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Chapter 1. The Effects of Acidic Deposition on Forested Soils

Chapter 1. The Effects of Acidic Deposition on Forested Soils

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WAYNE P. ROBARGE AND DALE W. JOHNSON



cloudwater (Weathers et al., 1988; Reisinger and Imhoff, 1989; Saxena

et al., 1989; Sigmon et al., 1989; Aneja et al., 1991). Acidification of soils

near point sources of N and S acid-forming substances is well known,

thus one of the earliest concerns over acidic deposition occurring over a

regional basis was the potential for soil acidification (Ulrich, 1980). Since

then, a number of reviews have been written and models proposed to

explain how acidic deposition can increase soil acidity (Ulrich et al., 1980;

Reuss, 1983; van Breemen et al., 1983; Kohlmaier et al. , 1983/1984; Bloom

and Grigal, 1985; Cosby et al., 1985; Rechcigl and Sparks, 1985; Reuss and

Johnson, 1985; 1986; Tabatabai, 1985; Krause et al., 1986; Ulrich, 1987;

Binkley et al., 1989a; De Vries et al., 1989; Reuss and Walthall, 1989;

Reuss et al., 1990; Walker et al., 1990). Such models have been used to

develop sensitivity criteria to determine which soils on a regional basis will

be most sensitive to acidification by acidic deposition (McFee, 1983; Binkley et al., 1989a). These model reactions, based on soil chemistry theory

and combined with field observations of changes in soil chemical parameters and surface water chemistry of streams, have led to the general

conclusion that acidic inputs into such ecosystems are capable of increasing

soil acidity (Reuss et al., 1987).

Objections to this general conclusion have largely been based on the

argument that the effects of acidic deposition on forest soil systems can

only be evaluated from the standpoint of how acidic inputs interact with

the natural processes of soil acidification (Rosenquist, 1978; Krug and

Frink, 1983a,b; Tabatabai, 1985). Central to this point of view is the fact

that soil formation in humid temperature climates is an acidifying process,

and that a correlation between areas of high acidic deposition and the

presence of acidic soils and stream waters is not sufficient cause to conclude

that acidic inputs have increased soil acidity in these ecosystems. Many of

the temperate forests of northern Europe and eastern North America that

currently receive acidic inputs have undergone substantial changes in land

use policy during the past 200 years (see, e.g. ,Brand et al., 1986). As many

of these forests are now aggrading, the natural soil acidification that

accompanies such regrowth cannot be attributed to acidic deposition (Krug

and Frink, 1983a).

The discussion concerning the pros and cons of both hypotheses on the

effects of acidic deposition on soil acidity still continues (Havas et al., 1984;

N. M. Johnson et al., 1984; Krug and Isaacson, 1984; Norton et al., 1989),

but it has served to illustrate the limitations in our current understanding about the processes of soil acidification in forest soils. Without such

knowledge it will not be possible to understand fully the effects of acidic

deposition on these soils, either now, at current levels of input, or in the

future, as regulatory measures begin to reduce acidic inputs. It is becoming



ACID DEPOSITION ON FORESTED SOILS



3



clear that the rate at which forest soils are changing will require a rethinking of our approach to studying soil changes in these ecosystems (D. W.

Johnson et al., 1991a).

This review on the effect of acidic deposition on forest soils assumes that

it is no longer necessary to argue whether acidic inputs are having an

impact on soil systems. The levels of input are now well characterized and

it is highly unlikely that any natural ecosystem would fail to respond in

some way (at the very least by increased leaching) to such sustained inputs

of potential energy. Emphasis instead is on what effects may be possible

given the natural physical and chemical processes acting in forest ecosystems. Discussions of possible mechanisms will differ from earlier reviews in

that more emphasis will be placed on the limitations of such mechanisms,

both in terms of theory and the currently available body of knowledge.

Reference to agricultural and other intensively managed soil systems is

excluded because it is generally accepted that the influence from acidic

deposition on such soils will be minimal (McFee, 1983; Tabatabai, 1985).

Our specific approach is to first attempt to define the often used but

poorly understood term “soil acidification,” followed by a brief synopsis of

the dominant physical and chemical processes in forest ecosystems that

interact with, modify, and respond to acidic inputs. This information is

then used as background to discuss published case histories of soil change

in regions where acid deposition may be causal, and in regions where it is

not. Last, a set of recommendations is put forward concerning areas for

future research that are needed to properly characterize and develop

models for the effects of acidic deposition on forest soils.



11. SOIL ACIDIFICATION

Soil acidification refers to a complex set of processes that result in the

formation of an acid soil (pH < 7.0). Soil acidification, therefore, in the

broadest sense, can be considered as the summation of natural and anthropogenic processes that lower measured soil pH (Krug and Frink, 1983a). In

forest ecosystems, natural acidifying processes include base cation uptake

(by plants or microbes); natural leaching by carbonic, organic, or nitric

acid; and humus formation (Ulrich, 1980). Anthropogenic acidifying processes include biomass harvesting (which simulates increased uptake)

(Binkley et al., 1989a), land use conversion (Berdbn et af., 1987; Billet

el af., 1988, 1990b; D. W. Johnson et af., 1988b), fertilization (van Breeman el d., 1982), as well as atmospheric inputs of acidifying compounds

(Reuss and Johnson, 1986). Barring inputs of lime from anthropogenic



4



WAYNE P. ROBARGE AND DALE W. JOHNSON



sources, or marine influences, forest ecosystems developed on noncalcareous-bearing parent materials in humid environments will have acid

soils.

Attempts to measure soil acidification as a result of acidic inputs often

center on attempting to detect changes in soil pH (see, e.g., Tamm and

Hallbacken, 1986). Though this approach seems intuitively obvious, practical considerations ranging from suitable analytical methodology, spatial

variability, and determination of soil horizonation to changes in land use

patterns often severely limit the usefulness of this rather simple approach.

Soil acidification cannot be quantitatively described by a single index

parameter, even though it is often assumed that soil pH is such a parameter

(Matzner, 1989). Other changes in soils that may occur during soil acidification include loss of nutrients due to leaching; loss or reduction in the

availability of certain plant nutrients (such as phosphorus and molybdenum, which are more strongly retained in acid soils); an increase in the

solubility of toxic metals (primarily aluminum and manganese), which may

influence root growth and nutrient and water uptake; and a change in

microbial populations and activities (Binkley et al., 1989a). Such changes

will often be accompanied by changes in overall soil pH, but the degree of

change will be dependent on a combination of properties within a given soil

system.

A more quantitative measure of soil acidification can be obtained by

defining it as a decrease in the acid-neutralizing capacity (ANC) of a soil

(Berdtn et al., 1987; De Vries and Breeuwsma, 1987). This approach is

similar to that for aqueous systems (Stumm and Morgan, 198l), wherein

the ANC of a soil solution can be defined as the aqueous base equivalence

minus the strong acid equivalence as determined by strong acid titration

to a reference pH (typically pH 4.5) (van Breemen et al., 1984). The ANC

of the inorganic fraction of a mineral soil can be defined as the sum of

basic components minus the strongly acidic components (van Breemen

et al., 1984)

ANC = 6[A1203]+ 6[Fe203]+ 2[Fe0] + 4[Mn02]



+ 2[Mn0] + 2[Ca0] + 2[Mg0] + 2[Na20]

+ 2[K20] - 2[SO3] - 2[P205] - [HCl]



(1)

where the brackets denote molar concentrations. Note that the metal

oxides include the metal cations in the soil solids, as well as those on the

exchange complex and in the soil solution. A decrease in the cationic

components (such as CaO) or an increase in the acidic components (such as

SO3) will result in an increase in soil acidification (a decrease in ANC)



ACID DEPOSITION ON FORESTED SOILS



5



(van Breemen et al., 1984). A decrease in the cationic components could

occur through biomass uptake or leaching, whereas an increase in the acidic

components could result from inputs of SOT2.This approach emphasizes

the mass of acidic input, as equivalents of H+ or NO, and SO:-, rather

than the intensity of input as measured by pH (Binkley and Richter, 1987).

Thus, a detailed H+ budget for a forest ecosystem, which attempts to

quantify all of the proton-producing and -consuming processes within a

given ecosystem, offers a means of separating out soil acidification due to

acidic deposition from natural acidification processes (van Breeman el al.,

1983). Soil acidification could then be defined as the result of an irreversible flux of protons to the soil ecosystem. The limitations of this approach

are that such budgets are difficult to construct, and the budget estimates for

the various processes are often associated with a large degree of uncertainty both spatially and temporally (Binkley and Richter, 1987).

The components of the soil that comprise the ANC can be divided into

processes that are relatively fast (approach equilibrium rapidly), slow

(processes that are rate limited but for which the kinetics are known), or

very slow (may be essentially ignored as having an impact on the system)

(Furrer ef al., 1990). Those processes considered to be fast have an immediate impact on the composition of the soil solution, and are also

referred to as intensity factors (Reuss and Johnson, 1986). The soil components involved are predominately the soil solution and those soil surfaces

that react rapidly to changes in the soil solution (e.g., cation and anion

exchange capacity). Slow and very slow processes are referred to as capacity factors and essentially reflect an integration of changes in a soil system

over time. These processes include cation and anion plant uptake, mineralization, oxidation and reduction, and primary and secondary mineral

weathering (Furrer et al., 1990). Over the long term, the capacity tactors of

a soil will control the range in intensity tactors that are ooserved.

The ion pools that comprise the capacity factors greatly exceed those of

the intensity factors and the inputs from acidic deposition (Keuss and

Johnson, 1986; Reuss and Walthall, 1989). These relatively large pools of

ions in already acid soils are the basis for the assumption that the impact

from acidic deposition will be small compared to natural acidification

processes (Krug and Frink, 1983a), and that substantial periods of time will

be required before detectable changes in these bulk soil chemical properties will occur, if at all (Tabatabai, 1985).

There is a growing consensus, however, that the primary effect of acidic

deposition on forest soils is via the intensity factors and that substantial

changes in the capacity factors by acidic deposition are not necessary

to influence the composition of the soil solution (Reuss et al., 1987;



6



WAYNE P. ROBARGE AND DALE W. JOHNSON



D. W. Johnson et al., 1991a). This approach centers on the fact that the

dependence of intensity factors on capacity factors in a soil is often nonlinear, and that small changes may result in relatively large changes in soil

solution composition (Reuss and Johnson, 1986; Reuss and Walthall,

1989). It is also becoming apparent that the natural acidification processes

within a given ecosystem predispose that system to the way it will respond

to acidic inputs. It is argued that the change in the anion composition of the

soil solution caused by the introduction of the NO; and SO:- can account

for the observed changes in soil solution and surface water acidity without

the necessity for involving further soil acidification. Such a scenario would

mean that the effects of acidic deposition on forest ecosystems would occur

fairly rapidly over a variety of soil types in a relatively short period of time

once a critical loading of NO; and, in particular, SO:- is exceeded. It also

follows that reduction in inputs below the critical input would have an

immediate positive effect. Such responses have been observed both in the

field (Wright et af., 1988a) and in the laboratory (Dahlgren et al., 1990).

Soil acidification has long been cited as one of the effects of acidic

deposition on forest ecosystems. However, it has often not been made

clear that acidification of soils from acidic inputs must be viewed from the

standpoint of being superimposed upon natural acidification processes. A

quantitative measure of soil acidification can be obtained by using the

concept of the ANC of a soil, but the term itself does not necessarily imply

that there is a specific parameter (such as soil pH) or set of parameters that

can be used to measure soil acidification. It is becoming apparent that there

has probably been too much emphasis on change in soil pH and cation

depletion as a necessary and expected effect of acidic deposition on soil

systems (D. W. Johnson et al., 1991a). The influence of acidic deposition

on forest soils might be better understood by focusing on the reactions

of the mineral acid anions NO; and SO:- in soils-in particular, how

these anions interact with soil acidity already present from natural acidification processes.



111. FOREST SOILS

Model reactions of acidic deposition with soils often only emphasize

chemical reactions within a theoretical soil horizon. Actual forest ecosystems are infinitely more complicated and offer a number of physical as

well as chemical factors that must be considered in order to determine the

effect of acidic deposition on forest soils and surrounding surface waters.



ACID DEPOSITION ON FORESTED SOILS



7



A. PHYSICAL

FACTORS

1. Canopy Interactions



Acidic deposition reaches the forest ecosystem in the form of rainwater

(Schaefer and Reiners, 1989), as cloud and fog droplet impact on the forest

canopy (Lovett et al., 1982; Bruck et al., 1989; Reisinger and Irnhoff, 1989);

Saxena et al., 1989; Sigmon et al., 1989; Saxena and Lin, 1990), and as dry

deposition (Lindberg et al., 1986; Johnson and Lindberg, 1989; Murphy

and Sigmon, 1989). Dry deposition is the accumulation of particulates and

gases (such as HN03 and SO2) on the forest canopy in the absence of

precipitation (Davidson and Wu, 1989). Interaction of these three forms of

acidic deposition with the forest canopy changes their initial chemistry

before they finally reach the forest floor and the underlying mineral soil as

throughfall and stemflow. Throughfall is that fraction of wet deposition

that comes in contact with the canopy before reaching the forest floor,

whereas stemflow is that portion that drains down the branches and trunk

(Parker, 1990). In most forest ecosystems with an intact canopy, it is

throughfall and stemflow that are the major inputs of acidity and other ions

directly into the soil system.

The degree of interaction between acidic deposition and the forest

canopy is illustrated by the data in Table I, which compares the relative

chemical composition of throughfall to that of cloudwater and rainwater

in a high-evaluation spruce-fir ecosystem in the Black Mountains of

North Carolina (Bruck er al., 1989). The relative percentage of NO,

and SO:- between cloudwater, rainwater, and throughfall are almost

constant, with perhaps a slight decrease in NO, and a slight increase

in SOT2 in the throughfall. The largest change observed is a shift in

dominant cations, with H+ and NH,f replaced by K + , Ca2+, and Mg2+.

These data are representative of throughfall measurements in other

forest ecosystems that receive acidic deposition (Richter et al., 1983;

Lindberg et al., 1986; Bredemeier, 1988, J o s h et al., 1988; Percy, 1989;

Sigmon et al., 1989; Parker, 1990) and in studies using simulated acid rain

treatments (Scherbatskoy and Klein, 1983; Kelly and Strickland, 1986;

Kaupenjohann er al., 1988). More detailed information on throughfall

chemistry under a variety of forest canopies can be found in Parker (1983,

1990) and Bredemeier (1988). A review of the processes that control

throughfall chemistry can be found in Schaefer and Reiners (1989).

The release of base cations from the canopy is largely in response to the

fact that SO:- and, to a lesser extent, NO, are not adsorbed by the canopy

along with H+ and NHT. It is now known, through the use of 35S(Garten



WAYNE P. ROBARGE AND DALE W. JOHNSON



8



Table I

Total Ion Percent (pEq liter-') per Event for Cloudwater, Rainwater, and Throughfall

Samples Collected in 1986"



Throughfall (%)

Red spruce

Cloudwaterb

Ion



Fraser fir



Rainwater'



(%I



Site 1'



Site 2d



Site 1



Anions

c1NO;

s0:-



47.9

1.5

11.6

34.8



45.6

2.5

8.8

34.3



47.9

3.2

8.2

36.6



42.3

5.5

9.7

21.0



48.3

3.0

7.7

37.6



Cations



52.1

27.1

13.2

3.0

1.a

5.2

3.1

12.3



54.4

30.4

12.7

2.0

3.4

4.9

1.0

11.3



52.1

19.3

3.5

1.8

9.6

13.4

4.6

29.4



51.6

15.3

3.9

3.6

10.7

18.5

5.7

38.5



51.7

16.6

6.1

1.8

10.3

12.2

4.6

28.9



H+



Nb+

Na+



K+



Ca2+

Mg'+

Sum



"After Bruck et al. (1989); reprinted by permission of Kluwer Academic Publishers.

bCollected at site 1.

'Mt. Gibbes (2006 rn); from June 29, 1986 to Sept. 21, 1986.

d E a ~ face

t of Commissary Ridge (1760 m); from June 29, 1986 to August 15, 1986.



et af., 1988; Garten, 1990), that SO:- in particular is conserved within the

canopy, and that an increase in sulfate loading, either as H2S04 or

NH,HSOa in cloudwater or rainwater ( J o s h et af., 1988) or as SO2 in dry

deposition, will increase the concentration of base cations in throughfall

and stemflow (Parker, 1990). As an example, Johnson and Lindberg (1989)

estimates that between 40 and 60% of the base cations in throughfall

collected at the Walker Branch Watershed in East Tennessee during 19811983 was due to canopy exchange with deposited airborne acids. This

increase in base cation loss must come at the expense of the nutrient pool

within the canopy, which in turn may mean an increase in base cation

uptake. For the time period cited, this extra base cation uptake due to

"neutralization" of acidic input via the canopy equals a total H + input of

between 0.9 and 1.1 kmol( +) ha-' yr-' of internal acidification potential

within the rooting zone (Johnson and Lindberg, 1989). Calculations for the

Solling Forest in West Germany (Matzner, 1989) and for forests in the

Netherlands (van Breeman et af., 1986) yield similar results. Neutralization

of acidic inputs via the canopy, therefore, represents an indirect means of



ACID DEPOSITION ON FORESTED SOILS



9



increasing the acid load on a forest soil (Matzner, 1989; Ulrich, 1989). In

certain ecosystems, almost half of the H+ loading from acidic deposition

can be transferred to the soil system before the water transporting the

acidic anions enters the soil (Johnson and Lindberg, 1989; Matzner, 1989).

The acidity not neutralized by the canopy enters the soil via throughfall

and stemflow essentially as a salt solution dominated by Ca2+ and K+ salts

of SO:-, with a relatively minor contribution from the remaining mineral

acids (Johnson and Lindberg, 1989). The composition of this mixture is not

constant but varies considerably depending on a variety of factors, such as

season of the year (Parker, 1990), seasonal changes in deposition loading

(Johnson and Lindberg, 1989), the overall nutrient status of the ecosystem

(Leininger and Winner, 1988; Reynolds et al., 1989; Huettl et al., 1990;

Klumpp and Guderian, 1990), and stand age (Stevens, 1987). Throughfall

and stemflow composition will also vary depending on the relative health of

a stand (Alenas and Skarby, 1988). On a shorter time scale, throughfall

and stemflow composition will vary due to length of time between rainfall events [i.e., amount of dry deposition loading varies (Velthorst and

van Breemen, 198911, and even during storm events. The overall ionic

strength of throughfall and stemflow generally decreases significantly

during the course of an individual event (Kelly and Strickland, 1986;

Lovett et al., 1989) due to washoff of particulates from the leaf surfaces

(Schaefer and Reiners, 1989) and loss in ability of the canopy to buffer the

reactions with rainwater during the course of a storm (Parker, 1990).

Spatial variability in throughfall composition usually exceeds 25% when

expressed as the coefficient of variation due to the flow of water through

the canopy (Duijsings et al., 1986).

The forest canopy is both a modifier and conduit of acidic deposition into

the forest ecosystem. As illustrated above, the interaction between acidic

inputs and the canopy can have profound influences on both the pathways

and the chemistry of acidic substances that actually enter the forest floor

and the underlying mineral soil. These changes, together with varying

residence times within the different soil horizons, will influence the nature

of soil reactions that are likely to occur.

2. Soil Horizons

Acidic inputs entering a forest soil as throughfall and stemflow encounter a gradation in organic matter content extending from the forest floor,

which is essentially 100% organic matter, to the underlying parent material

of the mineral soil, which usually contains no innate source of organic

carbon. Depending on the interactions of the soil-forming factors (Buol et

al., 1980), the gradation in organic matter content may occur as distinct



10



WAYNE P. ROBARGE AND DALE W. JOHNSON



boundaries between soil horizons (e.g., Fig. 1) or as a gradual decrease in

organic matter content with depth. The mixing of organic matter and

mineral soil gives a range of reactive surfaces that will respond differently

to changes in the percolating soil solution, depending on the initial composition and rate of input of throughfall and stemflow, and on how the

chemical composition of this initial solution is changed as it passes through

each succeeding soil horizon. The nature of the chemical reactions that

may occur within each soil horizon will be discussed elsewhere in this

review, but Fig. 1 does serve to illustrate the point that generalizations

about a “soil’s” response to acidic inputs need to be properly defined in

terms of the scale of observation (Fernandez, 1989). System level studies

dealing with nutrient cycling avoid the issue of differences among soil

horizons by integrating observations across the entire soil pedon (Adriano

and Havas, 1989). In such studies, actual mechanisms within the plantsoil system are of secondary importance, especially in terms of element

cycling between the canopy and soil and between soil horizons within

the soil pedon due to natural processes. Studies addressing the specific

mechanisms of the effects of acidic deposition on tree growth cannot ignore

differences between soil horizons, as the rooting zones of trees are seldom

confined to a given soil horizon (Fernandez and Struchtemeyer, 1985;

Coutts, 1989; Fernandez, 1989).

The gradation of soil organic matter throughout the forest soil pedon

also means that attempts at measuring differences in soil properties over

time, even within the same morphological horizon, must be approached



Figure 1. Exchangeable cations (1 M NH4CI extractable) and pH (0.01 M CaCI,) from a

sampling (n = 23) of undisturbed well-drained and moderately well-drained pedons under

forest cover in Maine. After Fernandez (1989).



ACID DEPOSITION ON FORESTED SOILS



11



with due caution (Fernandez, 1989). A soil sample from a particular depth

within a given soil pedon can be expressed as the summation of its organic

matter component and its mineral soil component:

soil = organic matter



+ mineral soil



(2)

Because the ability of soil organic matter to retain metal ions greatly

exceeds that of the mineral soil fraction on a mass basis (Sposito, 1989),

relatively small changes in soil organic matter content can dominate the

overall physical and chemical properties of a given soil sample. Comparison of different soil samples from the same horizon and location over time,

therefore, requires attention not only to location on the landscape, but also

to the proportion of soil organic matter and mineral soil within the sample

itself. This is especially true if changes due to acidic inputs are restricted to

very narrow spatial scales within a soil (Fernandez, 1987,1989; Haun et af.,

1988). The extent of spatial variability in soil physical and chemical properties in forest ecosystems is well characterized and typically exceeds 25%

coefficient of variation (Mader, 1963; McFee and Stone, 1965; Ike and

Cutter, 1967; Ball and Williams, 1968; Troedsson and Tamm, 1969; Beckett and Webster, 1971; Quesnel and Lavkulich, 1980; Federer, 1982;

Neilsen and Hoyt, 1982; Arp, 1984; Arp and Krause, 1984; Riha et al.,

1986a,b; Wolfe et al., 1987; Bringmark, 1989; Pallant and Riha, 1990).

Attempts to measure differences in soil properties over time need to

account for the inherent variability in most soil systems (McBratney and

Webster, 1983; Webster and Burgess, 1984; Kratochvil et a f . , C. E. Johnson et af., 1990).



3. Forest Hydrology

The forest canopy together with the various soil horizons, parent material, and underlying bedrock make up the forest watershed. As discussed in

the previous two sections, the forest watershed can be considered as a

series of chemical reservoirs that interact with atmospheric inputs that are

transported with the drainage water (Schecher and Driscoll, 1989). The

degree of chemical interaction and the residence time of the drainage water

in each reaction zone determine the flux of ions through the watershed and

eventually into the stream-lake environment (Chen ef al., 1984). Residence time within a given soil horizon will depend on its position on the

landscape (Veneman and Bodine, 1982; Veneman et al., 1984; Roberge

and Plamondon, 1987) and the number and direction of water flow paths

present in a given volume of soil (Whipkey and Kirkby, 1978; Beven and

Germann, 1982). It should not be assumed that in most forest ecosystems

water movement will be confined to the vertical direction (Schecher and



WAYNE P. ROBARGE AND DALE W. JOHNSON



12



Driscoll, 1989). Lateral movement is common in forest ecosystems on

hillslopes (Jones, 1987) and can account for a substantial portion of flow,

especially on an event basis (Fig. 2) (Roberge and Plamondon, 1987;

Gaskin et al., 1989; Hopper et al., 1990).

Rapid movement through a given soil horizon will favor control of the

soil solution by intensity factors and a decrease in the ANC of the drainage

water (Chen et af., 1984). Longer retention within a soil horizon will

increase the ANC of the soil solution, primarily through soil mineral

weathering (Peters and Driscoll, 1987). Prolonged contact between inputs

in drainage water and the soil along preferred flow paths will result in

changes in the nearby soil that are not evident from analysis of the bulk soil



a] AVERAGE ANION FLUX



3

Y



P



TF



SF



FF



BA,



BAl



Btv



Btl



BCv



BCI



SAMPLING LEVEL



Figure 2. Average nutrient flux for all storms at each sampling level; P, precipitation; TF,

throughfall; SF, stemflow; FF, forest floor leachate; BA,, BA soil solution, vertical component; BA, , BA horizon soil solution, lateral component; Bt, , horizon soil solution, vertical

component; Bt,, Bt horizon soil solution, lateral component; BC,, BC horizon soil solution,

vertical component; BC, , BC horizon soil solution, lateral component. After Gaskin et al.

(1989).



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