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Input-output relations of major ions in European forest ecosystems
Agriculture, Ecosystems and Environment, 47 ( 1993) 175-184
175
Elsevier Science Publishers B.V., Amsterdam
Input-output relations of major ions in European forest ecosystems A. Liikewille*, M. Bredemeier, B. Ulrich Institute of Soil Science and Forest Nutrition, UniversityGi~ttingen, Biisgenweg2, W-3400 G6ttingen, Germany
Abstract
A framework for the assessment of the state and trend of forest ecosystems from ion budget data is presented. Atmospheric deposition is regarded as an input and seepage water as a system output. A simple compartment model, subdividing the forest ecosystem, is used to calculate ion balances and the corresponding internal H + production/consumptionprocesses. The 'buffer range' of the forest soil substantially determines the reaction of a system to input and internal generation of acidity. Forest soils in the carbonate buffer range are characterised by high seepage water pH (6.2-8.6). High concentration levels of HCO3 are balanced by alkali and alkali earth ions (Mb or 'base' cations). In the cation exchange buffer range, pH values remain above pH 4.2 and concentrations of dissolved aluminium are low. With increasing proton load the chemical characteristics of the soil solid phase are changing. The consequences are a decrease in acid neutralisationcapacity (ANC) and base saturation (BS), and the formation of interlayer A1. At BS values below 10-15%, H + is buffered almost solely by the dissolution and exchange of aluminium (and/or iron) compounds. Acidity, mainly as A13+and H +, is transported with seepage water and thus influences deeper soil and bedrock layers. Examples of the mentioned budget types are given, utilizing data from long-term case studies in Germany.
Introduction In most forest soils in Central Europe, acid deposition exceeds internal proton loading (Van Breemen et al., 1984; Bredemeier et al., 1990). Forest ecosystems can be regarded as three-dimensional components of the ecosphere which are subdivided into compartments. The parts are interconnected by fluxes of material. The element budgets discussed in this paper represent a process-functional approach in viewing ecosystems. Organisms and their environment are considered as a single integral system. Finally, this approach implies that macroscopic phenomena like energy flow and nutrient cycling are more fundamental than the biotic entities performing the functions (O'Neill et al., 1986). A mass balance equation defining fundamental pro*Corresponding author.
176
A. L iikewille et al. / Agriculture, Ecosystems and Environment 47 (1993) 175-184
ccsses like photosynthesis, ion uptake, build-up of organic substances and the opposite processes of respiration and mineralisation of organic substances can be used to describe main fluxes of material in forest ecosystems (Ulrich, 1989a)
aCO2 + x M + +yA - + ( y - x ) H + + z H 2 0 + energy(hv) ~- ( Ca H20zMx Ay)organic matter "~"a O 2
( 1)
where a, x, y, and z are stoichiometdc coefficients, M + represents cations, and A - anions of unit charge, and organic matter includes organisms as well as humus. Assessment of ion fluxes in forest ecosystems Figure 1 shows a simple compartment model, used to quantify the most important internal net proton production processes (Bredemeier et al., 1990). In 'steady state', there is almost no net turnover (production/consumption) of H + ions. The actual system state can be assessed by looking at deviations from quasi steady state conditions (Ulrich, 1989a,b, 1991 ). To investigate the relationship between deposition and seepage of certain substances, the deposition rates of major ions are regarded as an input into the forest ecosystem (arrows 1 and 2 in Fig. I ). The fluxes of ions with seepage water below the rooted soil layer represent a system output (arrow 8 in
t, ~..AVno, pher~ Depo~on Atmo,p h ~ BIomm pecenn~ friction
'.mc:t~mnt)
t
,4)'~
t
IP" T (,) '
. Humus Layer
(¢)
V
('/)
M i n e r a l Soil
(70)
V
(m)
Outla~-$~la~leWete¢
Fig. 1. Simple compartment model of a forest ecosystem.
A. Lflkewille et aL / Agriculture, Ecosystems and Environment 47 (1993) 175-184
177
Fig. 1 ).The assessment of these input and output fluxes across defined boundaries requires knowledge of the hydrology which can be calculated using mathematical models (Hauhs, 1986). Another important input into the ecosystem is the (irreversible) release of ions due to mineral weathering (arrow 10 in Fig. 1 ). The general reaction of the system to the input of acidic deposition (on a time scale of years) mainly depends upon the chemical soil state, the amount of Mb cations, NO~- and NH~ bound in the annual biomass increment, the rate of organic matter stored in the soil and the quantity of drainage water. The input/output relations of acidity in terrestrial soils are the result of an interaction of external and internal hydrogen ion sources with the soil solid phases. Internal proton generation takes place within the soil which, however, must be considered as a 'black box'. Consequently, these processes cannot be measured directly. Their rates are assessed by means of compiling complete ion budgets of the soil. The method to calculate internal H ~ transfer processes is described and discussed in detail by Bredemeier et al. (1990). Soil internal net H + generation may occur as a result of the following biogeochemical processes: biomass/humus aggregation; dissociation of carbonic acid and of organic acids; net nitrification of organic nitrogen from the N pool of the ecosystem; assimilation of a surplus of deposited NH~- over NO~- and/or nitrification of deposited NH~-; buffering of deposited H + on canopy surfaces with subsequent regeneration of foliage buffer capacity (i.e. the uptake of Mb cations from the soil). These processes, designated as internal net proton production (INP), contribute together with the free acidity (H + ) in throughfall (see Fig. 1, arrow 4) to the total proton load (TPL) of a forest soil. Microbially mediated reduction of sulphate and nitrate in the soil is neglected. The relation between TPL and H + buffering due to mineral weathering determines the trend of development of soil chemical properties.
Buffer ranges in soils
The relation between acid neutralisation capacity (ANC) and pH in a mineral soil affected by acidic deposition can be represented as a titration curve (Van Breemen et al., 1983; Prenzel, 1985; Van Breemen, 1991 ). Figure 2 shows pH as an intensity factor and ANC as a capacity factor. A ANC denotes the rate of soil acidification. The pH limits given in Table I were calculated by assuming chemical equilibria between ion species in the soil solution and well defined solid phases. Thus deviations from these fixed values may occur.
178
A. Lf~kewille et aL /Agriculture, Ecosystems and Environment 4 7 (1993) 175-184
pH; intensity factor A
7 -
6
5
4
3
H + added, A ANC; capacity factor
Fig. 2. Sequence of possible buffer reactions changing the composition of the soil solution; quick acidification ( ), slow acidification ( . . . . ). Table 1 Buffer ranges in soils Buffering substance/process
pH range
Dissolution of carbonates Exchange of Mb cations Exchange of aluminium ions, dissolution of A1compounds Aluminium, iron compounds Iron compounds
8.6>pH>6.2 6.2>pH>4.2 4.2>pH>3.8 3.8>pH>3.2 3.2>pH
The plateau A-B (Fig. 2) represents the carbonate buffer range. The dissolution of CaCO3 is the predominant buffering process. Ca 2+ and H C O £ are the major ions in the soil solution CaCO3 + 2H + ~ Ca 2+ + 2HCO~-
(2)
The slope B-C shows the exchange buffer range. Mb ions adsorbed at the cation exchange complex serve as a buffer, e.g. 2H + +Cae2~+ ~ 2 H ~ + C a 2+
(3)
Additionally, the reaction of H + with soil minerals leads to the formation of positively charged A1 ion species, which in turn displace Mb cations from the exchange complex (build-up of interlayer Ah n [AI(OH)x~3- x) ] ). Parallel to this process, base saturation decreases, e.g. KA1Si308 + H + + H 2 0 ~ K
+ +AI(OH)3
+ 3SIO2
(4)
A. Lfikewille et al. / Agriculture, Ecosystems and Environment 47 (1993) 175-184
KAISi30s + 4H + + Cae2~~ K~ + A13x+ + 2H20 + 3SIO2 -I-2Ca 2+
179
(5)
Equations (4) and (5) show that the rate of silicate weathering influences the effectiveness of the exchange buffer considerably. Therefore, it is a very important kinetical constraint. In the case of quick acidification of a soil with given exchange capacity, base saturation and CO2 partial pressure, the slopes of the curve (Fig. 2 ) is extremely sensitive to the release of Mb cations through silicate weathering. The pH of the soil solution is dependent on CO2 partial pressure, which is particularly influenced by respiration processes. Owing to the pK~ values of the H2CO3 dissociation equilibria, pH levels below 5.0 indicate negligible HCO£ concentrations in the soil solution, even at high Pco2 values (e.g. 3%; Reuss and Johnson, 1986; Reuss, 1991 ). The interval C-D (Fig. 2) represents the aluminium buffer range. As base saturation decreases below about 15%, the pH drops to values less than 4.54.2 and the A1 buffering takes effect. A13+ is then the prevalent cation in the soil solution A1(OH)3 + 3H + ~A13+ + 3H20
(6)
If soluble A1 compounds are depleted, or if their dissolution is restricted by kinetic constraints at a chosen time scale, the soil reaches the iron buffer range.
Acid neutralisation capacity (ANC) and sulphur retention Since pH is an intensity factor, the acid neutralisation capacity should also be considered. The ANC of mineral soil material can be defined as the sum of ANC of the solid phase plus ANC of the aqueous phase (Van Breemen, 1991) ANC (S) = ANC (s) +ANC (aq)
(7)
The ANC (s) includes, for example, CaO, MgO and 1(20 as components of silicates (reference pH of 5.0) or A1 compounds (reference pH of 3.0). If ANC(aq) in the soil solution is removed by the percolating water, total ANC(S) decreases (Ulrich, 1991; Van Breemen, 1991). Soil acidification can be defined as a loss of acid neutralisation capacity. The principle of charge balance implies that ANC (aq) or alkalinity may be defined as the sum Of Mb cations minus the sum of strong acid anions (Reuss and Johnson, 1986; Reuss, 1991 ). Organic anions are neglected ANC (aq) =2(Ca 2+ ) +2 (Mg 2+ ) + (Na + ) + (K + ) - 2 ( S O 2- ) - ( N O ; ) - ( e l - )
(8)
parentheses denote molar concentrations/activities and ANC(aq) is expressed in/zmolc 1-1.
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A. Liikewille et al. /Agriculture, Ecosystems and Environment 47 (1993) 175-184
Owing to exchange processes in the soil matrix, the concentration of M b cations in the soil solution is dependent on the amount of dissociated H2CO3 and, consequently, on CO2 partial pressure. ANC(aq) is a useful means of investigating acidification processes. In whole-catchment experiments, water not in contact with the bedrock matrix degasses (CO2), resulting in a considerable increase in stream/lake water pH, whereas alkalinity (as defined in Eq. (8) ) remains unchanged (Reuss et al., 1987; Liikewille, 1990). The same phenomenon must be considered if pH values, which were measured in potentially degassed soil solution samples, are interpreted. At high concentrations of dissolved A1 species, ANC (aq) becomes negative. The expediency of a calculation with negative capacity values ( - A N C ( a q ) ) is questionable. For an assessment of the ecological consequences of soil acidification, e.g. the effects on tree roots, the consideration of intensity values such as pH and concentrations of potentially toxic ions such as A13+ is more suitable (Ulrich, 1991 ). If the observed sulphur retention in soils is considered it is difficult to distinguish between SO2- adsorption processes and the precipitation of AI hydroxysulphate minerals. The following equations may be considered as model reactions for more complex processes of chemiadsorption and/or solid formation (Matzner and Prenzel, 1991 ). Formation of alunite 3Al(OH)3 + 1.5H2504 + 0.5K2 SO4--}KA13(OH)6 (504) 2 + 3H20
(9)
Formation ofjurbanite A1(OH) 3 "~-H2 S O 4
---}A1OHSO4
+ 2 HE O
( 10 )
Sulphur retention must be regarded as a delay mechanism: acidity, or base neutralisation capacity (BNC (s) ), is stored in the soils and may become effective under changing input and/or ecosystem-internal conditions. Results of long-term case studies
Data from almost 30 long-term case studies in Central Europe were used to group forest ecosystems into budget types. Detailed information about the study sites is given by Ulrich ( 1989a,b, 1991 ) and Bredemeier et al. (1990). The buffer ranges serve as a key to understand the relation between deposition and seepage output of major ions. Figures 3 and 4 summarise representative examples for the different budget types. The units for the input and output fluxes are kmolc ha-I year-~. The data are mean values of several years. The first example represents a forest ecosystem with soils in the carbonate buffer range (G6ttinger Wald, beech). Figure 3 shows that the internal H + load far exceeds the input by deposition. Since HCO£ is the major ion in the
A. Liikewille et al. /Agriculture, Ecosystems and Environment 47 (1993) 175-184
181
Case Study
_
C-ddtingerWald,Beech
- 12.9 ///-///////////////,
~
+ 10.0
"/////////////
L Heide, Oak
/////,/.//III//,
~\\\\%.~,\\\\\\\
Hamto,Beech i
i
-4
t
-2
i
0
•
H+ deposited
•
[]
Ma output
[ ] Mb output
H+ internal
2
kmol c -
ha-I.
•
S input
4
yr-1
[ ] S output
[]
N input
[]
N Output
Fig. 3. Input-output relations of major elements in different forest ecosystems in Central Europe.
Case Study
,J
L Heide, Pine
---,.l Soiling, Beech [*;,~"~',~\\\\ \ \ \ \ \ \ \ ~ L
Soiling, Spruce I
till/ill;';.;'7, ,\\\\\\\\\\\\\\\\\\\\\\7
k\\\\\\\\\\' i
-6
i
-4
-2
0
l~nol c • ha-1. •
H+ deposited
[ ] M.o.~,,
•
H+ in~rnal
[ ] Mbo°~t
•
i
2
i
4
yr-1
$ input
[ ] s o.~ot
[ ] N input
[] N ou~ut
Fig. 4. Input-output relations of major elements in different forest ecosystems in Central Europe.
182
A. Lffkewille et al. / Agriculture, Ecosystems and Environment 47 (1993) 175-184
seepage water, ANC (aq) is positive. Bicarbonate is balanced by Mb cations. The example indicates that nitrate leaching may exceed N input. The N H g N levels in seepage water are negligible. Usually, forest ecosystems represent a sink for nitrogen. Owing to increasing NO~--N and ammonium deposition rates, some ecosystems show a tendency for N saturation (Malanchuk and Nilsson, 1989). This eutrophication may cause nutrient imbalances. In the G~ttinger Wald the relation between the SO2- -S input and output rates may point to an actual sulphur retention. In soils of the cation exchange buffer range, in almost all ecosystems investigated (e.g.L. Heide, oak) bicarbonate concentrations are negligible (pH < 5.0 ). One exception is the Harste beech stand where base saturation of the exchange complex is still relatively high (34%; NH~ extraction). In this case ANC (aq) shows low, but positive values ( + 75 mmolc 1-~ ). In general, the external H + load exceeds the net internal production. Small amounts of Ma cations (AI 3+, Mn 2+ ) leave the system with drainage water. The Mb cations Ca 2+, Mg2+ and K + are mainly balanced by SO42- and NO~-. The silicate weathering rate in undisturbed soils cannot accurately be measured. Thus all values given in the literature are estimates, ranging from 2.0 to 0.02 kmolc ha -~ year -1 (soil layer 1 m). The rate mainly depends on the mineral composition of the soil (Sverdrup et al., 1989). For those examples discussed in this paper, with soils showing low to moderate contents ofweatherable minerals, the input rate of Mb cations through silicate weathering is estimated to range from 0.3 to 1.0 kmolc ha- ~year- 1. Figure 4 indicates that the total proton load exceeds these release rates by a large degree. When the A1 buffer becomes effective, the input of acidity (H + ) is balanced by an output of A13+, H + and Mn 2+, which are leached together with sulphate and nitrate (e.g. L Heide, pine; Solling, beech; SoUing, spruce). Thus, acidity is transported to deeper soil and bedrock layers. Conclusions
The summarised classification of different budget types of forest ecosystems shows that input/output relations of major ions are useful indicators of a system's state. Assuming a prospective deposition development, the system trend may be assessed (Ulrich, 1991 ). Today in most forest ecosystems in Central Europe, the total proton load (TPL), i.e. the input by acidic deposition plus the net internal H + production, exceeds the rate of Mb cation release through silicate weathering by a large degree. Thus the major part of the TPL is buffered by exchangeable AI ions and/or soluble AI compounds (aluminium buffer range). The weathering rate of the soil/bedrock matrix must be considered as one
A. Liikewille et al. /Agriculture. Ecosystems and Environment 47 (1993) 175-184
183
of the key parameters in setting critical loads for the input of acidic compounds (Sverdrup et al., 1989). Soil acidification results in high concentrations of dissolved substances such as AI3+ which may have toxic effects on organisms. It is one of the key factors for the explanation of the widespread damage to trees in Europe. The transport of acidity (mainly AI3+, H + ) to deeper soil and bedrock layers may result in an acidification of groundwater and surface waters. Increasing nitrogen (NO3--N and NH + -N) deposition rates may lead to the eutrophication of forest ecosystems and thus to nutrient imbalances and the pollution of drinking water. Nitrification of deposited NH~ is one evident example that ecosystem-internal H + production often can be attributed to anthropogenic causes.
References Bredemeier, M., Matzner, E. and Ulrich, B., 1990. Internal and external proton load of forest soils. J. Environ. Qual., 19: 469-477. Hauhs, M., 1986. A model of ion transport through a forested catchment at Lange Bramke, West Germany. Geoderma, 38:97-113. Liikewille, A., 1990. Reconstruction of water acidification in a forested catchment using the process-oriented model BEM, first results. Poster Presentation, Int. Conf. on Acidic Deposition, Glasgow, 16-21 September 1990. Malanchuk, J.L. and Nilsson, J. (Editors), 1989. The role of nitrogen in the acidification of soils and surface waters. Nordic Council of Ministers, NOR]) 1989:92. Matzner, E. and Prenzel, J., 1992. Acid deposition in the German Soiling area: Effects on soil solution chemistry and A1 mobilization. Water, Air Soil Pollut., 61:221-234. Prenzel, J., 1985. Veflauf und Ursachen der Bodenversauerung. Z. Dtsch. Geol. Ges., 136: 293302. Reuss, J.O., 1991. The transfer of acidity from soils to surface waters. In: B. Ulrich and M.E. Sumner (Editors), Soil Acidity. Springer, New York/Berlin/Heidelberg/Tokyo, pp. 203217. Reuss, J.O., Cosby, B.J. and Wright, R.F., 1987. Chemical processes governing soil and water acidification. Nature, 329: 27-32. Sverdrup, H., de Vries, W. and Henriksen, A. (Editors), 1989. Mapping of critical loads prepared for the Workshop and Task Force on mapping critical loads and levels at Bad Harzburg (1989), Organized by the Secretariat of the United Nations Economic Commission for Europe (UN-ECE), the Nordic Council of Ministers (NMR). Ulrich, B., 1983. Soil acidity and its relation to acid deposition. In: B. Ulrich and J. Pankrath (Editors), Effects of Accumulation of Air Pollutions in Forest Ecosystems, pp. 127-146. Ulrich, B., 1989a. Stoffhaushalte yon WaldSkosystemen. In: Forschungsbeirat Waldschiiden / Lultverfinreinigungen (Editor), Dritter Bericht, Teil B, Band II. Ulrich, B., 1989b. Effects of acidic precipitation on forest ecosystems in Europe. In: D.C. Adriano and A.H. Johnson (Editors), Acid Precipitation, Vol. 2, Biological and Ecological Effects, pp. 189-272. Ulrich, B., 1991. An ecosystem approach to soil acidification. In: B. Ulrich and M.E. Sumner (Editors), Soil Acidity. Springer, New York/Berlin/Heidelberg/Tokyo, pp. 28-79.
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A. Liikewille et al. /Agriculture, Ecosystems and Environment 47 (1993) 175-184
Van Breemen, N., 1991. Soil acidification and alkalinization. In: B. Ulrich and M.E. Sumner (Editors), Soil Acidity. Springer, New York/Berlin/Heidelberg/Tokyo, pp. 1-7. Van Breemen, N., Mulder, J. and Driscoll, C.T., 1983. Acidification and alkalinization of soils. Plant Soil, 75: 283-308. Van Breemen, N., Mulder, J. and Driscoll, C.T., 1984. Acid deposition and internal proton sources in acidification of soils and waters. Nature, 307: 599-604.
175
Elsevier Science Publishers B.V., Amsterdam
Input-output relations of major ions in European forest ecosystems A. Liikewille*, M. Bredemeier, B. Ulrich Institute of Soil Science and Forest Nutrition, UniversityGi~ttingen, Biisgenweg2, W-3400 G6ttingen, Germany
Abstract
A framework for the assessment of the state and trend of forest ecosystems from ion budget data is presented. Atmospheric deposition is regarded as an input and seepage water as a system output. A simple compartment model, subdividing the forest ecosystem, is used to calculate ion balances and the corresponding internal H + production/consumptionprocesses. The 'buffer range' of the forest soil substantially determines the reaction of a system to input and internal generation of acidity. Forest soils in the carbonate buffer range are characterised by high seepage water pH (6.2-8.6). High concentration levels of HCO3 are balanced by alkali and alkali earth ions (Mb or 'base' cations). In the cation exchange buffer range, pH values remain above pH 4.2 and concentrations of dissolved aluminium are low. With increasing proton load the chemical characteristics of the soil solid phase are changing. The consequences are a decrease in acid neutralisationcapacity (ANC) and base saturation (BS), and the formation of interlayer A1. At BS values below 10-15%, H + is buffered almost solely by the dissolution and exchange of aluminium (and/or iron) compounds. Acidity, mainly as A13+and H +, is transported with seepage water and thus influences deeper soil and bedrock layers. Examples of the mentioned budget types are given, utilizing data from long-term case studies in Germany.
Introduction In most forest soils in Central Europe, acid deposition exceeds internal proton loading (Van Breemen et al., 1984; Bredemeier et al., 1990). Forest ecosystems can be regarded as three-dimensional components of the ecosphere which are subdivided into compartments. The parts are interconnected by fluxes of material. The element budgets discussed in this paper represent a process-functional approach in viewing ecosystems. Organisms and their environment are considered as a single integral system. Finally, this approach implies that macroscopic phenomena like energy flow and nutrient cycling are more fundamental than the biotic entities performing the functions (O'Neill et al., 1986). A mass balance equation defining fundamental pro*Corresponding author.
176
A. L iikewille et al. / Agriculture, Ecosystems and Environment 47 (1993) 175-184
ccsses like photosynthesis, ion uptake, build-up of organic substances and the opposite processes of respiration and mineralisation of organic substances can be used to describe main fluxes of material in forest ecosystems (Ulrich, 1989a)
aCO2 + x M + +yA - + ( y - x ) H + + z H 2 0 + energy(hv) ~- ( Ca H20zMx Ay)organic matter "~"a O 2
( 1)
where a, x, y, and z are stoichiometdc coefficients, M + represents cations, and A - anions of unit charge, and organic matter includes organisms as well as humus. Assessment of ion fluxes in forest ecosystems Figure 1 shows a simple compartment model, used to quantify the most important internal net proton production processes (Bredemeier et al., 1990). In 'steady state', there is almost no net turnover (production/consumption) of H + ions. The actual system state can be assessed by looking at deviations from quasi steady state conditions (Ulrich, 1989a,b, 1991 ). To investigate the relationship between deposition and seepage of certain substances, the deposition rates of major ions are regarded as an input into the forest ecosystem (arrows 1 and 2 in Fig. I ). The fluxes of ions with seepage water below the rooted soil layer represent a system output (arrow 8 in
t, ~..AVno, pher~ Depo~on Atmo,p h ~ BIomm pecenn~ friction
'.mc:t~mnt)
t
,4)'~
t
IP" T (,) '
. Humus Layer
(¢)
V
('/)
M i n e r a l Soil
(70)
V
(m)
Outla~-$~la~leWete¢
Fig. 1. Simple compartment model of a forest ecosystem.
A. Lflkewille et aL / Agriculture, Ecosystems and Environment 47 (1993) 175-184
177
Fig. 1 ).The assessment of these input and output fluxes across defined boundaries requires knowledge of the hydrology which can be calculated using mathematical models (Hauhs, 1986). Another important input into the ecosystem is the (irreversible) release of ions due to mineral weathering (arrow 10 in Fig. 1 ). The general reaction of the system to the input of acidic deposition (on a time scale of years) mainly depends upon the chemical soil state, the amount of Mb cations, NO~- and NH~ bound in the annual biomass increment, the rate of organic matter stored in the soil and the quantity of drainage water. The input/output relations of acidity in terrestrial soils are the result of an interaction of external and internal hydrogen ion sources with the soil solid phases. Internal proton generation takes place within the soil which, however, must be considered as a 'black box'. Consequently, these processes cannot be measured directly. Their rates are assessed by means of compiling complete ion budgets of the soil. The method to calculate internal H ~ transfer processes is described and discussed in detail by Bredemeier et al. (1990). Soil internal net H + generation may occur as a result of the following biogeochemical processes: biomass/humus aggregation; dissociation of carbonic acid and of organic acids; net nitrification of organic nitrogen from the N pool of the ecosystem; assimilation of a surplus of deposited NH~- over NO~- and/or nitrification of deposited NH~-; buffering of deposited H + on canopy surfaces with subsequent regeneration of foliage buffer capacity (i.e. the uptake of Mb cations from the soil). These processes, designated as internal net proton production (INP), contribute together with the free acidity (H + ) in throughfall (see Fig. 1, arrow 4) to the total proton load (TPL) of a forest soil. Microbially mediated reduction of sulphate and nitrate in the soil is neglected. The relation between TPL and H + buffering due to mineral weathering determines the trend of development of soil chemical properties.
Buffer ranges in soils
The relation between acid neutralisation capacity (ANC) and pH in a mineral soil affected by acidic deposition can be represented as a titration curve (Van Breemen et al., 1983; Prenzel, 1985; Van Breemen, 1991 ). Figure 2 shows pH as an intensity factor and ANC as a capacity factor. A ANC denotes the rate of soil acidification. The pH limits given in Table I were calculated by assuming chemical equilibria between ion species in the soil solution and well defined solid phases. Thus deviations from these fixed values may occur.
178
A. Lf~kewille et aL /Agriculture, Ecosystems and Environment 4 7 (1993) 175-184
pH; intensity factor A
7 -
6
5
4
3
H + added, A ANC; capacity factor
Fig. 2. Sequence of possible buffer reactions changing the composition of the soil solution; quick acidification ( ), slow acidification ( . . . . ). Table 1 Buffer ranges in soils Buffering substance/process
pH range
Dissolution of carbonates Exchange of Mb cations Exchange of aluminium ions, dissolution of A1compounds Aluminium, iron compounds Iron compounds
8.6>pH>6.2 6.2>pH>4.2 4.2>pH>3.8 3.8>pH>3.2 3.2>pH
The plateau A-B (Fig. 2) represents the carbonate buffer range. The dissolution of CaCO3 is the predominant buffering process. Ca 2+ and H C O £ are the major ions in the soil solution CaCO3 + 2H + ~ Ca 2+ + 2HCO~-
(2)
The slope B-C shows the exchange buffer range. Mb ions adsorbed at the cation exchange complex serve as a buffer, e.g. 2H + +Cae2~+ ~ 2 H ~ + C a 2+
(3)
Additionally, the reaction of H + with soil minerals leads to the formation of positively charged A1 ion species, which in turn displace Mb cations from the exchange complex (build-up of interlayer Ah n [AI(OH)x~3- x) ] ). Parallel to this process, base saturation decreases, e.g. KA1Si308 + H + + H 2 0 ~ K
+ +AI(OH)3
+ 3SIO2
(4)
A. Lfikewille et al. / Agriculture, Ecosystems and Environment 47 (1993) 175-184
KAISi30s + 4H + + Cae2~~ K~ + A13x+ + 2H20 + 3SIO2 -I-2Ca 2+
179
(5)
Equations (4) and (5) show that the rate of silicate weathering influences the effectiveness of the exchange buffer considerably. Therefore, it is a very important kinetical constraint. In the case of quick acidification of a soil with given exchange capacity, base saturation and CO2 partial pressure, the slopes of the curve (Fig. 2 ) is extremely sensitive to the release of Mb cations through silicate weathering. The pH of the soil solution is dependent on CO2 partial pressure, which is particularly influenced by respiration processes. Owing to the pK~ values of the H2CO3 dissociation equilibria, pH levels below 5.0 indicate negligible HCO£ concentrations in the soil solution, even at high Pco2 values (e.g. 3%; Reuss and Johnson, 1986; Reuss, 1991 ). The interval C-D (Fig. 2) represents the aluminium buffer range. As base saturation decreases below about 15%, the pH drops to values less than 4.54.2 and the A1 buffering takes effect. A13+ is then the prevalent cation in the soil solution A1(OH)3 + 3H + ~A13+ + 3H20
(6)
If soluble A1 compounds are depleted, or if their dissolution is restricted by kinetic constraints at a chosen time scale, the soil reaches the iron buffer range.
Acid neutralisation capacity (ANC) and sulphur retention Since pH is an intensity factor, the acid neutralisation capacity should also be considered. The ANC of mineral soil material can be defined as the sum of ANC of the solid phase plus ANC of the aqueous phase (Van Breemen, 1991) ANC (S) = ANC (s) +ANC (aq)
(7)
The ANC (s) includes, for example, CaO, MgO and 1(20 as components of silicates (reference pH of 5.0) or A1 compounds (reference pH of 3.0). If ANC(aq) in the soil solution is removed by the percolating water, total ANC(S) decreases (Ulrich, 1991; Van Breemen, 1991). Soil acidification can be defined as a loss of acid neutralisation capacity. The principle of charge balance implies that ANC (aq) or alkalinity may be defined as the sum Of Mb cations minus the sum of strong acid anions (Reuss and Johnson, 1986; Reuss, 1991 ). Organic anions are neglected ANC (aq) =2(Ca 2+ ) +2 (Mg 2+ ) + (Na + ) + (K + ) - 2 ( S O 2- ) - ( N O ; ) - ( e l - )
(8)
parentheses denote molar concentrations/activities and ANC(aq) is expressed in/zmolc 1-1.
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A. Liikewille et al. /Agriculture, Ecosystems and Environment 47 (1993) 175-184
Owing to exchange processes in the soil matrix, the concentration of M b cations in the soil solution is dependent on the amount of dissociated H2CO3 and, consequently, on CO2 partial pressure. ANC(aq) is a useful means of investigating acidification processes. In whole-catchment experiments, water not in contact with the bedrock matrix degasses (CO2), resulting in a considerable increase in stream/lake water pH, whereas alkalinity (as defined in Eq. (8) ) remains unchanged (Reuss et al., 1987; Liikewille, 1990). The same phenomenon must be considered if pH values, which were measured in potentially degassed soil solution samples, are interpreted. At high concentrations of dissolved A1 species, ANC (aq) becomes negative. The expediency of a calculation with negative capacity values ( - A N C ( a q ) ) is questionable. For an assessment of the ecological consequences of soil acidification, e.g. the effects on tree roots, the consideration of intensity values such as pH and concentrations of potentially toxic ions such as A13+ is more suitable (Ulrich, 1991 ). If the observed sulphur retention in soils is considered it is difficult to distinguish between SO2- adsorption processes and the precipitation of AI hydroxysulphate minerals. The following equations may be considered as model reactions for more complex processes of chemiadsorption and/or solid formation (Matzner and Prenzel, 1991 ). Formation of alunite 3Al(OH)3 + 1.5H2504 + 0.5K2 SO4--}KA13(OH)6 (504) 2 + 3H20
(9)
Formation ofjurbanite A1(OH) 3 "~-H2 S O 4
---}A1OHSO4
+ 2 HE O
( 10 )
Sulphur retention must be regarded as a delay mechanism: acidity, or base neutralisation capacity (BNC (s) ), is stored in the soils and may become effective under changing input and/or ecosystem-internal conditions. Results of long-term case studies
Data from almost 30 long-term case studies in Central Europe were used to group forest ecosystems into budget types. Detailed information about the study sites is given by Ulrich ( 1989a,b, 1991 ) and Bredemeier et al. (1990). The buffer ranges serve as a key to understand the relation between deposition and seepage output of major ions. Figures 3 and 4 summarise representative examples for the different budget types. The units for the input and output fluxes are kmolc ha-I year-~. The data are mean values of several years. The first example represents a forest ecosystem with soils in the carbonate buffer range (G6ttinger Wald, beech). Figure 3 shows that the internal H + load far exceeds the input by deposition. Since HCO£ is the major ion in the
A. Liikewille et al. /Agriculture, Ecosystems and Environment 47 (1993) 175-184
181
Case Study
_
C-ddtingerWald,Beech
- 12.9 ///-///////////////,
~
+ 10.0
"/////////////
L Heide, Oak
/////,/.//III//,
~\\\\%.~,\\\\\\\
Hamto,Beech i
i
-4
t
-2
i
0
•
H+ deposited
•
[]
Ma output
[ ] Mb output
H+ internal
2
kmol c -
ha-I.
•
S input
4
yr-1
[ ] S output
[]
N input
[]
N Output
Fig. 3. Input-output relations of major elements in different forest ecosystems in Central Europe.
Case Study
,J
L Heide, Pine
---,.l Soiling, Beech [*;,~"~',~\\\\ \ \ \ \ \ \ \ ~ L
Soiling, Spruce I
till/ill;';.;'7, ,\\\\\\\\\\\\\\\\\\\\\\7
k\\\\\\\\\\' i
-6
i
-4
-2
0
l~nol c • ha-1. •
H+ deposited
[ ] M.o.~,,
•
H+ in~rnal
[ ] Mbo°~t
•
i
2
i
4
yr-1
$ input
[ ] s o.~ot
[ ] N input
[] N ou~ut
Fig. 4. Input-output relations of major elements in different forest ecosystems in Central Europe.
182
A. Lffkewille et al. / Agriculture, Ecosystems and Environment 47 (1993) 175-184
seepage water, ANC (aq) is positive. Bicarbonate is balanced by Mb cations. The example indicates that nitrate leaching may exceed N input. The N H g N levels in seepage water are negligible. Usually, forest ecosystems represent a sink for nitrogen. Owing to increasing NO~--N and ammonium deposition rates, some ecosystems show a tendency for N saturation (Malanchuk and Nilsson, 1989). This eutrophication may cause nutrient imbalances. In the G~ttinger Wald the relation between the SO2- -S input and output rates may point to an actual sulphur retention. In soils of the cation exchange buffer range, in almost all ecosystems investigated (e.g.L. Heide, oak) bicarbonate concentrations are negligible (pH < 5.0 ). One exception is the Harste beech stand where base saturation of the exchange complex is still relatively high (34%; NH~ extraction). In this case ANC (aq) shows low, but positive values ( + 75 mmolc 1-~ ). In general, the external H + load exceeds the net internal production. Small amounts of Ma cations (AI 3+, Mn 2+ ) leave the system with drainage water. The Mb cations Ca 2+, Mg2+ and K + are mainly balanced by SO42- and NO~-. The silicate weathering rate in undisturbed soils cannot accurately be measured. Thus all values given in the literature are estimates, ranging from 2.0 to 0.02 kmolc ha -~ year -1 (soil layer 1 m). The rate mainly depends on the mineral composition of the soil (Sverdrup et al., 1989). For those examples discussed in this paper, with soils showing low to moderate contents ofweatherable minerals, the input rate of Mb cations through silicate weathering is estimated to range from 0.3 to 1.0 kmolc ha- ~year- 1. Figure 4 indicates that the total proton load exceeds these release rates by a large degree. When the A1 buffer becomes effective, the input of acidity (H + ) is balanced by an output of A13+, H + and Mn 2+, which are leached together with sulphate and nitrate (e.g. L Heide, pine; Solling, beech; SoUing, spruce). Thus, acidity is transported to deeper soil and bedrock layers. Conclusions
The summarised classification of different budget types of forest ecosystems shows that input/output relations of major ions are useful indicators of a system's state. Assuming a prospective deposition development, the system trend may be assessed (Ulrich, 1991 ). Today in most forest ecosystems in Central Europe, the total proton load (TPL), i.e. the input by acidic deposition plus the net internal H + production, exceeds the rate of Mb cation release through silicate weathering by a large degree. Thus the major part of the TPL is buffered by exchangeable AI ions and/or soluble AI compounds (aluminium buffer range). The weathering rate of the soil/bedrock matrix must be considered as one
A. Liikewille et al. /Agriculture. Ecosystems and Environment 47 (1993) 175-184
183
of the key parameters in setting critical loads for the input of acidic compounds (Sverdrup et al., 1989). Soil acidification results in high concentrations of dissolved substances such as AI3+ which may have toxic effects on organisms. It is one of the key factors for the explanation of the widespread damage to trees in Europe. The transport of acidity (mainly AI3+, H + ) to deeper soil and bedrock layers may result in an acidification of groundwater and surface waters. Increasing nitrogen (NO3--N and NH + -N) deposition rates may lead to the eutrophication of forest ecosystems and thus to nutrient imbalances and the pollution of drinking water. Nitrification of deposited NH~ is one evident example that ecosystem-internal H + production often can be attributed to anthropogenic causes.
References Bredemeier, M., Matzner, E. and Ulrich, B., 1990. Internal and external proton load of forest soils. J. Environ. Qual., 19: 469-477. Hauhs, M., 1986. A model of ion transport through a forested catchment at Lange Bramke, West Germany. Geoderma, 38:97-113. Liikewille, A., 1990. Reconstruction of water acidification in a forested catchment using the process-oriented model BEM, first results. Poster Presentation, Int. Conf. on Acidic Deposition, Glasgow, 16-21 September 1990. Malanchuk, J.L. and Nilsson, J. (Editors), 1989. The role of nitrogen in the acidification of soils and surface waters. Nordic Council of Ministers, NOR]) 1989:92. Matzner, E. and Prenzel, J., 1992. Acid deposition in the German Soiling area: Effects on soil solution chemistry and A1 mobilization. Water, Air Soil Pollut., 61:221-234. Prenzel, J., 1985. Veflauf und Ursachen der Bodenversauerung. Z. Dtsch. Geol. Ges., 136: 293302. Reuss, J.O., 1991. The transfer of acidity from soils to surface waters. In: B. Ulrich and M.E. Sumner (Editors), Soil Acidity. Springer, New York/Berlin/Heidelberg/Tokyo, pp. 203217. Reuss, J.O., Cosby, B.J. and Wright, R.F., 1987. Chemical processes governing soil and water acidification. Nature, 329: 27-32. Sverdrup, H., de Vries, W. and Henriksen, A. (Editors), 1989. Mapping of critical loads prepared for the Workshop and Task Force on mapping critical loads and levels at Bad Harzburg (1989), Organized by the Secretariat of the United Nations Economic Commission for Europe (UN-ECE), the Nordic Council of Ministers (NMR). Ulrich, B., 1983. Soil acidity and its relation to acid deposition. In: B. Ulrich and J. Pankrath (Editors), Effects of Accumulation of Air Pollutions in Forest Ecosystems, pp. 127-146. Ulrich, B., 1989a. Stoffhaushalte yon WaldSkosystemen. In: Forschungsbeirat Waldschiiden / Lultverfinreinigungen (Editor), Dritter Bericht, Teil B, Band II. Ulrich, B., 1989b. Effects of acidic precipitation on forest ecosystems in Europe. In: D.C. Adriano and A.H. Johnson (Editors), Acid Precipitation, Vol. 2, Biological and Ecological Effects, pp. 189-272. Ulrich, B., 1991. An ecosystem approach to soil acidification. In: B. Ulrich and M.E. Sumner (Editors), Soil Acidity. Springer, New York/Berlin/Heidelberg/Tokyo, pp. 28-79.
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Van Breemen, N., 1991. Soil acidification and alkalinization. In: B. Ulrich and M.E. Sumner (Editors), Soil Acidity. Springer, New York/Berlin/Heidelberg/Tokyo, pp. 1-7. Van Breemen, N., Mulder, J. and Driscoll, C.T., 1983. Acidification and alkalinization of soils. Plant Soil, 75: 283-308. Van Breemen, N., Mulder, J. and Driscoll, C.T., 1984. Acid deposition and internal proton sources in acidification of soils and waters. Nature, 307: 599-604.