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Acidification

Background

Acidification is a natural process. The term is used to describe the loss of nutrient bases (calcium, magnesium and potassium) through the process of leaching and their replacement by acidic elements (hydrogen and aluminium). However, acidification is commonly associated with atmospheric pollution arising from anthropogenically derived sulphur (S) and nitrogen (N) as NOx or ammonia. Anthropogenically derived pollutant deposition enhances the rates of acidification, which may then exceed the natural neutralising capacity of soils. The environmental impacts of acidification are one of the major contemporary environmental issues, both in the UK and globally. Nevertheless, even with no anthropogenic contributions to N and S deposition many soils, particularly those derived from acid parent materials, would still acidify over time. However, under such conditions soils would eventually reach a steady-state situation with any unneutralised acidity exported and neutralised further down the catchment. Acidification affects all aspects of the natural environment: soils, waters, flora and fauna.

Soils will acidify if there is (i) a source of H⁺ ions to replace base cations removed by ion exchange processes (ii) a means of removing the displaced base cations, achieved by a mobile anion such as sulphate (SO₄^2-) or nitrate (NO₃^-) (Billett, 1994). Weathering of parent material is the main way in which cations are replenished, but other soil processes such as adsorption and microbial reduction of SO₄ can also help to ameliorate acidification.

An important consequence of acidification is an enhanced level of aluminium (Al₃⁺) in the soil solution. This increased Al₃⁺ mobility can in many cases have significant effects on ecosystems (Bareham, 1996). High levels of soluble Al₃⁺ at low pH values disrupt cell wall structure in plant roots and inhibit nutrient uptake (Kennedy, 1992). Al₃⁺ can also kill earthworms at high concentrations and leach into water, affecting aquatic life (e.g. Battarbee, 1984). The aluminium chemistry of soils is very complex (see, for example, Abrahamsen (1984); Marschner (1991)) and effects on plants can vary even among ecotypes.

Acid precipitation may increase human exposure to toxic metals via food and drinking water with potentially adverse effects on health (Oskarsson et al., 1996). Despite there being a well established correlation between metal exposure and health effects in general, the precise contribution of acidification to this remains to be shown.

Acidification trends

Acid deposition and the effect of acidification on soils and plants have been observed since the mid 19th century and studied, mainly in relation to lakes and rivers, since the 1920s (Erisman and Draaijers, 1995). It is only in the last 25 to 30 years that acidification has become perceived as a major environmental issue, due to the obvious forest decline in parts of Scandinavia and central Europe (e.g. Hallbacken and Tamm, 1990; Brydges and Wilson, 1991; Last, 1991; van Lynden, 1995). Although there is much evidence to suggest causal links between acid deposition and mineral soil damage, these links are often difficult to prove unequivocally at the regional scale, because pollution gradients are usually strongly associated with climatic gradients (Erisman and Draaijers, 1995). There is also the role of soil/parent material heterogeneity (White et al., 1995).

The acidification of soils and freshwaters in regions remote from major sources of pollutants may have been accelerated by the increased dispersion of acidic pollutants caused by changes in legislation, fuel usage and combustion technology, which have resulted in the increasing importance of high level emissions from power stations (Lee, 1998).

A number of studies report increases in soil acidity over various timescales. In Scotland, forest soils in the Alltcailleach catchment of the upper Dee have been monitored over a 40 year period (e.g. Billett et al., 1988; Billett et al., 1990). In 80% of the surface organic horizons, pH fell between 0.07 and 1.28 pH units; in 73% of the mineral horizons below 40 cm, the decrease was between 0.16 and 0.84 pH units. Acid deposition was thought to play only a minor role in this process, most of the change being attributable to tree growth and natural soil/vegetation changes. This highlights the complexities involved in making direct links between acid deposition and ecosystem change in naturally acid soils. Nevertheless, in a similar study in north Wales, pH decreased at all sites measured, with extent of acidification broadly inversely related to initial degree of acidification (Kuylenstierna and Chadwick, 1991). In this case, the majority of change was attributed to industrial emissions.

The natural pH of Scottish organic peat soils has declined by about 0.5 pH units as a result of acid deposition, leading to a drop of about 1.1 pH units in drainage waters. Peats with highest acidity and lowest base saturation often correlate with areas where atmospheric deposition is highest (Skiba et al., 1989; Cresser et al., 1993).

'Clean rain' experiments, whereby pollutant and nutrient inputs to an ecosystem are manipulated by means of a roof cover, are often used to remove the acidifying inputs of S and N (e.g. Beier et al., 1995). Boxman et al. (1995) found that the soil solution responded rapidly to removal of acidifying inputs, with substantial decreases in S, NO₃^-, and NH₄⁺. There was also improved nutritional balance of K⁺, Mg²⁺ and P. However, no changes in tree growth were observed over the course of the work, suggesting that there is probably a significant time lag between improvement in conditions and recovery. Filtration experiments carried out by the Forestry Commission have shown that exclusion of gaseous pollutants improved the growth of beech and Norway spruce at a site in southern England, but not at sites in the Pennines or Perthshire (DoE, 1993).

Assessment of acidification

The sensitivity of an ecosystem to acidification is determined by the rate of supply of base cations relative to their loss. If base cation supply from chemical weathering and atmospheric inputs does not equal or exceed loss to biomass and runoff, this will lead to a gradual decline in base saturation and a reduction in nutrients available to plants will occur (Langan et al., 1994). Forest and moorland ecosystems will eventually show signs of nutrient deficiencies and imbalances, with a consequent decline in growth.

Numerous schemes have been used to classify and assign different sensitivities to geographical areas and ecosystem types. One approach used for identifying acidification vulnerable surface waters in the UK has been to rank geology and soils according to their ability to neutralise incoming acidity. Other schemes have used chemical criteria, such as aluminium concentrations, to identify the potential onset of damage to acid sensitive ecosystems.

National and international abatement strategies aimed at lessening the role of atmospheric pollutants on acidification processes are utilising the work done on developing critical loads and their exceedance. This approach identifies areas in which geochemical controls.

Acidification and Nature Conservation

Although there have been several reviews of acidification of the natural environment by anthropogenic sources of S and N (e.g. Last and Watling, 1991; Moffat, 1991; Woodin and Farmer, 1991, 1993; Heij and Erisman, 1995), to date there have been relatively few attempts to quantify the specific damage to designated protected sites in Britain, nor to relate this to emission sources (Pearce, 1993). Despite difficulties in proving conclusively the ecological effects of air pollutants - lack of reliable historical data, difficulties in identifying individual causes of change, restricted extent of field surveys and uncertainties involved in extrapolation - the weight of evidence suggests that enhanced S and N deposition is causing damage to a wide variety of habitats, communities and species in Britain (Press et al., 1986).

Woodin and Farmer (1993) compared the distribution of key conservation sites as listed in the Nature Conservation Review (Ratcliffe, 1977) with pollutant deposition maps, and found that c.50% of the key upland grassland, peatland and heathland sites in Britain receive an annual input of >10 kg S ha^-1 in precipitation. Similarly, about 60% of key lowland grassland, heath and scrub sites receive more than 20 kg N ha^-1 annually. In both cases, damage to a range of ecosystems is an actual or potential issue.

In 1993 the Department of the Environment commissioned the statutory conservation agencies in Britain to examine the potential impacts upon nature conservation of several pollution reduction scenarios using critical loads maps. The work is described in separate reports for Scotland (Ross, 1993), England (Bissett and Farmer, 1993), Wales (Sketch and Bareham, 1993) and jointly for England and Wales in a further report (Farmer and Bareham, 1993).

Surface water acidification

Acidification of surface waters can result from direct deposition of pollutants into lakes and streams or, more commonly, through runoff and soil throughflow from the surrounding catchment. High mountain lakes are more susceptible to atmospheric inputs than lowland lakes, due to e.g. climate, shallow soils, small watersheds and rapid flushing rates (Mosello et al., 1995). Diatoms are very sensitive to acidification, and recorded changes in species assemblages from sediment cores can provide a great deal of information about the progression of lake acidification (Battarbee, 1984). Prior to the onset of the 19^th Century, diatom assemblages suggest a little or no change in water acidity. On a shorter timescale, there is evidence of changes in invertebrate (principally mayflies) abundance and salmonids in relation to different pollutant loadings.

Current knowledge and observations suggest that S concentrations in surface waters are reversible, given adequate time (Dise et al., 1994). The process may be significantly delayed in wetlands however, where S accumulated over long periods may be released only very slowly by oxidation. For N, both positive and negative changes tend to be episodic rather than gradual, although N levels in freshwaters have to date rarely been observed to decline under natural conditions - most conclusions are based on manipulation experiments.

Prediction of the recovery of freshwater ecosystems is largely based upon the simulation of future hydrogeochemical conditions that are compatible with healthy, reproducing biological populations. Direct simulation of biological population status has not often been attempted, because the procedure has to be carried out in conjunction with geochemical simulation models, and the two approaches are very different (Cosby et al., 1994).

Groundwater acidification may occur in severely affected regions. The extent of the effect depends on a number of factors, including e.g. land use, soil characteristics, flow patterns, precipitation characteristics, and depth to the groundwater table. Seasonal variations in acidity and the magnitude of acidification are generally lower in groundwater than in surface water. For instance, in Bavaria the pH of groundwater in granitic areas only occasionally drops below 4.0 and is usually between 4.0 and 5.0, whereas acidic surface waters in the same areas normally have pH values less than 4.0 in spring and over 5.0 in summer (Landers et al., 1994).

As an integral part of the freshwater ecosystem, groundwater can also contribute to acidification of rivers and lakes through direct discharge. However, a major difficulty associated with studying spatial and temporal changes in groundwater acidification is soil heterogeneity. In particular, surface flow path distribution is complex and affects residence times of infiltration water in different soil horizons.

Emission abatement

Future trends in S emissions, and subsequent deposition to terrestrial and aquatic ecosystems, will continue to decline substantially in Europe and North America, as emissions are cut to achieve the targets set by the Oslo Protocol. Other protocols to reduce emissions leading to acidification are the "multi-pollutant, multi-effect" protocol, or Gothenburg Protocol, and the National Emission Ceilings Directive (NECD), which came into force on 27 November 2001.

These reductions will lead to a reduction in the area of critical load exceedance. As suggested above, these reductions may not necessarily be directly reflected in benefits for habitats and biota. Even where positive changes do take place, there may be considerable time lags, before reduced acidity has any effect on ecosystem processes. In terms of predicting recovery rates, more information is needed on how different species react to changes, both individually and in their relationships with each other. There is also the concern that increased N emissions may cancel out any gains produced by reduced S levels, and this issue will require close monitoring in the future. Studies of the biological response to reductions of S deposition in the UK suggest recovery is geographically patchy.

Assessment of the long term trend in acidification, using dynamic geochemical models, suggests that changes in deposition will give rise to a slight recovery before the onset of further acidification. At a wider level, concern is related to the growing role of nitrogen in the acidification process.

Acidification impacts may also be ameliorated by liming.

References:

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