acidification effects on groundwater - prognosis of the...

Future Groundwater Resources at Risk (Proceedings of the Helsinki Conference, June 1994). IAHS Publ. no. 222, 1994.

Acidification effects on groundwater - prognosis of the risks for the future

GERT KNUTSSON Division of Land and Water Resources, Royal Institute of Technology, S-100 44 Stockholm, Sweden

Abstract The natural acidification of groundwater in some types of environment has accelerated by acid atmospheric emissions and cultivation during the last 200 years. The direct effects are seen in changes of the groundwater chemistry in four stages, (1) seasonal depression of pH, alkalinity and some cations in very shallow groundwater, (2) long-term increase of calcium, magnesium, nitrate and sulphate in shallow groundwater from areas with acid rock and soil and high acid load, (3) reduction of alkalinity and/or pH over time in shallow groundwater in non-calcareous sandy aquifers and high total acid deposition, (4) zero alkalinity, pH below 5, high contents of aluminium, nitrate and sulphate as well as notable contents of heavy metals. Indirect positive effects are known when groundwater with high alkalinity neutralizes the input of acid rain or meltwater in streams and lakes. Indirect negative effects on surface water are the discharge of very acid groundwater containing high content of aluminium, iron and manganese with strongly negative, sometimes toxic, impact on biota. Another negative effect is the corrosion of water pipes, which leads to leakage and high contents of heavy metals in drinking water. The prognosis for the future by modelling is very alarming: 85% reduction of sulphur and 60% of nitrogen are required to halt and reverse the ongoing acidification of shallow groundwater in sensitive areas in Sweden.

PROBLEMS AND OBJECTIVES

Acidification of soil and water is an on-going natural, internal process, known for a long time in some types of environments. But acidification can also be due to anthropogenic causes linked to the industrialization. This external acidification has an imperceptible course during a long initial phase, which is why the effects were not recognized until long afterwards: on surface water in the 1950s and 1960s and on groundwater and soil not until twenty years later. The scientific problem is still how to distinguish between natu-ral and anthropogenic causes of acidification. The objectives of this paper are to descrribe the two types of acidification and then to sort out the effects on groundwater of anthropogenic origin as well as to make an assessment of the total acidification effects and to present a tentative prognosis of the risks for the future.

NAMJRAL ACIDIFICATION

Natural acidification of soil and water is a very slow, biogeochemical process, caused

4 Gert Knutsson

by soil respiration (giving carbonic acid), dissociation of humic acids, oxidation of sulphur and nitrogen compounds, oxidation and hydrolysis of ferrous iron in the soil and uptake of cations from the soil in the vegetation. A contribution to the acidification comes from the precipitation, which in its natural state has a pH value around 5.6 owing to the content of carbonic acid in the atmosphere and natural emissions of sulphur from land, sea and volcanoes. The natural acidification is evident in areas with weathering-resistant soils and rocks, where the climate is humid and the dominating water movement as well as the transport of chemical components is downward, resulting in a runoff of base cations. The leaching processes form typical soils such as latérites in the humid tropics and podzols in the temperate and boreal climates. The soils in non-glaciated areas are mature in contrast to the young soils in "recently" glaciated areas, for example in northern Europe and Canada. However, in "poor" soils of these areas there has been a marked leaching of the uppermost layer of the mineral soil with a typical podzol profile as a result. Natural acidification of water is documented by diatom and pollen analyses of lake sediments. The pH-value in lake water has continually — but very slowly — decreased from 7 to 6 since the déglaciation 12-13 000 BC in western Sweden (Renberg & Hedberg, 1982). According to the accepted concept that most of the water passes through the ground — in areas with this type of climate — before it reaches streams and lakes, a slow decrease of pH in groundwater can also be assumed. In fact, groundwater with pH around 6 and very low alkalinity (10-20 mg HC03 l"1) is the natural state today in vast areas with coniferous forests and podzolic soils on coarse-grained deposits and acid rocks, for example in Finland, Norway, Sweden and on the precambrian shield of Canada.

The biogeochemical processes in the soil profile are, as mentioned above, of great importance for the chemical composition of groundwater. The final composition of groundwater is, however, a result of the geochemical processes during the percolation of water through the unsaturated zone and the flow in the groundwater zone. The discharge area has a crucial role as the redox conditions are very complex and shifting, the organic content is mostly high and the exchange with surface water is shifting. The flow pattern of water is also of great significance: recharge and discharge areas exist in local, medium-sized and regional flow systems. The turnover rate of water, which gives the contact time between water and minerals - and thereby the time for ion exchange — will determine the degree of neutralization of groundwater together with the weatherability and effective surface areas of the minerals. In general, the turnover time increases with depth and length of the flow paths and along with that also alkalinity and pH of the groundwater. However, there can be large differences in flow time and flow paths on a local scale, especially in the upper zones of the ground due to heterogeneities such as high-permeable surface layers, macropores and veins in till. Seasonal and perennial fluctuations in the groundwater level can also cause natural changes in the hydrochemical conditions. During wet seasons with high groundwater levels, most water flows rather quickly in high-permeable, surface layers and short distances in local flow systems. But during dry periods or long periods of winter with low groundwater levels, small amounts of water flow slowly in less permeable, deeper layers and longer systems (Jacks et al., 1984). These two different situations in groundwater flow will cause evident changes in time and place of the groundwater chemistry; the first situation will give lower pH and alkalinity than the second as the water flows in soil horizons, which are leached from bases and have a

Acidification effects on groundwater 5

high organic content with low pH. Still greater changes can occur during perennial fluctuations in climate. An extremely low groundwater level after a period of dry years can give rise to oxidation of sulphides in the "dried" zone of the ground, which incTeases the content of sulphate (an acidification parameter) in groundwater, when groundwater level rises again.

Another type of natural acidification is found in coastal regions, where precipitation has a high content of sea salts (mainly sodium chloride) during periods of storms. This can result in ion exchange between sodium and hydrogen or aluminium in the ground, which may cause an episodic acidification of ground and surface water. This phenomenon is well documented in the catchment for acidification research in Birlcenes, southern Norway (Statens Forurensningstillsyn, 1993).

ANTHROPOGENIC ACIDIFICATION

The natural acidification processes have been accelerated by two diffuse anthropogenic sources: atmospheric emissions and cultivation. Anthropogenic acidification by point sources such as leakage from mineral waste is not discussed in the paper.

Atmospheric emissions

Local acid atmospheric emissions, which damaged the vegetation, were observed around the copper mine and smelter at Falun, Sweden, as early as in the 1730s by the famous Swedish scientist Carl von Linné. The acidification of industrial areas in England was described by R. A. Smith, who first used the expression "acid rain" in 1872 (Environment' 82 Committee, 1982). Since that time, and particularly since the Second World War, there has been an increase in the emissions of compounds of carbon, nitrogen and sulphur from the use of fossil fuels in industry and traffic as well as for heating. The first systematic monitoring of the deposition of airborne substances started in Sweden in the 1940s and in 1956 the European Air Chemistry Network was established by the International Meteorological Institute in Stockholm. Using data from this network in combination with his own data, Svante Odén published his findings in 19 68 that the precipitation over Scandinavia had become increasingly acidified and that the acid deposition originated mainly from emissions in central and western Europe an«i the British Isles. At this time an increasing acidification of lakes was reported from the west coast of Sweden (Environment' 82 Committee, 1982). Such a phenomenon was, in fact, already described in the 1950s from Nova Scotia, Canada (Gorham, 1957, ref from Jacks, 1993). Investigation of lake sediments in northeastern USA and eastern Canada as well as in Scotland and Sweden have shown (by diatom aad pollen analyses, determination of heavy metals and soot particles) the historical development of acidification during the last 200 years, especially since 1850. Today th«e acid atmospheric emissions and acidification are also a regional problem in the industrialized countries in Asia, for example China and Japan.

The long-range transported emissions across national boundaries have political implications as the acidification effects have economic consequences in the form of corrosion of buildings, damage offish stocks, forest die-back and health hazards. An ecological consequence is that sensitive ecosystems in remote areas can be influenced

6 Gert Knutsson

by long-term effects from relatively low concentrations of acid atmospheric depositions. This is the acute situation in the southern part of the High Mountains of Scandinavia (see Indirect effects/Negative effects).

The emissions of ammonia from areas with intensive agriculture, especially the livestock industry, for example in Denmark and the Netherlands, have also a great, but more local impact on soils and groundwater. Ammonia is converted to nitrate, N03, via NH4 by nitrifying bacteria. Acidification of groundwater occurs, if the vegetation does not adsorb all the N03, the excess of which leaches to the groundwater. The total amount of nitrogen compounds from different sources has, however, to be considered, when the effect on groundwater is calculated.

Cultivation

Intensive cultivation in agriculture and forestry over a long period may cause acidification of soils by nutrient uptake of bases and the removal of crops and timber. In modern agriculture the loss of nutrients is compensated by addition of fertilizers. But if ammonium-bearing fertilizers such as ammonium sulphate are used, the effect will be an acidification of arable land. Instead, fertilizers including lime or limestone have to be added to avoid acidification on noncalcareous soils. The effect of modern forestry on acidification must also be considered. The plant uptake is not as great as in agriculture (only 10%) and it varies with tree species and age of trees. But if the soils are poor, the removal of the whole tree will have an acidification effect. Deforestation also reduces the capability of the vegetation to adsorb the atmospheric input of nitrogen with an increase of nitrate to groundwater as a result.

DATA SOURCES

Opportunities to study the effects of acidification on groundwater are limited. The data sources are normally not sufficient as the temporal and spatial variations in groundwater chemistry are considerable. Large and/or deep groundwater systems react very slowly, which is why long-term series of analyses (15-20 years) are required to be able to distinguish any changes or effects. Limited and shallow groundwater systems react rather quickly but they often have a large spatial variation due to differences in geology, topography and land-use. Thus a lot of detailed data is needed for assessment of general trends in a region.

Four main types of data sources can be used (see Knutsson, 1994, for a more detailed examination): (a) data from well archives and old investigations; (b) data from monitoring programmes; (c) data from well surveys and research projects; (d) data from modelling work. The archive data have to be interpreted with great caution. There can be inconsistencies in site locations, sampling and analysing methods, especially in pH measurements. Technical modifications in and around wells and changed pumping rates can create limitations in the evaluation of data. Historical data from municipal wells are more reliable than those from private wells as the procedures in sampling and analysing have


been more standardized and as the time periods of regular sampling have been fairly long, for example 15-50 years in Sweden.

The records from groundwater monitoring networks are of great interest for studying long-term changes and trends in different types of groundwater systems. Time series analyses with statistical methods (principal component analysis and regressions analysis) show, however, that the measurements, sampling and analyses have to be well coordinated, equidistant and frequent to allow such statistical treatments (Atidersson & Stokes, 1988).

Data from well surveys or inventories are mostly consistent, as the work is normally performed by using standardized methods in field and laboratory. Data from research projects are supposed to be very reliable and of great value as the records from water analyses often include parameters such as heavy metals, species of aluminium and isotopes. Modelling technique can be used to reconstruct and predict changes and effects but the results have to be verified by reliable field and laboratory data.

ACIDIFICATION EFFECTS

The effects of acidification on groundwater can be classified into direct effects on the groundwater itself and indirect effects of acid groundwater on surface water and biota as well as on constructions, above all on water pipe systems and thereby on drinking water, which may give health effects.

Direct effects on groundwater chemistry

Observations of acidification effects on groundwater were not reported until twenty yeajs later than those on surface water, at first locally in western Sweden, close to large emissions of sulphur oxides (Hultberg & Johansson, 1981); and shortly thereafter on a regional scale in western and southeastern Sweden (Jacks & Knutsson, 1981, 1982). Since then, comprehensive research and investigation in Sweden and several other countries have shown significant changes in the groundwater chemistry in many areas due to the increased acid load. The findings can be summarized as follows (cf Soveri, 1985; Soveri & Knutsson, 1994; Knutsson, 1994): (a) The first stage of acidification of groundwater is a seasonal depression of pH,

alkalinity and some cations but a peak in the content of sulphate during and after storm and snowmelt events (Fig. 1). These so called "acid surges" have been reported from very shallow groundwater systems, for example in superficial tills on hillsides, in Canada (Bottomley et al., 1986; Craig & Johnston, 1988) and Sweden (Knutsson, 1992). The "pre-event" groundwater seems to be diluted by percolating acid water with extremely low ion concentration but high concentration of sulphate. Waterflow in preferential pathways may also contribute to the rapid changes (Espeby, 1989).

(b) The second stage of acidification is characterized by a long-term increase of calcium and magnesium (=total hardness) as well as of sulphate and nitrate but a stable pH. Such signs are found in shallow groundwater from many areas with acid rocks, poor soils and high acid load. The relationship between total hardness and alkalinity can also be used as an early indicator of changes in groundwater

Gert Knutsson

6.0

5.6

5.2

4.4

4.0

jAïXiM^

T T T T T T T T T T T T T T T l T T T r r r 1 1987 1988

T r n n T " n " T T T r 1 1989

Fig. 1 Seasonal fluctuations of pH in three springs in stony sandy till, Lofsdalen, Sweden.

chemistry by acidification. The 1:1 relationship represents the undisturbed state of weathering due to carbonic acid. A ratio greater than 1 or a displaced regression line indicate an input of anthropogenic, strong acids (Jacks & Knutsson, 1981; Jacks et al, 1984). The relationship is often > 1 in southern, western and eastern Sweden (Fig. 2) and/or the regression line has been displaced over time.

Halland county Kronoberg county

Ca2' + Mg2* mg/l meq/l

Ca2% Mg2* mg/l meq/l

120 240

Ca2*+Mg2* Blekinge county mg/l meq/l

80-

40-

o-

- 4

-2% y

H7

V

I • y f o /

2 I h -

K /

1 _ 4 1 1 1

120 240

meq/l HC03-

mg/l

Fig. 2 Relationship between total hardness and alkalinity of groundwater from municipal wells in superficial deposits, south and southwest Sweden (from von Brômssen, 1989).


(c) The third stage of acidification of groundwater is the reduction of alkalinity and/or pH over time. Decreasing alkalinity, which is a useful definition of acidification, has been observed in several areas where the bedrock and soil consist of weathering resistant minerals (but on the other hand increasing alkalinity in calcareous rock and soil). Real acidification of groundwater in terms of a significant pH decrease over time is not so common. It is documented in shallow groundwater in non-calcareous sandy aquifers, where the total acid deposition and/or the local emissions from industry and farming are high. The following examples can be mentioned. A marked pH decrease of 1-2 pH units was observed in Belgium from 1959 to 1984 (Voet & Vangenechten, 1984) and a moderate pH decrease of maximum one pH-unit in some shallow wells and springs in southern Finland (Lahermo, 1994) and of 0.7 pH unit in southwest Denmark from 1950 to 1985 (Bâdsgârd Pedersen, 1985). A significant, but modest pH decrease was found in the research catchment of Birkenes, Norway from 1980 to 1992 (Fig. 3). There are trends of decreasing pH in water from shallow wells and springs as well as from some municipal wells in south and west Sweden (von Brômssen, 1989). The geology around these wells is dominated by sandy tills or sandy glaciofluvial sediments with a petrography of granitic composition.

pH Alkalinity 5.5 T 5,4 5,3-t 5,2 5,1

5 + 4,9

80 82 84 86 90 92 Fig. 3 Decreasing pH and alkalinity in groundwater from the research catchment of Birkenes, Norway (from Statens Forurensningstilsyn, 1993).

(d) The last stage of acidification is when alkalinity goes down to zero, pH drops to around or below 5 and the concentrations of aluminium, nitrate and sulphate are considerable high. Dissolution of some heavy metals is notable for example cadmium. This is a very critical stage of acidification: groundwater is not suitable as drinking water. Nitrate concentrations are somewhere higher than the EU standard of 50 mg N03 l"1. High concentrations of aluminium can be dangerous for kidney patients and may cause Alzheimer's disease (Martyn et al, 1989). Some of the heavy metals may be of toxic levels. Such bad situations have been documented in areas with very high acid load and acid rock and soil in Germany: Fichtelgebirge with pH values between 4 and 5 and concentrations of aluminium up to 4.8 mg T1 (Sager et al., 1990), Teutoburger Forest with pH values from 3.8 to 4.4, sulphate mean value of 40 mg l"1 and aluminium mean value of 5.0 mg l"1. At such low pH values precipitation of amorphous aluminium oxides starts, which implied coatings on water-filters,casing and pipes in municipal wells (Luckewille 8c van Breemen, 1992). High concentrations of nitrate and aluminium as well as relatively high concentrations of heavy metals such as cobalt, chromium, nickel and zinc are reported from the Netherlands (Arends et al, 1987). In Sweden the concentrations of heavy metals are still acceptable but concentrations of aluminium

10 Gert Knutsson

from 1.5 to 2.0 mg l"1 are found in a few cases (Bertills et al., 1989). A further anthropogenic acidification of Swedish soils would increase the figures and the frequency in groundwater, as well as the mobility of heavy metals, for example that of arsenic (Xu et al., 1991).

INDIRECT EFFECTS

Effects on surface water and biota

The current concept is that groundwater acts as a dynamic and important part of many stream and lake ecosystems (Vanek, 1987). Discharge of groundwater to fens, springs, streams and small lakes may have either positive or negative effects owing to the chemical status and the flow of groundwater in relation to those parameters in surface water. Comprehensive research and investigations as regards the interaction between groundwater and surface water have been carried out, above all in Canada, Scotland and USA.

Positive effects are recognized when the "pre-event" groundwater with higher pH and alkalinity than the acid rain or meltwater contributes a large proportion (40-90%) of the runoff in streams, whereby the acid water is neutralized (Bottomley et al., 1984). Even if the contribution of groundwater directly to lakes is much smaller (5-10%) than to streams, the well-buffered groundwater is the major source of e.g. alkalinity, calcium, iron, magnesium, potassium and sodium to lakes, and is of great importance for the delay in acidification of lakes (Kenoyer & Anderson, 1989; Cook et al, 1991). Modelling studies demonstrated a delay effect on some lakes for more than 100 years (Anderson & Bowser, 1986). A very small extra input of groundwater could also be used for restoration of recently acidified lakes (Cook et al, 1991). Detailed investigations of a lake situated in a sandy, "silicate" aquifer in glaciated terrain, northern Wisconsin, USA, also showed, that the flux of groundwater and dissolved solids to the lake was highly seasonal. Large pulses of groundwater inflow occurred after spring snowmelt and after rain in the autumn (Kenoyer & Anderson 1989) as in streams.

Negative effects on surface water and biota are reported in catchments with thin, coarse-textured weathering-resistant soils and/or bare acid bedrock. The alkalinity of the groundwater is many times too low to neutralize the acid precipitation because of short turnover rate of water and low weatherability of soils and rocks. The contribution of groundwater to surface water may also be smaller in such areas with limited, shallow aquifers and bare rocks than in areas with large, deep aquifers. The impact of acid, shallow groundwater in glaciated terrains on the acidification of surface water, especially first-order streams and small lakes, has been recognized, for example in the Adirondack Mountains, New York, USA (Chen et al, 1984), in Ontario, Canadian Precambrian Shield (Bottomley et al., 1986) and in many regions of Finland, Norway and Sweden. Surface water in these regions with a pH lower than 5 support few or no fish species and very few species of invertebrates.

The discharge of acid groundwater containing metals to surface water can also have a strongly negative, sometimes toxic, impact on biota. High concentrations of aluminium have been found in acidic streams in the Mid-Atlantic Highlands, USA


(Kaufmann et al., 1991) and in springs and streams in the Scandinavian High Mountains. A grey precipitate of aluminium, which poisoned the vegetation, was reported at several discharge areas in West Norway (Saether & Follestad, 1992). Moderate concentrations of aluminium and manganese and high concentration of iron were observed in springs and streams at Djursvallen, Lofsdalen, Sweden (Jacks et al., 1986; Knutsson, 1992). The heavy metals were released by an "acid surge" of groundwater during snowmelt and after heavy rains in the autumn (Fig. 1) as well as by flushing out of precipitates. Some parts of the aluminium were in quick-reacting inorganic form, which is toxic for fish and some crustaceans (Howells, 1990). The content of inorganic aluminium was highest in the spring water with the lowest pH (4.85) and in one acid stream. Driscoll et al. (1988) found that there is a change in aluminium speciation from largely organic to inorganic forms when water is acidified. The dominating part of manganese in the streams at Djursvallen could also be detected as free metal ions but only a very small part of iron, which instead was humic bound and in suspended form (Jacks et al., 1986). Both manganese and iron were transformed afteiwards, precipitated and found as concretions and coatings on the stream bed. The oxidation and hydrolysis in the stream produced acidity, which accelerated the acid surge. During such periods of "acid surges" in the 1970s and 1980s the fish population of trout and grayling was wiped out and the benthic fauna was badly damaged (Jacks et al., 1986; Melin, 1986). The vast areas of peatland may also have contributed to the acid surges of this catchment, as sphagnum peat can be an important source of hydrogen ions according to Wels et al. (1990) and the humic matter can support the high outflow of metals. A study of the hydrochemistry in crystalline areas of northern Fennoscandia sho ws high concentrations of iron and to some extent also aluminium and manganese in stream water from areas with organic soils. The metals are transported in complexed forms with humic matter (Lahermo, 1991).

Effects on constructions and drinking-water

The most obvious effect of acidification of groundwater is the corrosion of constructions, above all the internal corrosion of water pipe systems, both concrete and metals. The indoor water pipes consist in many countries of copper. Pitting corrosion starts wh«en pH drops below 6.5, alkalinity is below 60 mg HC03 l"1 and the sulphate concentration is higher than the alkalinity (von Brômssen, 1986). In hot water systems corrosion will rapidly lead to leakage of water with serious economic consequences ($ 20 million/year due to acidification in Sweden according to von Brômssen, 1988). Cojrosion in copper pipes is also a large problem in Finland and causes damages to the val tie of 50 million FIM a year (Màkinen, 1989). But the corrosion will also increase the copper content in the drinking water, which may be a risk to the human health: infant diarrhoea is suspected when the copper content in tap water exceeds 1 mg Cu l"1

and fatal liver cirrhosis has been observed as an effect of elevated copper content in drinking water in several countries, for example Germany (Oskarsson & Strinnô, 1990) and India (Bhave et al., 1987). The copper content of Swedish tap water from private wells is high: 54% contain > 1.0 mg l"1 and 25% > 3.0 mg l"1 in standing tap water from 630 wells, but 10% contain > 1.0 mg l"1 and 3% > 3.0 mg l"1 in running tap-water from about 300 wells (Bertills et al., 1989). A study of private wells in rural No va Scotia, Canada showed that in some 50% the copper content was > 1.0 mg I"1

12 Gert Knutsson

in standing tap-water from the investigated wells (Maessen et al, 1984). A more serious observation was that 20% exceeded the 0.05 mg l"1 limit for lead, which means that lead pipes were used in several houses. Still higher content of lead (several mg I'M) has been determined in tap water in Scotland, where lead pipes are commonly used and lead poisoning is a well-known problem (Jacks, 1993).

PROGNOSIS OF RISKS IN THE FUTURE

Groundwater is of outmost importance for the water supply in many countries and the use of groundwater is increasing because of pollution of surface water. At the same time the risks of groundwater contamination is also increasing, among others the risk of further acidification. This risk can be assessed by using different methods. The sensitivity of land and the vulnerability of groundwater to acidification can be mapped on a local or regional scale if there is sufficient information on geology, hydrology, land use and topography (Fig. 4, Jacks & Knutsson, 1982; Holmberg et al., 1990). The time required to acidify a groundwater system may be estimated by hydrochemical budget calculations (Jacks, 1994), which need a lot of data concerning the geochemical and physical properties of the soil as well as the dynamics of soil and groundwater. The necessity to consider all the factors which interact and contribute to the acidification of soil and groundwater, in different risk scenarios for the future, speaks for the use of a numerical, computer-based model. Several models for analysis and prognosis of groundwater acidification have been developed during the last decade (Bergstrôm & Lindstrom, 1989). The models vary in complexity but all types have to describe the interaction between hydrological and chemical processes and the anthropogenic impact

Vajfervik

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o

Ma I mo

Ystad JOkm

Fig. 4 The sensitivity to acidification of shallow groundwater in southern Sweden. The sensitivity was calculated by use of data on type of rock and soil, soil texture, lime content, runoff and relief (from Jacks & Knutsson, 1982) The tighter dots the more sensitive is the groundwater.


as well as to express the processes mathematically. It is, however, necessary to verify the result of most modelling work with reliable field and laboratory data and such data are usually scarce.

MAGIC is a process-oriented model of intermediate complexity with advanced hydrochemistry but a simple water balance routine. It is used to reconstruct and predict long-term trends in soil and water acidification at a catchment scale (Cosby et al., 1985). PULSE is a lumped conceptual model with a rather advanced water balance mo«del (vertically distributed) but a simple hydrochemical submodel (Bergstrôm et ah, 1985). An integrated dynamic model has recently been developed in Sweden (Sandén & Warfvinge, 1992). PULSE, which describes the water content and the water flux in the unsaturated zone, is used in combination with the dynamic soil chemistry model SAFE. This model is a version of a steady state soil chemistry model for acidification sensitivity assessment and weathering calculations called PROFILE (Sverdrup, 1990). Th«e integrated model PULSE + SAFE and the MAGIC model have been used for prediction of the response in the groundwater chemistry to different future acid deposition scenarios. Simulations with both models gave the same overall result: that the acid load must be drastically reduced to stop the acidification of shallow groundwater in sensitive areas. Reductions in acid deposition of 90% (according to the MAGIC model) or reductions of sulphuric acid by 85% and of the total nitrogen deposition by 50% (according to PULSE + SAFE) are required to halt and reverse the ongoing acidification of shallow groundwater in sand (Sandén & Warfvinge, 1992, Fig. 5). The sulphur emissions in Europe began to decrease during the 1980s as well as the content of sulphate in shallow groundwater. But in contrast the nitrogen emissions continued to increase, why the total acid deposition is still alarming. The modelling results as well as hydrochemical budget calculations from field research are-as, indicate that the supply of base cations in sandy till soils in very sensitive areas will be exhausted within one or two decades, if the high acid loads of today continue (Jacks, 1994). Recovery of base saturation in response to reduced acid deposition has been simulated from 1950 to 2070 for two forest soils in southern Sweden by use of the MEDAS-model (Holmberg, 1990). Base saturation in a deciduous forest was slightly

£~* Masbyn J 150

ta O 50

'E3

g o 3

z •a •3 -50 < 1850 1875 1900 1925 1950 1975 2000 2025 2050 Fig. S Calculated acid neutralizing capacity for different future deposition scenarios by modelling. A drastic reduction of the acid deposition is required to halt the ongoing acidification of soil and shallow groundwater in Sweden (from Sandén & Warfvinge, 1992).

1 1 1 1 i

No reduction RedS:30%N:10% Red S:60% N:30% Red S:85% N:50%

14 Gert Knutsson

recovered and than stabilized with a reduction of 60% but in a coniferous forest not even stabilized with a reduction of 80%.

Thus, the prognosis for the near future is very alarming and the acidification of soil and groundwater has to be mitigated. A suitable tool for management of acidification is the critical load concept. The critical load of acid deposition is defined as "the maximum amount of sulphur and nitrogen deposition that will not cause long term damage to ecosystem structure and function" (Nilsson & Grennfelt, 1988). Critical load has been used as a basis for discussions of emission control strategies on a regional scale (Nordic Council of Ministers, 1986). The concept is also useful for developing local strategies for liming of soil and water in order to mitigate the acidification. Critical loads for groundwater can be calculated by making a balance of acidity and alkalinity for a soil column down to the groundwater level. In a Swedish study the following criteria were applied: "The acidity of the water should stay above pH 6 equivalent to an air-equilibrated pH of approximately pH 6.5 and alkalinity above 100 /jBq l"1 at 2.0 m depth" (Sverdrup & Warfvinge, 1992). These limits were chosen as

CL Acidity (50%-tile) Excecdance (50%-tile) • 1.0 to 3.0

Fig. 6 The critical load for shallow groundwater (left) and the present exceedance of the critical load (right) in Sweden. The maps illustrate the critical limit at 2 m depth and the median value (50%-tile) of the sites within each grid (from Sandén & Warfvinge, 1992).


corrosion of water accelerates below pH 6 and alkalinity below 100 /xEq l"1, aluminium concentrations rise at pH below 6 as do the concentrations of heavy metals. The depth of the soil column means the limit for shallow wells and that deeper wells will be protected sufficiently. The calculations were made with the help of the PROFILE model in 1395 different points over the whole country. The results are presented as shaded grids (50 x 50 km) on a map (Fig. 6). It is obvious that some areas in Sweden are very sensitive to acidification, especially the province of Harjedalen in the southernmost part of the High Mountains (cf p. 11), which has very weathering-resistant bedrock (quartzite, quartz-porphyry), coarse-textured tills and podzolic soils.

Special maps are made to illustrate in which areas the critical loads are exceeded and how much. The present deposition largely exceeds the critical load in the major part of Sweden, especially in southwest Sweden (Fig. 6). This means that at first shallow groundwater with many private wells will be acidified and with time also deeper groundwater with public water wells. Calculations with the PROFILE model (Sverdrup & "Warfvinge, 1992) indicate that there must be a huge reduction in the deposition levels (85% of S02, 50% of NOx and 50% of NH4) to reach a sufficient level to obtain a sustainable quality of the shallow groundwater. Over a period of time other measures must be taken to restore soils and groundwater in the most sensitive or most overloaded areas, for example liming (which is already on the way) (Norrstrôm & Jacks, 1993), changes in land-use and methods of cultivation in agriculture and forestry. The critical load maps and the calculations of the steady state chemistry of soils and groundwater can be useful in these types of landscape management.

REFERENCES

Anderson, M. P. & Bowser, C. J. (1986) The role of groundwater in delaying lake acidification. Wat. Resour. Res. 22(7), 1101-1108.

Anidersson, I. & Stokes, J. (1988) Investigation of potential groundwater acidification in Sweden — using principal componentanalysis and regression analysis. Report 3417, National Swedish Environmental Protection Board, Solna, Sweden.

Arends, B. G., van der Sloot, H. A. & van Duijvenbooden, W. (1987) Influence of acidification on the concentration of less common trace elements in Dutch groundwater. In: Vulnerability of Soil and Groundwater to Pollutants (ed. by W. Van Duijvenbooden& H. G. Van Waegeningh). Proceedings and Information No. 38, National Instituteof Public Health and Environmental Hygiene. International Conference Noordwijkaan Zee, The Netherlands, 30 March-3 April 1987.

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