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 HC0 3 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,
N0 3 , via NH4 by nitrifying bacteria. Acidification of groundwater occurs, if the
vegetation does not adsorb all the N0 3 , 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
Acidification effects on groundwater
7
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
jA
ïXiM^
5.2
4.4
4.0
TTTTTTTTTTTTTTTlTTTrrr
1
1987
1988
TrnnT"n"TTTr
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.
Kronoberg county
Halland county
2
2
2
Ca ' + Mg *
mg/l meq/l
2
Ca % Mg *
mg/l meq/l
120
Ca 2 *+Mg 2 *
mg/l meq/l
80- - 4
240
Blekinge county
V
K /
40- -2%
y
fI o•
/y
H7
o-
I
2
h120
1_
4
1
240
1 1
meq/l
HC03mg/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).
Acidification effects on groundwater
9
(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
90 92
80 82 84 86
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 N0 3 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
Acidification effects on groundwater
11
(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 HC0 3 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
Gert Knutsson
12
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
Goteborg
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.
Acidification effects on groundwater
13
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
£~*
J
Masbyn
150
1111
i
No reduction
RedS:30%N:10%
Red S:60% N:30%
Red S:85% N:50%
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).
Gert Knutsson
14
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).
Acidification
effects on
groundwater
15
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 S0 2 , 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.
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