SCOPE 21 -The Major Biogeochemical Cycles and Their Interactions   

ABSTRACT

The industrialized areas of the world have been receiving acid deposition during the last few decades as a result of nitrogen and sulphur emissions originating from the burning of fossil fuels. The acidification of aquatic and terrestrial ecosystems caused by acid deposition will result in changes in those biogeochemical processes that are pH dependent. Thus, acidification represents interactions of man-made perturbations of the S and N cycles with the cycles of C, N, P, and S. Changes that may take place in acidified soils and fresh-waters are: a decrease in the rate of organic matter degradation by micro-organisms in both terrestrial and aquatic systems; a shift to the product N₂O in denitrification; an enhancement of aluminium concentrations which may cause phosphorus to be sequestered in a form that is unavailable for plant uptake.

12.1 INTRODUCTION

The industrialized areas of the world have been receiving acid precipitation during the last few decades as a result of nitrogen and sulphur emissions originating from the burning of fossil fuels (Likens et al., 1981; OECD, 1977). Poorly buffered lakes receiving such acid inputs have become noticeably more acid with consequent damaging effects on the biota, while terrestrial systems, being more resistant to pH changes, have been less influenced. The effects of acid precipitation on terrestrial and aquatic ecosystems have been extensively studied as illustrated in the proceedings of three recent symposia (Drablos and Tollan, 1980; Hutchinson and Havas, 1980; Shriner et al., 1980).

The objective of this paper is to discuss the influence of enhanced hydrogen ion loading on several processes within the biogeochemical cycles of C, N, P, and S. Implicit throughout this paper is the interaction of a man-made perturbation of the N and S cycles with the cycles of C, N, P, and S.

12.2 ESTIMATES OF GLOBAL HYDROGEN ION FLUXES 

A. Anthropogenic Hydrogen Ion Fluxes

A few simple calculations will illustrate the magnitude of man's impact on global H⁺ fluxes. The area of industrial regions receiving pH <4.5 precipitation are presented in Table 12.1; they total about 2% of the earth's land area. With a yearly average precipitation of 0.8 m these regions receive ~25 meq H⁺ m^-2 yr^-1 and a annual H⁺ input of 8 x 10¹⁰ eq H⁺.

Table 12.1 Areas of industrialized regions receiving precipitation having an annual pH less than or equal to 4.5*

Location	Area
		(x 1012 m2)
Norway		0.1
Sweden		0.2
United Kingdom		0.2
Europe (remainder)		0.6
Canada		0.6
U.S.A.		0.9
Asia and other		0.20.6
Total*		2.83.2
		~ 2% of Earth land area
*Compiled from pH isopleth plots in OECD (1977), Hutchinson and Havas (1980) and Drablos and Tollan (1980). Due to the scarcity of data points in isopleth plots the total is accurate only to a factor of 2.

The emission of about 100 Tg S yr^-1 from the burning of fossil fuel and from other industrial processes (Cullis and Hirschler, 1980) yields 6.5 x 10¹² eq H⁺ yr^-1 if it is assumed that all S is oxidized to sulphuric acid. This is a maximum estimate of H⁺ as some of the sulphur will not be oxidized to H₂SO₄ and some H⁺ will be neutralized by ammonia, dust particles or fly ash. Similar assumptions for the 20 Tg N yr^-1 emitted as nitrogen oxides during combustion (Söderlund and Svensson, 1976) yields an additional 1.4 x 10¹² eq H⁺ yr^-1. Due to the relatively short residence times of the precursors of acid deposition, nitric and sulphuric acid will be primarily deposited near (< 10⁶ m) sources while areas farther away will receive relatively little acidity from industrial sources. If it is assumed that 2/3 of this H⁺ derived from N and S emissions is deposited over 10% of the world's area, as is the case for sulphur (Granat et al., 1976; Galloway and Whelpdale, 1980), and if it is deposited wet, then an H⁺ loading of 100 meq m^-2 yr^-1 and a precipitation pH of ~4 results; the remainder of the world would receive 6 meq H⁺ m^-2 yr^-1 at a pH of ~5.1.

Some of the emitted H⁺, N and S may be removed from the atmosphere by dry deposition, which includes gaseous adsorption, aerosol impaction, and gravitational settling of particles. In terms of many of the effects on soils and aquatic systems, the H⁺ liberated by the oxidation of dry deposited S and N will be the same as that of the H⁺ in precipitation.

Dry deposition is difficult to collect directly and therefore must be estimated either via deposition velocities and atmospheric concentrations (cf. Fowler, 1980) or by difference in an ecosystem mass balance (cf. Likens et al., 1977). Because these techniques involve a large effort, the distribution of dry deposition has only been determined over limited areas (Granat et al., 1976; Galloway and Whelpdale, 1980). Dry deposition of H⁺ and sulphur may represent an input to terrestrial ecosystems of the same magnitude as that from precipitation (Likens et al., 1977).

B. Hydrogen Ion Fluxes from Natural Sources

Acid deposition is only one of many H⁺ sources in terrestrial and aquatic systems. Hydrogen ion may be generated by: dissolution of CO₂ from root respiration and microbial decomposition; ammonium and other cation uptake by roots; nitrification; and oxidation of sulphur and nitrogen containing organic matter (Reuss, 1977; Bache, 1980). Several budgets of H⁺ produced from these processes in forested ecosystems have been made (Andersson et al., 1980; Sollins et al., 1980; Ulrich et al., 1980).

An indication of net H+ production in terrestrial systems may be obtained from the amount of cation leaching and weathering. This method can be used because hydrogen ion in soils is quantitatively taken up either in cation exchange or weathering reactions (see section 12.3.1; Lerman, 1979). Global cation leaching and weathering can be estimated from the world average cation load of rivers. Correcting for cations derived from sulphate and chloride mineral dissolution, the cation load in rivers from H⁺ caused leaching and weathering is 1.2 eq m^-3 (Holland, 1978). From an average river flow of 0.4 x 10¹⁴ m³ yr^-1 (Baumgartner and Reichel, 1975) an apparent H⁺ loading of 320 meq m^-2 yr^-1 results (Table 12.2). Cation leaching and weathering rates in areas being affected by acid deposition are in the range 20200 meq m^-2 yr^-1 (Likens et al., 1977; Abrahamsen and Dollard, 1979). This figure is lower than the global average because these regions are composed of crystalline bedrock and its derivatives. Thus the acidity from anthropogenic sources(~25 meq m^-2 yr^-1)may contribute 10% to an amount equal to the total H⁺ production in areas being acidified (Table 12.2).

Table 12.2 Hydrogen ion fluxes from precipitation and from terrestrial H⁺ production

Source	H+ Flux
	(meq m-2 yr-1)
Precipitation
pH 4.5	25
pH 4.0	100
Terrestrial H+ Production
Global Average*	320
Acid Impacted Areas†	20200
*Based on the cation load of rivers, corrected for cations derived from gypsum, halite and other chloride and sulphate minerals (Holland, 1978). See text for discussion.
†Based on cation weathering rates from Likens et al. (1977) and Abrahamsen and Dollard (1979).

12.3 DISCUSSION OF RESULTS FROM ACIDIFICATION EXPERIMENTS

In the past few years there have been a number of experiments designed to quantify the impact of acid precipitation (see, for example, articles in Drablos and Tollan, 1980; Hutchinson and Havas, 1980; and Shriner et al., 1980). The experiments have been primarily designed to give short-term results; rarely have they been run for periods long enough to examine feedback mechanisms. The following is a summary of some of the observed short term effects that may influence the biogeochemical cycles, if not on a global scale then certainly on a regional or smaller scale. The importance of these effects on the global cycles will be estimated. However, the extent to which this can be done is limited by our incomplete understanding of the processes involved.

12.3.1 Effects on Leaching in Terrestrial Ecosystems

The chemical composition of the various inorganic and organic constituents in soil affects the ability of that soil to support plant life and affects the quality of water percolating from that soil into streams and lakes. The increased leaching of the nutrient elements P, Ca, Mg, and K from soil caused by acid deposition may, in the long run, be detrimental to plant growth. The entrainment of these elements along with Al and other toxic metals may have a profound effect on waterways into which the ground-water flows.

The ability of acid precipitation to remove soil components depends on a number of factors including: pH and salt content of precipitation; amount of precipitation and dry deposition; cation exchange capacity and base saturation of soils (i.e. the fraction of basic cations on the exchange sites); surface run-off; and vegetative cover.

Hydrogen ions in rain and soil solutions may be taken up by cation exchange or weathering reactions in soil. The efficiency of H⁺ exchange with other cations depends upon the pH and the base saturation of the soil (Wiklander and Anderson, 1972). Above a soil pH of 4 to 5 there is an exchange of an equivalent of cation from the soil for an equivalent of H⁺ ion in the soil solution and results in an acidification of the soil and a neutralization of the soil solution. Below this range in soil pH only a fraction of the H+ is exchanged and yields partial neutralization of the soil solution.

Because cation exchange is an equilibrium process the relative amounts of H⁺ and other cations in the precipitation and in the soil solution determine whether the soil or the run-off becomes acidified (Wiklander, 1979). If the H⁺:cation ratio in precipitation is the same as that ratio in the soil solution then no net cation exchange and thus no acidification will occur. If this ratio is greater in precipitation than in the soil solution then the soil itself will be acidified and the run-off will become more alkaline. A lower ratio in precipitation than in the soil solution will result in an exchange of H⁺ from the soil sites for cations in the percolating solution. This causes an acidification of the soil solution.

Another factor in cation exchange process in soils is the anion associated with H⁺ in the precipitation. Wiklander (1980) has shown that polyvalent anions enhance the adsorption and decrease the leaching of cations. The effect follows the trend:

Cl NO3 < SO42 < H2PO4 < HPO42

Thus, for example, a switch to nitric acid from sulphuric acid in precipitation may yield more leaching. However, in most terrestrial systems N is limiting to growth and plants will retain nitrate against loss.

Sulphate is preferentially adsorbed by acid sesquioxides in mature soils (Couto et al., 1979; Johnson et al., 1980). Acid deposition of the form H₂SO₄ to sesquioxide-containing soils will result in immobilization of the sulphate and, due to electroneutrality, immobilization of the charge balancing cation and therefore no cation leaching. When sulphate adsorption equilibrium is attained, leaching of sulphate and the exchange of H⁺ for base cations will begin (Reuss, 1980). An important exception to sulphate adsorption by sesquioxides occurs in soils that have both high sesquioxide and organic matter contents. Sulphate adsorption is decreased in this type of soil apparently due to preferred adsorption of organic molecules (Johnson et al., 1980).

When hydrogen ions are taken up by weathering reactions, soil minerals are transformed and metal cations are liberated. The released ions may enter the soil solution, be taken up by plants or be adsorbed on exchange sites. Aluminium (Al³⁺) is one such weathering produced ion common in the solution associated with acid soils (Bache, 1980). Ulrich et al. (1980) suggest that acid deposition in West Germany has caused an increase in aluminium in the soil solution to levels that may be toxic to roots and has caused crown die-back and seedling failure. An additional factor in this study may be the effect of severe drought, which occurred during the same time period. In lakes fed by drainage with high levels of Al³⁺, fish and plant mortality has taken place (see section 12.3.5).

The removal of cations from forest soils in an area of Sweden receiving acid deposition has been documented by Troedsson (1980). In the analysis of exchangeable cations in 2500 plots during the period 19611971, statistically significant decreases in Ca²⁺_, Mg²⁺ and K⁺ and an increase in H⁺ and Al³⁺ in the humus layer of the soil were observed. This study gives an indication that the pH of soil subjected to acid deposition may be declining, although the natural changes in cation content and pH were not quantified.

In work done on an iron podzol treated with ground water acidified to pH 3, decreases were detected in exchangeable Ca and Mg (Stuanes, 1980). The Mg²⁺ and Ca²⁺, in the pH 3 leachate were enriched by a factor of 50 relative to the control. After 5 years of experimental acidification Farrell et al. (1980) found that the exchangeable cation content in a sandy soil decreased by 50% as a result of a pH 2 treatment. Comparison of ions leached from lysimeters containing soils from coniferous and deciduous forests and treated with simulated rain at pH 5.7 and 4.0 revealed 2560% higher Ca²⁺ and Mg²⁺, in the leachate for both soil types at the higher acid addition (Cronan, 1980).

The above discussion has concentrated primarily on the leaching of nutrients from base poor soils. Where acid deposition falls on calcareous soils the hydrogen ion will be neutralized and there will be a slight enhancement of Ca²⁺ and a decrease in bicarbonate concentration in the soil solution (Cole and Stewart, 1981). Nutrient elements, most importantly phosphorus, associated with these soils will also exhibit enhanced leaching and become more readily available for plant uptake or removal by percolating soil water (Lerman et al., 1977). This increased nutrient loading may result in eutrophication of the local receiving waters.

An opposite effect may take place with phosphorus in base-poor soils receiving acid deposition. Aluminium and iron minetals present in these soils may effectively sequester phosphate (Cole and Stewart, 1981), thus making it less available for plant uptake or flux from the soil.

To put these changes in phosphorus cycling into perspective it should be noted that the phosphorus cycle is dominated by a soil pool with a residence time of ~10³ yr (Pierrou, 1976). Increases in fluxes from that pool caused by acid deposition will have a minor effect on the amount of P in the soil over periods of tens of years. Man's use of P as a fertilizer is the main perturbation of the global P cycle (Pierrou, 1976; Wollast, Chapter 14, this volume).

The above examples have clearly shown that leaching is a complex function of a number of soil and deposition variables. Our current understanding does not allow more than a cursory conclusion that the observed leaching of nutrient elements will lead to an impoverishment of the soil's ability to support plant life. Changes suggested for acidified ecosystems have been both positivean enhancement of growth due to increased N loading (Abrahamsen, 1980; Tamm and Wiklander, 1980; Tveite,1980)and negativea decline in photosynthesis (NAS, 1981). Long term effects are uncertain, but the initial enhancement of adding N to N-deficient forests may give way to adverse changes in forest productivity as a result of leaching of Ca²⁺, Mg²⁺, K⁺, and P (Last et al., 1980).

12.3.2 Effects on Decomposition in Terrestrial Ecosystems

In terrestrial ecosystems the cycling of nutrients associated with litter is facilitated by decomposition by fungi, bacteria, and invertebrate animals. Once liberated from the litter these nutrients may be taken up by plants, enter the ground-water and water courses or be emitted into the atmosphere if they have a gas phase. Changes in the decomposing activities of these organisms, brought on by an increase in H⁺ deposition, will affect the rates of nutrient cycling within the terrestrial system (Tamm and Cowling, 1977). Changes in the flux of decomposition products (e.g. CO₂, NH₃, CO, CH₄, N₂O, H₂S) from the soil may affect the atmospheric portion of the various biogeochemical cycles.

During the natural evolution of base-poor soils through which water percolates, the soil becomes more acid (Bache, 1980). Microbiologists in the past have studied the biochemical changes of soils undergoing acidification, particularly in regard to agricultural land (Alexander, 1980b). It is important to point out that the natural acidification of soils is a long term process. The addition of acid deposition will clearly cause an increase in the rate of acidification (cf. section 12.2B). The rapidity with which acidification occurs is likely to have a large impact on both the cycling of nutrients and the ability of that soil to support plant life.

Experiments designed to ascertain the effects of acidification on the decomposition of organic matter have been done mainly in controlled field and laboratory experiments and lysimeter studies. A summary of a few of the experiments, along with the experimental conditions, are presented in Table 12.3.

The average pH of the simulated precipitation used in the cited studies ranged from 4.3 to 3.0. These treatments are more concentrated than pH 4.5 precipitation by up to a factor of 30. Many of the studies used a range of pH treatments extending down to as low as pH 2.5; the results from these higher H⁺ additions have not been included in Table 12.3.

Table 12.3 The effects of acid addition on decomposition of soil organic matter

			Soil pH
Material	Reference*	pH	Control	After	Acid 100
		treatment		treatment	Control
Glucose	1	4.1	7.1	7.1	80105
			5.4	5.1	97
			4.1	4.0	97106
	1	3.2	7.0	5.5	3854
			6.3	5.3	4370
			5.6	4.3	2952
			4.3	3.6	5766
Litter	2	3.0	4	4	115
Leaf Litter	3		4.6	3.0	63
Needle Litter	4		4.6	4.2	88
	4		4.6	4.1	81
	5	3.0			109
	6	3.0	5.3	5.1	98
Humus	5	4.3	4.5	4.3	65
		3.5	2.9	3.6	55
*(1) Alexander (1980b); (2) Roberts et al. (1980); (3) Francis et al. (1980); (4) Bååth et al. (1980), Lohm (1980);
(5) Abrahamsen et al. (1980); (6) Hovland et al. (1980).
†The ratio of the change in decomposition during acid treatment to that in control.

Two further comments on these results are necessary. First, the actual precipitation may be modified upon passage through the canopy of forests. Both neutralization of the incident precipitation (Eaton et al., 1973) and acidification (Mayer and Ulrich, 1980) may take place. Thus in forested areas the soil may receive an amount of H+ that is different from that at the precipitation collection sites. Second, the source of H+ in these experiments was sulphuric acid. Acid deposition also has a nitric acid component (Galloway and Likens, 1981). Uptake of nitrate by plants is accompanied by the release of a bicarbonate ion (Nye and Tinker, 1977, pp 164-167), which would neutralize H⁺, thus decreasing the net hydrogen ion loading.

Even though the pH treatments in Table 12.3 are not entirely realistic, they do give an indication of decreases in decomposition resulting from acidification. In general, a large H⁺ treatment produced larger changes in decomposition. From the pH 3.8 to 4.3 treatments a ~10% (range 35% to 6%) decrease in organic matter decomposition took place. The significance of this decrease in decomposition is clear when one considers that parts of Europe and North America are now receiving precipitation having an annual average pH as low as 4.2 (OECD, 1977; Gravenhorst et al., 1980).

A decrease in decomposition will cause a lower CO₂ flux from the soils to the atmosphere and an increase in soil organic carbon pool size. Taking a mean soil CO₂ evolution rate of 750 g C m^-2 yr^-1 (Schlesinger, 1977) over the 3 x 10¹² m² affected (Table 13.1) yields 2 x 10¹⁵ g C yr^-1. This is about 4% of the global CO₂ flux from soil decomposition (Bolin et al., 1979). If in the future as a result of acidification the decomposition decreases by 10% this would cause a change in CO₂ evolved of 0.2 x 10¹⁵ g C yr^-1 for a 0.4% decrease in global soil CO₂ evolution. Even though this would be a small perturbation of the total soil CO₂ flux, it amounts to some 10% of the uncertainty in the atmospheric CO₂ budget (Bolin et al., 1979; Bolin, chapter 2, this volume).

With a decrease in decomposition and CO₂ evolution rate, the soil organic matter will increase in size. The component of soil organic matter that will be influenced the most by acidification will be the litter, which has the fastest turn-over time (~10 y, Schlesinger, 1977; Bolin, 1981). The litter will exhibit the greatest change in pool size while the longer lived components of the soil organic matter will only be influenced on a time scale of hundreds of years. However, after the initial decrease in soil CO₂ flux bacteria or fungi decomposing at a reduced rate on a greater litter pool may return the CO₂ flux to pre-perturbation levels (Bolin, 1981). This, of course, assumes that there are no lasting toxic effects from, for example, heavy metals, Al³⁺, or H+.

12.3.3 Effects on N-Cycle Processes in Terrestrial Ecosystems

The availability of nitrogen is often the limiting factor in terrestrial plant growth and hence food production. In parts of the world undergoing acidification, the deposition of nitrate in acid precipitation is increasing and may be causing an increase in acidity (Galloway and Likens, 1981). Short term growth rates in terrestrial ecosystems have increased as a result of anthropogenic nitrogen, and perhaps also sulphur, sources (Abrahamsen, 1980; Tamm and Wiklander, 1980; Tveite, 1980) although long term adverse effects associated with acid precipitation, such as limited availability of other nutrients, may ultimately inhibit growth (Last et al., 1980; NAS, 1981).

Acid deposition also effects the nitrogen cycle in that the rates of microbiological transformations of the various N species (N₂-fixation, nitirification, denitrification) may be decreased in acidified habitats. These changes are well documented in the literature of the agricultural sciences (see Alexander, 1980a, for a recent review) although the impacts of acidification on N-cycle processes in other terrestrial systems are not as well known. For example nodulation in cultivated leguminous plants, a necessary precursor to symbiotic N₂-fixation, does not occur in soils having a pH below 4.6 for soyabeans, pH 6.5 for alfalfa and pH 5.2 for clover (Alexander, 1980a).

Recently, there have been a number of attempts to ascertain the effects of increasing acidity on microbiological N transformations in other terrestrial systems. Simulated acid rain of pH 3.2 inhibited Rhizobium nodulation (Shriner, 1977) and N₂-fixation was lowered in epiphytic lichens exposed to synthetic acid rain having pH 4 or lower (Denison et al., 1977). Francis and co-workers (1980) observed complete inhibition of N₂-fixation at soil pHs of 3.6 and 4.7 compared to samples incubated at a soil of pH 5.6. These short term acidification experiments were all performed with N-fixers adapted to higher pH. Over long periods of time other acid resistant strains may be able to nodulate these legumes or the original strains may, with time, acclimate to the lowered pH (Alexander, 1980a).

The lowered activity of legumes may also be a result of changes in the availability of nutrients or toxins. Molybdenum, a necessary trace nutrient for legumes, is transformed into an inaccessible form at lowered pH, while Mn, Al and Fe, which may be toxic to legumes, are in a more soluble form at lower pH (Alexander, 1980a).

Not as much simulation work has been done on the other N-cycle processes, nitrification and denitrification. In short term experiments the abundance and activity of nitrifying bacteria has been shown to decrease as the pH of the medium is lowered, with the activity ceasing below pH 4 (Alexander, 1980a). The number of denitrifying bacteria is inversely related to the pH of soils, although denitrification rates have not been shown to change appreciably upon acidification (Alexander, 1980a, b). In work on a successional sequence of northern hardwood forest stands Melillo et al. (1981) observed that available nitrate and not soil pH, which was in the range pH 3.53.9, may be the dominant factor in denitrification. Thus in naturally acid forest soils, acid resistant strains of denitrifying bacteria may predominate and such factors as temperature, moisture and nitrate content may be more important determinants for the denitrification rates.

With elevated soil nitrification concentrations, the ratio of the products of denitrification, N₂:N₂O, may be a function of the acidity of the medium in which this process occurs, with more N₂O formed at lower soil pH. Firestone et al. (1980) observed a 6-fold increase in N₂O formation for a soil pH decrease from 6.5 to 4.9 at 10 ppm nitrate concentration. Melillo et al. (1981) found that N₂O was the only significant product of denitrification in forest soils having a pH 3.53.9.

The possible impact of soil acidification on the `global N₂O flux into the atmosphere can be seen in the following example, in which a doubling of N₂O emissions from the affected areas will be assumed. Approximately 30% of the global soil N₂O flux originates from forested and boreal ecosystems (~80 x 10¹² m²), with most of the balance emanating from agricultural land (Söderlund and Svensson, 1976). Doubling the N₂O flux from the acid affected areas (Table 12.1) results in a ~2% (2 x (2/80) x 0.3) enhancement of the global flux. Recent work (Isaksen and Stordal, 1981; Crutzen, Chapter 3, this volume) suggests that changes in the N₂O flux of this magnitude will have only a negligible impact on stratospheric ozone concentrations.

12.3.4 Effects on Decomposition in Aquatic Systems

One of the observed effects of lake acidification caused by acid precipitation has been the decrease in the decomposition of autochtonous and allochtonous organic matter (Grahn et al., 1974; Hendrey et al., 1976). As reported in these two papers, this effect was caused by a decline in the number and populations of bacterial species and the rise to predominance of the slower metabolizing fungal species. The decrease in decomposition resulted in a decrease in the turn-over of nutrients within the lakes, which in turn caused a lowered primary production.

Experiments designed to confirm this decrease in decomposition as well as observations of natural lakes have yielded results that are both contradictory and supporting. Gahnström et al. (1980) found that glucose turn-over rates were similar in acidified and non-acidified lakes, although liming of one acid (pH 4.7) lake to pH 6.6 resulted in a significant increase in glucose decomposition. In another comparision of lakes, Andersson et al. (1978) observed that O₂ uptake, and hence decomposition, in sediments from acid and circumneutral lakes was approximately equal. Acid additions to sediments from the circumneutral lakes did not noticeably change these decomposition rates. Sediment pore water pH profiles revealed that although the overlying water was acidic (pH 5) in the Andersson et al. (1978) experiment, sediment buffering maintained a pH greater than 6.5 at 3 cm depth and below in the sediment column.

Traaen (1980) performed both long and short term acid addition experiments and observed that decomposition of glucose, glutamic acid and homogenized leaf litter were greater at pH 7 than pH 5.2 and that decomposition was greater at pH 5.2 than 4.0.

In an experimental acidification of a whole lake using sulphuric acid Schindler (1980) and Schindler et al. (1980) decreased epilimnion pH from 6.6 to 5.6 in 4 years and caused a shift in algal species and a drastic decrease in populations of some animal species. Algal primary productivity remained the same as in pre-acidification years. Rates of decomposition in the hypolimnion were unchanged after the acidification (Schindler, 1980), but the hypolimnion was maintained at pH 6.6 by the products of anoxic microbial activity in the sediments (Cook, 1981). A change in the species of decomposing organisms has occurred in this lake as a result of the increased sulphur loading. Prior to the experiment, methanogenic bacteria dominated the anoxic decomposition of organic matter while sulphate reducing bacteria accounted for less than 20% of the decomposition. As a direct result of higher sulphate concentrations, sulphate reducing bacteria have increased in importance to the point where they now dominate (Schindler et al., 1980). This is to be expected because given enough sulphate and organic matter, bacteria can derive more energy from sulphate reduction than from methanogenesis (see for example Froelich et al., 1979). This species shift will result in a decrease in CH₄ flux to the atmosphere from lake surfaces; H₂S rarely fluxes into the atmosphere from lakes due to its rapid oxidation in water containing oxygen.

An increase in S loading to low-sulphur forest soils, bogs and swamps that have anoxic zones may lead to a similar shift from methanogenic bacteria to sulphate reducing bacteria due to the greater energy derived from SO₄^2-reduction. If this takes place then a decrease in CH₄ flux and increase in reduced S flux to the atmosphere may ensue. As soils, bogs and swamps produce less than 5% of the CH₄ (Ehhalt, 1974) and reduced S fluxes to the atmosphere (Granat et al., 1976) these enhanced S loading effects on the global C and S cycles will be small but changes in the fluxes of CH₄ and H₂S may be of significance in the affected regions.

A similar conclusion is reached regarding the effect of anthropogenic-acid-induced changes in overall decomposition in lakes and streams. As only a total of 2 x 10¹² m², or 1.3% of the land area, is taken up by lakes and streams these changes will not influence the global cycles to an appreciable extent but will only yield effects within the affected lakes or streams.

12.3.5 Effects of Aluminium in Aquatic Systems

Aluminium is an ion mobilized from bedrock and soils by acidification that is important to nutrient cycling in fresh water ecosystems. Monomeric aluminium may complex phosphorus making it unavailable for algal uptake. It may also cause a change in natural nutrient cycling by removing the top predatorfishfrom the food chain.

Soluble aluminium may exist in three forms: either monomeric (A1³⁺); in inorganic complexes with OH, SO₄² F; or in organic complexes (Driscoll et al., 1980; Driscoll, 1980). In Driscoll's (1980) study, organically complexed Al was the most abundant ion at all pH while the monomeric form increased in proportion to the others at lower pH. In the pH range 4.55.5, Al³+ acts as a precipitant for Humic matter (Almer et al., 1978) thus removing it from the water column. Humic material plays an important role in lakes as a source for some minor nutrients and as a chelating agent for toxic metals.

Aluminium may also precipitate phosphorus with a minimum solubility at pH 5.5 to 6.0 (Stumm and Morgan, 1970; Almer et al., 1978). Thus, Al may limit the availability of P, although at pH <5.0 phosphorus becomes more soluble. Almer et al. (1978) investigated a suite of lakes and found a trend of biomass with pH in which biomass has a minimum at pH 5.15.6. They relate this to the solubility of aluminium phosphate and suggest that enhanced levels of Al³+ may limit the accessibility of phosphorus in this pH range. However, because of the inverse relationship between Al³⁺ and PO₄^3- in the solubility expression further elevation of aluminium concentration as a consequence of acidification may decrease the PO₄^3-  concentration at lower pHs (Hendrey, 1980).

In acidified lakes, hydrogen ion is toxic to fish because it upsets the salt balance, and Al³+ is toxic to fish because it precipitates as an aluminium hydroxide when it comes in contact with the mucus layer in fish gills (Muniz and Leivestad, 1980a, b). When the top predator in a food chain is lost, zooplankton and other invertebrate plankton predators on the next trophic level become larger and more abundant due to the lack of grazing (Stenson et al., 1978; Henrikson et al., 1980). In their study of a lake in which all the fish had been removed, Stenson et al. (1978) observed an increase in plankton grazing, a decrease in numbers of algae and a decrease in primary production. A larger, less efficient algae became more abundant, causing the decline in algal numbers and primary production. Such larger organisms recycle nutrients at a slower rate, thus compounding the effect of a decrease in photosynthesis.

The decrease in nutrient cycling resulting from fish mortality may lead to a net retention of organic matter and nutrients in lake ecosystems. Due to the small fraction of the globe covered by lakes this effect will only be of local importance and not influence the global nutrient cycles.

12.4 DISCUSSION

A summary of the effects of acid deposition are collected in Table 12.4. The effects fall into two main categories: either those resulting from changes in bacterial decomposition or those stemming from changes in leaching. Nearly all of the secondary effects have been observed in simulation experiments.

One might expect from an examination of Table 12.4 that a tertiary effecta decrease in productivity in both terrestrial and aquatic systemswould be apparent. Decreases in nutrient abundance as a result of enhanced leaching and decreased decomposition of organic matter would seem to lead to a decrease in productivity, although no such phenomenon has been convincingly demonstrated within an ecosystem (Last et al., 1980). Enhanced nitrogen, and perhaps also sulphur, abundances may counterbalance some of the adverse effects of acid deposition, but it is more likely a matter of the ability of the ecosystem, in particular the soil, to buffer against changes in acidity. The complexity of terrestrial and aquatic ecosystems makes the task of understanding their responses difficult.

The addition of calcium carbonate to acidified ecosystems has been widely endorsed as a means of counteracting acid deposition. Hydrogen ion is taken up in a reaction yielding HCO₃^- and CO₂, the former of which then increases the alkalinity and hence the acid neutralizing capacity of these systems. The removal of excess H⁺ in this way will counteract many of the effects listed in Table 12.4, although the deposition will still be acidic. Enhanced nitrogen and sulphur deposition would still occur, and the effects of increased loading of these elements would be the same as those before liming.

Table 12.4 Summary of possible effects of acid deposition on the cycles of C, N, S, and P

Despite the limitations in our knowledge of acid deposition, it is worthwhile to attempt a prediction of future distributions of acid deposition and the probable impact on soils and the fresh-water system. Let us assume that in the year 2025 A. D. 2.8 times the fossil fuel will be burned compared to that in 1975 (Table 1.3, chapter 1, this volume) a range of estimates is from 215 times the 1975 consumption rate for annual increases in consumption of 1.56.5 % (Bolin et al., 1979). For the purposes of this calculation a proportionality between fossil fuel burning and SO₂ and NO_X emissions will be assumed.

Due to the competition between NO_X and SO₂ for oxidants, NO_X will be oxidized more quickly and the concentration of SO₂ will increase, causing more dry deposition and a greater dispersal of sulphur (Rodhe et al., 1981). Also the oxidation of SO₂ is inversely related to water droplet pH (Penkett et al., 1979; Taylor et al., Chapter 4, this volume), thus delaying the water phase oxidation of SO₂ at low cloud water pH. These factors will probably result in an increase in the area influenced by acid precipitation, not a increase in the H⁺ concentration in the areas now being influenced (Rodhe, 1981). Future industrialization of Third World countries will probably cause enhanced SO₂ emissions in areas not currently receiving high anthropogenic inputs. Thus the area receiving precipitation having pH <4.5 will increase in the future, barring any regulated decrease in N and S emisions in the industrialized world.

Many of the changes brought on by acid deposition could cause perturbations in the global biogeochemical cycles of 0.1 to 2% as described above; effects on the cycles in industrialized regions will certainly be higher. With an increase in the area affected, these perturbations will also increase, although our ability to predict the changes awaits the results of further studies of the long term effects of acid deposition.

The total mass of S associated with the estimated available fosil fuel resources (Bolin et al., 1979) is 50 x 10¹⁵ g, if a 1% S content is assumed. This represents the total input of 3 x 10¹⁵ eq H⁺ if all this fuel is burned and all the S associated with it is emitted and converted to sulphuric acid; our current H⁺ emission is 8 x 10¹² eq H⁺ yr^-1 (section 12.2). The cumulative, long term effects of this potentially-emitted acid upon the fresh-water and soils can not be estimated because we do not have a full understanding of the medium to long-term responses of soil and fresh-water systems to acidification.

The above calculations, although speculative, are a first attempt at assessing the impact of acid deposition on the cycles of C, N, P, and S. Clearly, more work needs to be performed on the impact of dry as well as wet deposition and the behavior of SO₂ and NO_X in the atmosphere before greater confidence may be placed in determinations of effects on the cycles of C, N, P, and S and predictions of future effects.

ACKNOWLEDGEMENTS

This work has benefitted from discussions with many people, especially J. Melillo, P. Vitousek, H. Rodhe, F. Andersson and B. Bolin. Comments from T. C. Hutchinson, G. Abrahamsen and J. Galloway on an earlier draft are gratefully acknowledged. Financial support was provided by a post-doctoral fellowship from the A. W. Mellon Foundation.

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