Advances in agronomy volume 47

DVANCES I N

igronomy

V O L U M E 47


Advisory Board
Martin Alexander

Eugene J. Kamprath

Cornell University

North Carolina State University

Kenneth J. Frey

Larry P.Wilding

Iowa State University

Texas A&M University

Prepared in cooperation with the
American Society of Agronomy Monographs Committee

S. H. Anderson
L. P. Bush
R. N. Carrow

M. A. Tabatabai, Chairman
G. L. Horst
R. J. Lwmoore
R. H. Miller

G. A. Peterson

c.w. Stuber
S. R. Yates


D V A N C E S I N

onomy
VOLUME
47
Edited by

Donald L. Sparks
Department of Plant and Soil Sciences
University of Delaware
Newark, Delaware

ACADEMIC PRESS, INC.
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1


Contents
CONTRIBUTORS
...............................................
PREFACE......................................................

ix

xi

THE
EFFECTS
OF ACIDIC
DEPOSITION
ON FORESTED
SOILS
Wayne P. Robarge and Dale W.Johnson
I . Introduction .............................................
1
I1. Soil Acidification .........................................
I11. Forest Soils ..............................................
IV. Case Studies of Soil Change ................................
V. FutureResearch ..........................................
VI. Conclusions .............................................
References ..............................................

3
6
42
63
65
66

FINGERPRINTING
CROP
VARIETIES
J . S. C. Smith and 0. S. Smith

I . Introduction .............................................
I1. Characters Used to Fingerprint Cultivated Varieties . . . . . . . . . . . .
I11. Discriminational Ability of Fingerprinting Techniques . . . . . . . . . .
Iv. Usage of Fingerprints .....................................
V. NewTechniques .........................................
References ..............................................

85
86
108
119
125
127

TRANSPORT
OF CHEMICALS
THROUGH SOIL:
MECHANISMS.
MODELS.
AND FIELD
APPLICATIONS
William A.Jury and Hannes Fluhler
I . Introduction .............................................
I1. Transport and Transformation Processes ......................
I11. Analysis of Process Assumptions .............................
Iv. Field Studies of Solute Transport ............................
V. Concluding Remarks ......................................
References ..............................................
V

142
143
169
175
191
195


vi

CONTENTS

EVOLUTION
OF CORN
Walton C. Galinat
I. Introduction: Teosinte Is the Wild Corn ......................
I1. Transformation by Domestication and Isolation . . . . . . . . . . . . . . . .

I11. Time Required for Transformation under Domestication . . . . . . . .
IV Multiple Domestications in the Origin of Corn . . . . . . . . . . . . . . . .
V. Interpathway Heterosis ....................................
VI. Husk Enclosure of the Ear .................................
VII . Current Direction of Corn Evolution and Where It Is Going . . . . .
VIII . Summary and Conclusions .................................
References ..............................................

203
205
212
214

221
222
224
227
229

USEOF SURFACE
COMPLEXATION
MODELS
IN SOIL
CHEMICAL
SYSTEMS
Sabine Goldberg
I. Introduction .............................................
I1. Description of Models .....................................

I11. Application of Models to Protonation-Dissociation Reactions
on Oxides. Clay Minerals. and Soils ..........................
Tv: Application of Models to Metal Ion Adsorption Reactions on
Oxides. Clay Minerals. and Soils ............................
V. Application of Models to Inorganic Anion Adsorption Reactions
on Oxides. Clay Minerals. and Soils ..........................
VI. Application of Models to Organic Ligand Adsorption Reactions
onoxides ...............................................
VII . Application of Models to Competitive Adsorption Reactions
on Oxides ...............................................
VIII . Incorporation of Surface Complexation Models into
Computer Codes .........................................
IX. Summary ...............................................
References ..............................................

234
235
251
274
289
308
312
319
320
32 1


CONTENTS

vii

MODELING
THE TRANSPORT
AND RETENTION
OF INORGANICS
IN SOILS
H . M . Selim
Introduction .............................................
Transport Equations ......................................
Equilibrium Retention Models ..............................
Kinetic Retention Models ..................................
Multiple-Reaction Models .................................
Transport and Ion Exchange ................................
Transport in Layered Soil ..................................

References ..............................................

331
333
337
342
350
367
377
381

INDEX........................................................

385

I.
I1.
I11.

N.
V.

VI.
VII .


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Contributors

Numbers in parentheses indicate the pages on which the authors’ contributions begin.

HANNES FLUHLER (141), Institute of Terrestrial Ecology, Swiss Federal
Institute of Technology (ETH), Zurich, Switzerland
WALTON C. GALINAT (203), Eastern Agricultural Center, University
of Massachusetts-Amherst, Waltham, Massachusetts 02 1 54
SABINE GOLDBERG (233), USDA-ARS, U.S. Salinity Laboratory, Riverside, California 92501
DALE W. JOHNSON (l), Desert Research Institute, Reno, Nevada 89J12
WILLIAM A. JURY (141), Department of Soil and Environmental Sciences,
University of California, Riverside, Riverside, California 92J21
WAYNE P. ROBARGE (l), Department of Soil Science, North Carolina
State University,Raleigh, North Carolina 2 769s
H. M. SELIM (3 3 I), Agronomy Department, Louisiana State University,
Baton Rouge, Louisiana 70803
J. S. C. SMITH (85), Plant Breeding Division, Pioneer Hi-Bred International, Inc., Johnston, Iowa 50131
0.S. SMITH (85), Plant Breeding Division, Pioneer Hi-Bred International,
Inc., Johnston, Iowa 50131

ix


This Page Intentionally Left Blank


Preface
Environmental quality and biotechnology are currently two major research areas in the crop and soil sciences. Thus it is appropriate that a
substantial portion of this volume of Advances in Agronomy is concerned
with these topics.
Three chapters deal with aspects of environmental quality. Chapter 1 discusses the effects of acidic deposition on forest soils, emphasizing physical
and chemical processes in forest ecosystems that interact with, modify, and

respond to acidic inputs; presenting case histories of soil changes in regions
where acidic deposition may be causal and in regions where it is not; and giving recommendations concerning future research that is needed to more accurately characterize and develop models for the effects of acidic deposition on
forest soils. Chapters 3 and 6 examine modeling of organic and inorganic
chemical transport in soils. Chapter 3 provides a detailed discussion on the
mechanisms, models, and field applications of chemical transport in soils,
with particular emphasis placed on the importance of approaches that accurately describe chemical transport in heterogeneous field soils. Chapter 6 describes the features of models that govern retention of inorganic solutes in
soils. Single, multiple, and multicomponent or competitive models that are
based on equilibrium or kinetics are presented.
Many equilibrium-based models have been promulgated in the literature to
describe reactions at the solid-liquid interface. One group of these models is
microscopically based and is referred to as surface complexation models.
These models are the subject of Chapter 5. This chapter critically reviews
five common surface complexation models of the mineral-solution interface
and their use in describing soil chemical systems. Common characteristics
and adjustable parameters are covered, the application of models to ion adsorption on soil constituents and soils is presented, and the incorporation of
surface complexation models into computer codes is discussed.
Chapters 2 and 4 deal with the crop sciences. Chapter 2 is concerned with
fingerprinting crop varieties. Topics that are discussed include evolution in
the scope and technology of fingerprinting, characters used in fingerprinting
cultivated varieties, discriminational ability of fingerprinting techniques, use
of fingerprints, and new techniques. Chapter 4 deals with the evolution of
corn. It critically evaluates previous research in this area and assesses current
thinking on this very important crop.
I deeply appreciate the fine contributions from these authors.

DONALD
L. SPARKS
xi

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THE
EFFECTS
OF ACIDIC
DEPOSITION
ON FORESTED
SOILS
Wayne P. Robarge' and Dale W. Johnson2
'Department of Soil Science,
North Carolina State University,
Raleigh, North Carolina 27695
?Desert Research Institute,
Reno, Nevada 895 12

I. Introduction
11. Soil Acidification
111. Forest Soils
A. Physical Factors
B. Chemical Factors
IV.Case Studies of Soil Change
A. Categories of Soil Change and Methods for Measurement
B. Intensity-Type Changes
C. Capacity Changes Due to Leaching
D. Capacity Changes Due to Uptake
E. Seasonal Variations in Exchangeable Cations
F. Interactions of Uptake and Leaching
V. Future Research
VI. Conclusions

References

I. INTRODUCTION
Acidic deposition is composed primarily of N and S acid-forming compounds that undergo gas-phase oxidation and aqueous-phase reactions in
the atmosphere to form nitric and sulfuric acids (Tanner, 1989). Partial
neutralization of these acids by NH3 in the atmosphere results in H + ,
NH;, NO;, and SO:- being the dominant ions in acidic rainwater and

1
Advances m Apmny, Volunr 47
Copyighr Q 1992 by Academic Press, Inc. All rights of reproduction in my form reserved.


2

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).