sources on total-deposition field It is not easy to say how a single source affects the total deposition field. A simple answer can only be given if the thorny question of proportionality (see Section 1.11.5) is by-passed and if it is assumed that each source produces a deposition field independent of every other source. The following result will therefore be just as good, or just as bad, as that assumption. Considering sulphur dioxide and sulphate alone, the results of the EMEP model (which is a linear model and, as shown in Figure 1.5, gives a reasonable fit to observations) support a very simple relationship between the total sulphur deposition D (in grams per square metre per year) and the source strength E (in million tonnes per year):

THE FATE OF AIRBORNE POLLUTION

x’ (km) D/E

40 1.5

70 0.85

100 0.58

200 0.27

400 0.12

700 0.05

1000 0.02

21

2000 0.004

where x’ is the distance x from the source (in km) modified by a simple directional factor to allow for the general west-to-east mean airflow: Direction from source to receptor:

N 1.0

NE 0.6

E 0.7

SE 0.9

S 1.2

SW 1.3

W 1.3

NW 1.1

The exponential term in the formula represents the statistical effect of wet and dry deposition and implies an x’ length scale of 1000 km. A correction to the above expression for D may be made, wherever wet deposition predominates, by multiplying by R/ (R is the local average rainfall and the European average (~700 mm/year)). 1.11.4 Effect of tall stacks on distant depositions Tall stacks increase depositions at a distance, but not by very much. Experience tells us that emitting a given amount of pollution from a much taller stack has a dramatic effect in reducing associated ground-level concentrations in the locality. Since local depositions must therefore also be reduced to the benefit of the neighbourhood, more material must be left airborne to be deposited further afield. Models help to quantify this effect. Without going into details, they imply that building tall stacks for major emitters, leaving smaller sources unaltered, increases the long-range transport of SO2 only by some 10–15% and that the distances at which given associated depositions occur are typically increased by just some 50 km. 1.11.5 Relationship of depositions to emissions Next we consider whether depositions are proportional to emissions; or, expressing the question more fully, within acceptable limits of accuracy, is the deposition of a pollutant, averaged over a time appropriate to potential ecological damage, that results from any distant source region linearly proportional to the emission strength of that total source? This is a very important question when considering the likely benefits, and the cost-to-benefit ratio, that might accrue from emission control measures. Because of the extreme complexity of the processes involved and the conflicting evidence, the working group cannot at this time express a totally unequivocal viewpoint on this question. Some points can be made, however. (a) The chemical transformations that affect the main pollutants involved in acid depositions are intrinsically non-linear. (b) The means by which sulphur dioxide and some other species are taken up into cloud droplets is not fully understood, but presumably must depend on the droplets’ pH and other factors, and simple proportionality cannot be expected.

22

INTRODUCTION

(c) Non-proportionality between the air concentration of a pollutant affected by prolonged precipitation and the corresponding wet deposition rate affects the alongwind distribution of the deposition without affecting its integral, namely the total deposition. Because of the spatially quasirandom nature of rain, ensemble averaging over many rain events at a fixed receptor is virtually equivalent to averaging over this deposition distribution. Ensemble averaging therefore minimises the non-proportionality aspects evident in any single rain event. Questions of proportionality between depositions and emissions (as distinct from air concentrations) are complicated by the contribution of so-called ‘background’ sources (which are assumed to be outside Europe). This contribution, not obviously affected by changes in European emissions, is nevertheless of importance in many remote-area depositions. (d) Linear models (that assume proportionality) yield longterm deposition fields for sulphur over Europe and North America that are in fair accord with measured or inferred fields, especially when the additional ‘background’ contribution is included. (e) Comparison of historic records of depositions with trends in emissions produces confusing and conflicting evidence. Some of the reasons for this are, of course, the sparsity and possible unrepresentativeness of such records, the errors that probably plague many such records owing to dubious analytical and sampling techniques employed in earlier years, and changes in trajectory-climatology that, over several years, can change the relative importance to a receptor of one source region relative to another. 1.11.6 Is ozone a regional or long-range problem? Ozone appears to be one of the causes of extensive treedamage in Germany and elsewhere in central Europe. Some of this ozone is of natural origin, being created in the stratosphere and carried down into the boundary layer by large-scale atmospheric motions. However, ozone is also created by photochemical reactions which take place in polluted airmasses; that is, within the very processes that are generating acidifying species. Resulting levels of ozone tend therefore to peak and be maintained in large continental summer anticyclones, characterised by light winds, little long-range transport, strong sunshine and industrial haze. Therefore, in these normal cases, ozone is largely a regional problem, although occasionally subsequent advection occurs when high concentrations may be experienced far afield. 1.12 Recommendations In spite of considerable advances in our understanding, many scientific questions in this area remain unanswered or only partially answered. These should be identified and solutions sought. In particular, the key question—whether or not there is proportionality between emissions and corresponding depositions— should be pursued positively. Its solution will require the combined efforts of theoreticians and experimentalists, developing, on the one hand, models of various complexity that simulate explicitly or parametrically the involved chemical transformations that take place in the atmosphere both in and out of cloud and the consequences of appropriate time averaging on point depositions and, on the other hand, sensibly designed and well-executed field experiments—which are very likely to yield answers to very specific questions.

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23

Field experiments are often expensive and may require sophisticated aircraft and tracers. Nevertheless, without them, modelling is in danger of stagnating and may lose that element of realism which is essential if well-founded and far-reaching policy decisions are to be made. Cooperation between organisations within the U.K. and Europe, which have resources available in terms of scientists and technology, has been a pleasing hallmark of the study of acid deposition so far. This should continue, and wherever possible be extended, to ensure that future developments are not hindered.

THE WATT COMMITTEE ON ENERGY REPORT NUMBER 14

Section 2

Vegetation and soils W.O.Binns

This paper presents the work of Sub-group 2 (Soil and Vegetation) of the Watt Committee working group on Acid Rain. Membership of Sub-group 2

Dr W.O.Binns (Chairman) Dr J.N.B.Bell Dr M.J.Chadwick Dr J.A.Lee Professor T.A. Mansfield Professor K.Mellanby Dr H.G.Miller Dr R.A.Skeffington Dr M.H.Unsworth (observer)

Vegetation and soils

2.1 Historical Aspects Observations of rainfall chemistry date back to the early 18th century, but the earliest acount of atmospheric sulphur pollution was presented by R.A.Smith in 1852; he recorded that ‘all the rain was found to contain sulphuric acid in proportion as it approached the town’ (Manchester), thus demonstrating one of the features of acid precipitation. It seems likely that a similar condition had affected several other towns in northern England for at least the previous 50 years, though there are few data. However, in 1913, estimates of sulphur deposition in and near Leeds gave a rate of 170 kg ha−1 per year at 5 km east of the city centre and 57–108 kg at 11 km. At about the same date, a deposition of 107 kg S ha−1 per year with a rainfall of 685 mm was found at Garforth, 11 km east of Leeds. Acidity was measured as H2SO4 and at 22.5 kg ha−1 implies a pH level of about 4.2. This sulphur deposition was about five times that measured at Rothamsted, a rural site in southern England from which a continuous set of rain observations has been made since 1853. No truly rural observations are available from northern England, but the evidence from the Leeds area suggests that the countryside around these northern towns must have been subjected to some extent to acid

Figure 2.1 Mean annual sulphur dioxide concentrations (µg m−3) in two South Pennine towns, 1952–78 (Ferguson and Lee, 1983).

26

HISTORICAL ASPECTS

deposition for a considerable number of years. Further support for this can be provided by an examination of the peat profiles of the Pennine hills. Extensive blanket peat deposits cover much of the Pennines around the margins of the northern industrial towns. It is uncertain to what extent the sulphur components of coal smoke are accumulated in peat, but the profiles show the widespread accumulation of soot and heavy metals, such as lead, within peat formed since the Industrial Revolution. These pollutants would certainly have been accompanied by sulphur dioxide and it can be assumed, therefore, that much if not all of the region has been affected by man-made sulphur emissions, as well as by other pollutants, for at least two centuries. The current acidity of precipitation in rural Britain is shown in Figures 2.2 and 2.3. During the twentieth century extensive monitoring systems have been established. By the outbreak of the First World War in 1914, the systematic operation of deposit gauges had begun at a number of sites, notably in Manchester and Sheffield. Smoke abatement led to a 50% reduction of total deposited matter in some of these gauges by 1927, but pollution continued to be a problem. Statutory restrictions in the 1930s on such practices as firing colliery spoil may have helped further to reduce total deposited matter, but urban sulphate deposition rates differed little between 1929 and 1954, perhaps because the removal of sulphur dioxide from emissions is more complicated than the removal of smoke (Table 2.1). In the early 1950s, increasing effort was applied to the measurement of sulphur dioxide, and the Warren Spring Laboratory of the Department of Industry became responsible for co-ordinating the national monitoring of air pollution. After 1954, rain chemistry was no longer included in the national survey; but, by that date, the sulphur dioxide concentrations at the 91 mainly urban sites that formed the national network had fallen 10% below the 1939 concentrations, and during the next 25 years concentrations fell rapidly in at least several northern towns (Figure 2.1). The few data that exist suggest that this rapid fall in sulphur dioxide concentration in air is not simply relatable to the decrease in the acidity of rain, there being suggestions that the rain pH in northern England may have decreased in the last two to three decades. Comparisons are difficult, however, because differences in methodology may have resulted in differing contamination of deposit gauges by dry deposition. The decline in soot contamination may also have influenced the acidity as measured in deposit gauges. Other studies have shown rain becoming steadily more acidic in England over recent decades, despite a general fall in rural sulphur dioxide Table 2.1 Monthly weighted average sulphate—S in rain in southern Pennines, England MP

L

1929–34 − 2.9 1934–39 8.1 3.3 1939–44 7.1 4.9 1944–49 7.0 4.2 1949–54 8.2 4.7 Measurements in mg dm−3. MP—Manchester, Phillips Park; L—Leeds, Garforth; SS—Sheffield, Surrey Street; SA—Sheffield, Attercliffe; M—Marsden From Ferguson and Lee, 1983.

SS

SA

M

6.7 6.7 5.5 6.8 7.4

6.0 6.7 8.1 8.8 11.9

− 4.5 3.2 4.0 3.1

concentrations. There is evidence from Rothamsted, at least, of a marked increase in nitrate deposition. There is no doubt that much if not all of Great Britain is subjected to acid precipitation at the present time. There is little direct experimental evidence for the adverse effects of this long history of atmospheric pollution on plant growth and distribution, but there has been no shortage of observations and correlations

VEGETATION AND SOILS

27

Figure 2.2 Rainfall weighted annual average hydrogen-ion concentrations (Barrett et al., 1983).

suggesting such a relationship. Thus the dearth of lichen species in the Manchester area in the 1850s was linked with the ‘influx of factory smoke’, and a hundred years later, in Newcastle, this correlation was shown to be with SO2 concentration rather than with smoke; the fact that asbestos tiles were shown to support certain lichen species further into Newcastle than sandstone walls suggests some role for acidity per se in this phenomenon. The disappearance of a considerable number of moss species from the Todmorden area was noted in the 1860s and connected with ‘the super-abundance of factory smoke’. At the turn of the century the disappearance of a number of species from the Halifax flora was correlated with atmospheric pollution. What cannot be doubted is that in the southern Pennines, particularly in the blanket bogs, there has been a large and widespread vegetation change in the last 200 years. This has resulted in the extensive loss of the dominant Sphagnum species and many of the associated angiosperms which are characteristic of such communities. This change is chronicled in the peat and is correlated with the appearance of soot in the peat profile. The sensitivity of mosses and liverworts to SO2 or its solution products has been well demonstrated, and Sphagnum species have been shown to differ in their sensitivity to these pollutants—the most widespread species in the southern Pennines today, S. recurvum, being the most resistant. A field experiment in which a relatively unpolluted Welsh bog surface was modified by approximately weekly additions of sulphate or dissolved and unreacted sulphur dioxide showed a similar result, the Sphagnum species being killed at the highest SO2 concentration while the cotton grasses (Eriophorum species) were apparently unaffected. This treatment thus effectively reproduced the history of change in the southern Pennines which

28

HISTORICAL ASPECTS

Figure 2.3 Deposited acidity (Barrett et al., 1983).

had probably occurred as a result of atmospheric pollution and where cotton grasses are now the exclusive dominants over many thousands of hectares. In attempting to pinpoint an important and extensive vegetation type that is likely to be particularly sensitive to atmospheric pollutants, blanket bog is probably the natural choice for three reasons. First, blanket peatlands are dependent on wet and dry deposition from the atmosphere for the major input of ions. Second, they are normally dominated by a carpet of mosses and liverworts which is continuously exposed to wet and dry deposition from the atmosphere because the leaves lack a cuticle. Third, the species are adapted to a low solute supply. Although past atmospheric pollution may have caused major changes in the southern Pennine vegetation, it is interesting to speculate as to whether current deposition is affecting plant growth and metabolism. The increasing contribution of nitrate to acid precipitation provides one possible means of assessing this. Plants utilise nitrate by reducing it first to ammonia, and the enzymes responsible for this are substrateinducible. The first of these, nitrate reductase, is easily measured, and it has been demonstrated that current nitrate deposition in the southern Pennines rapidly induced nitratereductase activity in two species of Sphagnum transplanted from an unpolluted bog surface. This does not necessarily demonstrate an adverse effect of nitrate, but it does show a direct metabolic response of plants in the field to an enriched atmospheric input and points to a profitable area for further study. There is no unequivocal evidence that present-day precipitation is adversely affecting the growth of plants in British semi-natural communities, but

VEGETATION AND SOILS

29

current bulk deposition of sulphate, nitrate, ammonium and hydrogen ions in the southern Pennines is greater than in similar bog systems, e.g. in North Wales, and may contribute to the continuing absence of Sphagnum species and the failure of transplants into the region. Similar effects may be expected elsewhere in the country where wet habitats exist fairly closely down-wind of major pollution sources, as for example in South Wales. 2.1.1 Conclusions There is a long history of atmospheric pollution in Great Britain, and the vegetation changes associated with it are well documented. Rainfall is currently acid in rural Britain, and it is possible that some districts may be becoming more acid. 2.2 Effects of Ambient Air Pollution on Metabolism and Growth of Plants Until recently, interest in the widespread effects of ambient air pollution in Europe had been almost entirely restricted to sulphur dioxide. Over the last decade, there has been an increasing awareness that nitrogen oxides usually accompany SO2 at similar or greater concentrations (volume for volume), whereas ozone appears at elevated Table 2.2 Effects of ambient air on shoot dry weight of grass species in chamber experiments in Sheffield Year

Species

1973/4 Lolium perenne*

Lolium

Mean SO2 concentration, µg m−3

Duration, d Reduction in ambient compared with clean air, %

P<

Reference

70

56

36

0.001

Crittenden and Read, 1978

59 69 63 67

131 86 116 56

20 25 26 36

0.01 0.001 0.001 0.001 Crittenden and Read, 1979

42

0.001

14 25 14 23 39

0.05 0.05 0.05 0.05 0.01

multiflorum† Dactylis 45 72 glomerata‡ 1976/7 Lolium 45 28 perenne 44 28 Lolium 45 28 multiflorum 44 28 Dactylis 45 28 glomerata * Perennial ryegrass. †Annual ryegrass. ‡Cocksfoot.

Awang, 1979

30

HISTORICAL ASPECTS

levels episodically under suitable meteorological conditions in summer time. Both O3 and acid precipitation, which is derived from chemical transformations of SO2 and NOx, penetrate into areas remote from the origin of their precursors. 2.2.1 Effects of urban pollution on plant growth Studies of the impact of current ambient air pollution on the performance of plants have produced results that are difficult to reconcile with those of experiments in which plants are subjected to artificial fumigations. Experiments in a Sheffield suburb demonstrated remarkable improvement in the growth of several species of grasses when ambient air was purified by passage through activated charcoal (Table 2.2). The mean concentrations of SO2 in these experiments were, in general, considerably less than have been shown to influence plant growth in fumigation experiments. It has been suggested that the apparent discrepancy between ambient air and fumigation experiments has two possible explanations. First, pollutant levels fluctuate continuously in ambient air, and peak concentrations concealed within the overall mean may increase injury compared with the same mean concentrations administered continuously in fumigation experiments. Second, other pollutants present in the field may also contribute to reductions in plant growth. Recent work has suggested, however, that there is only a relatively small impact of short-term peak concentrations of SO2 interspersed with low or zero concentrations, compared with uniform levels. The peaks were up to 2140 µg m−3 compared with a uniform 160 µg m−3 for Timothy grass, and up to 750 µg m−3 compared with 100 µg m−3 for Scots pine. Support for the second possibility—that other pollutants are contributing to injury in urban areas—is provided by demonstrations of marked synergistic effects on the growth of several grass and tree species when mixtures of NO2 and SO2 are present in the range 166–194 µg m−3 and 118–139 µg m−3 respectively. Attempts have been made to simulate urban atmospheres in a series of fumigation experiments in which grasses were subjected to NO2/SO2 mixtures at fluctuating levels based on monitoring data from central London. Surprisingly, in general, only minimal effects of mixtures of mean concentrations of 85 µg m−3 SO2 and 46 µg m−3 NO2 were observed in experiments of up to 7 months duration, although small but significant stimulations and reductions in growth were recorded on occasions (Table 2.3). Thus there remains a discrepancy between the results of experiments where plants are subjected to ambient urban air and those where fumigations have been conducted in a manner that is supposed to simulate urban conditions. Physiological effects of SO2 and NO2 To appreciate the possible reasons for these discrepancies, it is necessary to consider the ways in which the pollutants interfere with the physiology and metabolism of the plant. SO2 and NO2 are both sufficiently soluble to dissolve readily in the extracellular water once they have entered a leaf. Although sulphur and nitrogen are essential mineral nutrients, the uncontrolled deposition of SO2 and NO2 can cause physiological and biochemical disturbances because the cells are exposed to higher levels of some chemical species (e.g. sulphite and nitrite) than occurs during normal metabolism. The function of the cuticle on the surface of the aerial parts of plants is mainly that of reducing excessive loss of water to the atmosphere, but it can also serve as an effective barrier to the entry of pollutants when stomatal pores are closed. From purely physical considerations we would expect pollutant uptake to be highly correlated with stomatal opening, and this has been shown for some tree species but not for others. The situation is further complicated by the fact that stomata themselves are affected by the presence of some pollutants. The main responses to SO2 and NO2 so far identified are:

VEGETATION AND SOILS

31

Figure 2.4 Percentage decreases in dry weight of Poa pratensis (smooth-stalked meadow grass) seedlings exposed to SO2 (40 pptm) NO2 (40 pptm) for between 4 and 50 days (Whitmore, 1982).

(1) SO2 can cause the stomata of some species to open. This effect, which can occur with concentrations as low as 50 µg m−3, is known to be the result of damage to epidermal and subsidiary cells. SO2 is thought to enter the leaf via the stomata, then dissolve in the extracellular water surrounding the epidermal and subsidiary cells which are adjacent to the substomatal cavity. Loss of turgor in these cells removes the physical pressure on the guard cells, allowing wider stomatal opening. (2) SO2 has also been reported to cause stomatal closure, especially when the concentrations used were high (1430 µg m−3) or when the relative humidity surrounding the Table 2.3 Effects on grass species of SO2/NO2 mixtures designed to simulate central London pollution characteristics Species

Mean SO2 concentration, µg m−3

Mean NO2 concentration, µgm−3

Duration, days Changes in shoot dry weight compared with controls

Lolium perenne (perennial ryegrass) Dactylis glomerate (cocksfoot) Phleum pratense (Timothy) From Lane, 1983.

87

46

206

−8%

84

48

181

+12%

87

46

206

−11 %

leaves was low; 100 µg m−3 SO2 has been shown to cause stomatal opening at high humidity and closure at low humidity in broad bean, sunflower and tobacco. (3) Effects of NO2 on stomata have been little investigated, though there is evidence that in runner beans it may cause a temporary stimulation of opening not unlike that induced by SO2 in the same species; a mixture of SO2 and NO2 (286 and 205 µg m−3 respectively), however, depressed opening from the second day of a 5-day fumigation. There is little published information on the effects of lower concentrations of NO2. Stomata regulate the gaseous exchange between leaves and the atmosphere and ideally determine the best compromise between the opposing priorities of permitting an inward flux of CO2 and restricting the outward flux of H2O. Even a small disturbance of the normal functioning of stomata is likely to impair the physiological efficiency of a leaf, and a major disturbance will lead either to waterdeficit stress or to a severely restricted supply of CO2 for photosynthesis. Some of the visible symptoms of both acute and chronic injury by SO2

32

HISTORICAL ASPECTS

are of a kind that may be explained in terms of stomatal malfunctioning. Excessive transpiration can rapidly lead to the death of leaf tissues, even in a plant that has an adequate supply of water in the soil. Such effects may be particularly important for the upper branches of trees whose supply of water is strictly limited by the resistance of the xylem conduits. In cases where pollutants cause stomatal closure, the interference with the supply of CO2 for photosynthesis must limit growth and may also be associated with premature senescence of leaves, and it is believed that such closure is actually a cause of senescence. We would expect the impact of changes in stomatal behaviour to vary according to the environment in which a plant is growing. Any deleterious effects of SO2 and NO2 that derive from stomatal responses must interact strongly with the environment, particularly soil water potential and atmospheric humidity. Discrepancies between the results of different experiments are therefore not surprising. 2.2.2 Morphological changes in polluted plants Some of the growth responses to SO2 and NO2 are also of a nature that would be expected to cause problems to plants growing in unfavourable conditions. Several studies have shown that there is a different allocation of dry matter in polluted plants, more material going to support the growth of the shoots and less to the roots. This material is used to support the development of an increased leaf area, and so the potential transpiration is higher when there is a smaller root system to explore the soil for water. There is recent evidence of a major interaction between SO2 and freezing tolerance, with more frost injury after exposure to SO2. It is not surprising that such a change occurs in plants with an increased shoot:root ratio. It is likely that the changed morphology of polluted plants will also affect their tolerance of other environmental stresses, but this possibility has not been investigated in detail. Clearly, the impact of a given dose of SO2 and NO2 must be considered in relation to ambient climatic conditions, and adverse circumstances might considerably lower the threshold concentrations at which injury occurs. The ambient versus filtered air experiments and the fumigations already mentioned (Tables 2.2 and 2.3) were performed in outdoor chambers that were subject to the influence of variable climatic conditions. Thus each experiment was essentially unique and reflected the impact of air pollution under a specific set of climatic conditions that will have fluctuated markedly throughout the duration of the investigation. Can a dose-response relationship be defined? In the face of such variability in the results of experiments it is important to demonstrate that under closely controlled environmental conditions a clear dose-response relationship can be obtained. Considerable success was achieved using young plants of smooth-stalked meadow grass exposed to SO2+NO2 for periods between 4 and 50 days (Figure 2.4*). After small doses (in terms of concentration x time) there was a stimulation of dry weight, and we can presume that in experiments in which stimulatory effects of the pollutants were found, the doses were in this region. The extent of the stimulatory part of the dose-response curve is Table 2.4 Effects on plants of addition of O3 to SO2/NO2 mixtures Gas concentraions µg m−3 Species

SO2

NO2

O3

Duration

Parameter measured

SO2+NO2

Effect of pollutant O3

SO2 + N2 +O3

Reference

Populus× interamer

56

42

62

42 days

Numbers of fallen leaves

+430% compared

−

+770% compared

J.Mooi (1984)

VEGETATION AND SOILS

33

Gas concentraions µg m−3 Species

icana cv. Donk* Pisum sativunrf Raphanus sativus‡

SO2

NO2

O3

Duration

293

400

220

5 hours

1068

764

800

6 hours

cv. Cherry Belle 3 hours per week for 4 weeks 400 287 300 12 hours per week for 4 weeks * Poplar. †Pea. ‡Radish. §Tomato. ||Dwarf bean. Lycopersi con esculentu m§ Phaseolus vulgaris||

800

573

600

Parameter measured

SO2+NO2 with clean air 0

Foliar necrosis, % Foliar 11% necrosis, % Hypocotyl −11% dry weight compared with clean air Leaf dry −11% weight compared with clean air Fruit fresh −20% weight compared with clean air

Effect of pollutant O3

SO2 + N2 +O3

2%

with clean air 18%

45%

57%

–20% compared with clean air –19% compared with clean air −27% compared with clean air

–35% compared with clean air –27% compared with clean air −27% compared with clean air

Reference

Fujiwara (1973) Reinert and Gray (1981)

Reinert and Heck (1982)

thought to depend on the environmental conditions. Figure 2.4 shows that the change from a stimulatory to inhibitory effect occurs over a fairly small range of doses, and it is important to note that in many experiments there is considerable error associated with the definition of dose. Effects of SO2 and NO2 on photosynthesis Inhibition of photosynthesis is one of the most frequently reported effects of SO2 on plants. Although SO2-induced stomatal closure must directly inhibit photosynthesis, there is evidence also of direct metabolic effects of SO2 and its products in plant cells. Structural changes within chloroplasts seen under the electron microsope are the first visible signs of perturbation and are thought to indicate ionic disturbances or pH changes. These are reversible after short fumigations, but recovery time increases after longer treatments or exposure to higher concentrations. Explanations of the toxicity of SO2+NO2 Studies of the pH gradients generated across photosynthetic membranes have recently shown surprisingly toxic effects of treatments with combinations of sulphite and nitrite ions. There were no significant effects of sulphite or nitrite given separately in concentrations of up to 1 mM, but the two ions together at concentrations of 0.1–0.5 mM greatly affected the ability of the membranes to maintain a pH gradient. When NOx enters a plant and dissolves in extra-cellular water it forms both nitrate and nitrite ions. In normal nitrogen metabolism, nitrate is converted by nitrate reductase to nitrite and then to ammonia by nitrite reductase. Fumigations with NOx have been shown to increase both nitrate and nitrite reductases in the leaves of several plant species. The reduction of nitrite to ammonia and subsequently to amino acids is probably an important detoxification mechanism and the presence of SO2 with NO2 has been shown to

34

HISTORICAL ASPECTS

Figure 2.5 Effects of variation in SO2 exposure concentration on depression of net photosynthesis in three species (Black, 1982).

prevent the rise in nitrite reductase. Accumulation of nitrite (a highly toxic ion) in leaves exposed to SO2 +NO2 could thus be a cause of increased damage that sometimes occurs when the two gases are present together in polluted air. SO2 and net photosynthesis Because of the complex nature of the effects of SO2 on various processes involved in photosynthesis, it is not surprising that it is impossible to describe a precise dose-response relationship. Figure 2.5 shows the relationship between the depression of photosynthesis and SO2 concentration for three different species. Various factors (e.g. boundary layer resistance, CO2 concentration, light intensity, relative humidity and temperature) can result in responses as variable as those in Figure 2.5. With the information available at the present time it is difficult to estimate the consequences of different levels of SO2 pollution on crops and natural vegetation. We are, however, now able to point to certain stages of growth (e.g. seedlings of grasses) when there is greater sensitivity to SO2 and to combinations of environmental factors which predispose plants to injury. For example, grasses are more sensitive to SO2 when light intensity is low and are more susceptible to frost injury after exposure to SO2. NO2 and net photosynthesis There has been much less research into the effects of NO2 on photosynthesis and the results from various laboratories are even more variable than those for SO2 in Figure 2.5. Inhibition of photosynthesis appears to begin at NO2 concentrations between 200 and 500 µg m−3. At lower concentrations NO2 applied on its own can stimulate growth, but it is uncertain how often this occurs in the field where NO2 is usually accompanied by SO2. 2.2.3 Importance of ozone In addition to SO2 and NO2, it is necessary to consider a third pollutant—ozone (O3)—when interpreting the results of experiments that demonstrate adverse effects of ambient air on plant growth. Although O3 *

To avoid confusion between different usages of the term ‘billion’, concentrations are expressed here as follows: One part per million=1×10−6 One part per thousand million (American billion)=1×10−9 Forty parts per thousand million=40×10−9

VEGETATION AND SOILS

35

Figure 2.6 Range of pH values likely to be encountered in rainfall and those causing effects on plants.

pollution occurs episodically and is essentially confined to the summer months, it can occur at phytotoxic concentrations over large areas of countryside, as well as in cities. When an O3-sensitive cultivar of tobacco was exposed outdoors throughout the summer of 1977, at sites scattered throughout the British Isles, O3 injury was detected at all of them, on at least some occasions, with the exception of the far north of Scotland. Every summer since 1976, the research group at Imperial College, London, has grown plants in chambers with ambient or purified air at rural locations in southeast England as part of a programme designed to determine the impact of ambient O3 levels on British crop and wild species. Consistent reductions in growth of O3-sensitive cultivars of a range of species, including radish, dwarf bean, spinach and barley, have been detected in the ambient air chambers following the occurrence of elevated levels of O3; in contrast, little or no effect has been seen on O3-tolerant cultivars on the same occasions. There have also been several incidents when ambient O3 has induced visible injury on the foliage of the experimental plants, including pea and radish. Considerable attention has been given to the combined impacts of O3 and SO2 on plants, with synergistic, antagonistic, and additive effects all being demonstrated, depending on species and fumigation conditions. However, in view of the fact that SO2 pollution is normally accompanied by similar concentrations of NOx and that Oa occurs episodically, it seems more appropriate to examine the impact on plant growth of the addition of O3 to an SO2/NO2 mixture. Surprisingly, very few investigations of this type have been performed. Table 2.4 shows the results of a selection of such experiments. Until recently all reported cases showed either an additive or synergistic effect in response to the addition of O3 to SO2/NO2. These included a remarkably large increase in senescence of poplar caused by the addition of a level of Oa within the maximum natural background concentration to an SO2/NO2 mixture at concentrations that are representative of large areas of western Europe. However, recent work has highlighted the complexities of interactions between O3 and SO2/NO2 mixtures; experiments with a range of species have demonstrated additive, synergistic, and antagonistic effects on growth. Tentative conclusions are that at lower concentrations of Oa, the effects of NO2/SO2 mixtures are generally enhanced, whereas suppression of such effects tends to take place at higher O3 concentrations. It is thus apparent that there is a major gap in our understanding of the effects on plant growth of the mixtures of the three most widespread gaseous phytotoxic pollutants occurring in ambient air. However, there seems to be little doubt that, when the impact of ambient air on plant growth has been studied in summer, the presence of O3 will have influenced the results on at least some occasions.

36

HISTORICAL ASPECTS

2.2.4 Acid precipitation and plant growth The final category of pollutant that may contribute to widespread vegetation injury in the field is acid precipitation. Claims in the early 1970s that this was causing a fall in forest productivity in Scandinavia have not been substantiated. In the U.S.A., however, there are clear indications of a decrease in three pine species in New Jersey and of Red spruce in the mountains of Vermont; but it should be pointed out that, over the period of the growth reductions, O3 concentrations have probably increased, thus confounding any possible impacts of acid precipitation. Many workers have investigated the effects Table 2.5 Some field measurements and calculations of acid-producing processes Process

Rate of acid generation (eq ha Situation per year)

‘Background’ rain, pH 5 1071 Acid precipitation 8562 Dry deposition of SO2 S2− Oxidation

9803 2500

Plant uptake Respiration Nitrification

40–5000 6100 8140

Calculated 1066 mm rain Tillingbourne, Surrey, England Calculated for 20 µg m−3 SO2 Oxidation of 2.7 cm peat per year Forest tree uptake Forest soil Devegetation, Hubbard Brook

Reference

Skeffington, 1983

Brown, 1980 Nilsson et al., 1982 Sollins et al., 1980 Bormann and Likens, 1979

1No

weak acid contribution mm rain, 750 eq ha−1 strong acid and 106 eq ha−1 weak acid. 3Deposition velocity 5 mm s−1

21066

of watering plants with artificial rain adjusted to various pH levels. Significantly, no effect has been shown on tree growth unless the pH level is lower than is normally found under ambient conditions. However, such effects have now been demonstrated on agricultural crops watered with artificial rain at pH 4.0; for example, a small reduction in seed weight of soya bean and, in beet, an inhibition of growth accompanied by the appearance of a few lesions on the foliage. These observations of effects on agricultural species raise the possibility that acid precipitation may contribute to growth reductions in the field caused by gaseous air pollutants. It is surprising, in view of the considerable interest in effects of acid precipitation, that its interactions with gaseous air pollutants have scarcely been studied. There is only one report of the impact of simultaneous exposure of plants to SO2 alone and acid precipitation, during which soya bean was subjected to a mean of 570 µg m−3 SO2 or artificial rain at pH 3.1 or both. Acid rain on its own increased seed dry weight, whereas SO2 alone reduced seed yield; in the mixture, there was no significant acid rain/SO2 interaction, with the positive and negative effects of the two pollutants apparently balancing out. During the last year, attention has become focused on possible interactions between O3 and acid precipitation in reducing the growth of trees. It has been suggested that O3 is contributing to die-back of Silver fir and Norway spruce in southern German forests, the die-back having generally been attributed to acid precipitation. It has been pointed out that the areas of forest die-back are not associated with particularly low-pH rain, but coincide rather with the regions of occurrence of both elevated O3 levels and frequent mists, the latter representing a possibly important, but

VEGETATION AND SOILS

37

little studied, route by which sulphur and nitrogen compounds derived from the atmosphere may impinge on foliage. A very preliminary experiment demonstrated that pretreatment of spruce with an O3 fumigation significantly increased the rate of leaching of magnesium from foliage treated with artificial rain of pH 3.1; this is consistent with measurements that have shown severe magnesium deficiency in affected trees in the field. Very recently, some further evidence of synergistic interactions of O3 and acid precipitation, this time on growth itself, has been produced by exposing soya bean plants, grown in chambers ventilated with ambient air, or air from which O3 had been removed by filtration, to simulated rain of pH 4.0, 3.4, and 2.8. As the pH of the ‘rain’ was reduced, the O3-induced reductions in growth increased. However, other recent investigations of the effects of a mixture of 0.2 ppm of SO2 and 0.1 ppm O3 on the growth response of soya bean to acid precipitation showed no significant acid precipitation/gas interactions. 2.2.5 Conclusions Investigations of the effects of ambient air pollution on plants to date have revealed that the situation is very complicated. It is apparent that SO2, NO2, O3 and acid precipitation all have the potential to contribute to the injury of plants. However, the interactions between these pollutants and also with other environmental stresses are little understood: synergistic, antagonistic and purely additive effects have been recorded under different circumstances. The most serious gaps in our knowledge are the significance of ambient levels of acid precipitation when accompanied by gaseous pollutants and of the interactions between pollutants and the many natural stresses imposed by the environment. 2.3 Effects of Acid Precipitation on Plants and Soils 2.3.1 Relevant concepts An acid is a substance capable of donating hydrogen ions to a base. The total acidity of a solution such as rainwater includes dissociated acids, the ‘free acidity’, together with undissociated weak acids, and as such can only be determined by titration with a strong base. Total acidity is rarely measured, but, rather, a glass electrode is used to measure hydrogen-ion activity, expressed in equivalents of hydrogen-ion per litre (H+) or as pH (-log10H). In dilute solutions this hydrogen-ion activity approximates to the concentration of dissociated H+. In the absence of pollution, rain would include elements derived from sea spray and terrestrial dust together with naturally occurring gases—notably CO2 but also SO2 and NO2. As such, rain has always been somewhat acid, perhaps having a median pH value around 5. With the rise of industrial activity, substantial amounts of the precursors of the strong acids HCI, HNO3 and H2SO4 are being added to the atmosphere, thus increasing the acidity in rainwater. These gaseous precursors (HCI, NOx and SO2) may also

38

HISTORICAL ASPECTS

Table 2.6 Acid content of incident rainwater, throughfall and stemflow measured over five years at various stands of Sitka spruce in Scotland and Cumbria eq H+ ha−1 per year Leanachan Incident rain 440 Throughfall+stemflow 140 Change, % −68 From H.G.Miller and J.D.Miller (1984). Measurements in eq H+ ha−1 per year.

Kilmichael

Strathyre

Elibank

Kershope

Fetteresso

970 390 −60

480 380 −21

750 280 −63

750 660 −12

640 590 −8

Table 2.7 Mean pH values of water at various points in different ecosystems for the year November 1980 to November 1981 at Glen Tanar, northeast Scotland Vegetation cover Heather Throughfall – Stemflow – Beneath ground vegetation 4.5 Beneath humus 4.0 Below rooting zone 5.6 *No ground vegetation grew beneath the spruce. pH value of bulk precipitation 4.6. pH value of filter gauge water 4.4. From Miller, 1983.

Scots pine Old

Young

3.9 3.1 5.5 3.6 4.4

4.7 3.5 5.5 3.8 4.7

Sitka spruce

Japanese larch

Birch

5.4 3.9

4.8 3.7 4.7 4.6 4.9

4.8 3.9 6.4 5.1 6.6

*

4.4 5.5

reach the soil and vegetation through dry deposition. The residence time of NOx, and in particular HCI, in the atmosphere is fairly short, so SO2 is the dominant gas away from the immediate vicinity of emissions. Close to emission sources, dry deposition exceeds wet deposition; but, on moving away from the source, total acid deposition decreases and there is a progressive shift towards a predominance of wet deposition. Figure 2.6 shows the range of pH values encountered today. The total man-made acidity deposited in many rural areas is less than the acidity generated naturally within terrestrial ecosystems through processes such as root uptake of an excess of cations over anions, organic matter decomposition, nitrification (i.e. the microbial conversion of ) and root and microbial respiration. However, pollution-derived acidity comprises strong acids, usually sulphuric or nitric acids, whereas much of the naturally-generated acidity either represents the direct transfer of hydrogen ions from roots to exchange sites, without the involvement of an anion, or is present as weak carbonic and organic acids. Particular exceptions include the sulphate resulting from natural mineralisation of organic forms of these elements—as can happen following drainage of organic soils—and excessive nitrification, such as may occur in warm, dry years. Examples of possible contributions from these various sources are given in Table 2.5. Both sulphates and nitrates are relatively mobile in the soil, though nitrate is usually effectively retained within the ecosystem by plant uptake. It follows that hydrogen ions, or other cations, associated with sulphate, or rarely with nitrate, can pass readily through the vegetation canopy and the soil to appear in drainage water.

VEGETATION AND SOILS

39

Figure 2.7 Representation of soil processes consequent on acid rain.

When discussing acidification and acid reactions in soil it should be borne in mind that, in this mixture of solid and liquid, the concept of acidity involves both hydrogen and aluminium ions and embraces not only the ions in solution but also those that can be displaced from the charged cation exchange surfaces on soil clays and decomposing organic matter. The hydrogen and aluminium ions on the exchange surfaces, therefore, represent a dynamic reserve of soil acidity. If soil acidification is to occur, levels of exchangeable hydrogen and aluminium must increase at the expense either of adsorbed basic cations (Na+, , K+ and, predominantly, Ca2+ and Mg2+) or by occupying new exchange sites. In the former case there is a loss of available nutrients from the soil. Whether this will occur depends (a) on the equilibrium between cations in solution and those on the exchange surface which controls the nature of any exchange and (b) on the presence of a mobile anion, such as sulphate, to transport the resulting cation out of the soil. A schematic illustration of these processes is shown in Figure 2.7. Irrespective of its origin, acidity can react with vegetation or soil such that it is neutralised, exchanged or buffered. Neutralisation is the destruction of acidity, as when hydrogen ions react with soil minerals to produce salts plus water. Many factors influence rates of neutralisation, notably the form of the minerals but also aspects of the soil physical environment, including hydrology and temperature. The reaction rates are dependent on hydrogen-ion concentration rather than simply the amount of hydrogen ions received. Hydrogen-ion exchange is the predominant reaction on vegetation and soil surfaces, the hydrogen ions replacing base cations, mostly calcium and magnesium, on the cation exchange sites. As in all such chemical systems, the direction of the exchange depends on the relative concentrations in solution and on the exchange surface, as for reaction to occur there must be disequilibrium between their hydrogen-ion concentrations. Thus hydrogen-ion concentration is the determinant of the direction of the reaction, although the extent of the exchange is a function of the amount of acid in solution. This exchange increases the acidity retained on the solid but decreases the acidity of the surrounding solution. Within a terrestrial ecosystem, particularly within the soil, there are various buffering mechanisms that serve to resist the change in acidity of the system as further acidity is added. Classic buffering is provided by weak acids, such as carbonic acid and most organic acids, whose dissociation (hydrogen-ion activity) reduces in response to external addition of hydrogen ions. In addition, soil is characterised by certain 'reversible neutralisation' reactions that may properly be regarded as buffer systems. A particularly important example is the dissolution (which consumes hydrogen ions) and precipitation (which releases hydrogen ions) of the abundant aluminium in the soil over the pH range 5.0 to 3.5. This system essentially prevents mineral soils becoming more acid than about pH 4 and means that acid soils are rich in soluble and exchangeable aluminium ions.

40

HISTORICAL ASPECTS

The sensitivity of a soil to acid precipitation has been defined in various ways. On the one hand, the most sensitive may be considered to be non-calcareous sandy soils of about pH 6, because they show the greatest pH change in response to unit acid input. On the other hand, acid soils with low base cation content may show the greatest proportional loss of nutrient ions, although the pH change will be minimal. 2.3.2 Impact of acid precipitation Direct damage to vegetation, in the form of lesions or necrosis, seems very unlikely at the levels of acidity presently encountered in Great Britain. Reviews of the literature suggest that necrosis only occurs at pH values of 3.3 and less, which is well below the values in even badly polluted rain. Certainly there is no definitive observation of such damage outside the laboratory. There has been some suggestion that at the existing levels of acidity there may be some interference with reproduction in those plants with exposed reproductive systems, such as bracken, but no such effect has been observed in the field. The only established effect of increasing rainwater acidity on a vegetation canopy is an accelerated leaching of many cations from foliage. This is generally not accompanied by a reduction in foliar nutrient content or in plant growth. It has already been mentioned that damaged spruce in the forests of southern Germany appear to be losing excessive quantities of magnesium and, to a lesser degree, calcium, to the extent that the trees become severely magnesium-deficient and eventually die. Although there may be an acid rain or acid fog component involved, it is considered that the primary factor is ozone damage to leafcell membranes, the ozone being produced from gaseous NOx by sunlight in the presence of gaseous hydrocarbons. As such, this damage can be considered to be a direct effect of gaseous pollutants rather than a response to acid rain. Figure 2.8 shows a scheme to explain the phenomena and mechanisms of forest decline. As the normal cation leaching from leaves is, in large measure, due to hydrogen-ion exchange, there is a commensurate reduction in the acidity of rainwater passing through the canopy and further reductions have been observed as the rainwater passes through the ground vegetation layer (Tables 2.6 and 2.7). Cation exchange in the leaf-litter layer beneath a hardwood forest can continue the reduction in acidity, although a more common observation is an increase in acidity at this stage. The pattern and extent of pH changes as water passes through an ecosystem, however, appear to vary with the dominant plant species present, as shown in Table 2.7. It should be borne in mind that the original root uptake of the base cations involved in reducing throughfall acidity would have entailed a reverse flux of hydrogen ions into the soil. Hydrogen-ion exchange at a leaf surface, therefore, entails a net transfer of acidity from rainwater to the soil exchange surfaces. There is then a further direct exchange of soil base cations for hydrogen ions as acidified water drains through the soil profile (Table 2.7). Thus, whatever the site of hydrogen ion exchange, it is accompanied by soil acidification in the sense that the soil exchange complex shows decreased base status and increased amounts of exchangeable hydrogen. In an unpolluted soil, leaching of exchanged base cations down the profile is associated with organic ions and bicarbonate. The’ former are largely decomposed in the lower soil horizons, precipitating the cations, but there is a loss of cations into drainage water as bicarbonate salts. Where there is a supply of sulphate anions, however, as in acid rain districts, the exchanged base cations are not reprecipitated but are transported, together with the mobile sulphate anion, into the groundwater and on into drains and streams. The introduction of strong acids in rain may accelerate acidification of soil and loss of nutrient cations. In certain circumstances, acidification can also increase mineral decomposition rates, giving an increased release

VEGETATION AND SOILS

41

Figure 2.8 Hypothesis to explain central European forest decline as postulated by Bosch et al. (1983) and Rehfuess (1983).

of fresh calcium and magnesium ions from soil minerals. Whether this is sufficient to replace the loss of base cations from exchange sites clearly depends on the availability and rate of weathering of the soil minerals. The theoretical neutralisation capacity of weathering granite is almost infinite compared to the size of current acid inputs. However, there is evidence that reaction products of weathering can accumulate, thus reducing the rate of weathering. Thus theoretical rates of mineral weathering may not be attained in the field and in some soils may not provide sufficient neutralizing capacity. In situations where the acid inputs are large, this may bring about some deterioration in the surface horizons of the soil profile. It was the fear that soil acidification and reduction in nutrient status might be occurring that prompted early suggestions from Sweden of a possible decline in forest growth. Demonstration of any such effect from measurement of tree rings is difficult because of the confounding effect of cyclical variations in climate. Recent attempts, including a reworking of the Swedish data originally used as evidence for a decline, have not revealed any growth reduction that could be unambiguously ascribed to acid rain. An alternative mechanism that has been proposed is that acid deposition may be harming plant growth through increasing the solubility of various metals in the soil, notably aluminium, until they become toxic to roots. However, there is no convincing evidence at the moment that the aluminium levels found in the field, even at low soilcalcium levels, are toxic to roots of mature trees. Metals other than aluminium, particularly heavy metals, may also be mobilised in soils, especially where these are deposited in significant quantities close to pollution sources. Most tend to have a high affinity for organic matter and are not very mobile in soil, but their solubility can increase with decreasing soil pH. Undoubtedly, toxic effects occur where pollution levels are very high, as in the vicinity of smelters and in some industrial environments, but such results should not be extrapolated to other situations; there is at present no evidence of heavy metal damage to roots at more distant sites following acidification. There is also the possibility that damage might occur through an effect of acid rain on soil microbial activity. As acid rain is unlikely to have any effect on the acidity of an already acid soil, changes in soil microbial populations are not likely to be extensive. Experiments to determine the direct effects of artificial acid precipitation on microbial activity in soil have produced conflicting results ranging from stimulations to

42

HISTORICAL ASPECTS

depressions. An increase in nitrogen availability has been found in many artificial acidification experiments. It has also been claimed that the rate of decomposition of organic matter may be retarded, which could have an adverse effect on tree growth. The lack of confirming evidence suggests that this is an area that requires more research. Some discussion has centred on possible nutritional benefits from the elevated levels of nitrate and sulphate now being introduced in rain. Certainly, many upland ecosystems might be expected to show a slight response to the nitrogen input in the long term, but the sulphur is unlikely to be of any additional value. In some agricultural situations, however, it is believed that pollution-derived sulphur is preventing the development of the deficiencies of this element that could be expected now that single superphosphate and ammonium sulphate are no longer used in preparing compound fertilisers. Growth responses to added sulphur have been detected in grass and cereal crops growing on sandy soils with low organic matter in northern Scotland where pollution inputs are low. However, fertiliser manufacturers have been quick to respond by marketing sulphur-containing fertilisers: this would seem to be a preferable approach to reliance on uncontrolled applications in acid rain. 2.3.3 Conclusions Direct effects of acid rain per se on vegetation are not to be expected (as distinct from the direct effects of gaseous pollutants discussed earlier), though there are increased losses of base cations from the system. Most soils are able to compensate for this by mineral decomposition. In areas where gas concentrations are low, therefore, negative growth effects would not be expected despite high acidity in rain, and any small positive effects would be of such an uncontrolled random nature that much significance should not be attached to them. 2.4 The 'Ulrich Hypothesis' on Effect of Acid Deposition on Forest Ecosystems Recent observations on tree-growth morphology, tree death and the replacement of forest areas by scrub and herb vegetation, particularly in the Federal Republic of Germany,* have led to the formulation of an ecosystemorientated hypothesis on the effect of acid depositions on forest ecosystems. The hypothesis, stated most directly by Ulrich, is really a set of interrelated observations and explanations that attempt to put observed forest injury and degradation into a conceptual framework, elements of which may be further measured or tested experimentally. It may be set against other hypotheses that attempt to explain observations in terms of synergistic effects of air pollutants (NO2, SO2 and O3) or effects of pollutants other than SO2, particularly O3, acting singly. Ulrich looks at forest ecosystems as a whole and distinguishes between interception deposition (gaseous deposition and impaction of aerosols, including fog) and precipitation deposition (rain, snow and paniculate 'rain'). In terms of ecosystem response, his view distinguishes between an accumulation phase and effect phases. It does not imply that these occur concurrently, and suggests that, after some decades of deposition, forest ecosystems are now entering the effect phases. It visualises a concentration effect on plants, which through repair mechanisms can minimise these effects, and a total deposition effect on soils which, through a lack of similar repair mechanisms, gradually builds up and creates soil changes that affect vegetation adversely.

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During the accumulation phase, pollutant deposition, due to the nutrient component from pollution, stimulates growth in ecosystems: this is followed by the effect phases where the accumulation of H+ ions decreases soil pH, releases Ca2+, Mg2+, K+ etc. into leachate water and ultimately results in increased concentrations of AI3+, Mn3+ and Fe3+ in soil solutions. Accumulation of aluminium hydroxy salts causes coating of clay minerals and reduces cation-exchange capacity (and hence soil-buffering capacity), and a reduction of Ca:AI ratios below 2 causes growth depression in the vegetation. The overall effect sequence has been summarised as: Phase 1 Elements accumulate causing growth enhancement. Phase 2 Hydrogen ions accumulate in the soil and microorganisms and roots are damaged. Phase 3 Leaves and bark accumulate ions to toxic concentrations. Phase 4 Trees die and forests are replaced by moorland and acid heaths. It does not seem reasonable at this stage to expect either that Ulrich's hypothesis will be falsified or that a lack of falsification will emerge. Rather, one would expect to look for consistencies in the observed associations from place to place; begin to detect some kind of dose-response relationship; and recognise elements of ecological and edaphic plausibility (can mechanisms be proposed by which causal agents could be mediated?). Analogous systems that are of some interest in connection with the proposed effects of soil acidification are those of acid sulphate soils and pyritic colliery spoils. Both include mechanisms whereby acidity is continually generated within the substrate and a series of buffering mechanisms operates and is exhausted sequentially. Three or four buffering-capacity components may be recognised. Where carbonate minerals are present (calcium carbonate, ankerite, siderite) acid buffering occurs at slightly below neutral to alkaline pH values. The buffering capacity for each 1 per cent CaCOa present is 20 meq 100 g−1 soil and 24 meq 100 g−1 for each 1 per cent MgCO3. However, total carbonate content is an inadequate parameter of potential buffering capacity (as would be mass-balance considerations of the potential buffering capacity of granite). The reason for this is threefold. First, any mixtures within carbonates (FeCO3 components, for example) would have a net neutralising capacity of nil as the ferrous iron liberated from the carbonate is oxidised to ferric iron and precipitated as Fe(OH)3. Second, thin soil sections show a restriction of the neutralising ability of carbonate particles as reaction products accumulate around the surface, ‘sealing’ off any further neutralising ability. Third, the pH range over which the buffering capacity contributed by carbonates can be effective is limited by the partial pressure of CO2 in the soil atmosphere, assuming that the buffering reactions are in equilibrium with solid carbonate in the soil. Soil pH is also buffered by adsorption of H+ (or H3O+) on to the cation exchange surfaces at the expense of exchangeable bases which are displaced into solution. Where iron and aluminium hydroxy salts occur they can form ‘coatings’ that hinder the exchange capacity, severely reducing this component of the buffering capacity. Weak acids (carbonic acids and many organic acids), by dissociation, form another acid buffering mechanism, at lower pH values, and the fourth component of the buffering system, at rather low soil pH values, is provided by H3O+ and aluminosilicate mineral reactions. When exchangeable bases on cationexchange surfaces are exchanged for from solution, the adsorbed H3O+ migrates into the clay lattice, displacing AI3+ ions. The behaviour of such AI3+ is determined by the distribution of aluminium ions

*See

footnote on page 4.

44

HISTORICAL ASPECTS

Figure 2.9 Zone map of Great Britain showing distribution of lichens. Zones: 0–2, no lichens or very sparse; 3–4, leafy lichens; 5, some shrubby lichens; 6, lichen Usnea and other species indicating very low SO2 levels (Gilbert, 1974).

between the cation-exchange surfaces and solution and by the degree of hydrolysis of the aluminium ions in solution. The buffering capacity of the hydrolysis reaction depends upon the total amount of aluminium in solution. At a fixed concentration, the effect of any increase in total ionic strength is to reduce the buffering capacity to added H3O+ . Reactions such as these may keep soils buffered at pH values well below 5, possibly as low as pH 3.5. The almost ‘step-wise’ character of soil acidity buffering components renders plausible the build-up effect due to total acid inputs over relatively long periods of time. 2.4.1 Conclusions Ulrich’s hypothesis has not been widely tested and consistent observations from a variety of areas are not available. Knowledge of the range of tolerances of tree species and provenances to different species of aluminium ions is not in any way complete, although variation in sensitivity to general substrate acidity by tree species is well documented; however, the aluminium concentrations encountered in forest soils where dieback has occurred do not seem to be high enough to cause damage to tree roots. Direct effects of acid rain per se on vegetation are less likely to occur than the direct effect of gaseous pollutants but there will be increased losses of base cations from the system. Magnesium deficiency certainly occurs in damaged forest

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Figure 2.10 Dry deposition of sulphur over Great Britain (Barrett et al., 1983).

stands but at present it is not possible to say unequivocally whether it arises through damage to the needles and leaves of trees or to their roots. It has yet to be demonstrated that negative growth effects occur where gas concentrations are low but there is high acidity in rain. 2.5 Lichens as Indicators of Atmospheric Sulphur Pollution Lichens may be used as biological indicators of 'atmospheric pollution'. They have been used to indicate high fluoride levels near smelters, and their presence or absence has been correlated with general but often unquantified pollution; but their greatest value has been for the estimation of levels of atmospheric sulphur dioxide. It has been shown that, where SO2 is the main gaseous atmospheric pollutant, the mean annual levels can be estimated with reasonable accuracy by the presence or absence of different species of lichens. A really thorough investigation requires an expert, but even school-children can make useful observations. Thus, shrubby lichens are only found where SO2 levels are very low, leafy species can withstand somewhat higher levels, and poor lichen cover is

46

HISTORICAL ASPECTS

Table 2.8 ‘New’ forest damage in F.R.G. according to extent of damage Area, ha×106

Damage category

Forest area, %

Total damaged area, %

1.Slightly damaged 1.846 25 2. Damaged 0.635 8.5 3. Extensively damaged 0.064 0.9 Total (1+2+3) 2.545 34 Total forest area 7.406 100 From Forest Damage Inventory, 1983. F.R.G., Anon., 1983.

72.5 25 2.5 100

generally a sign of high SO2 levels. A zone map of Britain based on lichens observed by school-children (Figure 2.9) agrees closely with the measurements of sulphur dioxide levels, whether of emission or deposition (Figure 2.10). Quite different results are obtained when the levels of acid rain are studied. Figures 2.2 and 2.3 gave estimates of the deposition of hydrogen ions in rain in rural areas of Britain. The lack of agreement with Figure 2.9 suggests that lichens are not affected by acid rain—a conclusion supported by observations in Scandinavia, where they flourish in areas where damage to fisheries is reported. It Table 2.9 ‘New’ forest damage according to federal states in F.R.G. Forest area according to species State

Area,

ha×106

Proportion of total forest area,%

Damaged area (damage category 1+2+3) Area, ha×106 Proportion of total forest,%

Schleswig-Holstein 0.137 2 0.016 Niedersachsen 0.977 13 0.165 Nordrhein0.855 12 0.295 Westfalen Hessen 0.834 11 0.120 Rheinland-Pfalz 0.771 10 0.180 Baden1.303 18 0.645 Württemberg Bayern 2.444 33 1.115 Saarland 0.085 1 0.009 Total F.R.G. 7.406 100 2.454 From Forest Damage Inventory, 1983. F.R.G., Anon., 1983.

Proportion of total damage area,%

12 17 35

0.6 6 12

14 23 49

5 7 25

46 11 34

44 0.4 100

will be noted that in Scotland lichens flourish in Galloway and parts of the Highlands where the deposition of hydrogen ions is high, and are absent from the area between Glasgow and Edinburgh where SO2 is high and wet deposition low. It thus appears that lichens may be used to distinguish between gaseous SO2 and the acidic substances in precipitation.

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2.6 Central European Forest Die-back and its Relevance to Britain The recent phenomenon of ‘forest decline’ in central Europe has already been mentioned in Section 1 and in the preceding sub-sections of this section. The first reports— of decline in European Silver fir—did not attract much attention outside Germany, and even in that country itself it caused no serious alarm because the same phenomenon had been recorded on a number of occasions during the past 100 years or so. However, from about 1980 onwards not only did the damage to fir become very much worse, but there were also reports of decline in Norway spruce; this was all the more disquieting since it was already known that there were large areas of damaged—and indeed dead—spruce in nearby Czechoslovakia. A survey (much of it based on estimates rather than inventory) in 1982 suggested that 8% of West German forests were damaged, mostly lightly, the first signs being the loss of needles in the middle and upper crown together with yellowing of older needles. The 1983 survey (Tables 2.8, 2.9 and 2.10) shows that 34% of the forest is estimated as damaged, but it is probable that the bases of the two surveys are not quite the same. Only Bavaria and Baden-Württemberg carried out an inventory by sampling; North-Rhine Westphalia used a simple inventory and the other States made estimates. It will be seen from Table 2.10 that Norway spruce accounts for nearly half the damaged area. Damage was first seen and has been worst at high elevation; varies greatly between individual trees; affects trees in full light worse than those in shade; appears to be worse on older trees (over 60 years old) than on younger ones; appears to occur regardless of soil type and geology; has arisen at the same time in widely separated parts of Germany; and is worse in evergreen than in deciduous trees. Most of the areas affected have only low levels of SO2 and are reasonably well supplied with lichens. These facts have led the majority of German scientists to accept the hypothesis that damage is primarily from ozone. There is evidence from the Black Forest at least that concentrations of ozone are high and that the normal diurnal cycle of formation during the day and decay at night does not always operate in the mountains, the concentrations sometimes staying steady over a number of days. This ozone damage is followed by leaching of magnesium and calcium from the needles or leaves by acid mists or fogs (these have been discussed in Sections 1.1 and 1.5). There is no evidence that aluminium concentrations are higher under damaged crops than healthy ones. It must be emphasised that firm evidence is rather sparse, and the hypothesis may well be modified (or even dropped) as new facts, derived from experiments in Table 2.10 ‘New’ forest damage according to tree species in F.R.G. Forest area according to species Species Area,

ha×106

Damaged area as proportion of total (damage category 1+2+3)

Proportion of total forest area, Area, ha×106 %

Spruce 2.951 40 1.194 Pine 1.464 20 0.636 Fir 0.176 2 0.134 Beech 1.250 17 0.332 Oak 0.615 8 0.091 Others 0.950 13 0.158 Total 7.406 100 2.545 From Forest Damage Inventory, 1983. F.R.G., Anon, 1983.

Species area, %

Total damaged area, %

41 43 76 26 15 17 34

47 25 5 13 4 6 100

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HISTORICAL ASPECTS

the laboratory and forest, are produced. As has been mentioned earlier in this section, trees damaged through action of pollutants on the foliage might recover if the causes of damage were removed. Deciduous trees might well recover more quickly than evergreens. European Silver fir has never been planted in Britain except as an ornamental tree, and Norway spruce, although important in the past and still much planted as Christmas trees, forms only between 5 and 7% of Britain’s managed forests and less than 2% of current planting, taking state and private forestry together. Forests in Britain are at a much lower absolute elevation than in central Europe and, because of the different —maritime—climate, episodes of high ozone are less likely to persist than under continental conditions— though as mentioned earlier in this section, damage from O3 pollution on sensitive cultivars of plants has been recorded in most parts of the country. In addition, few coniferous forests in Britain are kept beyond 60 years of age, not only because it is more profitable to fell than to retain them, but because of the increasing risk of windthrow. Thus, although these differences make it difficult to extrapolate from continental Europe to Britain, the absence of any new forest damage or injury, not ascribable to known biological or climatic agencies, suggests that the same phenomenon is unlikely to be seen here. References Anon. 1983, Neuartige Waldschäden in der Bundesrepublik Deutschland. Bericht des Bundesministers für Ernährung, Landwirtschaft und Forsten zur Waldschadenserhebung 1983. Ashenden, T.W. and Mansfield, T.A. (1978). Extreme pollution sensitivity of grasses when SO2 and NO2 are present in the atmosphere together. Nature 273, 142–3. Ashmore, M.R., Bell, J.N.B., Dalpra, C. and Runeckles, V.C. (1980). Visible injury to crop species by ozone in the United Kingdom. Environmental Pollution (Ser. A) 21, 209–15. Ashmore, M.R., Bell, J.N.B. and Reily, C.L. (1978). A survey of ozone levels in the British Isles using indicator plants. Nature 276, 813–5. Awang, M.B. (1979). The effects of sulphur dioxide pollution on plant growth with special reference to Trifolium repens. Ph.D. thesis, University of Sheffield. Bache, B.W. (1983). The implication of rock weathering for acid neutralization. In: Ecological Effects of Acid Precipitation, 175–187. National Swedish Environment Protection Board, PM 1636. Barrett, C.F., Atkins, D.H.F., Cape, J.N., Fowler, D., Irwin, J.G., Kallend, A.S., Martin, A., Pitman, J.I., Scriven, R.A. and Tuck, A.F. (1983). Acid deposition in the United Kingdom. Warren Spring Laboratory, Stevenage, Herts. Binns, W.O. and Redfern, D.B. (1983). Acid rain and forest decline in W. Germany. Forestry Commission Research and Development Paper 131. Black, V.J. and Unsworth, M.H. (1980). Stomatal responses to sulphur dioxide and vapour pressure deficit. Journal of Experimental Botany 31. 667–77. Black, V.J. (1982). Effects of sulphur dioxide on physiological processes in plants. In: Effects of Gaseous Air Pollution in Agriculture and Horticulture (Eds M.H.Unsworth and O.P. Ormrod), pp. 67–91. Butterworth, London. Bormann, F. and Likens, G.E. (1979). Pattern and process in a forested ecosystem. Springer-Verlag. Bosch, C., Pfannkuch, E., Baum, U. and Rehfuess, K.E. (1983). Über die Erkrankung der Fichte (Picea abies Karst.) in den Hochlagen des Bayerischen Waldes. Forstw. Centralblatt 102, 167–81. van Breemen, N. (1972). Soil forming processes in acid sulphate soils. In: Acid Sulphate Soils 1 (Ed. H. Dost), pp. 66–128. ILRI, Wageningen. Brimblecombe P. and Pitman J. (1980). Long-term deposit at Rothamsted, Southern England. Tellus 32, 261–7. Brown, K.A. (1980). The distribution of sulphur compounds in a peat bog in relation to streamwater chemistry. C.E.R.L. Report RD/L/N 1 50/80. Chadwick, M.J., Cornwell, S.M. and Palmer, M.E. (1969). Exchangeable acidity in unburnt colliery spoil. Nature 222, 161–2.

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Crittenden, P.D. and Read, D.J. (1978). The effects of air pollution on plant growth, with special reference to sulphur dioxide. II. Growth studies with Lolium perenne L. New Phytologist 80, 49–65. Crittenden, P.D. and Read, D.J. (1979). Ibid. III. Growth studies with Lolium multiflorum Lam. and Dactylis glomerata L. New Phytologist 83. 645–51. Crowther C. and Ruston A.G. (1911–12). The nature, distribution and effects upon vegetation of atmospheric impurities in and near an industrial town. J. agric. Sci. 4, 25–65. Crowther C. and Stewart D.W. (1912–1913). The distribution of atmospheric impurities in the neighbourhood of an industrial city. J. agric. Sci. 5, 391–408. Davison, A.W. and Bailey, I.F. (1982). SO2 pollution reduces the freezing resistance of ryegrass. Nature 297, 400–2. Evans, L.S., Conway, C.A. and Lewin, K.F. (1980). Yield responses of field-grown soybeans exposed to simulated acid rain. In: Proceedings of an International Conference on the Ecological Impact of Acid Precipitation (Eds D.Drablos and A. Tollan), pp. 162–3. SNSF, Oslo. Ferguson, N.P., Lee, J.A. and Bell, J.N.B. (1978). Effects of sulphur pollutants on the growth of Sphagnum species. Envir. Pollut. 16, 151–61 Ferguson, N.P. and Lee, J.A. (1980). Some effects of bisulphate and sulphate on the growth of Sphagnum species in the field Envir. Pollut. 21 A, 59–71. Ferguson, P. and Lee, J.A. (1983). Past and present sulphur pollution in the southern Pennines. Atmos. Environ. 17, 1131–7. Fujiwara, T. (1973). Effects of nitrogen oxides on plants. Kogai To Taisaku 9, 253–7. Garrels, R.M. and Christ, C.L. (1965). Solutions, Minerals and Equilibria. Harper, New York. Garsed, S.G., Mueller, P.W. and Rutter, A.J. (1982). An experimental design for studying the effects of fluctuating concentrations of SO2 on plants. In: Effects of Gaseous Air Pollution in Agriculture and Horticulture (Eds M.H.Unsworth and O.P.Ormrod), pp. 455–7. Butterworth, London. Gilbert, O.L. (1974). Air pollution survey by school children. Environmental Pollution 6, 175–80. IERE (1981). Impacts of acid deposition on vegetation, soil, drainage water and streams. In: Effects of SO2 and its Derivatives on Health and Ecology. Vol. 2. Natural Ecosystems, Agriculture, Forestry and Fisheries. International Electric Research Exchange. Jacobson, J.S. (1980). The influence of rainfall composition on the yield and quality of agricultural crops. In: Proc. Int. Conf. on Ecological Impact of Acid Precipitation, Norway, 1980, pp. 41–6. SNSF-project, Osla-Aas. Jones, T. and Mansfield, T.A. (1982). Studies on dry matter partitioning and distribution of 14C-labelled assimilates in plants of Phleum pratense exposed to SO2 pollution. Environmental Pollution (Ser. A), 28, 199–207. Lane, P.I. (1983). Ambient levels of sulphur and nitrogen oxides in the U.K. and their effects on crop growth. Ph.D. thesis. University of London. Livett, E.A., Lee, J.A. and Tallis, J.H. (1979). Lead, zinc and copper analyses of British blanket peats. J. Ecol. 67, 865–91. Mansfield, T.A. and Freer-Smith, P.H. (1984). The role of stomata in resistance mechanisms. In: Proceedings of First International Symposium on Gaseous Air Pollutants and Plant Metabolism (Eds M.J.Koziol and F.R.Whatley). Butterworth, London. Martin, A. (1979). A survey of the acidity of rainwater over large areas of Great Britain. Sci. Total Envir. 13, 119–30. Miller, H.G. and Miller, J.D. (1984). Personal communication. Miller, H.G. (1983). Studies of proton flux in forests and heaths in Scotland. In: Effects of accumulation of air pollutants in forest ecosystems, 183–93 (Eds Ulrich and Pankrath). Reidel. Mooi, J. (1984). Personal communication. Nilsson, S.I., Miller, H.G. and Miller, J.D. (1982). Forest growth as a possible cause of soil and water acidification—an examination of the concept. Oikos, 39, 40–9. Palmer, M.E. (1978). Acidity and nutrient availability in colliery spoil. In: Environmental Management of Mineral Wastes (Eds G.T.Goodman and M.J.Chadwick), pp. 85–126. Sijthoff and Noordhoff, Alphen aan den Rijn. Press, M.C. and Lee, J.A. (1982). Nitrate reductase activity of Sphagnum species in the South Pennines. New Phytologist 92, 487–94.

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Prinz, B., Krause, G.H.M. and Stratmann, H. (1982). Forest damage in the Federal Republic of Germany. LIS Report No. 28. Land Institute for Pollution Control of North-Rhine Westphalia, Essen. Rehfuess, K.E. (1983). Walderkrankungen und Immissionen—eine Zwischenbilanz. Allgemeine Forstzeitschrift (24) 18th June, 601–10. Reinert, R.A. and Gray, T.N. (1981). The response of radish to nitrogen dioxide, sulfur dioxide, and ozone, alone and in combination. Journal of Environmental Quality 10, 240–3. Reinert, R.A. and Heck, W.W. (1982). Effects of nitrogen dioxide in combination with sulfur dioxide and ozone on selected crops. In: Air Pollution by Nitrogen Oxides (Eds T. Schneider and L. Granat), pp. 533–46. Elsevier, Amsterdam. Richter, D.D., Johnson, D.W. and Todd, D.E. (1983). Atmospheric sulfur deposition, neutralization, and ion leaching in two deciduous forest ecosystems. J. Environ. Qual. 12, 263–70. Skeffington, R.A. (1983). Soil properties under three species of tree in Southern England in relation to acid deposition in throughfall. In: Effects of accumulation of air pollutants in forest ecosystems, 219–31 (Ed. Ulrich and Pankrath). Reidel. Smith, R.A. (1852). On the air and rain of Manchester. Mem. Proc. lit. Manchr. lit. phil. Soc. Series (2) 10. 207–17. Sollins, P., Grier, C.C., McCorison, S.M., Cromack, K., Fogel, R. and Frederiksen, R.L. (1980). The internal element cycle of an old-growth Douglas fir ecosystem in Western Oregon. Ecological Monographs 50, 261–85. Swedish Ministry of Agriculture (1982). Acidification Today and Tomorrow. Environment ‘82 Committee, Stockholm, Sweden. Troiano, J., Colavito, L.H., McCune, D.C. and Jacobson, J.S. (1983). Effects of acidity of simulated rain and its joint action with ambient ozone on measures of biomass and yield in soybean. Environmental and Experimental Botony (in press). Ulrich, B. (1983). An ecosystem orientated hypothesis on the effect of air pollution on forest ecosystems. In: Ecological Effects of Acid Deposition pp. 221–31. Report PM 1636, National Swedish Environment Protection Board, Solna. Unsworth, M.H., Biscoe, P.V. and Black, V.J. (1976). Analysis of gas exchange between plants and polluted atmospheres. In: Effects of Air Pollutants on Plants (ed. T.A.Mansfield), pp. 5–16. Cambridge University Press. Wellburn, A.R., Higginson, C., Robinson, D. and Walmsley, C. (1981). Biochemical explanations of more than additive inhibitory effects of low atmospheric levels of sulphur dioxide plus nitrogen dioxide upon plants. New Phytologist 88, 223–37. Whitmore, M.E. and Freer-Smith, P.M. (1982). Growth effects of SO2 and/or NO2 on woody plants and grasses observed during the spring and summer. Nature 300, 55–7. Whitmore, M.E. and Mansfield, T.A. (1983). Effects of long-term exposure to SO2 and NO2 on Poa pratensis and other grasses. Environmental Pollution (Ser. A), 31, 217–35. Whitmore, M.E. (1982). A study of the effects of SO2 and NO2 pollution on grasses with special reference to Poa pratensis L. Ph.D. Thesis, University of Lancaster.

THE WATT COMMITTEE ON ENERGY REPORT NUMBER 14

Section 3 Freshwater D.Hammerton

This paper presents the work of Sub-group 3 (Freshwater) of the Watt Committee working group on Acid Rain. Membership of Sub-group 3

D.Hammerton (Chairman) Dr. R.W.Battarbee T.R.Carrick D.H.Crawshaw R.Harriman A.V.Holden Dr P.S.Maitland C.Martin Dr H.G.Miller Dr J.H.Stoner

Freshwater

3.1 Introduction 3.1.1 Objectives This section of the Report examines the impact of acid deposition on freshwater ecosystems in relation to other factors such as geology, soil type, vegetation or land use. An attempt has been made to assess the scale of the problem and the degree of damage to the aquatic environment. Most of the information has been derived from the published literature, but it is supplemented by personal knowledge of members of the Subgroup and by the response to a questionnaire that was sent to all the water authorities in Great Britain. 3.1.2 Acidification mechanisms Naturally acid waters, it should be realised, are quite widespread in Great Britain. Such waters, often below pH 5.0, are mainly small pools and headwater streams which occur in association with peatlands in areas of slowweathering bedrock. These waters are characterised by their high organic content, and most, but not all, are fishless. They also have a characteristic flora and fauna of acid-tolerant species. Soil oxidation processes are particularly important in the large accumulations of organic nitrogen and sulphur compounds that occur in upland soils. These compounds are oxidised to inorganic sulphates and nitrates during long, dry periods and are transported in solution into drainage streams during the next wet period. Sulphate levels are often 2-3 times the mean levels and most of the major cations increase, including acidity, to balance the anion increase. This process occurs infrequently and is generally of a temporary nature, but it may be of great importance in some surface waters in the event of acid surges being produced. Sea-salt intrusions occur in coastal areas, especially during winter months. Concentrations of sodium chloride can increase by a factor of 5-10 times or more, and during these episodes the sodium ion acts as an effective regenerator of the cation exchange sites. High concentrations of calcium, magnesium and hydrogen ions are released into streams, but the frequency and extent of this process has not been established. Acid deposition occurs when, in sensitive catchment areas, the exchange of calcium and magnesium for hydrogen in the soil, on occasions, is inadequate to balance the high anion loadings, particularly sulphate; so the hydrogen ions and some soil-derived aluminium and manganese pass into drainage streams and acidify

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the aquatic system. Sulphate depositions could eventually reduce the availability of base cations in the uppermost soil horizons so that water-flow moving laterally to stream courses through these horizons is neutralised less than a flow that follows deeper pathways. Nitrate is also present in precipitation, but is less conservative than sulphate and, in nitrogen-deficient upland catchments, is considered to be of minor importance, except perhaps during snow-melt or storm-flow conditions. Land use changes—changes in land use or intensity of land management—have been suggested as causes of stream acidification. In Scandinavia, decreased grazing and increased intensity of forest management have both been put forward as factors. In the U.K. the greatest known change has been the major expansion of forests. An effect of afforestation on stream-water acidity has been observed in a number of localities. Possible processes leading to this effect include more efficient scavenging of dry and occult deposition, cation uptake of trees, greater evapotranspiration, improved soil aeration leading to greater oxidation of sulphur, changes in catchment hydrology or increased soil organic matter. Changes in agricultural practices, e.g. cessation of liming on upland pastures, may well affect stream acidity, although this has not been quantified. 3.2 Techniques for Assessing Acidification 3.2.1 Chemical records, especially of pH levels Although ideally chemical records are the most satisfactory way of assessing changes in acidification levels, since the measurements are direct, the method has limited applicability, since records exist only for a small number of surface waters over a generally short time-scale. Moreover, the difficulty of making pH measurements in low-conductivity waters suggests that earlier measurements may be inadequate. 3.2.2 Sediment records Lake sediments preserve a chemical and biological record of a lake’s history that can be revealed by sediment analysis—a method that has unique advantages. It is potentially useful at all sites, the record is usually continuous and accurate dating can enable the timing and rate of acidification to be assessed. The diatom record is most widely used for this purpose, and the quantification of the relationship between diatom assemblages and pH values has become increasingly accurate. Specific levels of past pH values can be reconstructed with a standard error better than ±0.5 pH units. 3.2.3 Fishery records Changes in fishery status have been used as presumed evidence of acidification, particularly where salmonids had declined, since these species fail to reproduce successfully in highly acid waters.

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3.2.4 Other biological records Plant and animal communities are sensitive to pH change, and changes in composition can be used to assess whether acidification has taken place. Museum collections can be particularly valuable in this context. 3.2.5 Chemical models In many areas of Europe and North America, where acid waters have been recorded, there are no long-term data for the assessment of chemical change. Indirect methods using chemical models have therefore been employed to make an assessment of the current acidification status of lakes and streams. The models are almost entirely based on comparisons of relationships in unpolluted areas. The early models used plots of alkalinity against calcium and magnesium, and revealed that waters in vulnerable areas subject to acid depositions showed significant deviations from a regression line describing unpolluted waters in such areas. In recent years Henriksen has developed a sophisticated model for predicting preacidification pH and alkalinity values of Norwegian waters: it has been adapted by Wright for calculating acidification values of North American and European waters. However, it should be pointed out that some of the basic assumptions of this model are being questioned by research workers in Europe and North America. 3.3 Effects of Acidification on Flora and Fauna The evidence for the effects of acidification on freshwater organisms is based almost entirely on comparative studies of freshwaters of differing acidity or experiments in lakes or streams. Few freshwaters have been closely monitored during the acidification process. 3.3.1 Algae and higher plants The available evidence shows that in more acid waters there is a marked reduction in the species diversity of algae and higher plants, while changes in community structure have been observed in phytoplankton, periphyton and macrophytes. In the phytoplankton, Chrysophyceae and Bacillariophyceae tend to be replaced by Dinophyceae and Cryptophyceae. In highly acid waters, dominance by chlorophyta and cyanophytan algal mats has been observed. The expansion of algal mats and Sphagnum moss in clear acid lakes can lead to the reduction of rosette macrophytes. 3.3.2 Zooplankton Among zooplankton, in general, there is a simplification of the community structure and dominance by a few species, copepoda usually replacing cladocera, for example. In particular, survival, reproduction and distribution of zooplankton appear to be adversely affected. Productivity is generally lower in acidic compared with similar non-acidic lakes.

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3.3.3 Zoobenthos Populations of zoobenthos are impoverished in acid environments. In particular, there is a loss of ‘hard water’ groups such as leeches, molluscs and crustaceans. Most mayflies and some stoneflies disappear as pH level declines, along with various other aquatic insects, e.g. some trichopterans, chironomids and dipterans. These changes may arise from the direct physiological impact of acids or metals, alteration in food supply or changes in predator-prey relationships. In particular, it is claimed that many of the reported effects of acidification result from the disappearance of fish. For example, the waterbug Glaenocorisa has recently colonised lakes in Sweden where acidification has eliminated fish. 3.3.4 Fish There is abundant evidence of the impact of increased acidity on fish. The effects include mortality, reproductive failure and reduced growth. It is not clear to what extent these changes are due to man-induced acidity, but there is a close geographical correlation with areas of known high acid deposition. Mortalities in acid waters have been associated with rapid change in pH values, as in snow melts or heavy autumn rains. In such cases the death of the fish is often due to a combination of low calcium concentration, low pH value and elevated levels of metals, particularly aluminium. Although adult fish can die directly as a result of acidity, more commonly recruitment is impaired and the species eventually disappears. In trout, for example, it has been shown that acid water below pH c. 4.5 inhibits production of the enzyme that breaks down the tough outer coating of the egg and the larval fish is trapped inside. Fish from acid lakes have been shown, in a number of studies, to contain elevated levels of mercury and other metals. The significance of this has not been evaluated. Acidity level and aluminium and calcium contents have now been identified as the critical parameters for egg and fish survival. At pH levels below 5.5, survival decreases with increasing concentrations of acidity and ionic species of aluminium. However, survival in waters with similar levels of acidity and ionic aluminium content increases with increasing calcium content (Howells et al., 1983). 3.3.5 Vertebrates other than fish There is evidence that amphibians and other higher vertebrates are affected by acidity. Amphibia may be particularly vulnerable, since most species in northern latitudes lay their eggs in meltwater ponds. Some birds may benefit from acidification through the loss of fish populations that prey on the invertebrate food of such birds as goldeneye, but it has been suggested in Sweden that others, e.g. pied flycatchers, may be laying thin-shelled eggs as a result of eating lake insects with a high aluminium content. 3.4 Evidence for Acidification outside U.K. The first evidence that poorly buffered freshwaters had become more acidic in recent decades was obtained in Scandinavia, specifically in southern Norway and Sweden, where declines in fish stocks had been reported since about 1950. The catchments exhibiting changes are characterised by granitic or similar

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bedrock from which their overlying soils are usually derived. In most areas that show evidence of acidification, the annual volume-weighted average pH in precipitation is below 4.5. Norway Studies in southern Norway have shown that the pH level in many lakes has decreased by 0.5–1. 0 units over two or three decades and in nine by more than 1.25 units. Lakes in this area had a lower pH for a given calcium level than those in geologically similar areas in the northwest where rainfall is less acidic. High aluminium levels have frequently been found in the more acid waters. Biological evidence has included the increased growth of filamentous algae in streams, the development of Sphagnum and algal mats over sediments, changes in diatom populations in sediment cores and particularly the loss of fish stocks. Salmon in seven southern rivers have been virtually eliminated, although catches in many other rivers have increased. Brown trout have declined seriously since the 1920s, and perch and char have been adversely affected. Losses in lakes have been particularly severe since 1960, the effects being first noticed in high-altitude lakes. Overall, in the south, 50% of the original fish populations has been lost and 60% of the remainder has declined. The loss of salmonid fish populations has been particularly associated with the loss of juveniles during spate conditions following spring snow-melt and heavy autumn rains after dry summers. Sweden Of 85000 Swedish lakes exceeding 1 ha in area, it is estimated that more than 18000 are acidic, 4000 with reportedly very serious biological consequences. In about 9000 lakes in southern and central Sweden fish stocks have been damaged to varying extents, and a further 1000 lakes in northern Sweden are believed to have been similarly affected. Finland Increased acidity in freshwaters following spring floods caused by snow-melt has been reported in Finland. Denmark Danish surface waters are mostly hard (very calcareous), but in Jutland there are a few lakes with alkalinities below 100µ eq l−1 (2 mg l−1) and with pH values down to 4.2. A recent report has suggested that there was a decrease of about 0.4 pH units between the late 1950s and the late 1970s in one lake in central Jutland. Belgium Acid precipitation is reported to have caused reduced pH levels in bog pools in northern Belgium. In one area the pH level decreased by one unit in about 10 years, and in another by about two units in 40 years. The ionic concentrations in the pools were higher than in similar pools in Scandinavia and Britain, and aluminium concentrations were between 0.3 and 8 mg l−1. German Democratic Republic Lake acidification has been measured in Thüringer Wald and Erzgebirge in East Germany, the lakes in the latter area suffering declines or complete destruction of trout stocks: pH values as low as 3.9 have been recorded. Czechoslovakia Some lakes in the Bohemian forest of Czechoslovakia are reported to have been acidified to a level of about pH 4.5. Netherlands From a study of changes in the diatom flora of Dutch moorland pools, it was deduced that pH values had fallen from an estimated 4-6 in 1920 to a measured range of 3.7–4.6 in 1978. Canada The first evidence of increasing acidity in Canadian lakes was reported in 1972 in the La Cloche mountains, Ontario, resulting from acidic emissions to the atmosphere from metal smelters at Sudbury, 65 km away. Hydrogen-ion concentrations in some lakes had increased 10- to 100-fold in a period of over 10 years with a deterioration in fish stocks. Later, other lakes in southcentral Ontario were shown to have increased in acidity. Fish populations in 38 lakes in the La Cloche Mountains, Ontario, are known to have been lost or reduced, and about 54 populations of different species have been lost. The disappearance of fish populations in Lumsden, George and OSA Lakes has also been well documented. The pH level fell by about two units

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in 10 years. Many other lakes in the Sudbury area of Ontario are believed to have lost their fish populations, but it should be noted that this is an area of heavy industrial pollution. In Nova Scotia, 16 lakes were reported in 1979 to have become significantly more acid since measurements had been made 21 years previously. Three rivers were also shown to have become more acid over 10–20 years while seven former salmon rivers no longer support salmon runs, all now having pH levels below 4.7. The pH level had decreased in the last 27 years and salmon catches in several other rivers had declined. The Laurentides Park area of Quebec contains over 2500 lakes, most of which are poorly buffered. A 1980 survey of many of these lakes, studied about 40 years earlier, concluded that pH levels had decreased by up to 0.95 units. United States of America A survey published in 1976 of 217 lakes in the Adirondack Mountains of New York, in the northeastern U.S.A., provided evidence of decreasing pH levels in 40 of the lakes which had previously been analysed during 1929-37. There was also a significant increase in the number of lakes without fish populations. Another survey of 396 Adirondacks lakes in 1979 included 138 which had been examined for pH level in 1930-34. Although measurements by pH-meter and a colorimetric comparator disagreed, colorimetric data compared with data obtained by the same method in the earlier period indicated a general decrease in pH since the earlier sampling. Other investigations of surface water in New Jersey and Pennsylvania suggested an increase in acidity. In California, evidence was found of a decrease in pH in two Sierra Nevada reservoirs between 1955 and 1980, and it was concluded that this was due to the increased acidity of atmospheric deposition, especially resulting from emissions of nitrogen oxides from automobile exhausts. In Maine a critical analysis of pH data from 1368 lakes between 1937 and 1974 showed that mean pH levels had decreased from 6.85 to 5.95, most of the change having taken place in the 1950s. Some sediment-core studies of diatom succession have been made in Adirondacks lakes and evidence was found of an increase in dominance by acid-tolerant species. Evidence for the loss of fish stocks in the U.S.A. was first obtained in the Adirondacks lakes. Of 40 highelevation lakes, only 10% had no fish in the 1930s, but by 1975 a further 17 lakes had lost their fish populations. Of all the Adirondacks lakes, about 180 (6%) have lost their fish stocks. Proof of complete absence of a fish population is not possible, but it is understood that survey techniques that once yielded fish no longer do so. No other areas of the U.S.A. have so far produced definite evidence of the loss of fish stocks due to acidification, although the loss of species in rivers and lakes in the eastern U.S.A. for undefined reasons has been documented. 3.5 Evidence for Acidification in U.K. 3.5.1 Chemical records There are few chemical records for acid lakes in Britain, and in most instances where information is available the data are limited in some way. The water sample may have been taken at an unrepresentative time of year, sampling was not regularly repeated over a period of years or the sample may have been taken too recently for any significant change to be observed. Information from the Galloway area of Scotland indicates a decline in pH level of between 1 and 2 pH units from values of pH>6 measured 20–30 years ago. A few lochs in the Trossachs area, north of Glasgow, also

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appear to have become more acid in the last 30 years. In the Cairngorm area of eastern Scotland there is little evidence of any significant decline in the pH levels of upland lochs from results of 1956 to the present day, although many of the lochs not sampled earlier now have pH values below pH 5.0. It is only in the case of the English Lake District lakes that routine monitoring has been continued over a significant period or lakes have been re-surveyed from time to time. Because of the recent publicity on acid rainfall and its impact on freshwaters, the Freshwater Biological Association published a major review article on this subject by D.W.Sutcliffe in its 1983 Annual Report. In his review, Sutcliffe examined long-term changes in Cumbrian water bodies between 1928 and 1980. He found that in productive lakes studied since 1939 or 1945, alkalinity, conductivity and pH had all increased markedly in the last 10–15 years and by as much as 50% in the case of alkalinity. In general, this appears to have been a result of increased productivity due to increasing nutrients from sewage and other sources. For unproductive upland lakes and tarns on Borrowdale Volcanics and Skiddaw Slates there are insufficient data to draw detailed conclusions, except that pH and alkalinity values have not altered markedly in the past 30 years. Although there is no evidence for any increase in acidity of atmospheric precipitation, Sutcliffe states that there is some evidence of enhanced alkalisation of these water bodies. In his concluding remarks Sutcliffe states that there is no evidence of an increase in acidity (lower pH or alkalinity) of water bodies in the Lake District over the last 50 years and that bulk precipitation of Wraymires has had a mean pH value of c. 4.4 for at least 28 years. 3.5.2 Sediment records Diatom analysis of the recent sediments in Galloway lakes, southwest Scotland, has shown that lochs on the granitic uplands have become more acid (Flower and Battarbee, 1983). Four lochs have now been considered: Loch Enoch, Round Loch of Glenhead, Loch Dee and Loch Grannoch (see Figure 3.1). The data allow the following conclusions. (1) (2) (3) (4)

All four lakes have been acidified by 0.5 to 1.2 pH units. Acidification has been a continuous process since about AD 1850–1900. Pre-acidification pH values were constant at between 5.5 and 6.1. In this case afforestation has not been a major factor, since lochs with unafforested catchments have been acidified and acidification in lochs with afforested catchments took place before the planting of trees. (5) The beginning and extent of acidification varies between lakes, depending on catchment characteristics. There is an apparent contradiction between the Galloway and Lake District evidence for acidification. It is clear, however, from the Galloway data that the Lake District records may not be old enough and that, for some lakes, the process of acidification may be too slow for any significant difference to be observed between 1950 and the present.

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Figure 3.1 Loch Dee, Galloway, Scotland, showing typical landscape with hills, forest and rough pasture.

3.5.3 Fishery records Evidence for changes in fishery status of waters in the U.K. due to acid deposition is extremely limited because of the paucity of historical information. A number of hill lochs in Galloway have lost their fish or have shown a decline in fish populations during the past 50 years. Trout populations in many of the adjacent spawning streams have disappeared or are virtually extinct. The most severely affected streams are those draining afforested catchments. The absence of fish has also been reported in acid streams in northern England and Wales, although historical information is generally unavailable. Apart from isolated incidents, there is no evidence that declining stocks are due to mortalities of adult fish and it is generally accepted that lack of recruitment, due to egg and fry mortality, is the rate-determining step. 3.5.4 Chemical models Using the models described in sub-section 3.2.5, acidification values were calculated for selected surface waters in the U.K. The lowest values were found for waters in northwest Scotland, where acid deposition rates are low. Waters in the Cairngorms area of Scotland and in Wales are acidified to a greater extent. The highest values are found for waters in areas close to major industrial conurbations, i.e. the Galloway and Trossachs areas of Scotland and the English Lake District, where depositions of sulphur are high and probably have been for many decades. 3.6 Extent and Distribution of Acid Waters in Great Britain In the U.K., as elsewhere in Europe, the acidification of surface waters is closely related to geology and soil type. Fortunately, the major proportion of the surface area comprises rocks and soils that are rich or moderately rich in calcium and are therefore well able to neutralise the acidity in rainfall. Acid lakes and streams have generally only been found in areas where the underlying geology comprises granites, slates and sandstones. However, it is only recently that the presence of acidified lakes and streams has been identified as a problem, mainly in Scotland and Wales. Many areas have still not been surveyed, and thus the full extent of the acidification of U.K. waters is not known.

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Figure 3.2 River Tywi catchment above Llyn Brianne, Wales.

3.6.1 Scotland In Scotland the biggest area of acidified waters is to be found in the Galloway hills of the southwest. Here some 57 lakes have been shown to be acidified in varying degrees, many of these being highly acidic, with sparse or non-existent fish populations which are associated with elevated levels of aluminium and heavy metals. A related survey of 39 streams in this area showed that many were also highly acidic. The Solway River Purification Board now considers that acidification is a serious problem and is currently involved in a joint research project with the Department of Agriculture and Fisheries for Scotland and the Forestry Commission to study the ecological effects of acid precipitation and afforestation together with the remedial use of lime and modified tree-planting regimes. Further north in Scotland, acid streams have been identified in the Duchray and Loch Chon catchment areas. Here afforestation has been shown by the Department of Agriculture and Fisheries for Scotland to be an important factor. Streams draining forested catchments were devoid of fish because of elevated levels of hydrogen ions and aluminium, while nearby moorland streams supported trout. The Clyde River Purification Board recently surveyed over 90 upland streams throughout its area and found that the 17 most acidic streams (pH 4.6–5.6) were in areas where the underlying rocks were predominantly granitic. The acid streams were all low in calcium, alkalinity and conductivity, whereas aluminium concentrations up to 0.4 mg l−1 were found in the most acid waters. The Highland River Purification Board in conjunction with the Water Research Centre has examined the relationship between pH and flow in the Findhorn catchment. A statistical analysis has shown that (1) there is a good correlation between flow and pH 7.5–8.0 at low flows to pH 6.0–6.5 at high flows, (2) there was a tendency towards increased acidity from 1976 to 1978 and there has since been a decline in acidity and (3) one exceptionally low pH measurement (4.0) coincided with a period of snow-melt. 3.6.2 Wales The Welsh Water Authority examined water-quality data for 13 streams in the upper Tywi catchment in west Wales and related this to rainwater quality between February 1981 and January 1983. Though rainfall in the area is amongst the cleanest in the U.K., episodic depositions of acidity were found to occur. ‘Excess’ sulphate concentrations were shown to be largely dependent on the collection of dry deposition by vegetation with the result that stream acidity and aluminium concentrations were determined by the

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combination of buffering capacity and bankside vegetation. The study showed that streams could be classified on the basis of water hardness and land use. Stream surveys and caged fish survival studies showed that unafforested streams with an average calcium content of >80 µ eq l−1 (mean pH 5.5–6.0) could support trout populations and a reasonably diverse invertebrate fauna. Unafforested streams with a mean calcium content of <80 µ eq l−1 (mean pH 5.0–5.5) could support only a sparse fish population and limited invertebrate diversity. Afforested streams with an average calcium content of < 130 µ eq l–1 (mean pH 4.5– 5.0) were incapable of supporting trout, and the invertebrate fauna was very restricted. It is likely that areas receiving similar wet and dry deposition can be classified on this basis, and that water-quality data can be used to predict the fishery and invertebrate status of streams and the likely effects of changed land use. The River Tywi (Figure 3.2) cannot now support salmonid fish and has a depleted invertebrate fauna due to acidity and high aluminium concentrations. Approximately 80% of the catchment was afforested in the 1960s. Extensive studies of this catchment have been carried out* and it is proposed to carry out land management studies over the next few years. The Welsh Water Authority is concerned that many other areas in Wales may be similarly affected and that Welsh fisheries may be particularly at risk, as these are based primarily on highly sensitive salmonid species and occur in areas overlying poorly weathering, nutrient-poor rocks and soils. The need for a widescale comprehensive waterquality and fisheries monitoring programme in Wales is underlined by the results of the survey of 29 lakes near Blaenau Festiniog in 1983 which indicated that, although some of these lakes may be naturally acidic, the conditions (especially pH and aluminium levels) may now be inimical to fish. Although some of these low-pH lakes are thought to contain trout, anecdotal evidence indicates that these stocks may be in decline. 3.6.3 England The North West Water Authority has surveyed fish stocks, benthic macro-invertebrates and water chemistry of 75 upland stream sites of salmonid nursery stream type within the Lake District National Park. Thirtyfour sites had mean pH values below 6.0, and many sites that were fishless or had very low salmonid densities had a (geometric) mean pH value of 5.6 or less. These sites were grouped in the upper parts of drainage systems associated with the western mountain mass of the Lake District where the underlying rock was of the Borrowdale Volcanic series and rainfall was high—in excess of 2200 mm/year. High aluminium levels (c. 1.0 mg l−1 were found at some sites. Distribution of Gammarus sp. and Baetis sp. showed clear evidence of restriction at low pH levels, and absence of the latter species seemed to be a good practical indicator of low-density or non-existent trout stocks. A major fish mortality on the River Esk in June 1980 under summer spate conditions is thought to have been due to pH depression to around 5.0 coupled with elevated levels of aluminium from more acid tributaries. Further studies are in progress involving continuous monitoring of pH level, temperature and river levels at two sites with fish movement at one. Rainfall analysis of individual ‘episodes’ is also being undertaken. A second, smaller mortality, involving 30–40 fresh-run sea trout, took place in September 1983. It was associated with a fall in pH level after heavy rainfall to around 4.3, and aluminium levels of around 0.25 mg l−1 were recorded.

*See

paper by Stoner, Gee and Wade, listed in References (page 40).

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In common with Yorkshire and Northumbrian Water Authorities, the Authority has a number of acid reservoirs, some with pH<4.0 and many with pH<6.0. Most of these receive drainage from Pennine peat moorlands, and the acidity has previously been assumed to be derived from humic acids. The South West Water Authority is not carrying out work specifically connected with acid rain, but has felt some concern over a general tendency for lower pH values to be observed in some of the rivers. Data on some Devon rivers over the past eight years indicate a downward trend but need to be treated with caution for a variety of reasons. A possible cause is a change in the farming subsidy for lime, although on the Dart the apparent fall in pH level is not supported by a reduction in calcium—which might be expected following a reduction in liming. Low pH values do occasionally occur in winter in the headwaters of rivers draining moorland areas, but this is more likely to be due to the underlying geology and soil type of the catchment. However, on the uppermost 12 km of the River Plym low pH levels are consistently recorded in winter; but there is no evidence that this is due to acid rain. The Yorkshire Water Authority has a number of acid streams and a considerable number of acid reservoirs. For example, 28 reservoirs have a pH value of 6.0 or less, 15 have a pH value of 5.0 or less and seven have below 4.5. However, most of these streams and reservoirs receive drainage from Pennine moorlands or acid mine waters, and the Authority does not consider that acid rainfall is causing any identifiable problems in the area. From routine monitoring the Northumbrian Water Authority has no evidence of acidity, but since the monitoring sites are some distances from the headwaters it would not expect these sites to exhibit acidity. It is known that some feeder streams to reservoirs can show acidity below pH 4 based on short-term sampling, but the nature of the catchments—peat moorland—has been considered to be the predominating if not sole origin of the acidity. In the remaining areas of England there is no evidence that acid waters constitute a problem. Indeed responses from the Anglian, Thames, Southern and Wessex Water Authorities have indicated that no acid waters below pH 7.0 have been found at any of the sites routinely monitored. However, it is known from other studies that some streams on sandstone catchments in the southeast are very acid and that some sites in the New Forest area of Hampshire were ‘acidified’ according to the Henriksen method for plotting pH level against calcium. Some of these sites, which in 1974 had pH values below 5.0 and low calcium concentrations, were shown in a survey carried out in 1982 to have become more acid, but the results must be treated with caution because of the small number of samples taken in 1974. The Severn Trent Water Authority reported slightly acid waters in two areas but, in general, has no evidence that acid rain is a problem within its area. 3.6.4 Results of survey This survey—based on a small number of published papers and the response to a questionnaire which was sent to all the Water Authorities—shows that, in the U.K., seriously acid surface waters are restricted to a small number of catchments situated in the west of Scotland, the English Lake District and west Wales. In these catchments the underlying geology comprises granite intrusions and Ordovician and Silurian Shales. Outside these areas some acid streams and reservoirs have been noted, for example, on sandstone and millstone grit. However, it should be noted that some of the sensitive catchments have not been monitored by Water Authorities because they are in what have previously been regarded as unpolluted areas. Except for Galloway and the English Lake District there are records for few sites in the U.K. that could be used to assess whether acidification has taken or is taking place. In Galloway, pH records, fisheries

FRESHWATER

63

records and the diatom record in lake sediments all show that certain lakes on granitic upland catchments have become more acid. In the Lake District, on the other hand, the available evidence (mainly chemical records) shows no evidence of significant change in upland tarns over the last 30 years. The apparent exception to this view provided by the Freshwater Biological Association’s chemical records of the last 50 years for lakes in the English Lake District is perfectly understandable when considered on a longer timescale. The likely onset of acidification was probably around 100–130 years ago. The Lake District, being closer to industrial sources including local iron works, etc., would have been one of the first areas to suffer. Evidence from existing distributions of the limpet Ancylus fluviatilis in the River Duddon catchment supports the hypothesis that acidification has taken place, though of course this does not help with an assessment of the timescale. When seen in the correct context, therefore, the Lake District evidence is in accord with other findings and reinforces the view that for several of the more sensitive lakes a small increase in acid input could produce rapid pH changes with the accompanying serious biological damage. 3.7 Conclusions (1) Many individual studies that provide the evidence for acidification of surface waters, loss of fish stocks and other biological effects can be criticised on a variety of grounds, such as poor methodology (especially in earlier years), unreliability of early records, inconsistencies and contradictions. Nevertheless, it is our view that, when all this is taken into account, the weight of the evidence supports the view that acidification of certain surface waters in Europe and North America has been accelerated since the industrial revolution owing to a variety of causes of which acid precipitation from man-made sources is important. (2) Surface waters in areas of high annual sulphate deposition whose calcium content is below 100 µ eq l −1 (2 mg l−1) are likely to be more acidic and contain elevated levels of aluminium. In such areas the degree of acidification is likely to be greater in surface waters draining areas that have been planted with coniferous forests. (3) In the United Kingdom the most vulnerable areas appear to be the upland areas of west Scotland and midWales where a combination of slow-weathering bedrock and base-poor soils affords little neutralising capacity. (4) The limited survey work carried out to date in the U.K. suggests that acidified waters are either fishless or support only sparse salmonid fish populations and limited invertebrate diversity. Acid ‘pulses’ with associated high levels of aluminium are likely to have been responsible for recent mortalities of freshrun sea trout in the West Cumbria Esk and Duddon catchments. These mortalities are seen within the context of a serious decline in sea trout runs in the last decade. (5) The lack of historic water-quality data with which to compare current data makes trend analysis very difficult for most of the U.K. From the available data, it appears that some lochs in southwest Scotland have become more acid in the past 30 years, but this trend has not been observed in the English Lake District. Diatom analysis of sediments has confirmed that several lochs in southwest Scotland have become more acid since 1850 and that the fall in pH level has accelerated in the last 50 years.

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EFFECTS ON FLORA AND FAUNA

3.8 Recommendations (1) Nationally co-ordinated monitoring programmes of upland surface waters should be instigated as a matter of urgency in order that the extent and degree of acidification can be defined and a baseline established against which future trends can be measured. (2) Surface-water monitoring programmes should be correlated where possible with rainfall data from the proposed national monitoring network. If necessary, additional rainfall sampling sites should be selected and monitored in accordance with the proposed protocol for measurements of precipitation composition (United Kingdom Review Group, 1983). (3) Research into the effects of land use on water quality and ecology in vulnerable areas should be supported with a view to developing management techniques capable of ameliorating present levels of acid deposition. (4) Predictive techniques capable of quantifying the effects of different levels of acid deposition on surfacewater quality in vulnerable areas should be established. (5) Further investigations are required into the historic timescale of acidification at selected sites using sedimentcore analysis. (6) There is an urgent need to monitor rainfall and streamflow in selected catchments with continuous recording of pH level and temperature in conjunction with automatic sampling to measure changes in water chemistry during episodic events. This would need to be correlated with detailed biological surveys in order to assess the impact of such events on aquatic ecosystems. Select References Flower, R.J. and Battarbee, R.W. Diatom evidence for recent acidification of two Scottish lochs. Nature 305, 130–33, 1983. Howells, D.G., Brown, D.J.A. and Sadler, K. Effects of acidity, calcium and aluminium on fish survival and productivity —A review. J. Sci. Food Agriculture 34, SS9, 1983. Stoner, J.H., Gee, A.S. and Wade, K.R. The effects of acidification on the ecology of streams in the Upper Tywi catchment in West Wales. Environmental Pollution Ser. A,35, 125–57, 1984. Sutcliffe, D.W.Acid precipitation and its effects on aquatic systems in the English Lake District (Cumbria). 51st Annual Report, pp. 30–62, Freshwater Biological Association. 1983. United Kingdom Review Group on Acid Rain: Report: “Acid Deposition in the United Kingdom” Warren Spring Laboratory, December 1983. Wright, R.F.Predicting acidification of North American lakes. Report 4, Norwegian Institute for Water Research, Blindern, Oslo, 1983.

THE WATT COMMITTEE ON ENERGY REPORT NUMBER 14

Section 4 Remedial Strategies C.J.Davies

This paper presents the work of Sub-group 4 (Remedial Strategies) of the Watt Committee working group on Acid Rain. Membership of Sub-group 4

C.J.Davies (Chairman) A.J.Clarke N.H.Highton G.S.Parkinson Dr M.A.Plint

Remedial strategies

Discussion of a remedial strategy to alleviate effects attributed to acid rain requires a transition from the purely scientific arena of the earlier chapters into an area where technical, economic, legal and political factors must also be considered, and this is done here in four sub-sections. Sub-section 4.1 addresses the question of how acid emissions could be reduced: it presents a survey of the estimated performance and costs (where available) of alternative means of emission reduction and of the scale of the changes required in the United Kingdom in the event of reductions of 30–50% compared with SO2 levels of 1980. The paper draws heavily on recent figures from the energy industries concerned. Sub-section 4.2 reports recent work at the University of York into how the increased generation costs of a large programme of emission reduction would feed through the electricity supply system and be reflected in increased prices to the consumer. Sub-section 4.3 provides a summary of the current legal framework governing emissions in the U.K. It includes discussion of likely international and European developments as they effect the U.K. The final sub-section addresses the crucially important question of how we can decide whether a major programme of emission reduction is worthwhile. Even a definition of ‘worthwhile’ is difficult, but most generally it concerns the balance between the expected benefits to society from reduced emissions and the benefits foregone in meeting the costs of the programme. It needs to take account of very many factors. No attempt is made here to propose an optimum policy; discussion is limited to possible methods of approach and identification of key uncertainties. 4.1 Technologies for Reducing Emissions 4.1.1 Components of a strategy Much of the discussion on the possible reduction of levels of ‘acid rain’ often focuses immediately on emission reduction by emission-reducing technologies at the power station. In practice a combination of several approaches could be more effective in dealing with acid rain’s effects and could be much cheaper. Figure 4.1 illustrates the range of approaches available. In some areas it may be possible to act directly upon the damage. The practice of liming lakes and rivers is very well established. The Organisation for Economic Cooperation and Development has estimated that the liming of affected lakes in Norway and Sweden would cost $60000000-$100000000 per year. This is just a few per cent of the extra capital investment (excluding running costs) that the European Commission estimates could be required to fit flue-gas desulphurisation just to new Community large coalfired plants

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67

between now and 2000. Similarly, in agriculture where soil acidity from natural causes is a problem, liming is a long-established practice. There is at present little evidence of damage to crops from wet or dry deposition of sulphur; however, should man-made acidity prove in some circumstances to be serious, liming could be extended. Individual approaches have to be developed for each damage area, and ultimately there may be limits to what can be achieved. However, there is a strong case, on economic grounds, for more thorough investigation of appropriate techniques—which may, anyway, be required as a complement to emission reduction, to bring damaged areas back into production. Reduction in emissions of sulphur and nitrogen oxides can be achieved through reductions in energy use, substitution of energy sources, or application of emission-reducing technologies. A possible complementary approach would be to consider limiting emission of other substances which have been implicated in the chemistry of acid rain production. There is a great deal of scientific uncertainty about the role, for example, of hydrocarbons and ozone, and they may have an important influence on the effectiveness of any control actions adopted. Through ‘conservation’ and substitution the U.K. has made very significant reductions in sulphur dioxide emissions over the last decade. There has been a slowing down of the trend of increasing energy consumption. The very modest growth in gross domestic product since the first oil crisis has been accompanied by a decrease of more than 10% in the energy use per unit of goods and services produced. Purposeful capital investment, for example in combined heat and power, offers scope for further significant gains in the efficiency of energy use. Low-sulphur energy substitution has come about through an increased use of gas, a switch to low-sulphur North Sea petroleum feedstocks and an increasing contribution by nuclear power. These switches have come about because of general economic forces rather than a deliberate environmental policy, and economic considerations are likely to be the driving force behind energy changes into the future. In principle, sulphur oxide emissions could be further reduced by a switch to low-sulphur coals, but two considerations are relevant. First, from U.K. coal production there is a limited and declining availability of low-sulphur coals. Second, as pollution control measures become more widespread, there will be a premium on low-sulphur coals in the world market. In the long term this premium could equate to the costs of using sulphur-retaining technologies. Whatever other measures are adopted, it is likely that, to achieve a reduction in emissions of the order that is being discussed internationally, an important role will be required of emission-reducing technologies: sulphur removal prior to combustion, during combustion or after combustion. In comparing the relative merits of alternative technologies, it is appropriate to look first at power generation. In 1982, power generation accounted for 65% of estimated U.K. sulphur dioxide emissions. More than 85% of these (over 50% of total national emissions) came from coal-based power generation. Most of this resulted from power generation at just 20 main sites. By contrast, the much lower emissions from industry originate from thousands of sites, and domestic emissions from millions of sites. The high loads and high load factors associated with power generation should lead towards a lower cost per kilogram of SO2 removed, compared with other energy users. The scale of the requirement for these technologies depends upon future CEGB emissions. In turn, forecasts of these emissions depend heavily on assumptions about economic growth (and consequent demand for electricity) and the contribution of nuclear power. Figure 4.2 presents figures for one central economic scenario (Case C of the CEGB Sizewell B submissions), and assumes that no special action is taken to reduce emissions. Between now and 1990 no new power stations, nuclear or fossil, not already under construction will come on stream and emissions are expected to be similar to 1980 levels. Looking to 2000, emissions would be reduced if there were a large growth in nuclear power or could be increased with

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TECHNOLOGIES FOR REDUCING EMISSIONS

Figure 4.1 Range of approaches to reduction of acid rain levels.

low nuclear growth and a consequent requirement for new fossil stations (up to six new 2000-MW fossilfuel stations are required in this scenario). However, on this longer timescale it is probably unrealistic to suppose that a new fossil station would be built that did not incorporate emission-reducing technologies. Future emissions would therefore be similar, whatever the balance between nuclear power and coal for future power stations. Assuming, then, that new stations would be ‘clean’ but that no action would be taken for existing stations, with the electricity demand envisaged in this economic scenario, emissions of SO2 from power generation would remain roughly at 1980 levels until 1990 and fall between 1990 and 2000. To summarise, any reductions of emissions required of the power generating sector before 1990–95 will have to come from coal selection and cleaning and from retrofitting of emission-reducing technologies. Beyond this date, emissions will be reduced through the introduction of ‘clean’ (either fossil or nuclear) power stations. 4.1.2 Process options in power generation Various technologies, both under development and established, could be considered for use in new power stations to reduce emissions of nitrogen and sulphur oxides. Some of these are available for retrofitting to existing stations, although this would always be more difficult and expensive than incorporation in a new station. Flue-gas desulphurisation Sulphur oxides are formed by the oxidation of organic and inorganic compounds in fossil fuels, and with conventional combustion technology most of the original sulphur present is released into the flue gases. Various technologies are available for flue-gas desulphurisation (fgd) —the removal of sulphur oxides from flue gases. Most existing plants worldwide use a wet lime or limestone system. These systems produce a thixotropic calcium sulphite sludge in large quantities, which presents considerable disposal problems; these processes are judged unsuitable to U.K. conditions. Instead, there is an interest in regenerative processes. The Wellman-Lord process, producing concentrated sulphuric acid or elemental sulphur as a by-product, is an example of these newer processes. Table 4.1 shows the estimated cost of this system, based on costed

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69

Figure 4.2 Projected CEGB emissions of SO2 with no emission-reducing technologies, MT per year. Source: CEGB.

design and engineering studies by the CEGB. The figures assume coal consumption of 4 Mt per year of 1. 9%-sulphur coal. A by-product value is assumed. The costs of the system have been estimated through simulation of the operational effects on the whole CEGB system. These figures suggest a cost for a new station of £190 per tonne of SO2 removed. The estimates of capital cost are well in line with the range, based largely on wetlime technology, that was reported at a recent OECD conference. Costing exercises carried out by the CEGB suggest that the cost differences amongst sites are more important than differences between processes. Fgd can be retrofitted, provided there is space on the power-station site. Capital costs would be increased. Low-NOx combustion Nitrogen oxide produced in a coal-fired power station comes partly from oxidation of organic nitrogen contained in the coal (fuel NOx) and partly from oxidation of atmospheric nitrogen at the high temperatures of combustion (thermal NOx). Fuel NOx is around 60–70% of the total NOx produced; however, less than a quarter of the nitrogen in the coal is oxidised, and the NOx produced in a given boiler has only a weak dependence on the nitrogen content of the coal (this clearly contrasts with the situation regarding sulphur oxides). Nitrogen-oxide formation is increased with increasing temperatures of combustion and oxygen-rich combustion environments. Low NO, burners provide for initial fuel-rich combustion conditions with a secondary oxidation stage. Various degrees of complexity of arrangement can be used, and reductions of NO, of up to 50% should be achieved by use of existing designs. In a new station the additional costs associated with a modest reduction in NOx should be small compared with fgd, but may result in some losses of efficiency owing to less complete combustion. However, larger flame envelopes are involved (essentially the fuel needs more space to burn without impinging upon and severely corroding the walls of the combusion chamber), and it has yet to be demonstrated that the technology is capable of being retrofitted to the range of U.K. power stations. Pressurised fluidised-bed combustion Operating a fluidised bed under pressure which is released through a turbine—pressurised fluidised-bed combustion (pfbc)—offers the possibility of high overall efficiency at a competitive capital cost. Sulphur

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oxides can be absorbed by incorporation of dolomite or limestone in the bed, and nitrogen oxide emissions are also lower than in conventional pulverised fuel boilers. The technology is being assessed in the U.K. by the NCB, CEGB and Department of Energy. An 80-MWth pressurised fluidisedbed combustion facility, owned by the U.K., U.S.A. and F.R.G. through the International Energy Agency, is nearing the end of its current experimental programme. Continuation of the programme is being discussed. Gasification combined-cycle technology Gasification combined-cycle technology—an advanced combustion technology involving gasification of coal—can be regarded as an alternative to pfbc, and it is unclear at the moment which of the systems will offer the balance of advantage; neither system could be retrofitted to an existing power station. The system is being assessed in the U.K. by the NCB, British Gas, CEGB and the Department of Energy. Catalytic reduction of NOx For the catalytic reduction of NO., ammonia is injected into the flue gas which, over a catalyst, produces nitrogen and water. Removal of 80% of the NO, can be achieved. The system is being used in Japan only. It is expensive, requires very large buildings and uses very large quantities of ammonia. Given the levels of reduction achievable through low-NOx combustion, a requirement for this technology is unlikely. Where space is available, it could be retrofitted. Direct limestone injection Limestone injected into flue gases can directly absorb some sulphur oxides. However, absorption efficiency is low, and very large quantities of limestone would be required. The CEGB assessment is that this system would be less atrractive than fgd. Coal washing With modern coal-mining methods, coal as mined contains a high proportion of ‘dirt’, of very low calorific value, which may contain sulphur in similar concentrations Table 4.1 Costs of flue-gas desulphurisation at 2-GW power station

Process Wellman-Lord1

SO2 recovery, Annual abatement,

Capital, Net operating costs,

Increase in cost of system,

Cost per tonne of SO2 abated,

%

£×106

£×106 per year

£×106 per year

£

117

4.4

26

190

152

1.5

31

228

−

−

t

90 136000 New 3×660 Wellman-Lord1 90 136000 Retrofit 4×500 Wet-lime2 − − 1Source: CEGB, 1983. 2Source: OECD, 1982.

96–276 −

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71

Figure 4.3 Estimated SO2 emissions in Europe and the USSR, 1982. Total emission in region, 61.44 t×106 per year; total emission in Western Europe, 20.75 t×106 per year; total emission in EEC states, 16.10 t×106 per year. From Highton and Chadwick, expert meeting II, 1982 Stockholm conference on the environment. National Swedish Environmental Protection Board, snv pm 1637, 1983. Reproduced by permission of the Central Electricity Generating Board, London.

to that in the coal itself. About two-thirds of the coals sent to the CEGB by the NCB are ‘washed’ to separate out most of the dirt. The objective is to increase the calorific value of the product, but some sulphur is incidentally removed. The NCB estimates that a further 56000 t of sulphur annually (about 4½% of CEGB emissions) would be removed by extending currently used coal-washing methods to all the deepmined coal sent to the CEGB. However, capital expenditure of £400000000 would be required, and the overall economics are very unfavourable compared with those of the removal of equivalent sulphur using fgd. The potential of washing processes not yet commercially proven is now being considered. The major existing coal-burning power stations will still be operating at high load factors through the 1990s. Large reductions in emissions over this timescale would require action at these stations. An optimal strategy for sulphur oxide removal could involve some retrofitting of fgd, concentration of lower-sulphur coals to non-fgd stations and some change in coal-cleaning practice. Studies are currently under way between the CEGB and NCB to identify suitable policies.

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Figure 4.4 Cause-effect relationships.

4.1.3 Industrial energy use Outside power generation, the industrial use of energy accounts for less than 20% of SO2 emissions. In broad terms, coal accounts for one-third of these and petroleum fuels for two-thirds. Fgd may be applicable at this level but probably at an unacceptable cost. Fluidised-bed combustion could retain sulphur in fuel oil or coal, and this technology is competitive with conventional combustion. However, present British shallowbed technology is not particularly suited for sulphur retention which would add to costs and would provide a waste-disposal problem. Its penetration, in any event, would be limited to the rate at which the industrial boiler stock was renewed. Desulphurisation of fuel oil could, in principle, contribute to emission reduction by both power generation and general industry. However, it is an expensive technology in terms of both capital and energy. Costs vary greatly between crude oils. For a Kuwait crude* atmospheric residual fuel oil (4.3% S), the cost of reduction to 1 % sulphur has been estimated at about £26 per tonne, equivalent to £400 per tonne of SO2 emission abated. For the vacuum residual fuel, the cost increases to £36 per tonne (taking into account a 16–19% energy loss), equivalent to £550 per tonne of SO2 abated. Given a declining market for heavy fuel oils (reflected in SO2 emissions reduced by more than 60% over the last ten years), the scope for this is now limited. Thus in the changed pattern of refining the majority of these residues no longer go to fuel oil, but are feedstock to conversion plants for conversion to lighter products such as gasoline and distillates. In this conversion a considerable removal of sulphur occurs—a step termed in the 1982 UNECE technical report as ‘process integrated desulphurisation’ as opposed to what they term ‘desulphurisation-onpurpose’. Looking to a national strategy, reduction of emissions from power generation is likely to provide the cheapest and most convenient way of achieving a prescribed level of overall reduction. There may be some scope for reducing industrial emissions by providing industry with any available low-sulphur fuel oils and coals, but this could only be effective nationally if fgd were used in power generation. However, while possibly providing the most economic policy overall, it may be judged inequitable for the electricity consumer to carry the full costs of a national abatement strategy.

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4.1.4 Environmental impacts Technologies to reduce acid emissions can themselves have environmental impacts. These can arise directly because of material requirements or waste products of the processes, or indirectly because the processes reduce thermal efficiencies. Nitrogen oxide abatement through low-NOx combustion seems to be the only system that would have minimal adverse environmental impact. Material input requirements vary greatly amongst the different kinds of processes. Several processes use limestone to capture SO2. If the limestone (calcium carbonate, CaCO3) were used with complete efficiency, each tonne of SO2 captured would require the use of something over 1.5 t of calcium carbonate. Existing technologies do not achieve this rate. A wet limestone fgd system has a requirement about 1.6 times this consumption of calcium carbonate. Early indications are that pressurised fluidised beds would require more limestone than fgd. To achieve 90% SO2 removal, direct limestone injection would require 6 times the stoichiometric rate, although this method would be impracticable. If these requirements are translated to a power station with an annual burn of 4000000 t of 2%-sulphur coal, about 400000 t of calcium carbonate would be required annually for a wet limestone system removing 90% of SO2. Direct injection, to achieve the same removal, would require about 1 500000 t per year. Not only would the quarrying and transport of these quantities of material have a significant environmental impact, but the waste-disposal problem they would represent could be intractable in the U.K. For this reason, the CEGB favours regenerative systems, where the net material input would be much smaller, or a system that would produce gypsum. Although this last-mentioned system would have a high material input, the product should be saleable and would displace an existing requirement for building material; or if not sold it could be dumped safely. Increased coal-washing and fgd result in an additional energy loss. This is highly variable for coalwashing, and may exceed 10 t of coal lost for each tonne of sulphur removed. The comparable figure for fgd is about two tonnes of coal for one tonne of sulphur. Coal-mining activity to make up for the coal lost would have an environmental impact—most directly, a solid-waste disposal problem. Many approaches to the removal of sulphur oxides result in an increase in CO2 emissions, although there is little reason to regard this as an environmental problem. Limestone processes which exchange CO2 for SO2 emit 44 t of CO2 for every 64 t of SO2 abated. It was suggested above that 2 t of extra coal would be required for each tonne of sulphur removal by fgd; this gives rise to a further 88 t of CO2 for each 64 t of SO2 abated. (This reasoning does not apply to coal-washing, where the carbon is lost before combustion.) If 50% of the current SO2 emissions by the CEGB (2600000 t per year) were removed by limebased fgd, the additional total of CO2 released would be equivalent to burning a little over 1 500000 t of coal per year; more than four-fifths of this would be from the actual burning of extra coal to make up for energy losses. 4.2 Impact of SO2 control on electricity prices This sub-section of the Report examines how a major programme of SO2 control for the electricity supply industry would affect prices that consumers pay for electricity.

*The figure of 2.5% is the sulphur content of ‘whole Kuwait crude’, the figure of 4.3% S being the sulphur content of the Kuwait atmospheric residue.

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4.2.1 Components of electricity tariffs The use of flue gas desulphurisation (fgd) would raise the cost of constructing and running an individual power station by 10–20%; but of the costs incurred by the industry that make up electricity prices, many would be unaffected by SO2 control. This can be illustrated by examining the range of cost components that comprise electricity tariffs. Table 4.2 shows these components and the direct effect of SO2 emission reductions based on fgd. In Table 4.2, peak and standard units relate to the cost of operating power stations to which fgd would not be fitted. Power stations supplying so-called shoulder units would be fitted with fgd only in an extreme combination of circumstances. Night units refer to the cost of operating large base-load power stations to which fgd would need to be fitted. This cost component would be increased by the running cost of fgd systems. Similarly basic capacity costs, which refer to the construction costs of large power stations, would be increased in accordance with the construction cost of fgd installations. Other costs, such as the fuel cost adjustments, service charges and all Area Board expenses, would remain unaffected. The weight attached to each of these cost components varies very considerably with each of the several electricity tariffs appropriate for different categories of consumer. 4.2.2 Estimated increase in tariffs Research at the University of York shows that for electricity purchased under the Domestic Unrestricted Tariff, which is the main tariff for households, an approximate halving of SO2 emissions from the electricity supply industry achieved over a period of about 20 years would raise prices by less than 4%. For large industrial consumers the effect would be an increase of slightly over 5% (Table 4.3). A reduction in emissions is meant here as a reduction compared with those emissions that would have occurred in the absence of control, as opposed to a reduction compared with 1980 emissions or any other such benchmark. The reason for this approach is that it allows control costs to be identified without the need to refer to a specific set of energy projections. The inherent difficulties in making such projections are underlined by the fact that in all countries forecasts of economic growth and energy use are routinely revised. The proportionate reduction from unconstrained emissions can be shown to be insensitive to different assumptions about the growth of electricity demand and the contribution of nuclear power. Table 4.2 Components of electricity tariffs Affected by fgd Yes Cost component* Peak units Standard units Shoulder units Night units Fuel cost adjustment Peak capacity

No

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Affected by fgd Yes

No

Basic capacity Service charge Area Board costs: 11–132 kV reinforcement Local authority rates and Electricity Council expenses Interest on working capital *See text.

In terms of absolute quantities, however, the tariff increases would indicate the effect of reducing projected CEGB emissions of SO2 from about 2 500000 t in the year 2000 to about 1250000 t. If the U.K. were to reduce emissions from all sources to one-third below the 1980 rate of 4700000 t, a reduction in CEGB emissions of this order could be required, depending on assumptions about emissions from other industrial sources. Such a programme for the electricity supply industry would require the inclusion of fgd on both new and existing large fossilfuelled power stations over a period of two decades. The estimated price increases do not include an allowance to make up for losses in generating capacity associated with implementing fgd. However, the cost associated with this would depend upon the timing and exact means used to replace the lost capacity. Nor has allowance been made for any continuation of technical developments that could reduce the cost of SO2 control. Table 4.3 Halving CEGB emissions* of SO2 in 2000: effect on electricity prices Increase, % Domestic electricity prices† Industrial electricity prices‡ * Central Electricity Generating Board of England and Wales. †Refers to domestic unrestricted tariff. ‡Refers to tariff for large industrial consumers.

3.8 5.2

4.2.3 Sensitivity to changing assumptions The estimated price increases to the consumer depend critically on assumptions about the construction and running costs of so-called ‘regenerable’ fgd installations of the type that would be likely to be used in the U.K. The analysis assumed that construction costs would be in the region of £130000000 for a new coalfired power station and would be 20% greater, on average, in cases where retrofitting was required. This is in line with recent CEGB figures, referred to in Section 4.1 above. Overall running costs depend on how power stations fitted with these control devices would be operated in the electricity system as a whole, and £25000000 per year is representative. Estimates of construction costs for fgd, both for British conditions and from countries with more experience in this field, have shown wide variation; in extreme cases, from half to over double these costs. Site-specific factors may be most important here, but if these extremes applied across the board, the tariff

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increases shown in Table 4.3 would become, for the most costly probable case, 8.5%, and for the least costly case 1.5%. The argument that has been sketched here depends on one further assumption. The effect of an addition of 4–5% to electricity tariffs might only apply if the electricity supply industry were to adhere to standard marginal-cost pricing principles so that the effect on tariffs would reflect the impact of fgd on the marginal cost of electricity suppy. A possible alternative would be to spread the cost of construction and operation of fgd incurred in each year over all tariffs to cover these extra costs exactly. Recalculation on this basis depends on the year looked at and, particularly, on the number of fgd installations that would need to be under construction in that year. However, in general the effect of halving emissions would still be to raise tariffs by close to 4%. 4.2.4 Comparison with past increases It is illuminating to compare these projected increases with the domestic and industrial changes in the real (or inflation-adjusted) prices of electricity that have occurred over the last decade in response to rapidly increasing world energy prices (Table 4.4). Domestic and industrial electricity prices increased in Table 4.4 British electricity price index (1980=100), adjusted for inflation Year

Domestic

Industrial

1974 1979 1980 1981 1982

77 92 100 108 110 Average rate of change, +4.6

70 84 100 109 112 % per year +6.1

1974–82

real terms by 43% and 60%, respectively, in the period 1974–82. This puts the average annual rate of increase at 4.6% and 6%. The effect of a substantial programme of SO2 control based on the accumulation of fgd installations on both new and existing power stations would be a once-and-for-all increase in electricity prices. The increase that would be faced by domestic and industrial consumers would be of a similar order to that of the annual price increases that have occurred over the last decade. 4.3 Existing and prospective legislation governing British emissions 4.3.1 United Kingdom legislation The release of emissions from British industrial plants has been controlled by Act of Parliament since 1906. The Alkali etc. Works Regulation Act (as incorporated into the Health and Safety at Work Act 1974)

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requires the use of ‘…the best practicable means for preventing the escape of noxious or offensive gases by the exit flue of any apparatus used in any process carried out in the work and for preventing the discharge, whether directly or indirectly, of such gases into the atmosphere and for rendering such gases, where discharged, harmless and inoffensive…’. Sulphur and nitrogen oxides are included in the list of ‘noxious and offensive gases’ over which the Inspectorate has jurisdiction. The Act applies to larger plants and probably covers over 80% of emissions of SO2 and more than half of NO. emissions in the U.K. It applies to all CEGB generating stations. It includes no explicit reference to cost or to need, and in principle could imply very exacting standards. In practice its application has been flexible, and cost-benefit ratios and the financial impact of a control measure are taken into account in deciding what is ‘practicable’. Because of the emphasis of the Act on individual plants, it is not a suitable vehicle for implementing a comprehensive national strategy. Under this Act, the Industrial Air Pollution Inspectorate (formerly, the Alkali Inspectorate) could certainly require a future power station to be fitted with fgd. Apart from this possible requirement, the CEGB has, since its formation, been subject to the condition (in its statutory consent under the Electric Lighting Act 1909) that space must be left for the later provision of sulphur-removal equipment, should the Secretary of State require this to be fitted. In practice, it was difficult to define the space requirement, and there might be severe difficulty now in accommodating modern fgd designs at many existing stations. In any event, in over 30 years this condition has never been invoked. Other British legislative measures include, in the Control of Pollution Act, powers to limit the sulphur content of fuel and heating oils. This power is likely to be useful for the limitation of emissions from small plants for which other control technologies are inappropriate. In general, it is designed to be used to improve air quality in particular areas, like central city areas, rather than as part of a national attempt to reduce overall emissions. The Clean Air Act also can be used to ensure satisfactory air quality in particular localities at risk. 4.3.2 European Community legislation An EEC Directive of 1980 relates to ambient air quality. Air-quality limit and guide values for SO2 and suspended particulates are already in force. NOx limit values are being prepared but are not yet agreed. Two major new directives are under discussion. The general objective of the European Commission is towards a stabilisation of emissions of a number of frequently occurring pollutants, including oxides of sulphur and nitrogen, followed by a progressive reduction. Emission standards for certain types of source are seen as one of the means of achieving this. The Commission is undoubtedly motivated by a concern about possible harmful effects of emissions. However, harmonisation of procedures is, of itself, seen as desirable. Differing levels of protection against air pollution are seen as creating unequal conditions of competition for the enterprises concerned, thus having a direct effect on the functioning of the common market. The first of the draft directives referred to, COM(83)173, seeks to establish harmonised procedures for plant approvals under which air-pollution controls should be determined. The directive would require authorising agencies to follow developments as regards the best available technology, but the conditions required for authorisation must take into account what is economically feasible for the plant in question. This draft is unlikely to present any difficulties for the U.K. It appears not to conflict with the existing procedures for plant authorisations, and the criteria, through the ‘economic feasibility’ reference, could be interpreted similarly to ‘best practicable means’. However, the draft also seeks to establish the principle that

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the Council of Ministers (acting on proposals from the Commission) can set emission-limit values for a specific plant. This principle of setting emission-limit values on a Community-wide basis is contested by several member states. (In March 1984 this directive was agreed. A clause incorporating consideration of the nature, quantity and harmfulness of the emissions has been included.) Without waiting for determination of this issue, the Commission has been preparing a further draft directive, setting emission-limit values of sulphur and nitrogen oxides and particulates for a large range of combustion plant. The most recent draft of this directive, COM(83)704, was issued only in December 1983; there has been no opportunity to include any commentary in this paper. The draft proposes both plantemission limits and overall national emission limits. For nitrogen oxides and sulphur dioxides the plantemission limits are only for new plant authorisations. They are more strict for plant above 300 MW; plant authorised after 1995 will have to meet more exacting standards than plant authorised between 1985 and 1995. These plant-emission standards are supplemented by proposed reductions in overall national emissions to be achieved by 1995. Proposed reductions compared with 1980 annual mean levels are: 60% for SO2; 40% for NOx; and 40% for ‘dust’. Thus, although the apparent main objective of the draft is towards introducing emission limits for new plants, to achieve levels of reduction of this order would certainly require a significant amount of retrofitting of control technologies to existing plant. Ironically—in view of current U.K. strategy—it is proposed that stack height should not exceed 200 m, except in special topographical conditions: it was pointed out in Section 1 of the Report that stack height has only a small effect on emissions at a distance, though it has been very important in controlling urban emissions. Reasons given by the Commission for proceeding with this directive arise from the increasing international concern with acid rain. 4.3.3 Convention on long-range transboundary air pollution: United Nations Economic Commission for Europe In 1979, under the auspices of the United Nations Economic Commission for Europe (UNECE), a Convention was drawn up under which signatories, including the U.K., undertook to ‘limit and as far as possible gradually reduce and prevent air pollution, including long-range air pollution’. The Convention includes signatories from the centrally-planned European economies. A Nordic proposal for a 30% reduction of emissions by 1993–95 compared with 1980 has received some support. Although not accepted it has led to general agreement on the need to decrease effectively the total annual emissions of sulphur compounds by the mid-1990s. It should be noted that the selection of 1980 as the base year both in these proposals and in the EEC proposals, referred to above, gives no credit to the emission reduction achieved by the U.K. in the period 1970–80. 4.3.4 Legislation elsewhere Prospective European legislation discussed above is based, to a large degree, on existing legislation in one member state or another. Several countries throughout the world have set national ambient quality standards for SO2 based on considerations of health protection, though the actual figures vary widely. Many countries have adopted emission control (usually at major sources alone, such as power stations and smelters) as the sole regulatory means of achieving air-quality standards. Limits on sulphur content of fossil fuels have also been applied either in place of or in addition to control of SO2 emission rates.

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Ambient standards were adopted in the U.S.A. in 1970, based principally on criteria reflecting effects on human health. Plant emission standards originally introduced to improve air quality have been greatly tightened. In addition, there is a mandatory requirement to incorporate fgd on all stations, including those burning lowsulphur fuels and those that could meet emission requirements. The reason for this is mainly political; it prevented a recession in the eastern U.S. high-sulphur coalmining states, which could have been caused by a substitution of low-sulphur coals from the western U.S.A. German and Dutch legislation provides a combination of national ambient air-quality standards, limits on fuel-sulphur content and limits on SO2 emission rates. In these countries, approval for new plants hinges on gaining acceptance from local authorities and the public in the vicinity of the plant site. These local demands are frequently for controls that are effective at well below the standards set nationally, and operators of plant are forced to comply with controls at the very limits of technical achievability. A similar local approval procedure operates in Japan, where 90% sulphur removal by fgd is regarded as inevitable for new plants. 4.4 Scientific support for an acid rain strategy Acid rain has become a focus of public concern, nationally and internationally; the formulation of an appropriate response is commanding high-level government attention. The response will need to take account of scientific, technical and economic considerations, as well as international political pressures. This sub-section considers how the present state of knowledge, as expressed in this and the preceding sections, can contribute to policy. Internationally-favoured policies for alleviating acid rain lean towards a programme of reductions in national emissions of sulphur and nitrogen oxides, sometimes supplemented by emission standards for large plant. Inherent in these policies are the following assumptions. (1) A reduction in national emissions would be likely to lead to reduction of acid deposition in areas where damage can occur; (2) A reduction in acid deposition in these areas would be likely to reduce actual or potential damage; and (3) The resulting avoidance of damage would be sufficiently worthwhile to justify the cost of the control strategy. Observations will be made on each of these assumptions. 4.4.1 The relationship between emissions and deposition In discussing transmission of pollution, it is useful to distinguish between long-range effects and purely local phenomena—for example, high concentrations of pollutants in city centres, or around industrial estates, or due to high levels of road traffic. In these local cases damage is attributed mainly to primary gaseous pollutants. Local remedies are certain to be most effective and cheapest, and it is unhelpful to consider them as part of ‘the acid rain problem’ and as appropriate objects for international legislation. Within the U.K., almost without exception, the problems of urban pollution have been conquered, and the expensive controls that are being discussed internationally will lead to only minimal further benefits. Looking to longer-range transmissions (which may be designated ‘secondary acid rain’ and have been discussed in more detail in Section 1 of this Report), it can be said that the relationship between changes in emissions at one point and deposition at another depends, among other things, on the distance between the

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points, the prevailing wind patterns and windspeed, humidity, temperature and sunlight. The presence of other substances which may act as catalysts, or be part of the chemical transformation (for example, ozone or hydrocarbons) may also be very important. Although most measurement has been related to sulphur dioxide emissions, combination effects of sulphur and nitrogen oxides need also to be considered. For example, a reduction of NOx may not only be beneficial directly, but may also substantially reduce the transformation of sulphur dioxides to sulphate and so indirectly reduce the long-range effects of sulphur emissions. Further research is required to allow those combination effects to be properly evaluated. Models are now available for the tabulation of imports and exports of SO2 for European nations. A tabulation for 1979, made by the UNECE EMEP programme for monitoring and evaluation of long-range transmission of air pollutants in Europe, explains why the U.K. has been subject to international pressures; it is a significant net exporter of SO2. Whereas more than three-quarters of deposition on to the U.K. is of domestic origin (and about half the remainder is background SO2 of indeterminate origin), the U.K. is, for example, the largest single source of deposition on Norway. It accounted in the 1978–79 budget tabulation for significantly greater deposition than that due to domestic Norwegian sources. However, it is interesting that the same tabulation estimates indeterminate background deposition on Norway at twice the U.K. contribution. It is relevant in policy considerations to consider how efficient control of U.K. emissions is if, for example, the primary objective is to reduce deposition on Scandinavia. Table 4.5 shows how great a reduction in emissions would be required to achieve a one-tonne reduction in deposition of sulphur in Sweden and Norway, respectively. Provided that opportunities remain for sulphur reduction in Scandinavia, Table 4.5 shows that they are likely to represent the cheapest and most direct means of reducing depositions in Scandinavia. These opportunities should continue to be pursued, even where they appear several times more expensive per tonne of sulphur removed than opportunities to reduce U.K. emissions. Table 4.5 Reduction in emission required to reduce sulphur deposition by 1 tonne Source of emission

Country of deposition

Sweden

Norway

Sweden Norway United Kingdom Source: EMEP

2 tonnes 6 tonnes 53 tonnes

20 tonnes 2 tonnes 46 tonnes

As concern broadens beyond particular issues and particular areas, the balance of advantage between local and international action may change. Because of the loose linkage that often exists between one country’s emissions and another’s deposition, any international collaboration should be on the largest possible scale. This is particularly true, given that it is still not known with any certainty which are the areas in Europe most at risk from acidification. International action, to be effective, must include the centrallyplanned economies of Europe. Figure 4.3 (page 45) illustrates this. Eastern Europe is responsible for twothirds of the European total, and the EEC States for about one-quarter (about the same as Eastern Europe excluding the Soviet Union). Action by any one country, or any limited group of countries, will have only a partial effect on total emissions. The most practical solution will lie in defining the scale of action that the greatest number of nations will be able to accept, rather than in piecemeal action by individual nations or groups. If the EEC alone succeeded in halving its emissions of SO2, it would be less effective than persuading all countries to reduce their emissions by 15%.

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4.4.2 Relationship between deposition and damage If there is scientific evidence linking acid deposition with areas of damage, it is reasonable to suppose that reduced deposition would lead to reduced or avoided damage. However, the form of the relationship (referred to as a ‘dose-response function’) is not well established for any of the damage areas. The sections of the Report that deal with water systems and forest systems demonstrate how far the scientific consensus is from defining the relationship. The different possibilities are illustrated in a qualitative way in Figure 4.4 (page 46). If cause and effect were directly related (Curve A), it can be inferred that, say, halving the emissions would halve the adverse ecological effects. On the other hand, if adverse effects increased more rapidly than emissions (Curve B), the desired halving of damage could be achieved with a proportionately smaller reduction in emissions. A third possibility is that most of the adverse effects arise from relatively low emissions and that further increases do not make matters appreciably worse (Curve C). In this case, action to halve the level of emissions would produce a very limited benefit. Because of its obvious importance in legislative decisions, the ‘linearity’ issue has become the main focus of discussion on both sides of the Atlantic. The evidence is conflicting: most a priori examinations of the problem suggest that either Curve A or Curve B is the more likely way for the environment to react to stress from polution. But, in Europe, the relative insensitivity of rainfall acidity to large changes in emission rates over the last 15 years or so tends to support Curve C, or, alternatively, to suggest that the initial assumption as to the cause of the adverse effects may be in error. The National Academy of Sciences (NAS) in the U.S.A. recently published a report which concluded, albeit rather tentatively, that over a region as large as the northeastern U.S.A. the relationship between causes and effects of acid rain would be linear. It acknowledged, however, that European data supported a contrary conclusion. The report has been followed by many counter-reports, articles and letters from research organisations and individual scientists hotly contesting the NAS conclusion on linearity, to the extent that the U.S. Administration may well postpone a decision on emission controls until major experiments have been undertaken to settle the linearity question. A related issue is the question of whether the affected lakes and forests will revert to their former state once the causes of acid rain are abated or whether the damage is irreversible. The alternatives are as follows. (1) The effects of pollution have been to reduce the capacity or opportunity of waters and soils to neutralise acid inputs (and possibly also to deplete them of essential biological nutrients) to the extent that, even if the cause is removed, the waters and soils will remain affected for a very long time. (2) Alternatively, the effects of pollution are principally episodic in nature, causing temporary overloads of the natural ecosystems which lead to the observed adverse effects, but the ecosystems are not permanently damaged. Once again, as is evident from the preceding sections of this Report, scientific evidence is unable to provide clear answers. The importance of the ‘reversibility’ issue lies in the timescales of the results to be expected from remedial measures. If the damage caused is episodic and reversible, then quite rapid improvement may follow in the affected ecosystems. If irreversible damage has been caused, recovery will be extremely slow and dependent on the weathering of rocks to replace the constituents lost by acidification. In this case, emission controls may act to prevent the spread of the affected areas but will not cure those already affected. Management of the lake catchments and forests will be necessary to restore their productivity, including the use of fertilisers, acidneutralising compounds, surface cultivation, restocking with hardy species etc. These forms of management may well be advisable as a ‘first aid’ measure in any event, since the benefit of reducing emissions is bound to take many years to materialise.

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Much of the concern of those questioning the wisdom of stringent control policies has not been so much over the cost-effectiveness of those policies, but simply over their effectiveness irrespective of costs. They judge that not enough is known about the dose-response relationships (particularly in response to decreased deposition) to enable control policies to be designed (in terms of level of reduction, pollutants controlled, and perhaps, other measures) that would be reasonably likely to achieve their desired objectives. 4.4.3 Relationship between benefits and costs An economist would argue that the effectiveness of a remedial strategy should be assessed by the balance between benefits accruing and costs incurred. Where the costs of a reduction in the level of emissions are less than the benefits associated with the reduction, reducing the emissions is a good thing. In a general sense this proposition is easy to defend. However, its usefulness is limited by the enormous task of applying it to so complex a policy area as emission control. The problems start with the scientific uncertainties already discussed. These preclude a reliable estimation of benefits in physical terms. However, even when a physical benefit can be quantified, it may be very difficult to translate it to a monetary scale— necessary for aggregating benefits and comparing them with control costs. Only where a good is regularly traded in the market-place can a straightforward money value be assigned. For environmental benefits this is almost never the case, and ever more roundabout means must be found to attach a value to, say, a human life or a beautiful view. Even when this is done, difficulties arise, because costs and benefits may occur over different timescales (bringing in problems of discounting) and they will be distributed differently over nations and over interest groups within nations. Despite these difficulties, there is great interest, in the U.S.A. and internationally, in using cost-benefit analysis to support environmental poicy. A recent study by OECD weighed the costs across Europe of some measure of SO2 abatement against the benefits accruing over a wide number of areas, including human health and materials as well as the condition of water and forests. This suggested that some increased measure of control would be cost-effective in Europe. However, the many debatable assumptions that had to be made (and are frankly admitted in the study) and the high no-control emissions assumed for the future reduce the value of its conclusions. The major interest and purpose of the study was to demonstrate a methodology. As part of the scientific framework supporting the UNECE Convention on Long-Range Transboundary Air Pollution, use of a significantly modified cost-benefit approach has been recommended and received political support. A new methodology combining features of cost-benefit analysis and decision analysis has been proposed. This is well adapted to handle the uncertainties that are inevitably present, and can be used to highlight the most important uncertainties (from a decision-support point of view) and to order research priorities. The U.K. has been prominent in supporting this methodology within the UNECE. These attempts to provide a general framework are worthy of support. Although there is a danger that they could be misconstrued as presenting certainties where none exist, they represent the best means that we have for arriving at a rational deployment of society’s resources. They provide means of evaluating, in policy terms, advances in scientific knowledge, and may soon at least be able to tell us what control strategies are clearly worthwhile, or clearly not worthwhile, even if that cannot point to the ‘best’ policy. The problem of designing a cost-efficient strategy— defined as a minimum cost for a given reduction in emissions—is an important step in the search for fully cost-effective strategies—where benefits at least equal costs. One characteristic of a least-cost solution is that it is likely to vary between national situations. Thorough investigation is required to identify the most cost-efficient means of reducing emissions in a

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national setting, and these are likely to change as the energy system evolves through time. It is unrealistic to expect international legislation to capture these complexities. What is required is a form of legislation that allows full use of local discretion. The appropriate level for such legislation is in terms of target aggregate national emissions for periods ahead. The longer the timescale allowed, the lower should be the cost of achieving the objective. Staged objectives would also help ensure that the energy system adapts in ways consistent with longer-term aims. More specific stipulations, for example stipulations about new plant as against existing, or large plant as against small, can only increase the costs of reaching the same overall objective. 4.5 Summary To summarise the state of the overall scientific debate as it relates to policy, several points can be made. (1) The scientific evidence supporting a relationship between acid deposition and damage over a range of activities is insufficient to estimate the benefits that would result from a given level of emission reduction. In particular, the understanding of combination effects between pollutants is deficient. (2) Preliminary evidence indicates that retrofitting fgd to power stations on a large scale would give a once-and-forall increase in electricity tariffs of 4–5%. (3) All approaches to damage reduction should be considered, including direct action, such as liming of lakes and local action to alleviate high local levels of emission. (4) There is an advantage in achieving the widest consensus possible in international action, rather than agreeing more severe limitations within some small group. (5) International legislation should be formulated in general terms of phased reductions in national aggregate emissions, rather than include more specific stipulations. (6) The use of cost-benefit analysis, and related techniques, can be a valuable aid in devising a rational deployment of resources to both environmental research and environmental control. Bibliography General reviews/control options/emission projections Acid rain: a review of the phenomenon in the EEC and Europe. Environmental Resources Ltd., London, 1983. Coal and environmental protection. OECD, Paris, 1983. EMEP: The cooperative programme for monitoring and evaluation of long-range transmissions of air pollutants in Europe. ECE, United Nations, Geneva, 1981. SO2 emission trends and control options in Western Europe. CONCAWE (Oil Companies’ European Organization for Environmental and Health Protection), No. 1, The Hague, 1982. C.H.Davies: Scenarios and electricity demand. In CEGB proof of evidence to the Sizewill B Public Enquiry, P5. Central Electricity Generating Board, London, Nov 1982. Control technologies and costs J.Bethelheim, J.R.P.Cooper, W.S.Kyte and D.T.H.Rowlands: The integration of a regenerable flue-gas desulphurisation plant on to a 2000 MW coal-fired power station site in the United Kingdom . In Management and Conservation of Resources, Institution of Chemical Engineers Syposium Series 72, C1–16, London, 1982. Coal and the Environment, Chapter 7. Commission of Energy and Environment, HMSO, London, 1981.

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Trends in residual fuel oil demands and conversion capacities. CONCAWE, No. 14, The Hague, 1983. Direct desulphurisation of residual petroleum oil: investments and operating costs. CONCAWE, No. 5, The Hague, 1981. A.J.Clarke: Contribution in: The Ecological Effects of Deposited Sulphur and Nitrogen Compounds. Phil. Trans. Ser. B, 305, 259–577, 1984. N.H.Highton and M.G.Webb: Pollution abatement costs in the electricity supply industry in England and Wales. Discussion Paper 65, Institute of Social and Economic Research, University of York, 30 pages, York, 1981. N.H.Highton and M.G.Webb: The effects on electricity prices in England and Wales of notional sulphur dioxide emission standards for power stations. Discussion Paper 67, Institute of Social and Economic Research, University of York, 26 pages, York, 1981. Legislation

Coal and the environment, Chapter 4. Commission on Energy and the Environment , HMSO, London, 1981. Air Pollution. Paper 12, Technology and the Environment, Dept of Industry, London, 1982. Proposal for a directive on the limitation of emissions of pollutants into the air from large combustion plants. COM (83)704, EEC, Brussels, Dec 1983. Cost-benefit analysis Costs and benefits of sulphur dioxide control. OECD, Paris, 1981. Cost-benefit analysis and environmental regulations. Conservation Foundation, Washington, D.C., 1983.

Appendix 1 Fifteenth Consultative Council Meeting of the Watt Committee on Energy Acid Rain

On 1st December 1983 the Watt Committee on Energy held the fifteenth in the series of its Consultative Council meetings. The theme was ‘Acid Rain’, and the meeting was held at the Institution of Civil Engineers, London. Those present were the secretaries and appointed representatives of the member institutions of the Watt Committee and others with professional interests in the subject of the meeting. Papers (listed below) were presented informally by members of the Watt Committee working group on Acid Rain, and there were several periods of discussion. The contents of the papers form part of the present Watt Committee Report. As published here, the papers have been expanded and revised to present additional information for which there was insufficient time at the meeting and to take account of questions that arose in discussion. Session 1 Chairman: Professor K.Mellanby CBE (Institute of Biology) Official Opening by Dr J.H.Chesters OBE FEng FRS (Chairman, Watt Committee on Energy) Professor Mellanby: Introduction Dr F.B.Smith (Royal Meteorological Society): Production, transformation, transport and deposition Dr A.W.C.Keddie (Warren Spring Laboratory): Emissions. Dr A.T.Cocks (Royal Society of Chemistry): Transformation, Synergistic Effects; Measurement. J.A. Garland (Royal Meteorological Society): Deposition Processes. Dr B.E.A.Fisher (Institute of Mathematics and its Applications): Modelling; Episodes. Dr F.B.Smith: Summary. Dr W.O.Binns (Institute of Chartered Foresters): Soil and vegetation Historical Aspects. Effects on Plant Metabolism and Growth. Vegetation and Soils. The Ulrich Hypothesis. The Central European Situation, Session 2 Chairman: Sir John Mason CB, FRS (Royal Society) D.Hammerton (Institution of Water Engineers and Scientists): Freshwater Dr J.H.Stoner (Institution of Water Engineers and Scientists): Water Quality and Fishery Aspects. Dr R.W. Battarbee (University College London): Sediment Records. D. Hammerton: Distribution of Acid Waters in the U.K.; Conclusions. Session 3 Chairman: Sir Morris Sugden FRS* C.J.Davies (Operational Research Society): Remedial strategies A.J.Clarke (Central Electricity Research Laboratory): Legislation. C.J.Davies: Control Technologies. N.H. Highton (Beijer Institute, Stockholm): Impact on Electricity Costs. Concluding remarks: Sir John Mason.

* The Watt Committee notes with regret that Sir Morris Sugden died on 3 January 1984.

Appendix 2 Abbreviations Used in this Report

General BAPMoN CEGB CERL EACN EEC

EMEP LRTAP NAS NCB NILU NORDFORSK OECD WMO UN UNEP UNECE

Background Atmospheric Pollution Monitoring Network Central Electricity Generating Board Central Electricity Research Laboratory (CEGB) European Atmospheric Chemistry Network European Economic Community (‘the Community’, whose power is exercised by the Commission of the European Communities—‘the European Commission’—under the policy direction of the Council of Ministers subject to the powers of the European Parliament) European Monitoring and Evaluation Programme Long-range transport of air pollution National Academy of Sciences (U.S.A.) National Coal Board Norwegian Institute for Air Research Scandinavian Council for Applied Research Organisation for Economic Co-operation and Development World Meteorological Organisation United Nations United Nations Environment Programme United Nations Economic Commission for Europe

Chemical Formulae and Other Expressions CO CO2 fgd H2O H2O2 H2SO4

carbon monoxide carbon dioxide flue-gas desulphurisation water hydrogen peroxide sulphuric acid

87

NO2 O3 OH pf pH SO2 NMHC

nitrogen oxides (this expression is used because both NO and NO2 may be present; see particularly Sections 1.1 and 1.2 of this Report) ozone hydroxyl radical (see Section 1.3) pulverised fuel expression for acidity (see Section 2.3.1) sulphur dioxide non-methane hydrocarbons

Units ha kg µg t km µm l W MW GW MV

hprtarp (1 ha=10 000 m2=2 471 acres) kilogram (1 kg=1000 g=2. 205 Ib) microgram (1 g= 1 000000 µg) tonne (metric) (1 t= 1000 kg=0.984 ton) kilometre (1 km=1000 m=0.621 mile) micron (1 m=1 000 mm=1 000 000 µm) litre (1 l=10−3 m3=1.76 pint=0.22 gallon (U.K.)) watt megawatt (1 MW= 1000 kW= 1 000000 W) gigawatt (1 GW= 1 000000000 W) megavolt (1 MV= 1000 kV= 1 000000 V)

Note: See page 21 for use of ‘billion’ and related expressions in this Report.

THE WATT COMMITTEE ON ENERGY

Member Institutions June 1984

* British Association for the Advancement of Science British Ceramic Society * British Nuclear Energy Society * Chartered Institute of Building * Chartered Institution of Building Services Chartered Institute of Patent Agents * Chartered Institute of Transport * Combustion Institute (British Section) * Geological Society of London * Hotel Catering and Institutional Management Association * Institute of Biology Institute of British Foundrymen Institute of Ceramics * Institute of Chartered Foresters * Institute of Cost and Management Accountants * Institute of Energy * Institute of Home Economics * Institute of Hospital Engineering Institute of Internal Auditors (British Chapter) * Institute of Marine Engineers Institute of Mathematics and its Applications * Institute of Petroleum * Institute of Physics Institute of Management Services * Institute of Purchasing and Supply * Institute of Refrigeration Institute of Wastes Management * Institution of Agricultural Engineers * Institution of Chemical Engineers * Institution of Civil Engineers * Institution of Electrical and Electronics Incorporated Engineers * Institution of Electrical Engineers * Institution of Electronic and Radio Engineers

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Institution of Engineering Designers * Institution of Gas Engineers Institution of Geologists * Institution of Mechanical Engineers * Institution of Metallurgists * Institution of Mining and Metallurgy Institution of Mining Engineers * Institution of Municipal Engineers * Institution of Nuclear Engineers * Institution of Plant Engineers * Institution of Production Engineers * Institution of Public Health Engineers Institution of Structural Engineers * Institution of Water Engineers and Scientists * International Solar Energy Society * Metals Society Operational Research Society * Plastics and Rubber Institute * Royal Aeronautical Society * Royal Geographical Society * Royal Institute of British Architects * Royal Institution * Royal Institution of Chartered Surveyors * Royal Institution of Naval Architects * Royal Meteorological Society * Royal Society of Arts * Royal Society of Chemistry * Royal Town Planning Institute * Society of Business Economists Society of Chemical Industry * Society of Dyers and Colourists Textile Institute

*Denotes present and past members of The Watt Committee Executive

THE WATT COMMITTEE ON ENERGY

GENERAL OBJECTIVE The objective is to promote and assist research and development and other scientific or technological work concerning all aspects of energy and to disseminate knowledge generally concerning energy for the benefit of the public at large. TERMS OF REFERENCE The Watt Committee on Energy, being a Committee representing professional people interested in energy topics through their various institutions, has the following terms of reference:— 1. To make the maximum practical use of the skills and knowledge available in the member institutions to assist in the solution of both present and future energy problems, concentrating on the U.K. aspects of winning, conversion, transmission and utilisation of energy and recognising also overseas implications. 2. To contribute by all possible means to the formulation of national energy policies. 3. To prepare statements from time to time on the energy situation for publication as an official view of The Watt Committee on Energy in the journals of all the participating institutions. These statements would also form the basis for representation to the general public, commerce, industry and local and central government. 4. To identify those areas in the field of energy in which co-operation between the various professional institutions could be really useful. To tackle particular problems as they arise and publish the results of investigations carried out if suitable. There would also, wherever possible, be a follow-up. 5. To review existing research into energy problems and recommend, in collaboration with others, areas needing further investigation, research and development. 6. To co-ordinate future conferences, courses and the like being organised by the participating institutions both to avoid overlapping and to maximise co-operation and impact on the general public. EXECUTIVE COMMITTEE as at June 1984: Chairman Dr J.H.Chesters, O.B.E., F.Eng., F.R.S. Deputy Chairman G.K.C.Pardoe E.A.Alcock, Institute of Cost and Management Dr A.F.Jackson, Institution of Electrical Engineers Accountants (Hon. Treasurer) Dr E.G.Masdin, Institution of Chemical Engineers Professor W.O.Alexander, Institution of Dr J.R.Miiford, Royal Meteorological Society Metallurgists W.B.Pascall, Royal Institute of British Architects H.W.Baker, Institution of Civil Engineers W.Ridley, Society of Business Economists H.Brown, Institution of Plant Engineers Dr P.A.A.Scott, Royal Society of Chemistry

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EXECUTIVE COMMITTEE as at June 1984: Chairman Dr J.H.Chesters, O.B.E., F.Eng., F.R.S. Deputy Chairman G.K.C.Pardoe R.F.C.Butler, International Solar Energy Society A.Silverleaf, Royal Institution of Naval Architects (U.K. Section) Professor I.C.Cheeseman, Chartered Institute of Professor A.J.Smith, Geological Society of London Transport Professor J.Swithenbank, Institute of Energy A.Cluer, Institute of Petroleum E.L.Walker, Institute of Purchasing and Supply Professor A.W.Crook, Institution of Mechanical N.G.Worley, British Nuclear Energy Society Engineers R.S.Hackett, Institution of Gas Engineers J.G.Mordue, Secretary Note: A part of the executive rotates on an annual basis at 30th April each year. The following institutions were members of the executive for the years shown:

1983/84 British Association for the Advancement of Science Hotel Catering and Institutional Management Association Institute of Chartered Foresters Institution of Agricultural Engineers Institution of Electronic and Radio Engineers Institution of Water Engineers and Scientists 1984/85 Geological Society of London Institute of Purchasing and Supply Institution of Metallurgists International Solar Energy Society (U.K. Section) Royal Institution of Naval Architects Royal Meteorological Society

Dr J.G.Colling wood M.J.Moore J.N.R.Jeffers J.C.Weeks R.F.C.Butler C.Cash Professor A.J.Smith E.L.Walker Professor W.O.Alexander R.F.C.Butler A.Silverleaf Dr J.R.Miiford