Post-Agricultural Succession in the Neotropics
Randall W. Myster
Editor
Post-Agricultural Succession
in the Neotropics
Randall W. Myster
Ecosystems Studies
Institute for Tropical Ecosystems Studies
University of Puerto Rico at Rio Piedras
PO Box 21910
San Juan, PR 00931–1910
rmyster@lternet.edu
ISBN: 978-0-387-33641-1
e-ISBN: 978-0-387-33642-8
Library of Congress Control Number: 2007937949
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To Olia, for all her love and support
“Seek simplicity, but distrust it.”
Alfred North Whitehead
Preface
It was 1985, and I remember walking into the office of Peter Morin at Rutgers
University. He told me of some boxes with computer punch cards that had been
littering his office for years and of his desire to get rid of them. Because of my
background in computer science, I said I would take a look. The cards were in bad
shape: moldy, broken in places, and full of cockroaches that were living on what was
left of the rubber bands that once had held them together. This was the Buell-Small
successional data set. I cleaned up the cards, ran them through an old card-reader,
and finally transferred the data onto magnetic tape which I could access using my
office computer.
While I was at Rutgers, I also managed to add five years to the data set by
going out to Hutcheson Memorial Forest with Steward Pickett and others where
we sampled the old fields that comprised the Buell-Small study. Accompanying
us was a kindly and cheerful old woman, the last surviving member of the team of
scientists who had the foresight to set up those permanent plots in the late 1950s and
early 1960s. Her name was Helen Buell. She had an infectious sense of humor—her
bumper-sticker read “I brake for Verbascum”—which came in handy as we crawled
through rose bushes and prickly raspberries. I have heard she is gone now, but I
would like to remember her here.
Inspired by the work with the Buell-Small data set, I set up post-agricultural
plots in the Neotropics that are reported on in several papers and in the first chapter
of this book. I believe, as many others do, that such plots are critical to a complete
understanding of vegetative changes after agriculture. I hope that those data and
other data analyzed and interpreted by the authors of the present book will be a
valuable contribution to this scientific undertaking, as has been true for the data
from the plots set up so many years ago in the temperate fields of New Jersey.
R. W. M.
vii
Contents
Part I Patterns of Living and Non-living Components
of Old Fields After Abandonment
1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
Randall W. Myster
2 Trends in Above and Belowground Carbon with Forest Regrowth
After Agricultural Abandonment in the Neotropics . . . . . . . . . . . . . . .
Erika Marín-Spiotta, Daniela F. Cusack, Rebecca Ostertag
and Whendee L. Silver
3
22
3 Post Sugar Cane Succession in Moist Alluvial Sites in Puerto Rico . . .
Oscar Abelleira Martínez and Ariel E. Lugo
73
4 Soil Fungi and Macrofauna in the Neotropics . . . . . . . . . . . . . . . . . . . .
Yiqing Li and Grizelle González
93
5 Earthworms and Post-Agricultural Succession . . . . . . . . . . . . . . . . . . . 115
Grizelle González, Ching Yu Huang and Shu-Chun Chuang
Part II Mechanisms and Tolerances that Cause Plant-Plant
Replacements Producing Old Field Vegetation Patterns
6 Factors Affecting the Species Richness and Composition
of Neotropical Secondary Succession: A Case Study of Abandoned
Agricultural Land in Panama . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 141
Elaine R. Hooper
7 Seed Dispersal by Cattle: Natural History and Applications
to Neotropical Forest Restoration and Agroforestry . . . . . . . . . . . . . . . 165
Clara Luz Miceli-Méndez, Bruce G. Ferguson
and Neptalí Ramírez-Marcial
ix
x
Contents
8 The Roles of Disperser Behavior and Physical Habitat Structure
in Regeneration of Post-Agricultural Fields . . . . . . . . . . . . . . . . . . . . . 192
K. Greg Murray, Kathy Winnett-Murray, Jason Roberts,
Katherine Horjus, William A. Haber, Willow Zuchowski,
Mark Kuhlmann and Tammy M. Long-Robinson
9 Neotropic Post-Dispersal Seed Predation . . . . . . . . . . . . . . . . . . . . . . . . 216
Randall W. Myster
10 Facilitation Versus Competition in Neotropical Old-Fields:
A Case Study After Pinus taeda Cultivation in Brazil . . . . . . . . . . . . . 221
Gislene Ganade, Lessandra Zanini and Ingo Hübel
11 Recruitment of Dry Forest Tree Species in Central
Brazil Pastures . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 231
Aldicir Scariot, Daniel L. M. Vieira, Alexandre B. Sampaio,
Ernestino Guarino and Anderson Sevilha
Part III Restoration, Management, and the Future
of Post-Agricultural Areas
12 Microbial Biomass in Native Amazonian Ecosystems
and Its Adaptation to Deforestation and Pasture Introduction
and Management . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 247
Brigitte J. Feigl, Carlos Eduardo P. Cerri, Carlos C. Cerri
and Marisa C. Piccolo
13 Management for Sustainability and Restoration of Degraded
Pastures in the Neotropics . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 265
Florencia Montagnini
14 Conclusions, Synthesis, and Future Directions . . . . . . . . . . . . . . . . . . . . 296
Randall W. Myster
Index . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 305
List of Contributors
Carlos C. Cerri
Centro de Energia Nuclear na
Agricultura (CENA)
Universidade de Sao Paulo
CP.96, 13400-970
Piracicaba, (SP), Brazil
Carlos Eduardo P. Cerri
Centro de Energia Nuclear na
Agricultura (CENA)
Universidade de Sao Paulo
CP.96, 13400-970
Piracicaba, (SP), Brazil
Shu-Chun Chuang
Institute of Zoology
National Taiwan University
No. 1 Roosevelt Road
Section 4, Taipei 106
Taiwan
Daniela Cusack
Department of Environmental Science,
Policy & Management
University of California, Berkeley
137 Mulford Hall #3114
Berkeley, CA 94720-3114
Brigitte J. Feigl
Centro de Energia Nuclear na
Agricultura (CENA)
Universidade de Sao Paulo
CP.96, 13400-970
Piracicaba, (SP), Brazil
Bruce G. Ferguson
El Colegio de la Frontera Sur Carretera
Panamericana y Periférico Sur s-n
Barrio Ma. Auxiliadora San Cristóbal
de Las Casas
Chiapas
México CP 29260
Gislene Ganade
Laboratório de Ecologia
da Restauração Biologia
Universidade do Vale do Rio dos Sinos
CP 275, São Leopoldo
RS, Brasil 93022-000
Grizelle González
International Institute
of Tropical Forestry
USDA Forest Service
1201 Ceiba St.
San Juan, PR 00926-1115
Ernestino Guarino
Embrapa Recursos Genéticos e
Biotecnologia
Parque Estação Biológica - PqEB - Av.
W5 Norte (final)
Caixa Postal 02372
70770-900 Brasília DF – Brazil
William A. Haber
Apartado 50-5655
Monteverde de Puntarenas
Costa Rica
xi
xii
Elaine R. Hooper
Department of Forestry
and Environmental Studies
Yale University
210 Prospect St
New Haven, CT 06511
Katherine Horjus
Department of Ecology
and Evolutionary Biology
University of California
Santa Cruz, CA 95062
Ching Yu Huang
Institute of Ecology
University of Georgia
Athens, GA, USA
Ingo Hübel
Laboratório de Ecologia da
Restauração, Biologia
Universidade do Vale do Rio dos
Sinos, CP 275
São Leopoldo
RS, Brasil 93022-000
List of Contributors
Ariel E. Lugo
International Institute
of Tropical Forestry
USDA Forest Service
Ceiba Street 1201
Jardín Botánico Sur
Río Piedras PR 00926-1115
Erika Marín-Spiotta
Department of Environmental Science,
Policy & Management
Univeristy of California, Berkeley
Current address:
U.C. Santa Cruz, Rm. 2125,
USDA/ARS
800 Buchanan St.
Albany, CA 94710
Oscar Abelleira Martínez
International Institute
of Tropical Forestry
USDA Forest Service
Ceiba Street 1201
Jardín Botánico Sur
Río Piedras, PR 00926-1115
Mark Kuhlmann
Department of Biology
Hartwick College
West Street, Oneonta
New York 13820
Clara Luz Miceli-Méndez
El Colegio de la Frontera Sur Carretera
Panamericana y Periférico Sur s-n
Barrio Ma. Auxiliadora
San Cristóbal de Las Casas
Chiapas, México
CP 29260
Yiqing Li
College of Agriculture,
Forestry and Natural Resources
Management
University of Hawaii
200 W. Kawili St.
Hilo, HI 96720
Florencia Montagnini
School of Forestry
and Environmental Studies
Yale University
370 Prospect St
New Haven, CT 06511
Tammy M. Long-Robinson
12024 E. Pleasant Lake
Manchester, MI 48158
K. Greg Murray
Department of Biology
Hope College
Holland, MI 49423
List of Contributors
Randall W. Myster
Institute for Tropical Ecosystem Studies
University of Puerto Rico at Rio Piedras
P.O. Box 21910
San Juan,
PR 00931-1910
Rebecca Ostertag
Department of Biology
University of Hawaii at Hilo
200 W. Kawili St.
Hilo, HI 96720
Marisa C. Piccolo
Centro de Energia Nuclear na
Agricultura (CENA)
Universidade de Sao Paulo
CP.96, 13400-970
Piracicaba, (SP), Brazil
Neptalí Ramírez-Marcial
El Colegio de la Frontera Sur Carretera
Panamericana y Periférico Sur s-n
Barrio Ma. Auxiliadora
San Cristóbal de Las Casas
Chiapas, México
CP 29260
Jason Roberts
128 Nichols St.
Lewiston, ME 04240
Alexandre B. Sampaio
IBAMA
SCEN Trecho 2 - Ed. Sede - Caixa
Postal 09870
70818-9 – Brasília, DF – Brazil
Aldicir Scariot
Embrapa Recursos Genéticos e
Biotecnologia
Parque Estação Biológica - PqEB - Av.
W5 Norte (final)
Caixa Postal 02372
70770-900 Brasília DF – Brazil
xiii
Anderson C. Sevilha
Embrapa Recursos Genéticos e
Biotecnologia
Parque Estação Biológica - PqEB - Av.
W5 Norte (final)
Caixa Postal 02372
70770-900 Brasília DF – Brazil
Whendee L. Silver
Department of Environmental Science,
Policy & Management
University of California Berkeley
137 Mulford Hall #3114
Berkeley,
CA 94720-3114
Daniel L. M. Vieira
Embrapa Tabuleiros Costeiros
Av. Beira Mar, 3250 - jardins
Caixa Postal 44
49025-040- Aracaju, SE - Brazil
Kathy Winnett-Murray
Department of Biology
Hope College
Holland, MI 49423
Lessandra Zanini
Laboratório de Ecologia da Restauração
Biologia, Universidade do Vale do Rio
dos Sinos
CP 275, São Leopoldo
RS, Brasil 93022-000
Willow Zuchowski
Apartado 50-5655
Monteverde de Puntarenas
Costa Rica
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Part I
Patterns of Living and Non-living
Components of Old Fields After
Abandonment
Chapter 1
Introduction
Randall W. Myster
1.1 Rationale
Ecosystems were originally defined as units of the earth’s surface which includes
both organisms and the physical factors forming the environment (Tansley 1935).
As the science of ecosystem ecology developed, ecosystems came to be categorized by function and structure (Odum 1953), with an emphasis on integration and
indirect interaction (Muller 1997). While all ecosystem functions deal with energy
capture/transfer and nutrient cycling, plant-based productivity and decomposition
have come to be considered among the most critical (Watt 1947).
To develop a deeper understanding of ecosystem functions, we need to know the
pattern of an ecosystem’s abiotic and biotic responses which define function, on
the one hand, and the processes and mechanisms that regulate the responses, on the
other. In particular for the biota, it has long been recognized that plants and other
taxa respond not only to gradients (Whittaker 1975, Kessell 1979, Walter 1979)
but also to disturbances (i.e., disruptions of ecosystem structure that occur largely
on gradients: Gleason 1926, Pickett and White 1985). Consequently, the vegetation
mosaic of any landscape results both from the environmental variation expressed on
gradients and from historical disturbances (Risser 1987, Turner and Dale 1991).
Gradients are gradual, directional, and large-scale (from a plants’ perspective)
changes in environmental factors (Whittaker 1975, Hallet al. 1992). Although gradients are mainly a spatial phenomenon, they may also show seasonal variation
and even be modified over long time periods, for example, by changes in climate.
Individual plants of a given species exist within certain ranges of environmental conditions (Good 1953) that, when combined with others of the same species, translate
into distribution patterns (usually bell-shaped, Gaussian Curves) across gradients
for that species (Whittaker 1975). Even though these curves correspond with direct
gradients, such as elevation, slope, and aspect, they actually reflect species-specific
responses to indirect environmental factors, such as the plant resources of temperature, light, water, and soil nutrients, which correlate with the direct gradients. Such
plant growth responses exist in a hierarchy of survivorship, growth, allocation, and
architecture.
Unlike gradients, disturbances are local, discrete, and patchy (Pickett and White
1985). However, similar to gradients, they create combinations of resources and
R. W. Myster, Post-Agricultural Succession in the Neotropics.
C Springer 2008
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R. W. Myster
other influences that plants respond to. Indeed, plant distribution patterns over successional time, after a disturbance, are just as species-specific and bell-shaped as
they are over gradient space (Pickett 1982, Myster and Pickett 1988).
Any mechanism involving interactions with individuals of another species taxa
(e.g., dispersal, predation, herbivory) and any tolerance to a plant resource/abiotic
environmental cue (e.g., light, water) that affect any part of a plant’s life-cycle can
form a response curve. Importantly because plants interact with each other by affecting their local resource environment, interactions such as competition and facilitation are expressed as tolerances. Taken together, response curves for each species
define that species’ “niche” (Whittaker 1965, Grubb 1977, Parrish and Bazzaz 1982,
Bazzaz 1996) that can be quantified using yes/no questions at each life-stage and
for each tolerance or mechanism that affects it. Typical questions during a tolerance
investigation could be: Will this seed germinate at a given temperature? Will this
seedling have a positive carbon balance at a given light level? During a mechanism investigation, typical questions include: Will this seed be dispersed at a given
distance from the forest? Will this seedling survive at a given herbivore density?
Each time the answer is “yes”, a point that represents a particular individual plant
is marked along the natural variation scale of the abiotic/biotic environmental cue
being investigated, such as the level of temperature, light, distance, or predator density. When the responses of many individuals from the same life-stage and the same
species are also marked on the same natural variation scale and then counted, a
frequency diagram for the entire species is created.
Because these curves are frequency diagrams and because individual plants can
be chosen at random, the response curves/surfaces express probabilities (after normalization: Zar 1999) of how individual plants of that species will respond to the
natural variation in the working of a given mechanism or in the availability of a
given resource.1 These response curves give the individual response variation within
a plant species for a given life-stage and environmental cue, and are generally wider
for species that come early in successional time compared to species that appear
later (Bazzaz 1979, 1996). Usually, a few of the plant mechanisms or resources
dominate, making the probabilities associated with them key to how the individual
plants of that species use the strategies of capture, congruence, and capacitance
(Bazzaz 1996). Furthermore, the fact that the response curves have the same shape as
plant distribution patterns both over space (Whittaker 1975) and over time (Pickett
1982, Myster and Pickett 1988) suggests that the distribution of a plant species is an
outcome of that species’ niche probabilities.
Conducted for decades over many of the most common gradients on earth, studies of succession after agriculture (also called old field succession) have helped
ecologists gain insights into the processes that structure plant communities and into
the role of history and initial conditions in community development (Myster and
Pickett 1990b, 1994). Examination of old fields has led to the rise of important
ecological theories, such as the initial floristic composition hypothesis (Egler 1954)
1
As an example of a response surface which expresses probabilities, see Figure 3 in Burton and
Bazzaz (1991) for emergence of five common old field trees in the American temperate zone.
1 Introduction
5
and the resource ratio hypothesis (Tilman 1988). Using permanent plots established
at abandonment as the backbone of old field research (e.g., the Buell-Small old
field plot study in New Jersey USA: Buell et al. 1971, Myster and Pickett 1990a),
scientists have been successful in finding the pattern of response over time after
abandonment from crops (e.g., plant cover [Myster and Pickett 1992b] and diversity [Myster 2003b]) and in discovering many of the mechanisms that determine
the pattern (Myster 1993, Bazzaz 1996). Furthermore, studies of post-agricultural
succession have provided a framework for comparison and evaluation of various
theories of succession and community development (Finegan 1984).
A considerable amount of old field and pasture research has taken place in
the temperate regions of the United States (e.g., Oosting 1942, Bazzaz 1968, Buell
et al. 1971, Pickett 1982, Tilman 1988, Myster 1993) and Europe (Castro et al. 1986,
Miles 1979). The recovery of agricultural areas, however, is vital to ecosystems all
over the world (Borhidi 1988). In the tropics, for example, the cutting down of
wooded areas (Grainger 1988, Fearnside 1993) for agriculture and pasture (Skole
and Tucker 1993) is the major cause of deforestration (Buschbacher 1986, Uhl
et al. 1988, Fearnside 1993, Thomlinson et al. 1996). It should come as no surprise
then that recovery of Neotropical areas after agriculture (Donfack et al. 1995,
Fernandes and Sanford 1995, Quintana-Ascencio et al. 1996) is relevant to such
important issues as forest regeneration (Grainger 1988, Brown and Lugo 1990,
Myers 1991, Singh 1993, Skole and Tucker 1993), forest ecosystem restoration
(Brown and Lugo 1990, Lugo 1992), sustainability of agriculture (Brown and Lugo
1990, Serrao and Toledo 1990), maintenance of biodiversity (Brown and Lugo
1990, Aide and Cavelier 1994), and impacts of global climate change on forest
dynamics (Padoch and Vayda 1983, Hobbie 1992, Mabberley 1992, Keller et al.
1993). Neotropical areas recovering from agriculture may also serve as a buffer
between “primary” forest and more intensely human-influenced areas (Brown and
Lugo 1990). Last but not least, it is expected that with future increase of global
warming forest disturbance in general and old field succession in particular will
affect larger and larger parts of the earth (Bazzaz 1996).
The nature of the recovery of old fields depends on their disturbance “regime,”
which is characterized by severity, size, spatial location, and frequency (Myster
2001). The measurement of severity is best accomplished by sampling the loss of
plant biomass and determining whether such loss includes loss of plant meristems
(Raunkaier 1934, Pickett and White 1985, Tilman 1988, Myster 2001). Severity
of old fields is moderate, compared to the more severe landslides (Myster 2001),
because old fields maintain an intact soil profile, roots, and humus layer despite
the removal of above-ground vegetation. In size, old fields do not usually exceed a
few tens of hectares (they are larger, however, than many Neotropic treefall gaps:
Brokaw 1982), and crops that are planted in them often require specific locations
(e.g., Coffee grows well in the mountains while Banana and Sugarcane do best
in the valleys). Frequency of old field disturbance includes the number of times
a field has been cultivated, the duration of each cropping period, the time between
fallow periods, and the order of crop rotation. Usually fields are fallow longer than
they are in crop and can be either reused with different crops and fallow periods
until abandoned or used for 5-10 years until productivity declines. As the length of
6
R. W. Myster
time an area is in crop increases, the length of successional, or recovery, time may
also lengthen. For example, a short cropping period can lead to large contributions
from the seed bank and trees may regenerate quickly from stump or root sprouts,
whereas an increased cropping period brings about the decay of rootstocks and tree
stumps, thus providing for grass invasion and prolonged tree invasion. Among other
conditions that influence old field recovery are historical effects of the past crop (a
crop’s “signature”: Myster and Pickett 1988, 1990b, 1994), the year of abandonment, the season of abandonment, the percent of border with forest (Bazzaz 1968),
and a field’s plowing record (whether the field was plowed under or left fallow when
abandoned: Myster and Pickett 1990a).
Studies of post-agricultural succession in the temperate zone have demonstrated
the particular importance of past crops in determining old field patterns (e.g., alteration of successional pathways and species composition and abundance for up to 8
years after abandonment: Myster and Pickett 1988, 1990b, 1994). Hence, a major
focus of this book will be on the examination of fields after common Neotropic
crops, such as Maize (Zea mays), Sugarcane (Saccharum officinarum), Banana
(Musa spp.), Coffee (Coffea spp.), Cassava (Manihot esculenta), Beans (Phaseolus
spp.), and Rice (Oryza spp.), and after pasture grasses that usually establish themselves after cropping but can also be purposely planted (e.g., Setaria sphacelata).
Pastures will be treated as a kind of old field succession for several reasons. First,
they are usually converted from depleted agricultural fields when cows are allowed
to enter. Second, they may continue to have remnants of past crops growing in them
for some years. And third, they can be planted with native and exotic grasses or be
colonized by local grass species due to cattle activity (e.g., trampling of vegetation,
creation of hummocks, deposition of dung, and soil compaction).
Two forest-clearing techniques that are the most common in the Neotropics will
also be discussed in this book. Both methods, while increasing light quantity and
contact of rainfall with the soil, eventually decrease soil fertility. The first technique is mechanical clearing that aims at removing woody biomass (e.g., by using
a bulldozer) and brings about changes in physical properties of the soil. The other
technique is shifting, or “slash and burn” (Uhl 1987), agriculture that occurs when
natural vegetation is first cut down to the ground and then burnt. Although burning
produces a short pulse of nutrients like phosphorus, it generally volatilizes both
organic matter and nitrogen while raising soil temperature and, in the long term,
makes soil poor in both nutrients and the seed bank. Shifting agriculture also entails
a rotation of fields rather than crops that is accompanied by long fallow periods
(Grigg 1978). The decision as to when to rotate fields and for how long to leave them
fallow depends on such factors as weed and insect loads in the fields, soil fertility,
production, labor availability, and local dietary needs (Staver 1990). The implementation of either one of these forest-clearing methods also causes carbon loss, a
mixing of the soil horizons, and a higher degree of soil aeration (Bazzaz 1996).
Both mechanical clearing and shifting agriculture lower the level of advanced
regeneration from sapling banks and from root/stem resprouting (Uhl et al. 1988)
so that seed and seedling dynamics dominate regeneration after agriculture. Consequently, an understanding of old field succession demands knowledge of invasion
and establishment at the seed and seedling phases of the plant life-cycle (Grubb
1977, Myster 1993).
1 Introduction
7
Old field recovery goes through several structural and conceptual stages: (1)
domination by past crop and its various effects, (2) patch dynamics of grass, past
crop remnants, asexual shrubs, and other plants, (3) tree invasion, and (4) development of a closed-canopy “secondary” forest. Although individual plants may or may
not colonize bare ground during these stages, they always become involved in the
on-going process of plant-plant replacement (Busing 1996). Plants enter this process because they grow and interact asymmetrically (Bazzaz 1996)—one individual
is always gaining advantage over another. As a result, the plants that are growing
best replace the plants that are growing worst, are in senescence or dead. However,
plant-plant replacement does not necessarily involve two plants only, for more than
one plant can be replaced by just one or only one plant can be replaced by more than
one. Because individual plants are “planted” in the soil, they have a defined physical
space which can be occupied and reoccupied as plant-plant replacements proceed.
Every land plant in the world sooner or later assumes these roles-of the “replacer”
and the “replaced.”
Because plants move very slowly, we naturally do not see them and their communities changing. Yet, they are always both engaged in and are the product of the
dynamic process of plant-plant replacement. Plant-plant replacements determine
which plants are present at every point in space and time, for example, during a
typical “static” plant growth and allocation field experiment or a “snapshot” plot
sampling. Even in those cases, however, plant niches and their interaction with
local environmental cues will determine how those plants grow. Fundamentally,
neither species nor patches are replaced in communities. It is individual plants that
are replaced by other individual plants. I suggest that plant communities do not go
through continuous cycles of disturbance (regeneration) => stability (growth) =>
disturbance (regeneration) –like a corn field– but instead are undergoing plant-plant
replacements at all times.
After a closed canopy has developed, old field plants may eventually enter a
thinning state when dead plants in the canopy are replaced not with new individuals
but with modified growth of existing plants. There is a physiological limit to this
kind of growth, however, and at some point every plant will die and be replaced. Furthermore, although there can be a growth phase for perennial plants between replacing and being replaced, it is rare, transitory, and, consequently, over-emphasized in
plant ecological studies. This focus on individual plant-plant replacements eliminates the need for many of the old dichotomies of plant ecology, such as primary/secondary forests, natural/human disturbances, successional/nonsuccessional
communities, and pioneer/climax species. This conceptual model unifies various old
theories of plant community organization because plant-plant replacements define
not only the successional dynamics of old fields but the dynamics of any plant community.
Replacements are controlled by species tolerances and mechanisms that “activate” the niches of the species involved in the replacement. Tolerances include germination and growth while mechanisms encompass dispersal, seed pool, seed predation, seed pathogens, seed germination, seedling predation, seedling pathogens,
seedling herbivory, and/or seedling competition (e.g., Myster and McCarthy 1989,
Myster and Pickett 1993, Myster 1994, 2003a, 2003c, 2004a, 2006, 2007). The process of plant-plant replacement is at the heart of the dynamic nature of terrestrial
8
R. W. Myster
plant communities both over successional time and over space (e.g., at the edges of
plant distributions where gradients have a major effect on plants: Whittaker 1975,
Myster 2001).
For those who still need to be convinced of the dynamic nature and ubiquity
of plant-plant replacements, I suggest setting aside a plot of ground, marking a
few hundred plants, and coming back in a few years (longer for perennial plants)
to see how many individuals have been replaced. The data I have from permanent plots sampled for decades in a variety of both Neotemperate and Neotropical old fields clearly show that replacements are working everywhere and always.
Replacements are also a very common phenomenon in so called “primary” Neotropical forests (e.g., in the 50 ha plot located at Barro Colorado Island, Panama
http://ctfs.si.edu/datasets/bci). It seems that plants wait not to be counted (sensu
Harper 1977) but to be replaced!
The working model of plant-plant replacements entails a strong possibility that a
plant will be replaced by another plant (whether of the same or of a different species)
when the probabilities expressed in the species’ niche of that plant are lower than the
probabilities expressed in the species’ niche of the other plant, those probabilities
being activated by the combination of abiotic and biotic conditions at a given point
in space and time and for that particular life-stage in the plants’ development. I
suggest that it is the smallest probability—from among the many niche “slices” for
each species that are activated by local environmental conditions—that will be the
most limiting and, therefore, most critical in determining whether or not one plant
will replace another.
Replacements create changes in plant abundances, which may also create
changes in plant composition that lead to the emergence of other patterns at larger
spatial and temporal scales (e.g., successional rate and direction, patch dynamics,
plant distribution, exotic plant invasion, plant associations). Care must be taken,
however, with the observation of these plant patterns. For example, even though
individual plants continue to be replaced, preservation of species abundances
and/or species composition at a larger scale of organization may still occur, giving
the impression of a “coexistence” of species (when the identity or number of
plant species remain the same despite changes in plant abundances) or of a plant
community at “equilibrium” (when neither plant species nor plant abundances
change). Existence of high-level patterns that do not change very much over time
should not obscure the fact that such patterns are the product of a dynamic and ever
changing process of plant-plant replacement. Any plant pattern is only a snapshot
of the vegetation at a given instance in time and at a specific scale of observation,
whereas plant-plant replacements occur continuously.
Because plant niches express probabilities, plant-plant replacement itself must
be a probabilistic process. In fact, succession is probabilistic (Pickett 1982), for it
is made up of a multitude of plant-plant replacements that take place over time (for
example, when all the individuals of a given species leave a disturbed area and/or an
individual of a new species arrives, creating a change in species composition: Myster and Pickett 1994, Brokaw and Busing 2000). Again, the dynamics of any plant
community is probabilistic because a plant community—and any of its plant-based
structures and functions—is defined by its’ resident plants that have been brought
forth by a probabilistic plant-plant replacement process (Pickett 1982, Myster and
1 Introduction
9
Pickett 1988). Thus the ultimate reality of every plant community are the plant-plant
replacements, for they produce vegetation changes at any and all scales.
Along with succession, plant associations, and species coexistence/equilibrium,
distribution of plant species exemplifies how plant-plant replacements determine
patterns of a plant community at higher levels of organization. Since plant distribution patterns are shaped by plant-plant replacements, there is a strong possibility
that distribution of a species will increase if, under given conditions, the species’
niche probabilities are higher than those of another species’ at the edges of the two
species’ distribution where the species meet, or decrease if the first species’ probabilities are lower. For example, red cedar (Juniperus virginiana L.) is expanding its
range in the United States by replacing plants of various other species at the edges
of its distribution. The success of red cedar’s invasion finds its explanation in the
species’ regeneration niche (those parts of a plants’ niche that represent seeds and
seedlings: Grubb 1977) that includes higher probabilities in seed dispersal, extended
seed viability in the soil, lower seed predation, lower herbivory, and higher drought
resistance than many other species (reviewed in Myster 1993).
1.2 Species Associations in Permanent Plots
Studies in the temperate zone of the Americas show that the spatial and temporal
distribution of old field plants is not random and that it is associated with various
factors (such as the location of “nurse” trees and plant resource levels: Myster and
Pickett 1992a, Myster 1993). In the course of an investigation of these temporal and
spatial patterns, numerous questions present themselves: When do species appear
and when do they leave? How long are they present? Are there edge effects from the
forest? Do species clump together and, if they do, does this autocorrelation change
with time? Chronosequences supply some evidence to answer these questions. However, inasmuch as they place different fields sampled at different times sequentially
on a time line, this approach offers only, at best, a partial solution. The true answers
come from regularly sampled permanent vegetation plots that are established at the
abandonment of agricultural fields because such plots show the actual sequences of
plant abundance and species change over time (Pickett 1982) an important aspect of
the dynamic nature of plant communities (Myster 2001).
As case studies, let us consider plots in natural pasture in Puerto Rico and plots
in former Banana plantations, former Sugarcane plantations, and seeded pastures in
Ecuador. The first study site is a natural pasture that had been grazed for decades
before it was abandoned. It borders the Luquillo Experimental Forest (LEF) of
northeastern Puerto Rico, USA (18◦20’N, 65◦ 45’W: Aide et al. 1995, Liu and
Zou 2002), close to the town of Sabana. The LEF, a long-term ecological research
(LTER) site of the National Science Foundation (www.luq.lternet.edu), is a tropical montane wet forest with tabonuco (Dacryodes excelsa), ausubo (Manilkara
bidentata) and motillo (Sloanea berteriana) below 600m, palo colorado (Cyrilla
racemiflora) and palm (Prestoea montana) between 600m and 850m, and cloud
10
R. W. Myster
forest above 850m (Waide and Lugo 1992). The study pasture is located at the lower
elevations and receives between 2m and 5m of rainfall per year, with an average
temperature of 18◦C. Its soils are fertile and volcaniclastic in origin (Thomlinson
et al. 1996).
Twenty-five 5m ×2m contiguous plots were laid out on the border with the forest
(Myster 2003b), the long side parallel to the forest in order to maximize any edge
effects and to facilitate comparison to the Buell-Small plot study in New Jersey,
USA (Myster and Pickett 1990a). Since island diversity in plant species is greatly
reduced compared to mainland diversity, the plots were big enough to capture diversity in these agrosystems (Myster 2004a). The plots did not have any remnant trees
or sprouting tree roots at the beginning of the study, and their tree seed bank was
very small (Myster 2006). Starting in May of 1997 and continuing annually for
10 years, the plots were sampled for percent cover of each plant species. Percent
cover—an indication of a species’ ability to capture light and, therefore, to dominate
these areas in the process of becoming forested communities (Myster 2003b)—was
estimated visually in relation to each plot’s area.
The analysis of the data obtained from these plots showed that grass dominated
throughout the first 5 years after abandonment, constituting 75% of the plant cover
after 5 years, while forbs declined to 40% and ferns and woody species increased
to 45 and 15% respectively. Common trees included Syzygium jambos, Calophyllum calaba, and Tabebuia heterophylla. Common shrubs were Clidemia hirta and
Miconia spp., both of which invaded from the forest. The community parameters
that peaked within the first 5 years after abandonment were productivity (at 400
g/m2 /yr), total basal area (at 1,000 cm2 ), and richness (at 19 species: Myster 2003b).
The second study site comprises recovering Banana plantations, Sugarcane
plantations, and seeded pastures at the Maquipucuna Reserve, Ecuador (0◦05’N,
78◦ 37’W; www.maqui.org; Sarmiento 1997, Rhoades et al. 1998, Rhoades and
Coleman 1999, Zahawi and Augspurger 1999, Myster 2004a, 2004b). Maquipucuna
lies between 1,200m and 1,800m and is classified as tropical lower montane wet
forest (Edmisten 1970). It has deeply dissected drainages with steep slopes and
receives between 2m and 5m of rainfall a year. The temperature ranges between
14◦ C and 25◦ C, with an average temperature of 18◦ C. The reserve’s fertile andisol
soil is developed from volcanic ash deposits (Myster and Sarmiento 1998).
In June of 1996, 6 agricultural fields were selected: two Sugarcane (Saccharum officinarum) plantations, two Banana (Musa sp.) plantations, and two pastures
seeded in Setaria sphacelata.2 All 6 fields were 2 ha rectangles located within a few
hundred meters of each other at the lower elevations of the reserve, on the border
with the primary forest. In each field, twenty-five 5m × 2m contiguous plots were
laid out, with the long side on the forest border. This forest is also tropical montane
wet forest.
2
Although native to Africa, the Setaria grass is being planted more and more extensively in the
Neotropics because it can withstand heavy grazing by spreading asexually and can form large
crowns. Well accepted by cattle, the Setaria grass is tolerant of relatively cold temperatures and
short-term waterlogged conditions, both common at the Maquipucuna Reserve.
1 Introduction
11
Starting in 1997, these subplots were sampled annually for 10 years for percent cover of each plant species. Maquipucuna plant taxonomists, trained at the
University of Georgia, USA, where voucher specimens are kept on file (Myster
and Sarmiento 1998, Zahawi and Augspurger 1999), assisted in the identification
of species by using specimens located on site. The analysis of the data from the
first 5 years after abandonment showed that the resident graminoid dominated in
Sugarcane (at 50% after 5 years), in Banana (at 70%), and in pasture (at 90%), that
forbs and ferns were at low levels everywhere, and that woody species increased
in Sugarcane (at 45% after 5 years) and in Banana (at 25%) but remained sparse
in pasture (at 10%). Common trees included Acalypha pladichephalus, Costus sp.,
Vernonia patens, and Piper aduncum while common shrubs were Pilea sp. and
Miconia spp. After 5 years, both total basal area and species richness were greatest
after Sugarcane (at 700 cm2 and 59 species respectively), lower after Banana (at
600 cm2 and 49 species), and lowest after pastures (at 100 cm2 and 25 species:
Myster 2007).3
The plot data were used to implicate plant interactions and successional mechanisms by generating spearman coefficients of rank associations (SAS 1985, Milbau
et al. 2005). The analyses showed that the actual number of associations among
species was much lower than the potential number of 3,250 per field (all pairs of 26
species over the 10-year period). For example, the pasture in Puerto Rico had a total
of 362 associations, 55 of which were negative (Table 1.1), while one of the two
Banana fields in Ecuador had a total of 168 associations, 12 of them being negative
(Table 1.2).
Because controlled field experiments have confirmed that species that interact
significantly also form significant associations with each other (Tirado and Pugnaire 2005), the statistics from the case studies (Tables 1.1–1.7) point to a low
level of species interactions in these fields (as was also seen in temperate old field
plots in New Jersey, USA: Myster and Pickett 1992b). These results support the
view of communities as loosely-organized assemblages of species that are created
mainly by tolerance parts of individual species’ niches (Gleason 1926, Whittaker
1975) rather than as tightly-linked entities that are structured by species interactions.
Most of the significant associations in the study plot data were positive (unlike
those, mainly negative, associations that were computed from plots in temperate
old fields: Myster and Pickett 1992b). This suggests that facilitation in these earlysuccessional plots where plants are under stress (see Chapter 10) may be more
important than competition among those relatively few species that are actively
interacting. In the Puerto Rican pasture (Table 1.1), species that formed many positive associations included the trees Syzgium jambos, Guarea guidonia, Ocotea leucoxylon, and Prestoea montana. Among major inhibiting species were the grass
Panicum spp., the fern Gleichenia bifida, and the tree Tabebula heterophylla. In
3
The data from the plots in Puerto Rico (LTERDATB #97) and in Ecuador (LTERDATB#101) are
housed in the archives of the LEF LTER site. Visit the Luquillo web site (www.luq.lternet.edu) for
further details.
H
I
J
K
L
M
N
O
P
Q
R
S
T
U
V
W
X
Y
B C
D E
F
A +– +
– ++
–
B
++++++ ++ +++++++ –
++
C
+ +++
D
+
++
–+
E
F
G
+
++
+
++
+++
G
+
++
I J
++ –
++ ++++
+
++++ ++
+
+
++ ++
+
H
L
+++
+
–
M N
+++
O
++
+++ – –
++
++ ++
++–
++ +
+
++++
++
+
+
+++
+ +++ ++
+
+
+–
+++++++ ++ +
– ++
–
++++
–
––
+++++
++
+
++
++
+
+++++
+
+
+++++ ++
K
+
R
+
+
S
++ +
–– – + –
+
+
– – +++
Q
++
++++
+
+
+++ + +
+
++
++ +– +
++
+
++
+ ++++
+
–– –
++
–
+
+++
P
V
W
+
–
+
–+
+
++
+
++++
Z
–++
++++ ++
Y
–
++ –
+
+
+
++
+++++
++
+++
+
+
++
+
++++ +
+
++
++
–
+++
–
+
++
++
+++
++++ +++
+–
++
––
++++ ++++ ++
-++
X
+
+++++- +– – – ––++
++
+
++
++
–
++
+
+
+-
++ +++ – +–
+
+
–– +
++
+++ – – – + ++
U
+
+
+++++
+
++
+++
+
+
+
+
+
++
+
T
–
Table 1.1 Significant spearman rank correlation coefficients among all plant species in the Puerto Rican pasture over the first 10 years of succession. Significant
positive associations are indicated by a “+” and significant negative associations are indicated by a “−”. Each matrix entry contains all significant associations,
with a maximum of ten, between two species in a certain field. Plant species and families are indicated as follows: Bromelia spp. (A), Guarea guidonia (B),
Ocotea leucoxylon (C), Citrus frutus (D), Syzgium jambos (E), Desmodium spp. (F), Gleichenia bifida (G), Inga laurina (H), Citrus limon (I), Casearia sylvestris
(J), Prestoea montana (K), Calophyllum calaba (L), Miconia prasina (M), Eugenia pseudopsidium (N), Tabebuia heterophylla (O), Eugenia malaceensis (P), Piper
hisperdium (Q), Andira inermis (R), Psychotria brachiata (S), Miconia racemosa (T), Psychotria berteriana (U), Xanthosoma spp. (V), Clidemia hirta (W), Panicum
spp. (X), Myrcia splendens (Y) and Ocotea sintenisii (Z).
1 Introduction
13
Table 1.2 Significant spearman rank correlation coefficients among all plant species in the
Ecuador Left Banana Plantation over the first 10 years of succession. Significant positive
associations are indicated by a “+” and significant negative associations are indicated by a “−”.
Plant species and families are indicated as follows: Acalypha pladichephalus (A), Begonia spp.
(B), Geonoma undata (C), Cyathea spp. (D), Musa spp. (E), Pilea spp. (F), Anthurium spp. (G),
Trichipterix pilosissima (H), Nectandra spp. (I), Ochroma spp. (J), Baccharis spp. (K), Anthurium
spp. (L), Setaria spp. (M), Boconia frutescens (N), Piper aduncum (O), Erythrina megistophyla
(P), Vernonia patens (Q), Hedyosum spp. (R), Commelina diffusa (S), Althernantera spp. (T),
Siparuna piloso-lepidota (U), Solanum spp. (V), Vernonia spp. (W), Digitaria sanguinalis (X),
Inga spp. (Y) and Passiflora spp. (Z).
B C
A +
B
C
D
E
F
G
H
I
J
K
L
M
N
O
P
Q
R
S
T
U
V
W
X
Y
D
E
F
G
H
+
+
++
++ ++ ++ ++++ +
+
+++ + +
+++ +
+
+
+
++
+
I
J
K
L
+
+
– ++ –
++
++ –
+–
–
+– +
–
+
+
–
–
++ +
+++
+++
M N
O
P
Q
++ +
++
+++ +++
+
+++ ++ +
++
–
+++ ++
–
++ ++ ++
+
+
R
S
T
U V
W X
Y
++ +
+ +
+
+
+
+
+
+
+
Z
++
+
+
+
+ +
+
+ +
+
+
++
+
+
+
+
+
+
+
+
+
+
+++ +–
++
+
++ ++
+++ +
++
++
+
+ +
+++++
++
++ +++
+
+
one of the two Banana fields, key species with many positive associations included
Begonia spp., Trichipterix pilosissima, and Ochroma spp. (Table 1.2). However,
the past crop in that field (Musa spp.) did not play a significant role. In the other
Banana field, Begonia spp., Cuphea spp., and Brugmansia spp. were very active
(Table 1.3). In one of the two Sugarcane fields, key species included members of
the families Asteraceae, Verbenaceae, and Pupillionacia. The past crop (Saccharum officinarum), however, was not implicated (Table 1.4). In the other Sugarcane
field, Cuphea spp. and Piper aduncum were important, but the past crop was not
(Table 1.5). In the Ecuadorian pastures, negative associations were more common
than positive associations, especially when the planted pasture grass Setaria sphacelata was one of the species (Tables 1.6–1.7).
Although the results in species associations strongly suggest species interactions
(see Tirado and Pugnaire 2005 and references therein) among some tropical plant
species (as for some temperate plant species: Myster and Pickett 1992b), any further
comment would be premature, given the lack of tropical studies involving these
species (see Chapter 14). When augmented with the findings of future tropical
experiments, however, these results will allow greater synthesis of research effort
and more viable interpretations of post-agricultural succession in the Neotropics.
14
R. W. Myster
Table 1.3 Significant spearman rank correlation coefficients among all plant species in the
Ecuador Right Banana Plantation over the first 10 years of succession. Significant positive associations are indicated by a “+” and significant negative associations are indicated by a “−”. Plant
species and families are indicated as follows: Acalypha pladichephalus (A), Costus spp. (B), Musa
spp. (C), Solanum muricatum (D), Piperaceae (E), Setaria spp. (F), Tagetes terniflora (G), Begonia
spp. (H), Cuphea cartagenensis (I), Polypodiaceae (J), Vernonia patens (K), Brugmansia spp. (L),
Digitaria sanguinalis (M), Urticaceae (N), Chusquea spp. (O), Nectandra spp. (P), Piperaceae (Q),
Commelina diffusa (R), Erythrina megistophyla (S), Heliotropium spp. (T), Inga spp. (U), Musa
acuminate (V), Chenopodium album (W), Crataegus monogyna (X), Boconia frutescens (Y) and
Cecropia monostachyma (Z).
B C D E F G
H
+ –
– +
+++
+ – + ++
+
C
+
D
E
+++
A
B
F
G
H
I
J
K
L
M
N
O
P
Q
R
S
T
U
V
W
X
Y
–
I J K L
+ +
–
– +
++
++
++
+ +
– + +
+
–
++++ ++ + +
++ + +
+ + +
+–
+++
M N
+
O
P
+
Q
+
R
S
T
U V
W
–+
++
+
+
+
+
+
+
+
+
+
++
++ +
+
+ ++ +
+
++ +
+
+++ +++
+++
Z
+
+
+
+
Y
+
++ +
+
+
+
+
X
+
++ +
+
+
+
+
+
+
+
+
+
+
+
+
+
+++
+++ ++
++++ ++ +
+
+
+
+++ +
++ +
++
+
++
1.3 About this Book
Examined in this book are vital concerns of post-agricultural recovery in the
Neotropics, such as common clearing, planting, cultivation, harvesting and abandonment practices, the post-agricultural environment and disturbance regime, and
successional mechanisms. In addition, this book has advantages over other studies of tropical disturbances in that it focuses directly on post-agricultural areas and
includes real data of successional pattern and process while investigating the function and structure of old fields and presenting a conceptual framework that both
current and future Neotropical studies can use and add to.
All work that is reported in the book chapters has been done in the Neotropics,
the American landmass and associated islands that lie between the Tropic of Cancer
(23.5◦N) and the Tropic of Capricorn (23.5◦S). Between these latitudes, the sun
remains in the sky the longest over the course of a year compared to the rest of the
earth, which leads to a greatly reduced seasonality and prolonged growing period.
Responding to this geographic variation, the chapter authors share their expertise
1 Introduction
15
Table 1.4 Significant spearman rank correlation coefficients among all plant species in the
Ecuador Right Sugarcane Plantation over the first 10 years of succession. Significant positive
associations are indicated by a “+” and significant negative associations are indicated by a “−”.
Plant species and families are indicated as follows: Acalypha pladichephalus (A), Asteraceae (B),
Digitaria sanguinalis (C), Polypodiaceae (D), Nectandra spp. (E), Stachys micheliana (F), Piperaceae (G), Lantana camara (H), Verbenaceae (I), Erythrina megistophyla (J), Piper aduncum (K),
Rubus spp. (L), Commelina diffusa (M), Elephantopus mollis (N), Cecropia spp. (O), Costus spp.
(P), Miconia spp. (Q), Passifloraceae (R), Fabaceae (S), Chusquea spp. (T), Marantaceae (U), Pilea
spp. (V), Hieracium spp. (W), Sabicea spp. (X), Columnea spp. (Y) and Orchidaceae (Z).
A
B
C
D
E
F
G
H
I
J
K
L
M
N
O
P
Q
R
S
T
U
V
W
X
Y
B C D E F G
H I
J
+– + + ++
++
–
+ + + ++++ – ++
K L
+– +
M N
O
P Q R
S
T
U
+
–
+ ++
+
–+ –
+
–
++
++
+
+
+ +++
+
++ ++
+
W X
+
Y
Z
+ ++ +
++ ++ –
+++
+
+
+
V
+
+
+
+
+
+
++
+
+
+
+
+
+
+
+
++ +
++
++
+
+
++
+
+
+
–
+
++
+
+
+
+
+
+
+++ +
+
+
+
+
+
+++
+
+
+
++ +
++
+
++ +
+++ +
+
+
+
++
+
+
+
+
and relate studies over a wide range of the major gradients of temperature and rainfall found throughout the Neotropics.
The book is divided into three sections, each of which develops and expands
upon key points of post-agricultural succession in the Neotropics:
I. Patterns of living and non-living components of old fields after abandonment.
II. Mechanisms and tolerances that cause plant-plant replacements, producing old
field vegetation patterns.
III. Restoration, management, and the future of post-agricultural areas.
There are no clear-cut boundaries, however, among the three parts. The book
offers a lot of “cross-fertilization,” for most of its chapters discuss issues that formally belong to more than one section. This multilateral approach presents itself
as the best way for a comprehensive investigation of the complex phenomenon of
post-agricultural succession in the Neotropics.
We begin the book by exploring permanent plot data in areas that used to be
Neotropic Banana plantations, Sugarcane plantations, and pastures. We continue by
examining the above- and below-ground Carbon (C) dynamics after crop agriculture
and pasture and by discussing the mechanisms of C sequestration in these areas and
the methods of studying soil C dynamics. Next, we focus on species composition,
16
R. W. Myster
Table 1.5 Significant spearman rank correlation coefficients among all plant species in the
Ecuador Left Sugarcane Plantation over the first 10 years of succession. Significant positive associations are indicated by a “+” and significant negative associations are indicated by a “−”. Plant
species and families are indicated as follows: Musa spp. (A), Costus spp. (B), Cuphea cartagenensis
(C), Digitaria sanguinalis (D), Miconia spp. (E), Piper spp. (F), Rubus spp. (G), Sida rhombifolia
(H), Asteraceae (I), Baccharis spp. (J), Polypodiaceae (K), Lantana camara (L), Vernonia patens
(M), Acalypha pladichephalus (N), Solanum spp. (O), Saccharum officinarum (P), Piper aduncum
(Q), Verbenaceae (R), Commelina diffusa (S), Erythrina megistophyla (T), Nectandra spp (U),
Altus spp. (V), Orchidaceae (W), Polybotria spp. (X), Vernonia spp. (Y) and Polypodiaceae (Z).
A
B
C
D
E
F
G
H
I
J
K
L
M
N
O
P
Q
R
S
T
U
V
W
X
Y
B C D
E F
+
++ ++ ++
+++ + ++++
+ +++
+
G
+
H I J
–
+
K
L M N
–+
++ +
+++
++ + +
+
+–
+ + –
+
–
–
+
+
++ +
+
+
O
+
–
P Q
R
++ +++ +
+
S
–
+
T U V
+
+
+
+
–
+
W
+++
+
–+
–
+
+
+
+
+
++
–
+
+
+
+
+
++
+
+
Y Z
+
+
+
+
+
+
+
+
+
+
+
+ +
+
+
+
+
+
+
++++ +
+
++
++
X
+
+
+
++
+
+
++ +
+
++
+
++
++
++
+
+
+
++++
+
+
+
++++ ++
+
+
+
+
+ ++
+++
+++
Table 1.6 Significant spearman rank correlation coefficients among all plant species in the
Ecuador Left Pasture over the first 10 years of succession. Significant positive associations are
indicated by a “+” and significant negative associations are indicated by a “−”. Plant species and
families are indicated as follows: Setaria sphacelata (A), Acalypha pladichephalus (B), Xanthosoma sagittifolia (C), Heliocarpus americanus (D), Heliconia spp. (E), Ficus spp. (F), Solanceae
(G), Araceae (H), Citrus spp. (I), Vernonia patens (J), Begonia spp. (K) and Nectandra spp. (L).
A
B
C
D
E
F
G
H
I
J
K
L
B
----
C
--
D
-+
E
-
F
+++
G
H
I
J
+ ++
++
+
+
K
L
+
+
++
+
1 Introduction
17
Table 1.7 Significant spearman rank correlation coefficients among all plant species in the
Ecuador Right Pasture over the first 10 years of succession. Significant positive associations are
indicated by a “+” and significant negative associations are indicated by a “−”. Plant species and
families are indicated as follows: Setaria sphacelata (A), Erythrina megistophyla (B) and Polpodiaceae (C).
A
B
C
B
-----
C
---
soil C, and nutrients after pastures and after Sugarcane and tree plantations. The role
of alien species is introduced at this point to be developed in chapters that follow.
We finish the first section by concentrating on the living part of the soil: the fungi,
macrofauna, and earthworms. With all three kinds of organisms, the patterns of their
diversity and their abundance during succession are explored. Also examined are the
effects that these organisms may have on soils and on ecosystem functioning and the
possible distribution of these organisms among functional groups.
In the second section of the book, we discuss what is known about mechanisms
and tolerances that lead to plant-plant replacements during succession. First, we
look into the complete suite of plant mechanisms and tolerances in Neotropic wet
forest. Then we move on to seed dispersal by cattle and by birds. Raised in the
discussion are questions of how these types of dispersal affect viability and germination and how the animal vectors themselves affect and, in turn, are affected
by the vegetation structure in old fields. Next, studies that explore variation in the
working of seed predation are reviewed, and issues of facilitation and competition
are investigated in field experiments. Finally, plant mechanisms and tolerances in
Neotropic dry forest are examined.
In the final part of the book, we address the issues of restoration, management,
and future conditions of agricultural fields. Discussed are the role of microbes in
restoration and sustainability of cropping systems as well as the ways in which
agricultural practices affect microbial diversity. Also explored are different kinds
of pasture degradation, objectives in pasture restoration, and specific techniques,
such as intercropping and not clearing trees, from Costa Rican studies. We conclude the book by assessing the status quo of research in the area of Neotropical
succession after agriculture along with research perspectives and their relation to
ecological paradigms. Finally, specific avenues for future old field studies are presented.
We hope that these studies of old fields after agricultural disturbance will also
enhance an understanding of natural transition from grassland/meadow to forest
and a comprehension of fluctuations in the development of plant communities. This
knowledge will enable researchers and land managers not only to predict and manipulate the nature of land recovery but also to more skillfully utilize lands with various
kinds of vegetation cover, from lands on which woody vegetation is desired (e.g.,
forests, surface mines, road-sides) to lands on which woody vegetation should be
removed (e.g., rangelands, roadsides, utility corridors).
18
R. W. Myster
Acknowledgment I wish to thank Mr. Michael Malahy for his help in analysis of the plot data.
I also wish to thank Mr. Bert Witteveen for logistic help at Maquipucuna Reserve and Dr. Nick
Brokaw for logistic help at Luquillo Experimental Forest. Finally, I thank Ms. Eda Melendez
and her staff for their help with the LTER data sets. I received support from grant DEB-0218039
from the National Science Foundation to the Institute of Tropical Ecosystem Studies, University
of Puerto Rico, and the USDA Forest Service, International Institute of Tropical Forestry as part
of the Long-Term Ecological Research Program in the Luquillo Experimental Forest. Additional
support was provided by the Forest Service (U.S. Department of Agriculture) and the University
of Puerto Rico.
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Chapter 2
Trends in Above and Belowground Carbon
with Forest Regrowth After Agricultural
Abandonment in the Neotropics
Erika Marín-Spiotta, Daniela F. Cusack, Rebecca Ostertag
and Whendee L. Silver
2.1 Introduction
Increasing forest cover on lands which were recently forested (reforestation), as
well as on lands which have not supported forest growth in recent times (afforestation), has been proposed by the Intergovernmental Panel on Climate Change (IPCC)
to help mitigate anthropogenic C emissions from land-use change and fossil fuel
use (Brown et al. 1995a, Watson et al. 2000, Metz et al. 2001). Recent research
has suggested that these strategies would be most effective in the tropical latitudes
(Gibbard et al. 2005). Tropical forests have higher potential carbon (C) uptake rates
than forests in temperate or boreal biomes (Brown et al. 1995a, Watson et al. 2000).
Globally, 40% of terrestrial biomass C is in tropical forests (Dixon et al. 1994),
and 40% of this is in secondary forests (Brown and Lugo 1990). In addition to
providing opportunities for C sequestration, reforestation can lead to the recovery of
important forest ecosystem goods and services. These include, but are not limited to,
watershed protection, erosion control, regional climate stabilization, wood and nontimber products, and habitat for biodiversity (Brown and Lugo 1990, Guariguata
and Ostertag 2001, Naughton-Treves and Chapman 2002, De Walt et al. 2003).
Although deforestation is still a dominant trend across the tropics, secondary
forests resulting from human disturbance are becoming an increasingly important
forest cover type (Brown and Lugo 1990). Just as rates of deforestation are difficult
to establish with certainty (see Achard et al. 2002, DeFries et al. 2002, Eva et al.
2003, Fearnside and Laurance 2003), estimates for rates of secondary forest growth
also differ, although most agree that the trend is positive. In the 1980s, tropical secondary forests were estimated to cover more than 600 million hectares (ha) globally
with an annual rate of formation of 9 million hectares per year (ha/yr), and growing
(Brown and Lugo 1990). In 1993, the estimated area of tropical America covered
by secondary forests was 165 million ha (Weaver 1995 in Kammesheidt 2002). The
United Nations Food and Agriculture Organization in 1990 classified 33 million
ha of previously agricultural or pasture lands in Latin America as fallow (cited in
Kammesheidt 2002). These fallow lands are key components of forest regrowth
22
R. W. Myster, Post-Agricultural Succession in the Neotropics.
C Springer 2008
2 Agricultural Abandonment in the Neotropics
23
and succession toward secondary forest cover. In many regions across the tropics,
reforestation is in the form of agroforestry systems or plantations. Reforestation
also occurs due to natural regeneration of forest species on abandoned cropfields
or pastures, or by active planting and management which facilitate the process of
secondary succession.
The factors behind the increase in secondary forest growth in the tropics vary by
region, as do the initial drivers of deforestation, but typically include one of the following: shifts in economic bases from agriculture to industry, rural to urban migration across domestic and international borders, changes in land tenure policies, or
loss of site fertility (Rudel et al. 2002, Aide and Grau 2004, Soares-Filho et al. 2004,
Arroyo-Mora et al. 2005). Brown and Lugo (1990) identified regional differences
in the former land use type prior to secondary forest growth. In tropical America
and Africa, most secondary forests originated from agricultural fallows. In tropical
Asia, logging was a more important contributor to secondary forest formation than
agricultural abandonment. Tropical Africa had the smallest percentage of secondary
forests from logging. The authors also documented differences in the disturbance
history of the pre-conversion forest leading to secondary forests. At the time of
their review, 58% of secondary forests in tropical America were first generation
secondary forests, from previously undisturbed or primary forest. In contrast, 72–
76% of secondary forest areas in tropical Africa and Asia had experienced multiple
cycles of disturbance (Brown and Lugo 1990). However, regional trends can change
as economic policies are altered, and an updated study considering current trends in
globalization and its effect on reforestation would be a valuable undertaking.
In this chapter, we review the recent literature on the effects of reforestation on above and belowground C stocks in both natural and managed secondary
forests and plantations growing on former agricultural and pasturelands. We revisit
Silver et al.‘s (2000a) analysis of the potential for C sequestration with reforestation
in the tropics with new data published since their review. We limit our analysis to
studies from the Neotropics, where the bulk of studies are still located, though we
would like to point out that in recent years there has been an increase in publications
on biomass and structural changes during secondary forest and plantation development in the African and Asian tropics (e.g., de Jong et al. 2001, Lemenih et al. 2004,
Walker and Desanker 2004, Lawrence 2005a, Makana and Thomas 2006). We also
highlight advances made in the study of soil C dynamics with land-use change,
particularly, and provide suggestions for future research directions.
2.2 Reforestation in the Neotropics
Changes in economic policies and human migration have resulted in agricultural
abandonment and reforestation across the tropics. A prime example of the increase
in secondary forest cover during the last half-century is the Caribbean island of
Puerto Rico, where the most reliable data on secondary forest cover exists, along
with Costa Rica and the Brazilian Amazon (Kammesheidt 2002). Economic changes
in Puerto Rico, beginning in the late 1940s with the transition from being a colonial
24
E. Marín-Spiotta et al.
dependency to a free associated state (Dietz 1986), resulted in a large-scale abandonment of agricultural and pasturelands at an approximate rate of 10,000 ha/yr
from 1950 to 1990 (Grau et al. 2004). These abandoned lands generally reverted to
secondary forests through natural succession (Marrero 1950, Weaver and Birdsey
1990). As a consequence, the area of Puerto Rico under forest cover, which in the
mid-1940s was only 13%, increased to 30% by 1980 (Weaver and Birdsey 1990).
A similar trend related to forest regeneration on degraded pastures has occurred
in the moist lowlands of central and eastern Colombia (Etter et al. 2005). In the
Brazilian Amazon, secondary forest area is estimated to have increased five-fold
from 2,900,000 in 1978 to 16,100,000 ha in 2002 (Neeff et al. 2006). For example,
one municipality (Igarapé) in Pará, Brazil reached 73% secondary forest cover by
the late 1990s (Neeff et al. 2006). While often this increase in secondary forest
cover represents a net gain in total forest cover with the loss of agricultural and
other cleared lands, in many areas, such as the Brazilian Amazon, an increase in
secondary forest cover is directly related to the loss of mature or primary forest.
An increase in secondary forest cover on formerly cleared lands for agriculture or
pasture typically represents an increase in forest biomass and associated C stocks.
Regional C fluxes are often estimated as gains or losses of C stocks at the landscape scale (Houghton et al. 1991, Fearnside and Guimarães 1996). Carbon gain
or release at the landscape scale is estimated as the change in carbon content due to
changes in biomass of the vegetative cover. Aerial and satellite-based remote sensing
technologies have been used for estimating forest successional class, biomass, and
geographical extent in Costa Rica (Helmer et al. 2000) and the Brazilian Amazon
(e.g., Houghton et al. 2000, Neeff et al. 2006). Determination of the net carbon
balance of tropical forests requires certainty in biomass estimates of the different
land cover types, accurate estimates of the spatial extent at a landscape scale and on
the rates of conversion, and ways to extrapolate from plot-level studies to regional
scales (Powers et al. 2004). Secondary forest biomass, and hence their C stocks, are
affected by a suite of environmental and ecological factors which are discussed in
this chapter.
2.3 Tropical Forest Carbon Stocks and Their Measurement
The carbon budget of a tropical forest is affected by the balance of C inputs (primary
production) and C outputs (respiration, decomposition, erosion, leaching losses),
and can be expressed by a simple donor-controlled model:
∂C
= Inputs − kC
∂t
where ∂C/∂t is the rate of change in C content over time and kC represents the
output term, where C is the standing stock of carbon and k is a decay constant.
In non-steady state systems, C accumulates when the rate of inputs into a particular
pool exceeds outputs and/or the rate of outputs decreases, either by a reduction in the
2 Agricultural Abandonment in the Neotropics
25
standing stock or a decrease in turnover time (1/k). Attempts to increase C storage
in an ecosystem will often involve the transfer of C into pools with longer turnover
times, which increases the amount of time a C molecule is not in the atmosphere
where it can act as a greenhouse gas. The potential for tropical forests to remove C
from the atmosphere and store it effectively depends on net primary productivity and
the residence time of C in each reservoir. An actively growing forest functions as a
C sink as trees increase in height and diameter, and the amount of C fixed is greater
than C lost in respiration and in decomposition of litterfall. Carbon uptake by plants
can be defined as the net annual removal of C from the atmosphere per unit of land
area, which can measured by a change in biomass over time (Fearnside 1996).
A forest in C balance, or steady-state, may also store C if the residence time of
C in the largest stocks is long. The largest and longest-lived C reservoirs in tropical
forests are soil organic matter and standing live aboveground biomass (tree boles).
Tropical soils can contain more than twice as much C as in aboveground biomass
(Post et al. 1982) and have the potential to store C at great depths (Nepstad et al.
1994, Jobbágy and Jackson 2000). The residence time of C in tropical soils varies
greatly with soil depth and soil type. In the highly weathered Oxisols and Ultisols,
mean residence time of bulk soil C in the top 10 cm is on the order of decades
(Trumbore 1993, de Camargo et al. 1999, Amundson 2001), while residence times
of up to 20,000 years have been reported for mineral horizons at depth in some
volcanic soils in Hawaii (Torn et al. 1997). Substantial C stocks can be stored in
large trees (> 80 cm in diameter) which may live up to 1,000 years (Chambers et al.
1998), though 300–500 years may be more representative of an average lifetime for
large canopy and emergent trees (Fichtler et al. 2003).
A common method for studying the successional regeneration and temporal C
dynamics of secondary forest on abandoned lands is the use of chronosequences.
Using this method, forest stands of different ages are selected, using age as a proxy
for successional time. An important assumption of this method is that the patterns
observed across the chronosequence will be comparable to the patterns occurring at
one site over time. Another common method is paired-site comparisons, where secondary forest sites or plantations are compared to primary forests, typically in close
proximity. In this chapter we include data from both chronosequence and paired site
studies.
2.3.1 Aboveground Carbon Stocks
Carbon stocks in vegetation are generally calculated as 50% of dry biomass; or
45% for young secondary forests due to the presence of a larger proportion of
early succession trees with lower wood density than in mature forests (Fearnside
and Guimarães 1996, Alves et al. 1997). Plant biomass can be measured directly
by harvesting and weighing all components in a specified area, or calculated from
allometric equations using measured parameters such as diameter-at-breast-height,
wood volume, and stem height. The advantages and uncertainties associated with
different methods of measuring secondary tropical forest and plantation biomass
have been evaluated in detail elsewhere (e.g., Brown and Lugo 1984, Brown et al.
26
E. Marín-Spiotta et al.
1989, 1995b, Chambers et al. 2001, Keller et al. 2001, Losi et al. 2003, Brandeis
and Suárez Rozo 2005). Significant stores of C can also be found in root biomass
(Nepstad et al. 1994), in standing and fallen coarse woody debris (Delaney et al.
1998, Clark et al. 2002, Keller et al. 2004, Rice et al. 2004) and in lianas, particularly
in disturbed forests (Gerwing and Farias 2000, Read and Lawrence 2003a).
2.3.2 Belowground Carbon Stocks
Carbon in soils exists in organic forms: soil organic matter (SOM) and dissolved
organic carbon (DOC), or in inorganic forms: carbonates, with a minor proportion
as CO2 gas in soil pores or dissolved as carbonic acid in soil water. Most highly
weathered tropical soils are acidic, and so the major component of soil C is organic.
In the drier tropics or in young soils derived from limestone, karstic soils can contain significant amounts of inorganic C (Crowther 1980). In this chapter we discuss
changes in organic C pools, which are the focus of the majority of tropical studies
on highly weathered soils. Changes in C storage are affected by changes in the rate
of organic C inputs (e.g., above- and belowground litter, dissolved organic C) and
outputs (e.g., C mineralization, leaching losses, erosion). The residence time of C
in soils is regulated by a variety of physical, biological, and chemical mechanisms,
which affect microbial decomposition and losses of C by other means (reviewed in
Paul 1984, Oades 1988, Sollins et al. 1996, Krull et al. 2003, Mikutta et al. 2006).
Soil organic matter (SOM) is a heterogeneous mixture of plant, animal
and microbial residues in different stages of decomposition (Stevenson 1982,
Kögel-Knabner 1993). In addition to acting as one of the largest terrestrial reservoirs
of C (Post et al. 1982, Eswaran et al. 1993), SOM is an important contributor to
soil fertility as a source of plant nutrients and through its enhancing effects on soil
structure, its high water holding capacity, and buffering capacity (Tiessen et al.
1994). SOM also provides energy and nutrients for microbial metabolism (Lovley
et al. 1996, Scow 1997) and plays an important role in metal contaminant and
pesticide binding due to its high sorptive capacity (Balabane et al. 1999, Ahmad
et al. 2001, Farenhorst 2006). SOM concentrations were traditionally measured by
wet oxidation (Walkley and Black 1934). More commonly used today are the loss
of ignition method for SOM (Dean 1974), where the amount of C is calculated
as 50% of the SOM mass, and by gas chromatography, which is a much faster,
automated method (Byers et al. 1978, Nelson and Sommers 1996). The C content of
a particular soil profile is then calculated by multiplying % C by the bulk density of
the soil sampled and by the depth of the soil core and summed over the total depth
in question (Veldkamp 1994).
2.4 Trends in Carbon Stocks with Reforestation: Data Analysis
We surveyed the recent literature and updated secondary forest biomass and soil
C data presented in Silver et al.‘s (2000a) review. In that review, data of above
and belowground C accumulation from secondary forests across the tropics was
2 Agricultural Abandonment in the Neotropics
27
analyzed with respect to time by prior land use abandonment, prior land use type,
and annual rainfall. We extracted their data from studies in Neotropical sites and
added about 90 data points to their aboveground biomass and more than 100 data
points to their soil C database from new studies published since their review. The
abundance of studies on C dynamics with reforestation being published in the last 7
years testifies to the importance this topic has gained.
Data for this chapter were collected from chronosequence and paired site studies. We included forests and plantations growing on former croplands and pastures
resulting from unassisted as well as assisted, unmanaged and actively managed
regeneration. We tested for the effect of previous land use (agriculture, pasture or
cleared), present cover type (secondary forest versus plantation), and life zone. For
consistency, we followed the site classification used by Silver et al. (2000a): life
zones were based on annual precipitation, and included dry (<1,000 mm/yr), moist
(1,000–2,500 mm/yr), and wet (>2,500 mm/yr). While a lot more studies have been
published since Silver et al.‘s (2000a) review, some of the same methodological
constraints persist. For example, data are still very strongly skewed towards the
first two decades of regrowth and to the wet and moist life zones. In addition,
aboveground biomass estimates in individual studies are likely influenced by the
authors’ choice of allometric equation and plot size (for discussion of uncertainties
surrounding tropical forest biomass measurement and estimation, see Brown et al.
1995b, Alves et al. 1997, Chambers et al. 2001, Keller et al. 2001, Brandeis and
Suárez Rozo 2005).
To identify trends in aboveground C accumulation, we ran regressions on C by
life zone, past land use and cover type (planted versus secondary succession) over
time since abandonment. We use site age and time since abandonment interchangeably. To avoid the effect of uneven representation of past land uses across biomes
and vice versa in our dataset, we also tested the effect of land use within moist
forests only, which had the highest representation of all land uses (59 agricultural
sites, 21 cleared, 19 pasture), and the effect of life zone on agricultural lands only,
which had the highest representation of all life zones (17 wet forests, 59 moist, 28
dry). To fulfill assumptions of normality, we used the natural log of aboveground
C and site age. All secondary forest and plantation ages were used, but we did not
include mature or primary forests in our regressions with time as we are uncertain
of their real age.
To examine patterns in soil C with reforestation, we compiled data from studies
that reported either soil C content (on a per area basis) or both soil C concentrations
(on a per soil mass basis) and bulk density values. SOM values were converted
to C by multiplying by 0.5. Our more recent literature search increased the range
of soil depths reported (from 0.05 to 8.0 m) from the Silver et al. (2000a) review,
although most studies still sampled in the top 0.50 and 0.25 m. Following Silver
et al.‘s (2000a, 2000b) technique, we used regressions to determine soil C relationships with depth to standardize all studies to a common depth of 0.25 m. The
wet and dry life zone equations were derived from the literature dataset. Soil C in
the wet life zone decreased with depth according to the linear equation y = 1.29
+ 0.84x (r2 = 0.66, p < 0.0001, n = 98) where y= log C and x = log depth. For
the dry forests, the relationship was y = 2.34 + 0.59x (r2 = 0.86, p < 0.005, n = 7).
28
E. Marín-Spiotta et al.
The literature dataset gave a weak relationship (r2 = 0.28) for the moist forests, so
we used an equation derived from detailed soil C profiles sampled every 10 cm to
1.0 m depth in 18 moist forests in Puerto Rico, y = 2.36 + 0.52x (r2 = 0.66, p <
0.0001, n = 180) (Marín-Spiotta, unpublished data). These sites included primary
forests and secondary forests aged 10, 20, 30, 60 and 80 years and so the equation
incorporates age variability such as that found in our larger database. We used these
equations to calculate multipliers for the maximum depth or the depth closest to
0.25 m, whichever was smallest reported, for each site to adjust soil C stocks to the
standardized depth. We report soil C data for forests older than 3 years.
All statistical analyses were conducted on JMP IN version 5.1 software (SAS
Institute). We tested for relationships between soil C and forest age (years),
life zone (wet, moist, dry), and previous land use (pasture, agricultural crops,
cleared and abandoned), and present cover type (plantation, secondary forest). We
compared mean soil C stocks between secondary forests, plantations and mature
forests used as end-members in the cited studies, and between life-zone using
Wilcoxon/Kruskal-Wallis nonparametric test because C stocks were not normally
distributed and variances between sample sizes were unequal.
2.4.1 Aboveground Biomass Accumulation:
Results and Discussion
Aboveground C accumulation showed significant log-linear trends with time for all
life zones and past land uses (Table 2.1, p<0.0001). Only planted cover types did
not show a significant trend with time. Moist and wet forests had greater rates of C
accumulation than dry forests overall. During the first 20 years of succession moist
forests accumulated C fastest than wet forests, while wet forests accumulated more
total C in the long-term (Fig. 2.1). When comparing forests growing on agricultural
Table 2.1 Best fit regression equations for aboveground carbon (Mg/ha) with time following
tropical reforestation. See Appendix 1 for data
Parameter
Equation
r2
p
n
All data
Life zone
Moist forests
Wet forests
Dry forests
Past land use
Agriculture
Pasture
Cleared
Cover type
Plantations
Secondary forests
Aboveground C = −4.0 + 18.6*(ln AGE)
0.44
<0.01
190
Aboveground C = 3.0 + 17.7*(ln AGE)
Aboveground C = −15.5 + 24.0*(ln AGE)
Aboveground C = −2.9 + 5.9*(ln AGE)
0.53
0.48
0.48
<0.01
<0.01
<0.01
107
47
28
Aboveground C = −1.5 + 15.1*(ln AGE)
Aboveground C = −38.9 + 28.7*(ln AGE)
Aboveground C = −17.4 + 37.8*(ln AGE)
0.43
0.60
0.77
<0.01
<0.01
<0.01
104
41
25
n.s.
Aboveground C = −4.3 + 18.6*(ln AGE)
0.44
<0.01
10
182
Note: n.s., not significant
Aboveground C (Mg/ha)
2 Agricultural Abandonment in the Neotropics
29
100
Wet
Moist
Dry
0
0
10
20
30
40
50
60
70
80
Age (yrs)
Log curve fits are shown for aboveground C accumulation in secondary tropical
forests. + = dry forest, x = moist forest, o = wet forest.
Fig. 2.1 Aboveground carbon accumulation during secondary succession by life zone
land only to control for the effects of past land use, wet forests clearly had the highest rates of aboveground C accumulation, while dry forests had the lowest (Fig. 2.2).
Among past land uses, cleared lands showed the highest overall rates of aboveground C accumulation during 80 years of succession (Fig. 2.3). During the first 10
years of succession, agricultural land appeared to have had higher rates of aboveground C accumulation than pastures, but over longer succession pastures surpassed
agricultural lands (Fig. 2.3). This trend was maintained when analyzing only moist
forests, which had the most even representation of all past land uses (data not
140
Wet
Aboveground C (Mg/ha)
120
100
80
Moist
60
40
Dry
20
0
0
10
20
30
40
50
60
70
80
90
Age (yrs)
Log curve fits are shown for aboveground C accumulation in secondary tropical forests. +
= dry forest, x = moist forest, o = wet forest.
Fig. 2.2 Aboveground carbon accumulation by life zone for agricultural lands only
30
E. Marín-Spiotta et al.
Aboveground C (Mg/ha)
Cleared
100
Pasture
Agriculture
0
0
10
20
30
40
50
60
70
80
Age (yrs)
Log curve fits are shown for aboveground C accumulation in secondary tropical
forests. + = dry forest, x = moist forest, o = wet forest.
Fig. 2.3 Aboveground carbon accumulation during secondary succession by past land use
shown). In an ANOVA including all factors, time, life zone, past land use and cover
type were all significant (p < 0.05, r2 = 0.55).
Greater aboveground C accumulation in wet and moist forest life zones than dry
ones is consistent with the first two biomes being more productive overall (Brown
and Lugo 1982). The differentiation between wet and moist forests was much clearer
when examining only agricultural land than when all past land uses were included,
showing the importance of past land use for regeneration in these forests. This also
implies that wet and moist forests may have similar potentials for aboveground C
accumulation during secondary succession depending on the local mosaic of land
use history.
Our results comparing previous land uses were more surprising, and were in contrast to Silver et al. (2000a), who found that agricultural land had the highest rates
of C accumulation aboveground and that cleared areas had the lowest rates. This
difference could be related to a number of new data points added to the study that
had secondary succession on land that had been cleared and then allowed to grow
back within < 2 years in the Brazilian Amazon (Lucas et al. 2002), and a number
of new data points for land regenerating on agricultural land in a dry forest in the
Yucatan (Read and Lawrence 2003a). However, our test of past land use in moist
forests only showed that the trend for highest regeneration potential on recently
cleared land held, with agricultural and pasture land regenerating at slower rates.
Many of the cleared forests were allowed to grow back within a short time period
(Lucas et al. 2002), while other land uses like agriculture were maintained for over
50 years in some cases (Schroth et al. 2002). This implies that the length of land use
prior to abandonment could be as important, or more so, than the type of previous
land use for aboveground C accumulation.
2 Agricultural Abandonment in the Neotropics
31
2.4.2 Patterns in Soil Carbon with Reforestation:
Results and discussion
On average, moist forests had significantly higher soil C stocks to 0.25 m (95.0 ±
6.4 Mg/ha, mean ± 1 standard error, n = 81) than wet forests (67.2 ± 2.8 Mg/ha,
n = 100) and dry forests had intermediate values (67.7 ± 4.03 Mg/ha, n = 12). There
were no significant differences between forests growing on former agricultural lands
(95.7 ± 8.2 MgC/ha, n = 61) and those on abandoned pastures (71.9 ± 2.5 MgC/ha,
n = 83), or between forests ≥ 20 years old (86.6 ± 6.6 MgC/ha, n = 55) and those
> 20 years old (71.9 ± 3.1 MgC/ha, n = 109). Across all life-zones, there were no
significant differences in soil C between mature forests (90.8 ± 12.7 Mg/ha, n = 29),
plantations (72.1 ± 6.5 Mg/ha, n = 31) and all secondary forests grouped together
(80.6 ± 4.9 Mg/ha, n = 90).
Time since abandonment had a significant (p< 0.01) but very weak (r2 = 0.05)
effect on soil C stocks to 0.25 m (Table 2.2). The strongest relationship we found
between age and soil C was for the dry forests, although this life-zone had the
smallest number of data points. Wet forests, which comprised the larger sample
size, did not show any significant trend, similar to Silver et al.‘s (2000) findings. Past
land-use also had a significant effect on soil C, with forests on formerly cleared-only
land showing the strongest relationship (r2 = 0.41, p = 0.06) while those on former
agricultural sites showed a weaker relationship (Table 2.2 and Fig. 2.4). Sites regrowing on former pastures did not show any trend over time since abandonment.
With respect to current cover type, both secondary forests and plantations showed
significant, but very weak relationships between soil C stocks and site time.
Silver et al. (2000a) reported a slightly more robust relationship (r2 = 0.11,
p<0.05) between time and soil C for all ages with a smaller dataset (n = 57) compiled
from pantropical studies. While sites with different land use history and across the
Table 2.2 Best fit regression equations for soil carbon (Mg/ha) (in the top 0.25 m) with time
following tropical reforestation. See Appendix 2 for data
Parameter
Equation
r2
p
n
SOIL C = 42.8 + 9.3*(ln AGE)
0.05
<0.01
164
SOIL C = 34.9 + 9.3*(ln AGE)
n.s.
SOIL C = 15.0 + 17.0*(ln AGE)
0.09
0.01
0.64
<0.01
70
85
9
SOIL C = 27.1 + 17.2*(ln AGE)
n.s.
SOIL C = −23.4 + 35.6*(ln AGE)
0.12
<0.01
0.41
=0.06
57
83
9
SOIL C = 25.9 + 15.0*(ln AGE)
SOIL C = 38.1 + 10.2*(ln AGE)
0.15
0.05
<0.05
<0.05
31
90
All ages
All data
Life zone
Moist forests
Wet forests
Dry forests
Past land use
Agriculture
Pasture
Cleared
Cover type
Plantations
Secondary forests
Note: n.s., not significant
32
E. Marín-Spiotta et al.
250
Soil C (Mg/ha)
200
Cleared
150
Agriculture
100
50
0
0
10 20 30 40 50 60 70 80 90 100 110
Age (years)
Log curve fits are shown for changes in soil C in secondary tropical forests. Former
pastures showed no significant trend. + = dry forest, x = moist forest, o = wet forest.
Fig. 2.4 Soil carbon stocks during secondary succession by past land use
three climatic zones studied accumulated aboveground biomass C during secondary
succession, soil C stocks changed very little, or not at all, with time since abandonment. Soils appeared to respond very differently than aboveground biomass and
may be more resilient to changes in land use and plant cover. Formerly cleared sites
appeared to accumulate soil C faster than agricultural sites, similar to the pattern
observed in aboveground biomass. The low predictive power of time suggests that
there are other factors influencing rates and direction of soil C changes with secondary succession in the tropics. In the following sections we explore other factors
from individual studies that potentially determine the rate and magnitude of soil C
change during reforestation.
2.5 Land-Use Change and Tropical Forest Carbon
At a global scale, aboveground biomass and soil C stocks are influenced by climate (precipitation and temperature), soil and vegetation type, and ecosystem
age (Schlesinger 1977, Post et al. 1982, Eswaran et al. 1993, Torn et al. 1997,
Malhi et al. 1999, Jobbágy and Jackson 2000). Human activities, such as logging
and deforestation for agriculture or pasture establishment, mining, dam construction or urban development also affect C storage in forests (Parrotta and Knowles
1999, Caspersen et al. 2000, Houghton and Goodale 2004). The loss of aboveground C during deforestation has been documented extensively (e.g., McWilliam
et al. 1993, Fearnside 1996, Houghton et al. 2000, Hughes et al. 2000). After
the removal of a disturbance, C stocks may recover or attain a new steady-state
below or above pre-disturbance levels. In areas where climate is favorable to forest
2 Agricultural Abandonment in the Neotropics
33
growth and a seed source is available (Uhl 1987, Holl et al. 2000, Chazdon 2003),
natural regeneration of intensively used lands is possible. Many areas of tropical forest which had been previously considered to be untouched primary forest,
are instead located on sites of intense human agricultural activity and are in different stages of recovery from disturbance (Bush and Colinvaux 1994, Chazdon
2003).
Multiple environmental, ecological, and anthropogenic factors interact to determine the rate of natural or managed reforestation at any one place, making it difficult
to predict the success of reforestation at a particular site, or to estimate regional and
global C sequestration potentials of regrowing forests. At the broadest scale, climate
affects primary productivity and forest regrowth after a disturbance (Johnson et al.
2000, Anderson et al. 2006). Whether the driving forces behind the initial change
in land use and land cover are extrinsic (economic policy) or intrinsic (soil degradation) is likely to affect the regenerative potential of a site. Previous land-use history,
the type, intensity, and duration of use, can affect the successional trajectory of a
site, and its potential to recover from a disturbance (Nepstad et al. 1991, Aide et al.
2000, Silver et al. 2000). Furthermore, the same factors may have opposite effects on
rates of above and belowground C accumulation (Silver et al. 2000). The frequency
of a natural or anthropogenic disturbance, such as hurricanes or fire, may also slow
down forest regrowth (Zarin et al. 2005).
In the following sections, we discuss recent literature on the factors affecting
tree biomass, coarse woody debris, litterfall, roots, and soil C stocks during reforestation in the Neotropics. We refer the readers to Post and Kwon’s (2000) review
for the potential for C sequestration after agricultural abandonment in soils under
natural or replanted perennial vegetation (forests and grasslands) across different
global climatic zones. More recent reviews by Pregitzer and Euskirchen (2004) and
Anderson et al. (2006) explore ecosystem C and biomass C changes with forest age
globally by biome, although they do not distinguish between the type of disturbance
before succession.
2.6 Trends in Aboveground Carbon
2.6.1 Tree Biomass
2.6.1.1 Natural Regeneration
The success and rate of secondary forest regeneration depends on a multitude of
factors, including prior land use, climate, available seed source, and soil fertility.
While early secondary succession can be fast, with biomass accumulations estimated up to 25–50 t/ha on 5-year-old sites, and up to 75–150 t/ha on 15-yr-old sites
throughout the Amazon basin (Neeff 2005), subsequent recovery may be slower, and
can depend on species composition and forest type (Gehring et al. 2005). Typically,
34
E. Marín-Spiotta et al.
forest regeneration on former croplands and pastures results in net C sequestration aboveground due to greater C stocks in longer-lived trees with the accumulation of woody biomass. In Silver et al. (2000a), prior land use had a significant
effect on aboveground biomass accumulation, with C accumulation on abandoned
agricultural fields > abandoned pastures > previously cleared sites. Compared to
other land uses, grazing can lead to greater soil compaction (Spiel 1996) and slower
recovery of forest height and basal area (Aide et al. 1995, Zimmerman et al. 1995,
Stern et al. 2002, Chazdon 2003). In agricultural fields, different crops can have
“signatures,” such as root exudates, associated species, and effects of fertilizers
and herbicides, all of which can influence soil properties and affect subsequent
succession (Myster 2004). The distinction between prior land-uses is sometimes
confused because conversion of forest first to agriculture and then to pasture once
soils become nutrient depleted is common across tropical forests (Buschbacher
1986, Uhl et al. 1988, Fearnside 1993, Thomlinson et al. 1996, Carpenter et al.
2001). In some areas, like the Brazilian Amazon and in Colombia, direct conversion to cattle pasture was more common in the last 40 years (Fearnside and
Guimarães 1996, Etter et al. 2005), but today, soybean cultivation is one of the
main drivers of deforestation (Hecht 2005, Nepstad et al. 2006, Steward 2007).
Other factors affecting regeneration include disturbance history, landscape mosaic,
and physical environment (Lugo 1988, Murphy and Lugo 1995, Marcano-Vega
et al. 2002).
Land use history also has an effect on C storage through its influence on successional trajectories in secondary Neotropical forests. Although successional forests
may recover structural characteristics in as little as 20–30 years (Guariguata and
Ostertag 2001, Kennard 2002), species composition may remain distinct in old secondary forests (Zou et al. 1995, Aide et al. 2000, Lugo and Helmer 2004, MarínSpiotta et al. 2007). In a study of secondary dry tropical forests, Read and Lawrence
(2003a) related aboveground biomass to forest age, while forest structure was most
affected by water availability and anthropogenic disturbances. After a disturbance,
different regeneration pathways could affect biomass accumulation (Uhl et al. 1981,
Lucas et al. 2002). In a wet secondary forest chronosequence on abandoned pastures
in Puerto Rico, 80-year old secondary forests had greater aboveground biomass than
primary forests because of the dominance of woody species in the former, and a
heavy dominance of palms in primary forests (Marín-Spiotta et al. 2007). Shifts
in life-form and an increase in the abundance of palms in the seedling community
with secondary succession have been described in wet tropical forests in Costa Rica
(Capers et al. 2005), which may affect aboveground structure and biomass as the
forests age. A comparison of species composition scenarios for a forest on Barro
Colorado Island, Panama, by Bunker et al. (2005) found that future aboveground
C storage could vary by several hundred percent, depending on species composition. Thus, which species predominate during secondary succession may have a
large influence on the future C storage of these lands. A chronosequence study in
the Brazilian Amazon showed that a history of slash and burn agricultural had a
large effect on forest structural characteristics, affecting the full recovery of aboveground biomass (Gehring et al. 2005). Use of fire and mechanization during clearing
have been shown to slow down recovery of biomass (Uhl et al. 1981, 1988, Zarin
2 Agricultural Abandonment in the Neotropics
35
et al. 2005). The legacy of land-use history can persist through a natural disturbances
such as hurricanes (Boucher et al. 2000, Pascarella et al. 2004).
Forest growth generally occurs at different rates during early versus later stages
of succession. Early phases of succession are generally observed to be the times
of greatest C accumulation in aboveground biomass, as observed in high-altitude
tropical forests of Ecuador (Fehse et al. 2002a), and stem density has been seen
to peak at intermediate age classes in moist tropical forests (Saldarriaga et al.
1988, Mizrahi et al. 1997, Denslow and Guzmán 2000, Kennard 2002, Ruiz et al.
2005,). Silver et al. (2000a) found that overall aboveground biomass had significantly faster biomass accumulation during the first 20 years of succession
(6.17 Mg ha−1 yr−1 ) than the subsequent 60 years. Nonetheless, tropical forests
generally show relatively fast aboveground growth, up to 70% of mature forest
height and basal area can be reached in 25 years, as observed on prior agricultural lands in moist forests of the Bolivian Amazon (Peña-Claros 2003) and dry
Bolivian forests (Kennard 2002). Rapid early growth during secondary succession can lead to 50% recovery of aboveground biomass in a variety of biomes,
from moist (Gehring et al. 2005) to dry (Read and Lawrence 2003a), though Read
and Lawrence (2003a) estimated that mature-forest levels of biomass could take
over 100 years to reach in dry forests in the Yucatan because of slower growth
at later stages of succession. However, many secondary forests are much younger
than 25 years, and may never reach maturity before they are cleared again, with
great implications for long-term C sequestration. Neeff et al. (2006) estimated
that the average age of secondary forests in the Brazilian Amazon in 2002 was
5 years.
Soil fertility can be an important factor affecting forest regeneration. Tropical soils are typically old, highly weathered Ultisols and Oxisols that are poor in
mineral-derived nutrients such as phosphorous, calcium, and other base cations
(Brady and Weil 2002). Prior land uses such as repeated fire and long-term pasture
use may also lead to losses of nitrogen from soils. In a fertilization experiment on
abandoned pastures in the Brazilian Amazon, Davidson et al. (2004) found that both
grasses and early successional trees during early succession (6 years) responded
positively to nutrients, especially nitrogen. Another study in the Brazilian Amazon
(Gehring et al. 1999) also found a positive response of vegetation in secondary
succession to nutrient additions, but in this case phosphorous was the main limiting nutrient. Gehring et al. (1999) found that different woody species had varying
biomass responses to fertilization, but that grasses had the best response and gained
a competitive advantage with increased nutrients. Similarly, Uhl (1987) found that
nutrient additions favored only grasses on abandoned pastures in Venezuela. In all
cases, it is clear that soil nutrient availability can affect both the trajectory of succession and the amount of biomass accrual.
2.6.1.2 Plantations
Planting native and exotic timber species on abandoned pastures is a popular
approach for reforestation across the Neotropics. Because of C sequestered in
36
E. Marín-Spiotta et al.
plantations, the proven success of prior large-scale plantations, and a developed
professional forestry sector, Latin America has been identified as a potentially
key player in global C trading (Wright et al. 2000). Planting woody species can
greatly increase the rates of aboveground C accumulation relative to natural succession, especially in areas where herbaceous species such as grasses and ferns
can dominate unmanaged succession on abandoned pastures (Aide et al. 1995,
Cusack and Montagnini 2004, Myster 2004). Herbaceous species have much lower
aboveground C accumulation than woody species (Aide et al. 1995). Planted woody
species can improve physical conditions, shading out herbaceous species (Holl
1998, 1999), reducing soil temperatures, and increasing soil nutrient and physical
properties, increasing recruitment of other woody species in the understory and
promoting forest succession (Oliver and Larson 1996). Prior land use and extent of
pasture use can have a strong impact on subsequent C and biomass accumulation
in plantations as well, similar to issues discussed above for natural succession
(Silver et al. 2004).
There has been extensive study of how the species selected for planting affect
forest regeneration and C accumulation. Different species have different growth
patterns, making rates of aboveground C accumulation highly variable across plantation types. Aboveground C accumulation can range from as little as 0.8 to as
much as 15 Mg C ha−1 yr−1 during the first two decades of plantation establishment (Lugo et al. 1988). The species planted also have a strong effect on
later composition of the stand, with planted timber species an important component of forest species diversity 60 years after planting (Silver et al. 2004). Both
native and exotic species have been used to plant abandoned pastures. One advantage of some exotics is that they are fast growing and rapidly accumulate C. For
example, a 20-year old teak plantation (exotic species) in Panama averaged 120
Mg/ha in aboveground C (Kraenzel et al. 2003). Due to the importance of planted
species in the composition of the forest, there has been an emphasis on planting native species in monoculture and mixed stands. A study of 12 and 13-yr
old native species timber plantations in Costa Rican lowland forests found that
most species had greater C accumulation in mixed stands than in single species
stands. Across three experimental blocks with combinations of 9 native species,
mixed stands (64 ± 13 Mg C/ha) had greater C accumulation after 12+ years
than pure stands (50 ± 9 Mg C/ha) (p = 0.1) (Redondo-Brenes and Montagnini
2006). In this study, plantations with fast-growing species accumulated more C in
early stages of plantation development (<10 years), while plantations with slowergrowing species were projected to store more C in the long-term. Similarly, mixed
species plantations in Hawai’i with Eucalyptus and Albizia (an N-fixer) had greater
biomass accumulation after 17 years than monocultures of either species (Kaye
et al. 2000). Since species are planted, successional patterns and rates of aboveground biomass accumulation can differ from natural secondary regeneration. Silver
et al. (2004) found that timber trees grew fastest during the second 33 years than
the first 22 years, indicating that C accumulation is important during later stages
of succession. This contrasts with findings for natural succession, as discussed
above.
2 Agricultural Abandonment in the Neotropics
37
2.6.2 Litterfall
Despite a considerable amount of work on changes in vegetation structure during
tropical secondary forest development using chronosequences, where sites vary
in time since forest regrowth, most studies of litterfall do not use this approach
(but see Lawrence 2005b). More commonly, studies compare secondary forest of
a single age to plantations or to primary forest (e.g., Cuevas et al. 1991, Lugo
1992, Li et al. 2005), or compile data worldwide across forests of different ages
(e.g., Brown and Lugo 1990). Our understanding of litterfall nutrient inputs to
the forest floor and soil during tropical secondary succession seres is therefore
limited.
Secondary forests are generally categorized as establishing high rates of litterfall
relatively quickly, within the first 25 years of succession (Ramakrishnan and Toky
1983, Brown and Lugo 1990, Guariguata and Ostertag 2001, Lawrence 2005b), and
soon after plateau to a production rate of a primary forest (Ewel 1976). Moreover,
litterfall mass is likely to reflect development of biomass during succession (Ewel
1976), with young forests generally having high rates of litterfall in the first 20 years
after abandonment (Brown and Lugo 1990, Guariguata and Ostertag 2001). For
some recent litterfall studies, see McDonald and Healey 2000, Sánchez-de León
and Zou 2004, Silver et al. 2004, Li et al. 2005. Typical values for Neotropical
secondary forests range from 6–10 Mg ha/yr, but there is wide variation depending
on forest type, age, and climate. Avenues for further study might be developing
synthetic models to address how these environmental variables affect litter mass
and nutrients, seasonal patterns (Read and Lawrence 2003b), and more long-term
studies that directly follow a given site or chronosequences over time. Litterfall has
also been shown to influence forest tree seedling and sapling mortality in mature
forests (Clark and Clark 1991, Guariguata 1998, Drake and Pratt 2001, Gillman
et al. 2004), and its effects on secondary forest community structure would be worth
investigating.
2.6.3 Coarse Woody Debris
Very little work has been done on fine and coarse woody debris inputs, stocks,
or decomposition during tropical succession. Unfortunately, not all studies standardize the diameter ranges of coarse and fine woody debris, but the term usually
includes both standing dead and fallen debris (Harmon and Sexton 1996). During
secondary succession, coarse woody debris should be accumulating, but initially
may be quite low as decomposition rates and slow inputs may initially diminish
stocks after land clearing, and thus has been described as a U-shaped curve (Janisch
and Harmon 2002). It has been suggested that coarse woody debris does not accumulate in an ecosystem until trees are at least 10 cm diameter at breast height (Eaton
and Lawrence 2006).
38
E. Marín-Spiotta et al.
Typical stocks of woody debris over succession are unknown due to a paucity of
chronosequence studies. Eaton and Lawrence (2006) measured coarse woody debris
in forests of different age and land uses in Mexican dry forest. Woody debris stocks
in 1–16 year old secondary forests (agricultural fallows) were 5.35–8.61 Mg/ha,
with approximately 75% of the total woody debris as coarse in all sites. Using different methods, DeWalt et al. (2003) report downed coarse woody debris increased
with forest age, ranging from a volume of 6.5 m3 /ha in 20 year old forest to 35 m3 /ha
in old-growth forest. One complication in some of these studies is woody debris is
affected by that disturbance histories (e.g., Harmon et al. 1995) as well as by life
zone (Delaney et al. 1998), and these often co-vary with secondary forest age.
More data are available on stocks of woody debris in old-growth forests, although
studies vary in methodology and whether they report only fallen debris or include
standing dead trees. For mature Neotropical forests, values range from 2.43 Mg/ha
in a Venezuelan dry forest to 61.44 Mg/ha in floodplain forest (for some examples,
see Harmon et al. 1995, Delaney et al. 1998, Clark et al. 2002, 2003, Keller et al.
2004, Rice et al. 2004). In addition, the allocation patterns of woody debris are not
well studied. Estimates range from 13.5% to 33% of total aboveground biomass in
tropical forests (Harmon et al. 1995, Clark et al. 2002, Eaton and Lawrence 2006).
A significant portion of total woody debris biomass may also be in the fine woody
fraction (≤ 10 cm diameter); this was estimated to be 25% in Mexican dry forests
(Eaton and Lawrence 2006). These numbers indicate that woody debris can be a
significant C pool that is often overlooked.
2.7 Trends in Belowground Carbon
2.7.1 Roots
Root biomass during secondary succession has not received as much attention as
aboveground biomass. Like other areas, most studies of root biomass compare a
single aged forest to a mature forest, or a secondary forest to a plantation (generally
monocultures). Thus, there are almost no studies that examine root biomass along a
chronosequence (but see Jaramillo et al. 2003 for a 30 years chronosequence), and
how patterns of root biomass in mixed species plantation stands compare to those in
naturally regenerating stands is generally unknown.
It is hypothesized that early in succession more biomass should be allocated to
fine roots and leaves for resource capture, and thus, that ratios of roots to shoots
should be greater in younger secondary forests (Yan et al. 2006). Jaramillo et al.
(2003) showed that in 8–30 year old secondary forests there was very little biomass
allocated to roots below 40 cm in depth, suggesting that these forests may be
quite susceptible to disturbances. Hertel et al. (2003) found that secondary montane
forests in Costa Rica had shallower organic layers, lower root biomass, and substantially lower nutrients than mature forests. It appears that some forests such as this
montane one do not return to biomass levels after several decades of succession,
2 Agricultural Abandonment in the Neotropics
39
while other forests may (Carvalheiro and Nepstad 1996); however, as mentioned
previously our understanding of the recovery process is quite limited. For some
recent studies in secondary forests, see Hertel et al. 2003, Jaramillo et al. 2003,
Powers 2004, and Adachi et al. 2006. Unfortunately, biomass estimates are difficult
to compare among studies due to differences in the definition of fine roots, different sampling depths, and different procedures regarding the separation of live and
dead roots.
2.7.2 Soil Carbon
Changes in soil carbon stocks during succession are more difficult to observe than
changes in aboveground biomass, because the longer residence time of carbon in
the soil may outlast the duration of land use, and because of potential storage
at greater depths (beyond 1m) than those usually sampled (Nepstad et al. 1994,
Trumbore et al. 1995, Silver et al. 1996). Both net gains and no net change in soil
C have been reported with reforestation in the Neotropics (e.g. Mexico, Hughes
et al. 1999 and Ecuador, Rhoades et al. 2000). The wide variation in results may
be attributed to differences in land use type and intensity, soil properties, and time
since land-cover conversion, although the mechanisms for these observed patterns
have not been well defined. Land use change can also affect the turnover time and
distribution of different soil C fractions, even when changes in the bulk soil C pool
are undetectable (Bashkin and Binkley 1998, Binkley and Resh 1999, Marín-Spiotta
2006, see following section on advances in soil methods). In addition to serving as a
reservoir for C, soil organic matter (SOM) is a main source of plant nutrients and an
important contributor to soil fertility (Tiessen et al. 1994), thus understanding how
agricultural abandonment and reforestation affect soil C stocks is important not only
for C sequestation but also for restoration of degraded soils.
The type and intensity of prior land use may affect the direction and rate of soil
C content change during reforestation (Lugo et al. 1986). In a review of pantropical
secondary forests, Silver et al. (2000a) previously reported slower soil C accumulation rates on former croplands than on pastures, opposite to trends observed for
aboveground biomass C. Soil C accumulation in cultivated soils may be slower
because conventional agricultural practices typically leads to a severe decline in soil
C stocks (Mann 1986) relative to pastures, where belowground C inputs may equal
or rival those of forests (Fisher et al. 1994, Trumbore et al. 1995, Neill et al. 1997,
Schwendenmann and Pendall 2006). Some studies in Costa Rica (Guggenberger
and Zech 1999) and Puerto Rico (Lugo et al. 1986, Weaver et al. 1987, Silver et al.
2004) have reported increases in soil C associated with an increase in plant biomass
after pasture abandonment. Rhoades et al. (2000) reported that lower montane secondary forests in Ecuador on abandoned sugar cane and pasturelands accumulated
soil C at a rate of 1.9 Mg/ha/yr, leading to the recuperation of primary forest soil C
stocks in as little as 20 years. In contrast, a study in tropical wet Australia, found
long-lasting effects of soil physicochemical changes after pasture use that inhibited
recuperation of undisturbed forest soil C stocks (Rasiah et al. 2004). Even after 40
40
E. Marín-Spiotta et al.
years under abandoned pasture with no grazing, and 30 years of abandoned pasture
plus 10 years under replanted forest with native species, soil C stocks remained
significantly below those of undisturbed rainforest, likely due to aluminum toxicity
and low soil pH.
While aboveground biomass typically increases with secondary forest age
(Hughes et al. 1999, Read and Lawrence 2003a), soil C pools often show no
relationship with time since agricultural or pasture abandonment, as our data
analysis showed. In a seasonally dry region of Mexico, soil C stocks in the top
10 cm did not differ between primary forests, early-successional (10–15 years),
mid-successional (20–30 years), or late-successional (60 years) forests regrowing
on land formerly cultivated for maize by slash-and-burn (Saynes et al. 2005).
Microbial biomass C was greater in early- and mid-successional forests, and also
showed seasonal trends. Other studies reported no significant differences in soil C
pools between pasture, especially, and secondary or primary forests. In the Cayey
mountains of Puerto Rico, gains of new secondary forest C were compensated
by losses of pasture-derived C from the soil, resulting in no net change with
reforestation or during 80 years of secondary succession (Marín-Spiotta 2006).
Using density fractionation and radiocarbon modeling, Marín-Spiotta (2006) found
that the majority of soil C was associated with mineral surfaces, which are known
to have a stabilizing effect on soil C, and had long residence times, thus the bulk of
the C pool was resilient to land-use change. No significant change in soil C stocks
has also been reported during secondary succession after abandoned pastures in the
Brazilian Amazon (Feldpausch et al. 2004), and slash-and-burn agriculture in Los
Tuxtlas, Mexico (Hughes et al. 1999). Hughes et al. (1999) attributed their results
to the high C storage capacity of the young volcanic soils in their study sites. These
studies suggest that some tropical soils may be resilient to changes in land use and
to disturbances that dramatically affect aboveground biomass.
Soil type, in particular differences in texture and mineral content, not only have
an effect on bulk soil C content (Post et al. 1982, Eswaran et al. 1993, Neufeldt
et al. 2002), but also on its stability and response to land-use change. Using stable
C isotopes, López-Ulloa et al. (2005) found that secondary forests on Andisols in
northwestern Ecuador lost pasture-derived C and gained forest-derived-C at much
faster rates than similarly aged forests on Inceptisols, which appeared to be more
insensitive to land-use change.
Land-use history has important effects on soil C stocks, with type and intensity
potentially over-riding time since abandonment. A well-replicated study in Ecuador
found that the direction of changes in soil C stocks following pasture to forest conversion was best explained by the time a site had been under pasture use (de Koning
et al. 2003). Younger pasture soils (< 10 years) had on average 9.3 Mg/ha more
soil C than paired secondary forest and plantation sites, while pastures > 20 years
had lower soil C content than the forested sites. The difference between pasture and
reforested soil C content decreased with pasture age before conversion, pastures
20–30 years old had 18.8 Mg/ha less than forests, and pastures > 30 years had 15.8
Mg/ha less than forests.
The rate and direction of soil C change may vary at different stages of succession,
and with soil depth. In a cloud forest chronosequence growing on abandoned maize
2 Agricultural Abandonment in the Neotropics
41
fields in Oaxaca, Mexico, soils down to 40 cm had the greatest C accumulation rates
in the first 15 years of succession, which also coincided with the largest changes in
aboveground forest structure. In the next 30 years, the top 20 cm had a net loss
of soil C, followed by a smaller increase after 45 years of secondary forest cover
(Bautista-Cruz and del Castillo 2005). The lower depths (20–40 cm) showed no
patterns with forest age after the first 15 years.
Species composition and diversity can have effects on soil C, in addition to the
effects on aboveground biomass discussed previously. Reforestation of native cerrado with pine plantations in Brazil resulted in lower soil C stocks, while reforestation with Eucalyptus increased soil C (Neufeldt et al. 2002). In Hawaii, no net
soil C sequestration was documented in young (1 year, Binkley and Resh 1999) or
first-rotation (8-year, Binkley et al. 2004) Eucalyptus plantations growing on former
sugarcane fields, even after fertilization. Using the natural abundance 13 C method,
the authors measured equal rates of new plantation C3 inputs and former agricultural
C4 soil C losses. Mixed species plantations (Eucalyptus and Albizia, an N-fixer) did,
in contrast, accumulate soil C over 17 years, with soil C content positively related
to increasing percentage of Albizia stems (Kaye et al. 2000). The authors attributed
this to inhibited decomposition of residual sugarcane C in the presence of an Nfixer. Paul et al. (2002) also identified species effects on soil C stocks in plantations
established on former cultivated or pastoral lands. Their review suggested that soil
C accumulated under hardwoods and some softwoods and nitrogen fixing species,
while soil C was lost under some pine plantations.
2.8 Advances in Methods to Study Effects of Land-Use
Change on Soil Carbon
Detecting changes in bulk soil C pools with land-use change can be difficult
because of the large size of the pool. Differences in methodologies used across
studies, including a failure to correct for changes in bulk density associated with
changes in soil C concentrations (Veldkamp 1994), also contribute to the wide
range of reported results. Estimating changes in C pools by measuring changes
in inputs and outputs allows for a more precise assessment of soil C dynamics. Modeling and experimental studies on SOM decomposition and the effects
of cultivation on soil C have provided evidence that bulk soil C is a heterogeneous mixture of pools with different chemical properties and residence times in
the soil (Jenkinson and Rayner 1977, Parton et al. 1987, Trumbore 1993). As
shown in these studies, at least three distinct pools have been identified in soils
across biomes and climatic zones: an active or labile pool turning over in one
to 5 years, a slow or intermediate pool turning over on a decadal time scale,
and a passive pool with turnover time of centuries to millennia. A number of
physical and chemical separation methods have been developed to explain the
dynamics of the different soil C pools, or soil C fractions (e.g., Greenland and
Ford 1964, Elliott and Cambardella 1991, Golchin et al. 1994, Trumbore and
42
E. Marín-Spiotta et al.
Zheng 1996, Christensen 2001, Paul et al. 2006). Chemical fractionation typically separates soil C pools based on their solubility in water, strong acids, or
bases, while physical fractionation methods use differences in density, particle- or
aggregate-size.
When coupled with isotopic and other analytical and spectroscopic techniques,
fractionation approaches have been successful at describing short- and long-term
dynamics of soil C in response to land-use change and other disturbances. Enriched
isotope tracers (13 C, 14 C, 15 N) have been applied extensively to track the movement
of C or N atoms through different ecosystem components, especially leaf litterfall,
microbial biomass, and soil C pools (e.g., Swanston et al. 2002, 2005, Bird et al.
2003, Hanson et al. 2005). Differences in the natural abundance of 13 C in C3 and
C4 photosynthetic plants can be used to track soil C dynamics when a change in
vegetation occurs (Balesdent et al. 1987, Farquhar et al. 1989, Bernoux et al. 1998).
For example, in the tropics, most forage grasses in pastures are C4 plants, as are
sugarcane and maize crops, while woody plants and forest species are C3. The inputs
of different land cover types to SOM can be calculated from differences in 13 C /12 C
ratios (see Vitorello et al. 1989), which can be used as a natural tracer for describing
changes in soil C turnover with land-use change in the tropics (Martin et al. 1990,
Veldkamp 1994, Neill et al. 1997). In Puerto Rico, Marín-Spiotta (2006) used this
technique to show that even though bulk soil C stocks appeared unchanged during a
reforestation chronosequence, the SOM pool was not static.
The mean residence time of both above and belowground C stocks can also be
estimated from measurements of radiocarbon concentrations in SOM and modeling
the decay of 14 C emitted from nuclear weapons testing in the 1960s and incorporated into plant biomass during photosynthesis as 14 C−CO2 (Trumbore 1993, 2000,
Chambers et al. 1998). Using stable and radiocarbon isotope analyses of low and
high density soil C fractions, de Camargo et al. (1999) found that soil C dynamics in a secondary forest resembled those of a primary forest after 16 years, even
though aboveground C had still not recovered primary forest levels. In the same
Puerto Rican chronosequence mentioned earlier, Marín-Spiotta (2006) used 14 C to
estimate faster turnover rates of soil aggregates in pastures than in forests.
Spectroscopic techniques have greatly informed our understanding of the chemical makeup of SOM and of the processes leading to its formation and stabilization
in soils. Advances in solid-state cross-polarization 13 C- nuclear magnetic resonance
spectroscopy (NMR) spectroscopy allow for the study of SOM in situ without
chemical modification (Kögel-Knabner 2000, Preston 2001, Helfrich et al. 2006).
13 C-NMR can provide a first approximation of the relative abundance of organic
C functional groups, such as alkyl, O-alkyl, aromatic and carbonyl C, which can
be used as indicators for the presence of different compounds, such as carbohydrates, proteins, lignin, lipids, and charred materials (see Baldock et al. 2004).
This method, combined with physical fractionation, has been applied successfully
to identify chemical changes in different soil C fractions with land-use change
across a wide range of climates and soil types (Oades et al. 1988, Golchin et al.
1994, 1995, Guggenberger et al. 1995a, 1995b, Helfrich et al. 2006). Other methods
for SOM characterization are Fourier transform infrared spectroscopy (FTIR) and
pyrolysis gas chromatography/mass spectrometry (GC/MS). Nierop et al. (2001)
2 Agricultural Abandonment in the Neotropics
43
used pyGC/MS to determine the source and degree of decomposition of soil C
profiles under organic farming, conventional tilling and pasture. Coupled with isotope ratio MS, pyrolysis-GC/MS has shed light on the fate of individual compound
groups during the incorporation of plant litter into SOM (Gleixner et al. 1999).
Chromatographic techniques, such as XAD fractionation (Croué 2004), solidphase extraction on C-18 columns (Louchouarn et al. 2000), and high-pressure size
exclusion chromatography (Chin et al. 1994) are useful for separating soil extracts
and dissolved organic matter from plants and in soils by polarity and molecular
weight, properties which are related to chemical recalcitrance and decomposability.
Generally, low-molecular weight and hydrophobic compounds are thought to represent the most labile fractions of plant litter and SOM (carbohydrates, amino acids,
small proteins, organic acids), while the high-molecular weight and hydrophobic
compounds represent more recalcitrant materials. These types of compounds identified by chromatography columns have been shown to accumulate under different
vegetation types and land-use regimes (Sanger et al. 1997). These and other analytical techniques which can be used to enhance our understanding of soil C dynamics with land-use change have been reviewed recently by Kögel-Knabner (2000),
Northcott and Jones (2000) and Poirier et al. (2005).
Results from the studies mentioned above and others using these techniques suggest that bulk soil C measurements are often too coarse to detect changes over forest
successional time periods, and may lead to the erroneous conclusion that the soil C
pool has not responded to land-use change. The soil C fraction most sensitive to
land-use is typically the low density, particulate C fraction, also known as the labile
or active pool (Cambardella and Elliott 1993, Alvarez et al. 1998, Guggenberger and
Zech 1999, Compton and Boone 2000, Baisden et al. 2002, John et al. 2005, Helfrich
et al. 2006). While perhaps making up a small proportion of the total C pool, the
light fraction may cycle large amounts of C and thus dominate soil-atmosphere feedback interactions (McGroddy and Silver 2000). A number of studies have detected
changes in the distribution and chemistry of different soil C fractions, even when
the bulk soil C pool appeared unchanged in tropical, temperate and boreal soils
(Binkley and Resh 1999, Neff et al. 2002, Li et al. 2005, Marín-Spiotta 2006). The
implementation of physical and chemical fractionation, isotopic and spectroscopic
techniques, has increased our understanding of the composition and location of C
compounds in the soil organo-mineral matrix, and elucidated many of the chemical,
physical and biological stabilization mechanisms controlling root biomass and soil
C formation and turnover. The application of these methods to soil C studies in the
tropics, however, is still limited.
2.9 Future of Secondary Forests
The C sequestration potential of secondary forests depends on their future fate. As
the area of secondary forest cover increases, these will become the main source of
land available for future human use. In many regions, most secondary forests are in
a fallow stage, that is, they have only temporarily been taking out of the production
44
E. Marín-Spiotta et al.
cycle, so reconversion to non-forest land uses is highly likely (Kammesheidt 2002,
Schroth et al. 2002, Walker 2003, Soares-Filho et al. 2004). In Asia, the length
of time of swidden fallow secondary forest systems is decreasing, and these are
increasingly being converted to cash-crop and cash-tree plantations (de Jong et al.
2001). The age of secondary forests in the Colombian lowland landscape is also
increasingly becoming younger (Etter et al. 2005), implying that secondary forests
are experiencing multiple cycles of disturbance. With increasing frequency of disturbance, the potential for net loss of C as CO2 to the atmosphere increases. In regions
where the economy has shifted from agriculture to an industrial or service-based
economy, such as in Puerto Rico, conversion to agriculture or pasture is no longer
a likely threat to secondary forests. Instead, these are being increasingly cleared
for urban uses, such as residential and commercial development, or golf courses
(Thomlinson and Rivera 2000, Helmer 2004).
Secondary forests have been traditionally undervalued, particularly due to the
predominance of exotic species, and so they typically fall below the radar of conservationists and environmental planners. Although species composition of secondary
forests may not return to that of primary forests (Lugo and Helmer 2004), there is
increasingly more evidence that secondary forests can reach structural and potentially functional characteristics of primary forests in as little as 20 years (Brown and
Lugo 1990, Guariguata and Ostertag 2001). The important role of secondary forests
in C sequestration, in restoration of forest ecosystem goods and services, and as
habitat for biodiversity in fragmented landscapes (Dunn 2004, Veddeler et al. 2005)
is increasingly being recognized.
2.10 Summary and Recommendations for Future Studies
We compiled data on aboveground biomass and soil C stocks from Neotropical secondary forests and plantations growing on former agricultural or pasturelands and
examined the effect of time since abandonment on C accumulation. Our analysis
showed that the predictive power of time since abandonment on soil C stocks was
very low. We review a number of studies that did not find changes in soil C with
site age. Land-use history (type and intensity), frequency of disturbance, and the
use of fire, are potentially more important than regenerative time for determining
success, direction, and rates of C sequestration belowground. An identification of
these factors at the landscape level coupled with the creation or enforcement of
regional land use and zoning plans is important for predicting where resources for
reforestation should be focused. In many areas, forest regrowth will occur unassisted
in a relatively short timescale. When barriers exist to natural regeneration, conversion to plantations or managed restoration can facilitate the recovery of aboveground
biomass. The conservation value of secondary forests needs to be better explored at
the landscape level as well, and incorporated into environmental and agricultural
planning.
2 Agricultural Abandonment in the Neotropics
45
Studies that combine isotopic and molecular techniques with more traditional
ecological methods increase our ability to detect changes in soil C stocks and better
understand soil C dynamics with land-use change. The application of these techniques to studies in the tropics is still limited. Our understanding of the factors controlling root biomass and soil C with reforestation is a lot less than those on aboveground dynamics. There are many studies across the tropics on different aspects
of regeneration, species diversity, seedling survival, structural characteristics (stem
density, basal area), but a lot fewer examine soil C. We encourage interdisciplinary
collaborations between plant and community ecologists and biologists with soil scientists and chemists to provide a more comprehensive picture of forest dynamics
with reforestation. Soil studies should report carbon content on a per area basis or
carbon concentrations with bulk density values for the proper conversion to carbon content. Unfortunately, too many studies report %C (or %N) alone per depth,
when potential changes in bulk density, especially during land-use change, hinder
comparisons of concentrations across treatments.
Unlike for soil C, it is clear that the potential for C accumulation aboveground
is high during secondary succession in Neotropical forests, and C accumulation
follows a log-linear pattern, with rapid early accumulation. This C accumulation
reaches an eventual plateau, which appears to occur within 50 years in dry forests,
but can take over 100 years on wet and moist forests. Given similar land use histories, wet tropical forests clearly have the greatest potential for aboveground C
accumulation. Our data also suggest that lands that are cleared and rapidly returned
to successional processes have the potential to accumulate more C over time than
lands that have been in sustained agriculture or pasture. A previous study found
that agricultural lands had slightly greater potential for forest regeneration (Silver
et al. 2000a), and the authors suggested that these lands were selected for agriculture
because they were the richest in soil nutrients. This effect appears to be dependent
on the length of time land is held in agriculture, so it may be useful to focus reforestation efforts on areas with minimal time since conversion.
Most studies on aboveground dynamics with reforestation also report basal area,
and few report biomass stocks. We encourage authors to attempt to provide biomass
values using published allometric equations. While using allometric equations certainly has its caveats and sources of uncertainty, these have been well identified and
there is a very large number of biomass equations available for different forest types
and life zones, especially for Neotropical sites. The dry tropics are a particularly
understudied life zone. We also found almost twice as many studies conducted in the
first 20 years after abandonment, but our data shows that older forests may behave
differently, so it is important to include sites beyond the first two decades whenever
available. As the age of secondary forests decreases, these older secondary forests
are becoming a threatened habitat, like primary forests.
Life zone
M
W
M
M
M
M
M
M
M
M
M
M
M
W
M
M
W
W
M
M
M
Age (y)
0.3
0.5
0.7
0.8
0.8
0.8
0.8
0.8
0.8
0.8
0.8
0.8
0.8
0.8
1
1
1
1
1.3
1.4
1.9
A
A
A
A
A
A
A
A
A
A
A
A
A
A
A
A
C
A
A
A
A
Past Land use
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
Cover
Type
Mixed vs.
Single spp
Native vs.
exotic
2
1
2
2
5
6
50
5
4
2
7
7
7
2
4
3
3
4
4
7
6
Aboveground
Biomass C
(Mg/ha)
Peru
Mexico
Peru
Guatemala
Guatemala
Guatemala
Guatemala
Guatemala
Guatemala
Guatemala
Guatemala
Guatemala
Guatemala
Mexico
Guatemala
Brazil
Puerto Rico
Brazil
Brazil
Peru
Brazil
Location
Szott et al. 1994
Hughes et al. 1999
Szott et al. 1994
Tergas and Popenoe 1971
Tergas and Popenoe 1971
Tergas and Popenoe 1971
Tergas and Popenoe 1971
Tergas and Popenoe 1971
Tergas and Popenoe 1971
Tergas and Popenoe 1971
Tergas and Popenoe 1971
Tergas and Popenoe 1971
Tergas and Popenoe 1971
Williams-Linera 1983
Snedaker 1970
Uhl et al. 1988
Silver et al. 1996
Uhl 1987
Gehring et al. 1999
Szott et al. 1994
Gehring et al. 1999
Reference
Appendix 2.1 Aboveground C pools in secondary forests and plantations of known age following abandonment of agricultural and pasturelands and in nearby
mature forests in the Neotropics. Life zones are wet (W), moist (M), and dry (D) based on mean annual precipitation descriptions given by authors. Past land uses
are agriculture (A), pasture (P). Cover types are secondary forests (S) or plantations (P). Plantations are identified as mixed (M) or single species (S), and native
(N) or exotic (E). Aboveground C was calculated as 50% of dry biomass
Appendices
46
E. Marín-Spiotta et al.
2
2
2
2
2
2
2
2
2.3
2.4
2.5
2.5
2.8
3
3
3
3
3
3
3
3
3
3
3.5
3.5
3.5
3.6
4
4
4
4
4
D
M
M
M
M
M
M
W
M
M
M
M
M
D
D
M
M
M
M
W
W
W
W
M
M
M
M
D
D
M
M
M
A
A
A
S
C
A
A
A
A
A
P
P
C
A
A
A
A
C
A
A
C
C
C
P
P
P
A
A
A
A
A
P/A
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
6
4
6
9
30
36
7
6
6
12
8
4
17
3
5
19
17
50
11
10
4
0
6
8
15
4
20
4
3
19
21
24
Mexico
Brazil
Costa Rica
Colombia
Brazil
Brazil
Guatemala
Brazil
Brazil
Peru
Brazil
Brazil
French Guyana
Mexico
Mexico
Brazil
Brazil
Brazil
Guatemala
Brazil
Brazil
Brazil
Brazil
Brazil
Brazil
Brazil
Peru
Mexico
Mexico
Costa Rica
Mexico
Brazil
(continued)
Read and Lawrence 2003
Alves et al. 1997
Ewel 1971
Folster et al 1976
Gehring et al. 2005
Lucas et al. 2002
Snedaker 1970
Uhl 1987
Gehring et al. 1999
Szott et al. 1994
Uhl et al. 1988
Uhl et al. 1988
Maury-Lechon 1982
Read and Lawrence 2003
Read and Lawrence 2003
Alves et al. 1997
Alves et al. 1997
Gehring et al. 2005
Snedaker 1970
Uhl 1987
Uhl et al. 1982
Uhl et al. 1982
Uhl et al. 1982
Uhl et al. 1988
Uhl et al. 1988
Uhl et al. 1988
Szott et al. 1994
Read and Lawrence 2003
Read and Lawrence 2003
Ewel 1971
Hughes et al. 1999
Lucas et al. 2002
2 Agricultural Abandonment in the Neotropics
47
Life zone
M
M
W
W
M
M
D
D
D
D
D
M
M
M
M
W
W
D
M
M
W
W
D
M
M
M
D
M
Age (y)
4
4
4
4
4.4
4.5
5
5
5
5
5
5
5
5
5
5
5.5
6
6
6
5
5.5
6
6
6
6
7
7
Appendix 2.1 (continued)
A
P
A
A
A
P
A
P
A
A
C
A
A
A
A
P
A
A
A
A
A
A
A
P
Past Land use
S
S
P
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
Cover
Type
S
Mixed vs.
Single spp
E
Native vs.
exotic
14
9
20
14
27
27
4
9
16
8
6
19
34
65
18
17
16
5
3
21
17
16
5
3
21
22
3
4
Aboveground
Biomass C
(Mg/ha)
Guatemala
Brazil
Puerto Rico
Brazil
Peru
Brazil
Mexico
Mexico
Mexico
Mexico
Mexico
Brazil
Colombia
Brazil
Guatemala
Brazil
Puerto Rico
Mexico
Brazil
Costa Rica
Brazil
Puerto Rico
Mexico
Brazil
Costa Rica
Guatemala
Mexico
Brazil
Location
Snedaker 1970
Uhl et al. 1988
Lugo 1992
Uhl 1987
Szott et al. 1994
Uhl et al. 1988
Read and Lawrence 2003
Read and Lawrence 2003
Read and Lawrence 2003
Read and Lawrence 2003
Read and Lawrence 2003
Alves et al. 1997
Folster et al 1976
Gehring et al. 2005
Snedaker 1970
Uhl 1987
Lugo 1992
Read and Lawrence 2003
Davidson et al. 2004
Ewel 1971
Uhl 1987
Lugo 1992
Read and Lawrence 2003
Davidson et al. 2004
Ewel 1971
Snedaker 1970
Read and Lawrence 2003
Davidson et al. 2004
Reference
48
E. Marín-Spiotta et al.
7
7
7
7
7
7.5
8
8
8
8
8
8
8
8
8
8
8
8
8
9
9
9
9
9
9
9
9
9
9
9
9
M
M
M
W
W
M
D
D
D
D
M
M
M
M
M
M
M
W
W
D
M
M
M
M
M
M
M
M
M
M
M
C
C
A
P
A
P
A
A
A
A
P
P/A
A
P
P
P
P
A
P
A
A
C
C
C
C
C
C
C
C
A
A
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
75
35
23
3
22
19
7
5
6
16
4.8
52
33
44
43
16
2
11
49
5
45
67
65
57
59
57
62
56
62
22
36
Brazil
Brazil
Guatemala
Puerto Rico
Mexico
Brazil
Mexico
Mexico
Mexico
Mexico
Brazil
Brazil
Guatemala
Brazil
Brazil
Brazil
Brazil
Mexico
Mexico
Mexico
Brazil
Brazil
Brazil
Brazil
Brazil
Brazil
Brazil
Brazil
Brazil
Colombia/Venezuela
Guatemala
(continued)
Gehring et al. 2005
Lucas et al. 2002
Snedaker 1970
Aide et al. 1995
Williams-Linera 1983
Uhl et al. 1988
Read and Lawrence 2003
Read and Lawrence 2003
Read and Lawrence 2003
Read and Lawrence 2003
Davidson et al. 2004
Lucas et al. 2002
Snedaker 1970
Uhl et al. 1988
Uhl et al. 1988
Uhl et al. 1988
Uhl et al. 1988
Hughes et al. 1999
Hughes et al. 1999
Read and Lawrence 2003
Alves et al. 1997
Lucas et al. 2002
Lucas et al. 2002
Lucas et al. 2002
Lucas et al. 2002
Lucas et al. 2002
Lucas et al. 2002
Lucas et al. 2002
Lucas et al. 2002
Saldarriaga et al. 1988
Snedaker 1970
2 Agricultural Abandonment in the Neotropics
49
Life zone
D
M
M
W
W
W
M
M
W
D
D
D
M
M
W
W
W
M
W
D
M
D
M
M
M
W
Age (y)
10
10
10
10
10
10
11
11
11
12
12
12
12
12
12
13
13
14
14
15
15
16
16
16
16
16
C
A
A
C
A
P
P
P
P
A
A
A
A
A
C
A
P
P
P
A
A
A
P
A
Past Land use
Appendix 2.1 (continued)
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
P
P
S
S
S
S
S
S
S
S
S
Cover
Type
M
S
Mixed vs.
Single spp
N
N
Native vs.
exotic
10
80
27
7
2
1
32
26
18
11
14
15
55
41
3
32
25
27
40
9
23
13
52
101
58
136
Aboveground
Biomass C
(Mg/ha)
Mexico
Brazil
Guatemala
Puerto Rico
Puerto Rico
Puerto Rico
Brazil
Colombia/Venezuela
Puerto Rico
Brazil
Brazil
Brazil
Brazil
Colombia/Venezuela
Puerto Rico
Costa Rica
Costa Rica
Colombia/Venezuela
Brazil
Mexico
Ecuador
Brazil
Brazil
Colombia
Brazil
Mexico
Location
Read and Lawrence 2003
Gehring et al. 2005
Snedaker 1970
Marin-Spiotta et al. in press
Marin-Spiotta et al. in press
Marin-Spiotta et al. in press
Alves et al. 1997
Saldarriaga et al. 1988
Cuevas et al. 1991
Read and Lawrence 2003
Read and Lawrence 2003
Read and Lawrence 2003
Lucas et al. 2002
Saldarriaga et al. 1988
Aide et al. 1995
Redondo-Brenes and Montagnini, 2006
Redondo-Brenes and Montagnini, 2007
Saldarriaga et al. 1988
Schroth et al. 2002
Read and Lawrence 2003
Fehse et al 2002
Read and Lawrence 2003
Alves et al. 1997
Folster et al 1976
Lucas et al. 2002
Hughes et al. 1999
Reference
50
E. Marín-Spiotta et al.
17
17
17
17
18
18
18
18.5
20
20
20
20
20
20
20
20
20
20
20
21
22
22.5
22.5
24
25
25
25
25
26
30
30
30
M
W
W
W
D
D
M
W
M
M
M
M
M
W
W
W
W
W
W
M
W
W
W
D
D
D
M
W
W
M
M
W
A
A
A
C
P
A
C
A
A
A
A
P
P
P
C
P
C
A
A
A
A
A
A
P
A
A
S
P
P
P
S
S
S
P
S
S
S
S
S
P
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
M
S
E
E
E
E
E
22
50
85
51
13
7
72
85
100
31
49
32
42
60
38
127
75
84
43
109
15
36
55
24
23
15
115
44
100
146
27
121
Brazil
Hawai’I
Hawai’I
Puerto Rico
Mexico
Mexico
Brazil
Puerto Rico
Brazil
Colombia/Venezuela
Colombia/Venezuela
Colombia/Venezuela
Colombia/Venezuela
Panama
Mexico
Mexico
Puerto Rico
Puerto Rico
Puerto Rico
Brazil- AMZ
Puerto Rico
Puerto Rico
Puerto Rico
Mexico
Mexico
Mexico
Brazil
Puerto Rico
Mexico
Brazil
Colombia/Venezuela
Mexico
(continued)
de Koning et al. 1999
Kaye et al. 2000
Kaye et al. 2000
Lugo 1992
Read and Lawrence 2003
Read and Lawrence 2003
Alves et al. 1997
Lugo 1992
Gehring et al. 2005
Saldarriaga et al. 1988
Saldarriaga et al. 1988
Saldarriaga et al. 1988
Saldarriaga et al. 1988
Kraenzel et al. 2003
Hughes et al. 1999
Hughes et al. 1999
Marin-Spiotta et al. in press
Marin-Spiotta et al. in press
Marin-Spiotta et al. in press
Lucas et al. 2002
Aide et al. 1995
Lugo 1992
Lugo 1992
Read and Lawrence 2003
Read and Lawrence 2003
Read and Lawrence 2003
Gehring et al. 2005
Aide et al. 1995
Hughes et al. 1999
Lucas et al. 2002
Saldarriaga et al. 1988
Hughes et al. 1999
2 Agricultural Abandonment in the Neotropics
51
Life zone
W
W
W
M
M
M
M
W
W
W
M
M
M
M
W
W
W
W
W
W
Age (y)
30
30
30
35
35
40
44
49
50
50
54
60
60
60
60
60
60
60
60
60
Appendix 2.1 (continued)
P
A
A
A
P
P
P
P
P
P
A
P
P
P
A
A
A
Past Land use
S
S
S
S
S
S
S
P
S
S
S
S
S
S
P
S
S
S
S
S
Cover
Type
S
Mixed vs.
Single spp
both
E
Native vs.
exotic
39
48
45
55
54
80
83
62
137
40
79
58
99
69
40
33
54
87
77
83
Aboveground
Biomass C
(Mg/ha)
Puerto Rico
Puerto Rico
Puerto Rico
Colombia/Venezuela
Colombia/Venezuela
Colombia/Venezuela
Puerto Rico
Puerto Rico
Mexico
Puerto Rico
Puerto Rico
Colombia/Venezuela
Colombia/Venezuela
Colombia/Venezuela
Puerto Rico
Puerto Rico
Puerto Rico
Puerto Rico
Puerto Rico
Puerto Rico
Location
Marin-Spiotta et al. in press
Marin-Spiotta et al. in press
Marin-Spiotta et al. in press
Saldarriaga et al. 1988
Saldarriaga et al. 1988
Saldarriaga et al. 1988
Jordan and Farnsworth 1982
Lugo 1992
Hughes et al. 1999
Lugo 1992
Silver et al. 2004
Saldarriaga et al. 1988
Saldarriaga et al. 1988
Saldarriaga et al. 1988
Silver et al. 2004
Aide et al. 1995
Aide et al. 1995
Marin-Spiotta et al. in press
Marin-Spiotta et al. in press
Marin-Spiotta et al. in press
Reference
52
E. Marín-Spiotta et al.
80
80
80
80
80
80
80
Mature
Mature
Mature
Mature
Mature
Mature
Mature
Mature
Mature
Mature
Mature
Mature
Mature
M
M
M
M
W
W
W
D
D
D
D
D
D
D
D
M
M
W
W
W
A
A
A
A
P
P
P
A
A
A
A
A
A
A
A
P
C
P
P
P
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
67
89
72
71
93
152
156
30
35
31
41
33
39
34
30
132
229
84
92
66
Colombia/Venezuela
Colombia/Venezuela
Colombia/Venezuela
Colombia/Venezuela
Puerto Rico
Puerto Rico
Puerto Rico
Mexico
Mexico
Mexico
Mexico
Mexico
Mexico
Mexico
Mexico
Brazil
Brazil
Puerto Rico
Puerto Rico
Puerto Rico
Saldarriaga et al. 1988
Saldarriaga et al. 1988
Saldarriaga et al. 1988
Saldarriaga et al. 1988
Marin-Spiotta et al. in press
Marin-Spiotta et al. in press
Marin-Spiotta et al. in press
Read and Lawrence 2003
Read and Lawrence 2003
Read and Lawrence 2003
Read and Lawrence 2003
Read and Lawrence 2003
Read and Lawrence 2003
Read and Lawrence 2003
Read and Lawrence 2003
de Koning et al. 1999
Gehring et al. 2005
Marin-Spiotta et al. in press
Marin-Spiotta et al. in press
Marin-Spiotta et al. in press
2 Agricultural Abandonment in the Neotropics
53
Depth
(cm)
30
100
40
45
10
30
100
10
25
45
20
25
30
25
25
50
50
50
25
30
30
45
25
Age (y)
4
4
5
5
5
5
5.5
6
7
7
7
7
7
8
8
8
8
8
8
8
8
8
9
W
W
M
M
W
W
W
M
M
M
W
W
W
M
M
M
M
M
W
W
W
W
M
Life Zone
P
P
P
C
P
P
P
P
P
P
P
P
A
P
A
P
A
P
P
P/A
A
Past Land
Use
S
P
S
S
S
P
S
S
S/P
S
S
S/P
P
S/P
S/P
S
S
S
S/P
S
S
P
S/P
Cover type
116
44
45
65
14
77
39
51
62
68
75
52
85
55
52
65
68
70
60
71
120
135
59
Soil C
(Mg/ha)
100
14
47
64
30
66
12
110
62
67
90
52
73
55
52
61
64
66
60
61
103
82
59
Soil C 0–25 cm
(Mg/ha)
Mexico
Puerto Rico
Brazil
Brazil
Costa Rica
Panama
Puerto Rico
Dominican Republic
Ecuador
Brazil
Mexico
Ecuador
Costa Rica
Ecuador
Ecuador
Brazil
Brazil
Brazil
Ecuador
Mexico
Mexico
Hawaii
Ecuador
Location
Hughes et al. 1999
Lugo 1992
Sommer et al. 2000
Feldpausch et al. 2004
Reiners et al. 1994
Potvin et al. 2004
Lugo 1992
Templer et al. 2005
de Koning et al. 2003
Feldpausch et al. 2004
Williams-Linera 1983
de Koning et al. 2003
Powers 2004
de Koning et al. 2003
de Koning et al. 2003
Buschbacher et al. 1988
Buschbacher et al. 1988
Buschbacher et al. 1988
de Koning et al. 2003
Hughes et al. 1999
Hughes et al. 1999
Binkley et al. 2004
de Koning et al. 2003
Reference
Appendix 2.2 Soil C pools in secondary forests and plantations of known age following abandonment of agricultural and pasturelands and in nearby mature
forests in the Neotropics. Depth refers to maximum soil depth reported for the measurement of soil C content. Soil C 0-25 cm are soil C values adjusted to a
common depth using a regression approach (see text). Life zones are wet (W), moist (M), and dry (D) based on mean annual precipitation descriptions given by
authors. Past land uses are agriculture (A), pasture (P). Cover types are secondary forests (S) or plantations (P)
54
E. Marín-Spiotta et al.
9
9
9
9
9
9
9
9
10
10
10
10
10
10
10
10
10
10
10
10
10
11
11
11
11
11
11
11
12
12
12
10
10
15
15
15
15
15
15
25
25
25
25
30
30
30
30
20
25
25
25
25
25
10
25
25
25
30
30
40
10
10
W
W
W
W
W
W
W
W
M
M
M
M
M
M
M
M
W
W
W
W
W
D
M
M
W
W
W
W
M
W
W
P
P
A
A
A
A
A
A
P
P
P
P
P
P
P
A
A
A
P
P
A
C
P
P
P
P
P
C
A
P
C
S
S
P
P
P
P
P
P
S/P
S/P
S/P
S/P
S
S
S
S
S
P
S/P
S/P
P
S
S
S/P
S/P
S/P
P
S
S
S
S
13
19
32
35
34
37
38
41
69
55
76
80
103
67
84
57
32
70
73
99
60
50
23
46
71
69
41
42
44
21
22
28
41
49
54
52
57
58
63
69
55
76
80
126
82
103
70
39
70
73
99
60
50
50
46
71
69
35
36
46
45
47
Costa Rica
Costa Rica
Costa Rica
Costa Rica
Costa Rica
Costa Rica
Costa Rica
Costa Rica
Ecuador
Ecuador
Ecuador
Ecuador
Puerto Rico
Puerto Rico
Puerto Rico
Honduras
Brazil
Puerto Rico and US Virgin Islands
Ecuador
Ecuador
Puerto Rico and US Virgin Islands
Puerto Rico and US Virgin Islands
Brazil
Ecuador
Ecuador
Ecuador
Puerto Rico
Puerto Rico
Brazil
Costa Rica
Costa Rica
(continued)
Reiners et al. 1994
Reiners et al. 1994
Russell et al. 2004
Russell et al. 2004
Russell et al. 2004
Russell et al. 2004
Russell et al. 2004
Russell et al. 2004
de Koning et al. 2003
de Koning et al. 2003
de Koning et al. 2003
de Koning et al. 2003
Marin-Spiotta 2006
Marin-Spiotta 2006
Marin-Spiotta 2006
Paniagua et al. 1999
Johnson et al. 2001
Brown and Lugo 1990b
de Koning et al. 2003
de Koning et al. 2003
Brown and Lugo 1990b
Brown and Lugo 1990b
Salimon et al. 2004
de Koning et al. 2003
de Koning et al. 2003
de Koning et al. 2003
Cuevas et al. 1991
Cuevas et al. 1991
Sommer et al. 2000
Reiners et al. 1994
Reiners et al. 1994
2 Agricultural Abandonment in the Neotropics
55
12
12.5
13
13
14
15
15
15
15
15
15
15
15
15
15
15
15
15
16
16
16
16
16
16.5
17
17
17
17
25
10
45
25
25
20
20
20
25
25
25
25
25
5
10
25
25
25
25
25
30
30
30
30
10
100
25
25
W
D
M
W
M
M
M
M
M
M
M
M
M
W
W
W
W
W
M
W
W
W
W
W
M
M
W
W
Appendix 2.2 (continued)
Depth
Age (y)
(cm)
Life Zone
P
P
A
A
P
P
P
A
A
A
P
P
P
P
P
P
C
P
P
P
P
P
P
A
P
P
P
Past Land
Use
P
S
S
S/P
S/P
S
S
S
S/P
S/P
S/P
S/P
S/P
S
S
S/P
S/P
S/P
S/P
S/P
S
S
S
S
S
P
S/P
S/P
Cover type
72
39
74
74
78
98
162
136
63
100
66
64
116
33
21
64
108
95
40
104
77
92
44
79
25
93
93
75
Soil C
(Mg/ha)
72
67
73
74
78
148
245
205
63
100
66
64
116
128
45
64
108
95
40
104
66
79
38
68
54
61
93
75
Soil C 0–25 cm
(Mg/ha)
Hawaii
Mexico
Brazil
Ecuador
Ecuador
Mexico
Mexico
Mexico
Ecuador
Ecuador
Ecuador
Ecuador
Ecuador
Ecuador
Costa Rica
Ecuador
Ecuador
Ecuador
Ecuador
Ecuador
Costa Rica
Mexico
Costa Rica
Costa Rica
Brazil
Puerto Rico
Ecuador
Ecuador
Location
Bashkin and Binkley 1998
Saynes et al. 2005
Feldpausch et al. 2004
de Koning et al. 2003
de Koning et al. 2003
Bautista-Cruz and del Castillo 2005
Bautista-Cruz and del Castillo 2005
Bautista-Cruz and del Castillo 2005
de Koning et al. 2003
de Koning et al. 2003
de Koning et al. 2003
de Koning et al. 2003
de Koning et al. 2003
Fehse et al. 2002
Reiners et al. 1994
de Koning et al. 2003
de Koning et al. 2003
de Koning et al. 2003
de Koning et al. 2003
de Koning et al. 2003
Guariguata et al. 1997
Hughes et al. 1999
Werner 1984
Guariguata et al. 1997
de Camargo et al. 1999
Lugo 1992
de Koning et al. 2003
de Koning et al. 2003
Reference
56
E. Marín-Spiotta et al.
17
17
17
17
17
17
17
17.5
17.5
18
18.5
19
20
20
20
20
20
20
20
20
20
20
20
21
22.5
22.5
23
23
23
25
30
50
50
50
50
50
50
30
30
25
100
25
25
30
30
30
20
25
25
25
30
30
70
25
100
100
20
25
25
10
W
W
W
W
W
W
W
W
W
D
W
W
M
M
M
M
W
W
W
W
W
W
W
W
M
W
M
W
W
D
C
A
P
A
P
P
P
P
P/A
P
A
P
A
A
A
A
P
A
A
A
A
A
A
P/A
P/A
P
S
P
P
P
P
P
P
S
S
S/P
P
S
S/P
S
S
S
S
S/P
S
P
S
S
P
S/P
S
S
P
S
S
S
114
127
127
135
145
148
128
129
101
59
89
49
79
103
45
67
25
98
57
56
103
87
110
83
61
86
117
65
47
39
98
71
71
75
81
83
72
111
87
59
28
49
79
126
55
82
30
98
57
56
88
75
46
83
40
27
177
65
47
67
Costa Rica
Hawaii
Hawaii
Hawaii
Hawaii
Hawaii
Hawaii
Ecuador
Ecuador
Ecuador
Puerto Rico
Puerto Rico
Ecuador
Puerto Rico
Puerto Rico
Puerto Rico
Brazil
Ecuador
Puerto Rico
Puerto Rico
Mexico
Mexico
Panama
Ecuador
Puerto Rico
Puerto Rico
Brazil
Puerto Rico and US Virgin Islands
Puerto Rico
Mexico
(continued)
Guariguata et al. 1997
Kaye et al. 2000
Kaye et al. 2000
Kaye et al. 2000
Kaye et al. 2000
Kaye et al. 2000
Kaye et al. 2000
Rhoades et al. 2000
Rhoades et al. 2000
de Koning et al. 2003
Lugo 1992
Lugo et al. 1986
de Koning et al. 2003
Marin-Spiotta 2006
Marin-Spiotta 2006
Marin-Spiotta 2006
Johnson et al. 2001
de Koning et al. 2003
Li et al. 2005
Li et al. 2005
Hughes et al. 1999
Hughes et al. 1999
Kraenzel et al. 2003
de Koning et al. 2003
Lugo 1992
Lugo 1992
Smith et al. 1998
Brown and Lugo 1990b
Lugo et al. 1986
Saynes et al. 2005
2 Agricultural Abandonment in the Neotropics
57
25
25
25
25
25
26
26
30
30
30
30
30
30
30
31
35
35
36
36
36
40
40
40
40
40
42.5
45
45
45
49
10
25
25
25
25
25
30
25
25
30
30
30
25
30
30
25
25
20
20
20
25
40
20
25
25
25
20
20
20
100
W
W
W
W
W
M
W
M
M
M
M
M
W
W
W
D
D
M
M
M
M
M
W
W
W
W
M
M
M
W
Appendix 2.2 (continued)
Depth
Age (y) (cm) Life Zone
A
P
P
P
P
A
A
P
P
P
P
P
P
A
A
A
P
C
C
C
P
A
A
A
P
A
A
A
A
Past Land
Use
S
S/P
S/P
S/P
S/P
S
S
S/P
S
S
S
S
S/P
S
S
S
S
P
P
P
S
S
S
S
S
S
S
S
S
P
Cover type
90
57
54
58
65
40
82
57
23
79
56
42
58
117
43
75
77
92
93
95
70
50
22
75
60
83
130
90
135
45
Soil C
(Mg/ha)
194
57
54
58
65
40
70
57
23
97
68
51
58
100
37
75
77
139
140
143
70
53
27
75
60
83
196
136
204
14
Soil C 0–25 cm
(Mg/ha)
Jamaica
Ecuador
Ecuador
Ecuador
Ecuador
Puerto Rico and US Virgin Islands
Mexico
Ecuador
Puerto Rico
Puerto Rico
Puerto Rico
Puerto Rico
Ecuador
Mexico
Costa Rica
Puerto Rico and US Virgin Islands
Puerto Rico
Brazil
Brazil
Brazil
Puerto Rico and US Virgin Islands
Brazil
Brazil
Puerto Rico and US Virgin Islands
Puerto Rico
Puerto Rico and US Virgin Islands
Mexico
Mexico
Mexico
Puerto Rico
Location
McDonald and Healey 2000
de Koning et al. 2003
de Koning et al. 2003
de Koning et al. 2003
de Koning et al. 2003
Brown and Lugo 1990b
Hughes et al. 1999
de Koning et al. 2003
Lugo et al. 1986
Marin-Spiotta 2006
Marin-Spiotta 2006
Marin-Spiotta 2006
de Koning et al. 2003
Hughes et al. 1999
Werner 1984
Brown and Lugo 1990b
Lugo et al. 1986
Smith et al. 1998
Smith et al. 1998
Smith et al. 1998
Brown and Lugo 1990b
Sommer et al. 2000
Johnson et al. 2001
Brown and Lugo 1990b
Lugo et al. 1986
Brown and Lugo 1990b
Bautista-Cruz and del Castillo 2005
Bautista-Cruz and del Castillo 2005
Bautista-Cruz and del Castillo 2005
Lugo 1992
Reference
58
E. Marín-Spiotta et al.
50
50
50
51
55
60
60
60
60
61
61
75
75
75
80
80
80
100
Mature
Mature
Mature
Mature
Mature
Mature
Mature
Mature
Mature
Mature
Mature
Mature
25
25
100
25
25
10
30
30
30
10
30
20
20
20
30
30
30
25
10
25
25
10
10
10
20
20
20
20
30
30
D
D
W
W
M
D
M
M
M
M
M
M
M
M
M
M
M
M
D
D
D
M
M
M
M
M
M
M
M
M
A
A
A
P
P
P
P
P
A
A
A
P
P
P
A
A
P
P
S
S
P
S
S
S
S
S
S
S
S
S
S
S
S
S
S
90
90
81
80
80
41
61
64
60
43
78
135
105
126
97
49
59
80
36
60
45
26
18
23
189
106
210
126
70
44
90
90
25
80
80
70
75
78
73
93
95
204
159
190
119
60
72
80
62
60
45
56
39
50
285
160
317
190
86
54
Puerto Rico and US Virgin Islands
Puerto Rico
Puerto Rico
Puerto Rico and US Virgin Islands
Puerto Rico and US Virgin Islands
Mexico
Puerto Rico
Puerto Rico
Puerto Rico
Puerto Rico
Puerto Rico
Mexico
Mexico
Mexico
Puerto Rico
Puerto Rico
Puerto Rico
Puerto Rico and US Virgin Islands
Mexico
Puerto Rico and US Virgin Islands
Puerto Rico and US Virgin Islands
Brazil
Brazil
Brazil
Mexico
Brazil
Mexico
Mexico
Puerto Rico
Puerto Rico
(continued)
Brown and Lugo 1990b
Lugo et al. 1986
Lugo 1992
Brown and Lugo 1990b
Brown and Lugo 1990b
Saynes et al. 2005
Marin-Spiotta 2006
Marin-Spiotta 2006
Marin-Spiotta 2006
Silver et al. 2004
Silver et al. 2004
Bautista-Cruz and del Castillo 2005
Bautista-Cruz and del Castillo 2005
Bautista-Cruz and del Castillo 2005
Marin-Spiotta 2006
Marin-Spiotta 2006
Marin-Spiotta 2006
Brown and Lugo 1990b
Saynes et al. 2005
Brown and Lugo 1990b
Brown and Lugo 1990b
de Camargo et al. 1999
Salimon et al. 2004
Salimon et al. 2004
Bautista-Cruz and del Castillo 2005
Smith et al. 1998
Bautista-Cruz and del Castillo 2005
Bautista-Cruz and del Castillo 2005
Marin-Spiotta 2006
Marin-Spiotta 2006
2 Agricultural Abandonment in the Neotropics
59
Depth
(cm)
30
40
5
10
10
10
20
25
25
25
25
30
30
30
30
30
30
Age (y)
Mature
Mature
Mature
Mature
Mature
Mature
Mature
Mature
Mature
Mature
Mature
Mature
Mature
Mature
Mature
Mature
Mature
M
M
W
W
W
W
W
W
W
W
W
W
W
W
W
W
W
Life Zone
Appendix 2.2 (continued)
Past Land
Use
Cover type
73
57
29
14
17
28
23
89
110
85
62
87
91
116
90
109
49
Soil C
(Mg/ha)
89
60
112
30
37
60
28
89
110
85
62
75
78
100
77
94
42
Soil C 0–25 cm
(Mg/ha)
Puerto Rico
Brazil
Ecuador
Costa Rica
Costa Rica
Costa Rica
Brazil
USA-Hawaii
Puerto Rico and US Virgin Islands
Puerto Rico and US Virgin Islands
Puerto Rico
Costa Rica
Costa Rica
Costa Rica
Ecuador
Ecuador
Costa Rica
Location
Marin-Spiotta 2006
Sommer et al. 2000
Fehse et al. 2002
Reiners et al. 1994
Reiners et al. 1994
Reiners et al. 1994
Johnson et al. 2001
Bashkin and Binkley 1998
Brown and Lugo 1990b
Brown and Lugo 1990b
Lugo et al. 1986
Guariguata et al. 1997
Guariguata et al. 1997
Guariguata et al. 1997
Rhoades et al. 2000
Rhoades et al. 2000
Werner 1984
Reference
60
E. Marín-Spiotta et al.
2 Agricultural Abandonment in the Neotropics
61
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Forest Ecology and Management 78:147–157.
Chapter 3
Post Sugar Cane Succession in Moist Alluvial
Sites in Puerto Rico
Oscar Abelleira Martínez and Ariel E. Lugo
Abstract Sugar cane cultivation spanned over two centuries in Puerto Rico, at one
time covering 14% of the island’s total area. The most productive plantations were
located in alluvial valleys that were deforested before there was an opportunity to
record the species composition and structural development of the original native
forests. This has led to speculation about the nature of these forests. With the abandonment of sugar cane cultivation in the twentieth century, sugar cane fields have
been converted to pastures and a variety of secondary forests. One of these is a new
forest type dominated by Spathodea campanulata, an invasive, wind-dispersed, and
evergreen or nearly deciduous introduced species. We studied three stands of these
forests at ages 25–40 years old on moist alluvial sites and found high basal area,
intermediate tree density, high Complexity Index, and relatively low species richness. We recorded 19 tree species with dbh > 2.5 cm and 25 understory species
equivalent to a total of 31 species. Of these, 11 species were introduced and 20 were
native, including two endemic tree species. The large tree stratum (dbh > 10 cm)
had high dominance of S. campanulata, 6 tree species, and a mean height of 12 m.
Compared to large trees, small trees (> 2.5 <10 cm dbh) had higher species richness,
and less dominance, while the understory had the highest species richness and lowest dominance. In spite of centuries of cultivation, forest soils were nutrient-rich,
which allowed trees to attain large diameters, and produce nutrient-rich litter. The
structural parameters of alluvial S. campanulata forests are comparable with those
of remnant alluvial moist forests elsewhere on the north coast of Puerto Rico but
the species composition is different. Periodic inundation of these forests appears
to favor flood-tolerant species such as S. campanulata, and extends the period of
domination of this species relative to its behavior in other geologic substrates in
Puerto Rico. This process is exacerbated by long-term and extensive history of
cultivation that reduced the species pool of seed sources for alluvial forests in the
region. It will be difficult to restore the species complement of the original native
forests because the disturbance regime of alluvial sites in Puerto Rico will continue
to favor the presence, albeit reduced, of introduced species in the mature phases of
these forests.
R. W. Myster, Post-Agricultural Succession in the Neotropics.
C Springer 2008
73
74
O. Abelleira Martínez, A. E. Lugo
3.1 Introduction
Sugar cane was first planted in Puerto Rico in 1953 (Álvarez Nazario 1982,
Domínguez Cristóbal 2000). Considerable deforestation accompanied its development as a crop throughout the seventeenth and eighteenth centuries (Domínguez
Cristóbal 1989a, 1989b). Eventually, sugar cane became the major agricultural
crop in Puerto Rico for most of the nineteenth and twentieth century. At a peak
production of over 1 million tons of sugar in the 1950s, the area covered by sugar
cane plantations was over 120,000 ha or about 14% of the island’s land area (Picó
1969). Most of this cane was planted on alluvial soils on the coastal plains of Puerto
Rico. In many instances, drainage canals were dug to protect sugar cane from
flooding. These drainage works allowed cultivation to extend to wetland soils and
caused the displacement of coastal wetlands.
Alluvial valleys were the first lands to be significantly deforested in Puerto Rico
because they represented the best agricultural conditions available (flat lands, deep
fertile soils). As a result of this early activity, there is no documentation of the
structure and species composition of native vegetation in these valleys before they
were converted for agricultural use. The vegetation of alluvial valleys was probably arrayed along a hydroperiod gradient that included forested and non-forested
wetlands in flooded areas and moist tropical forests in locations with short or no
hydroperiod (Lugo 2005). Today, alluvial soils attract urban development, and forest
cover on alluvial soils is minimal as is the area protected (Table 3.1).
Sugar cane cultivation declined after the 1950s and was terminated in the 1980s.
The abandonment of sugar cane fields results in different succession pathways
depending on the location of the field and the timing of the abandonment. On
higher ground, invasions of guava trees (Psidium guajava) hindered initial efforts
by Spaniards to establish pastures on alluvial soils (Domínguez Cristóbal 2000).
Today, this tree still forms thickets on abandoned pasturelands. With the abandonment of the drainage canals, herbaceous wetlands as well as mangrove and Pterocarpus forests expanded in area in the coastal zone (Lugo 2006). The expansion
of mangroves is confined to saline soils and the expansion of Pterocarpus forests
has been limited to stands near rivers or other wetlands. Many sites not invaded
by guava trees allowed pastures to become a major land cover type by the 1970s
(Ramos and Lugo 1994). In moist transition dry alluvial valleys, mature vegetation
is dominated by Bucida buceras and Guazuma ulmifolia and secondary forests with
Table 3.1 Area of lowland alluvial habitat and forest cover and protected status of forests in lowland alluvial habitats in Puerto Rico (Helmer et al. 2002). Data are for 1991–1992. There are
smaller areas of alluvial habitat in montane conditions, but these areas were not included here
Statistic
Dry alluvial
Moist alluvial
All Puerto Rico
Zone area (ha)
Percent protected
Forest area (ha)
Forest area protected (ha)
Percent of zone area
Percent of island-wide land development
45,179
5.5
5,368
211
0.5
10.5
144,767
4.5
16,224
729
0.5
26.0
869,156
365,160
34,026
100
3 Post Sugar Cane Succession in Moist Alluvial Sites in Puerto Rico
75
mixtures of native and introduced species are developing on abandoned sugar cane
fields in those dry alluvial valleys (Lugo 2005). However, in moist alluvial valleys
of the north and east coast, abandoned sugar cane fields have been replaced in many
locations by forests dominated by Spathodea campanulata (Chinea 2002).
Spathodea campanulata belongs to the family Bignoniaceae and its native to
tropical Africa in a band around the equator between 12◦ N and 12◦S from sea level
to 1,200 m elevation in a wide variety of soils (Little and Wadsworth 1964, Francis
2000). Spathodea campanulata is shade intolerant, and produces lateral roots, light
seeds (125,000 seeds/kg), and buttresses (Little and Wadsworth 1964, Francis 2000).
It is also a pioneer species (Whitmore 1990) that was introduced to Puerto Rico for
its ornamental value and is now considered a fast-growing invasive species (Francis
2000, Chinea 2002). The earliest specimen in island herbaria is dated 1912 (UPR
Botanical Garden Herbarium). There is evidence that S. campanulata shows histological and biochemical adaptation to flooded soils (Medri and Correa 1985).
The invasion of deforested alluvial soils by S. campanulata gave us the opportunity to study post sugar cane succession on moist alluvial sites. This is important
because these sites have not been forested for centuries and there is much speculation about the potentially impressive structural development of native alluvial
forest stands (Lugo 2005). Gleason and Cook (1926) considered the forests on alluvial soils, termed playa forests by them, as the climax community for the northern coastal plain. They listed the following species as probable components of this
forest type: Calophyllum antillanum (synonymous with C. calaba) and Tabebuia
pallida (synonymous with T. heterophylla). Wadsworth (1950) added Manilkara
nitida (synonymous with M. bidentata), and Mastichodendron foetidissimum (synonymous with Sideroxylon foetidissimum) as likely timber species in these forests.
Wadsworth also estimated that alluvial forests attained 26 m in height.
Would S. campanulata forests mimic the native stands not seen in the Island since
the seventeenth century? The answer could be yes, at least structurally, if sites were
not degraded. On the other hand, Chinea (2002) noted that stands dominated by
S. campanulata are floristically far removed from the original vegetation on alluvial
sites and it is not clear how much time will be required for the re-establishment of
the original vegetation, or if that is even possible (see Lugo and Helmer 2004). Nevertheless, in this study we describe the structure and species composition of alluvial
forest stands dominated by S. campanulata, infer some aspects of their functioning,
and compare this succession with other post agricultural successions in Puerto Rico
(Aide et al. 2000, Chinea 2002) and elsewhere (Muthuramkumar et al. 2006).
3.2 The Alluvial Environment
Our study sites are located in the subtropical moist forest life zone (sensu
Holdridge 1967) with a mean annual temperature of about 25◦ C (NOAA
1998) and a mean annual rainfall between 1971 and 2000 of 1,443 mm
(http://www.srh.noaa.gov/sju/pr_mean_annual_pcp.jpg). The alluvial soils of the
Río Cibuco valley reach maximum depths of 85.3 m. According to Abruña et al.
(1977) these alluvial soils on the north coast constitute some of the finest soils of
76
O. Abelleira Martínez, A. E. Lugo
Table 3.2 Site characteristics, stand age, and land use history. Soil information is from Acevido
(1982). For the land use information we used air photos from 1963, 1971, 1977, and 1985. We
also interviewed people and used Ramos and Lugo (1994) for the 1977 information. Elevation data
were obtained in the field with a Trimble GPS unit at + 2 m precision
Site
Soil order
Cibuco I
Inceptisol
Cibuco II
Mollisol
Paso del Indio
Mollisol
Soil series
and slope (%)
Past agricultural use
and stand age (yr)
Coloso, clay
cienosa (0)
Toa Alta, clay
marga (0)
Toa Alta, clay
marga (0)
Sugar cane (27)
17.6
Sugar cane (38)
7.0
Sugar cane and
pasture (25)
15.7
Elevation (m)
Puerto Rico and are classified as providing prime agricultural lands, which means
they reach the highest standards for food production in terms of growth conditions
(water, nutrients, structure, etc.) and resistance to erosion. The Río Cibuco Valley
has been populated and farmed since the sixteenth century (Álvarez Nazario 1982).
The US Army Corps of Engineers (1973) documented the history of flooding
at Río Cibuco and Río Indio valleys and reported the following flood levels in the
vicinity of our sites for the standard flood expected under extreme storm events (m
above mean sea level): Río Indio- 16.8, Cibuco I- 22, and Cibuco II- 13. During
these events the river rises up to 5 m above its normal level at rates of up to 1 m
per hour, for time periods of up to 8 hours, and the critical stage lasting for up to 27
hours. Given the elevations of the floodplain at our sites (Table 3.2), the sites flood
between 1.1 and 6.0 m during peak floods. Historical floods are known since the
1770s when the town of El Naranjal, now Vega Baja, was established in the region.
These data show a recurrent flooding pattern that trees must adapt to because floods
can occur at any month of the year (Army Corps of Engineers 1973).
3.3 Site Selection and Methods
We selected three sites in the alluvial floodplains of Puerto Rico’s north coast
between Vega Baja and Vega Alta (Fig. 3.1). We focused on closed canopy secondary forests dominated by S. campanulata with 25–40 years of age and used high
resolution aerial photos to locate sites, and field visits to interview land owners about
past land uses of their lands. We strived for sites as similar as possible in terms of
soils, life zone conditions, topography, and previous land use (Table 3.2).
At each site, we selected 10 random points in each of two parallel 100-m transects. The transects were at least 15 m apart and were also randomly located. At each
point we identified to species the four trees with diameter at breast height (dbh)
> 10 cm nearest to the point (one in each of four imaginary quarters surrounding
the point) and measured their dbh at 1.37 m and distance to the point (Cottam and
Curtis 1956). We repeated this procedure for trees with dbh >2.5 <10 cm. The total
sampling was 160 trees per site (80 > 10cm dbh and 80 > 2.5 cm < 10 cm dbh). For
trees with multiple stems we measured each stem’s diameter at dbh for the estimate
of basal area but counted the tree once for density estimates. We refer to trees with
3 Post Sugar Cane Succession in Moist Alluvial Sites in Puerto Rico
77
Fig. 3.1 Location of study sites
dbh > 10 cm as large trees and the ones with lower dbh as small trees. We measured
the height of 10 large trees in each stand with a clinometer.
To characterize understory vegetation, we randomly selected four additional
points per site and established 5 × 5 m plots at each point. All woody plants with
basal diameter < 2.5 cm inside these plots were identified to species and counted.
These plants were mostly shrubs, small trees, saplings, and seedlings. We constructed species/area curves for each site using data from the four plots per site.
Each tree and understory species was classified as native or introduced. If native,
we further checked those that were endemic. We used Little and Wadsworth (1964)
and Little et al. (1974) for the nomenclature of species.
With the data collected from transects, we followed Cottam and Curtis (1956) to
estimate tree density, basal area, frequency and Importance Value (IV), expressed in
percent. For understory data, we divided the number of plants of each species by the
area of the plot to estimate stem density and estimated an IV in terms of frequency
and density, expressed in percent.
At each site we collected loose litter using five 25 × 25 cm randomly located
plots per site. The litter was sorted in the field into leaves, wood, and miscellaneous.
Where possible, we separated S. campanulata leaves from other leaves, and classified leaves into recently fallen, fragmented, and old. In the laboratory, samples
were oven dried at 60◦ C, weighed, ground, and analyzed for a suite of elements
(Al, Ca, Fe, Mg, Mn, Na, P, C, S, and N) as follows. A sub sample was oven dried
at 105◦ C for 24 hrs and a moisture factor calculated and applied to each analysis
78
O. Abelleira Martínez, A. E. Lugo
(Wilde 1979). Ground material was digested with concentrated HNO3 and 30%
H2 O2 and analyzed with a Spectro plasma emission spectrometer (Spectro Ciros
ICP) for P, K, Ca, Mg, Al, Mn, Fe, and Na using the digestion method recommended
by Luh-Huang and Schulte (1985). Total Nitrogen, total Carbon and total Sulfur
were analyzed using the dry combustion method by means of a LECO CNS-2000
Analyzer. The procedure used is a modified version of the method in LECO Corporation (1995). In the dry combustion method a small weighed sample is combusted
by heating it to a high temperature (1,300◦C) inside a resistance furnace and in a
stream of purified oxygen.
Precision for most analyses was assured by running samples of known chemical composition every forty determinations (for total C, total N and total S every
20 determinations). These control samples: citrus leaves (NBS-1572), peach leaves
(NIST-1547), and pine needles (NIST-1575) were obtained from the National Institute of Standards and Technology, USA. The calibration standards: tobacco leaves,
orchard leaves and alfalfa, used in the total C, N, and S analysis were obtained from
LECO Corporation (St. Joseph, MI).
We also collected soil samples at two depths: 0–15 and 15–30 cm. We dug five
pits per site at randomly located points. Sampling included known soil volumes for
bulk density determinations and soil for chemical analyses. Samples were transported to the laboratory in pre-weighted cloth bags and dried at 100◦ C to constant
weight. Chemical analyses included the same elements as above plus Loss on Ignition. Samples analyzed for total P, K, Ca, Mg, Al, Mn, and Fe were treated the
same as litter samples above. For exchangeable ions such as Ca, Mg, Na, and Al we
used the 1N KCl extraction method and for Fe, Mn, K, and P we used the OlsenEDTA(NH4 -EDTA-NaHCO3 ) procedure recommended by Anderson and Ingram
(1993). For total N, C, and S we used the same methods as in with litter above.
Soil pH was determined in a solution 1:1 (soil:water/soil:1N KCl) using an Orion
Ionalyzer Model 901 with a combination pH electrode (McLean 1982).
We only report the stocks of nutrients and organic matter in loose litter and soil.
To estimate element stocks in loose litter, we multiplied the mean dry weight per
unit area of litter component by the corresponding element concentration. For soil
element stocks, we multiplied the bulk density of each sample by depth of sampling and by the corresponding element concentration. Concentration and soil bulk
density data are available from the authors.
3.4 Forest Development After Abandonment
3.4.1 Species Composition
We recorded 18 tree species above the understory (Table 3.3) and 26 species in the
understory (Table 3.4) in the three alluvial stands. Of these, 13 species occurred
both as trees and in the understory, five occurred only as trees but not in the understory, and 13 were found only in the understory. All species in the understory were
tree or woody shrub species. Two species were endemic to Puerto Rico: Thespesia
3 Post Sugar Cane Succession in Moist Alluvial Sites in Puerto Rico
79
Table 3.3 Importance Value of tree species in alluvial stands of Spathodea campanulata in the
north coast of Puerto Rico. The sizes of trees correspond to the diameter at breast height (1.37 m)
Cibuco I
> 10
> 2.5 < 10
Cibuco II
Paso del Indio
> 10
> 10
> 2.5 < 10
Spathodea campanulata*
Cassia siamea*
Albizzia procera*
Terminala catappa*
Casearia guianensis
Guarea guidonia
Casearia decandra
Zanthoxylum martinicense
Spondias mombin
Erythrina poeppigiana*
Casearia sylvestris
Cupania americana
Eugenia biflora
Ocotea coriacea
Thespesia grandiflora**
Callophylum calaba
Cordia laevigata
Bucida buceras
91.8
89.9
> 2.5 < 10
(cm)
Species
90.0
4.4
3.7
1.8
49.9
6.9
91.4
83.3
6.1
12.8
11.0
2.6
1.8
1.4
1.3
1.2
1.2
1.2
1.2
1.2
6.7
5.5
4.0
5.5
5.7
1.8
2.5
1.8
1.8
2.2
3.6
2.1
1.8
*introduced
**endemic
grandiflora, which occurred both as large trees and in the understory, and Roystonea
borinquena, which was found only in the understory. Thus, a total of 31 species
were recorded in these alluvial stands. Five introduced species occurred as trees,
8 occurred in the understory, and two of these species were shared between strata.
There were a total of 11 introduced species in the list or 35% of the species tallied.
The species/area curve (Fig. 3.2) saturated with the plots sampled, suggesting that
the sampling was sufficient to characterize the floristic diversity of the understory.
The species composition varied among stands (Tables 3.3 and 3.4; Fig. 3.2). Cibuco
II had more tree species in the understory (Table 3.4), than either Paso del Indio or
Cibuco I, but Cibuco I had more tree species in the tree stratum than the other two
sites (Table 3.3).
Fig. 3.2 Species/area curves
for understory vegetation in
alluvial stands of Spathodea
campanulata in northern
Puerto Rico. All woody
plants with a dbh < 2 cm
basal diameter were included
Number of Species
20
16
12
8
Cibuco I
4
Cibuco II
Paso del Indio
0
0
50
100
Area Sampled (m2)
150
80
O. Abelleira Martínez, A. E. Lugo
Table 3.4 Importance Value for species in the understory of alluvial stands of Spathodea campanulata in northern Puerto Rico
Species
Cibuco I
Cibuco II
Paso del Indio
Andira inermis
Spathodea campanulata*
Guarea guidonia
Calophyllum calaba
Casearia guianensis
Ocotea coriacea
Roystonea borinquena**
Inga laurina
Melicoccus bijugatus*
Casearia sylvestris
Terminala catappa*
Casearia decandra
Delonix regia*
Cinnamomum elongatum
Eugenia monticola
Cochlospermum vitifolium*
Cupania americana
Guettarda scabra
Syzygium jambos*
Leucaena leucocephala*
Erythroxylon brevipes
Eugenia biflora
Thespesia grandiflora**
Bucida buceras
Psidum guajava*
Spondias mombin
40.7
15.8
11.9
10.3
10.3
4.2
3.4
3.4
13.0
4.7
26.1
6.5
1.9
7.3
1.6
26.2
13.1
4.5
14.6
3.3
7.0
6.9
6.8
3.3
3.1
1.9
1.9
1.9
1.6
1.6
1.6
1.6
6.5
4.5
3.3
7.7
3.3
3.3
3.3
3.3
3.3
*introduced
**endemic
Table 3.5 Structure of alluvial stands of Spathodea campanulata in the north coast of Puerto Rico.
Tree dimensions refer to diameter at breast height (1.37 m). The area sampled for understory was
100 m2 . Empty cells mean the information is not available
Site
Cibuco I
Cibuco II
Paso del Indio
Component
Trees >10 cm
Trees > 2.5 <10 cm
All trees
Understory
Trees >10 cm
Trees > 2.5 <10 cm
All trees
Understory
Trees >10 cm
Trees > 2.5 <10 cm
All trees
Understory
Stem density
(no/ha)
1,071
1,015
2,087
6,300
1,090
1,096
2,187
17,200
1,825
1,482
3,307
4,200
Basal area
(m 2 /ha)
77.0
2.6
79.5
104.5
2.5
106.9
82.0
4.8
86.8
Number of
species
4
15
15
8
3
5
6
19
3
6
8
14
3 Post Sugar Cane Succession in Moist Alluvial Sites in Puerto Rico
81
3.4.2 Forest and Community Structure
Tree density and basal area were very similar across sites (Table 3.5). Most of the
basal area of the stands was in the large trees, and the ranking of stands according
to basal area was the same as their ranking for species richness: highest in Cibuco
II followed by Paso del Indio and Cibuco I. The density of understory plants was
higher in Cibuco II than in the other two stands. We found no significant difference in tree height between stands. The height of the measured 30 trees averaged
12.1 + 0.5 m.
The S. campanulata forests had gradients of species dominance and species richness from the canopy to the understory (Fig. 3.3). Large trees had lower species
Importance Value (%)
A
100
Cibuco I
Cibuco II
Paso del Indio
10
1
0
1
2
3
4
5
Species Rank
Importance Value (%)
B
100
Cibuco I
Cibuco II
Paso del
Indio
10
1
0
5
10
15
Species Rank
Importance Value (%)
C
100
Cibuco I
Cibuco II
Paso del Indio
10
1
0
5
10
15
20
Species Rank
Fig. 3.3 Importance value curves for (a) tree species with > 10 cm diameter at breast height
(dbh), (b) tree species with dbh > 2 < 10 cm dbh, and (c) understory vegetation with basal diameter
< 2 cm in Spathodea campanulata forests in northern Puerto Rico
82
O. Abelleira Martínez, A. E. Lugo
Table 3.6 Biomass of litter components of Spathodea campanulata alluvial forest stands in northern Puerto Rico. Data corresponds to July, 2004. Mean and standard error are included, n = 10.
All data are in g/m2 and values are rounded off to the nearest gram and therefore might not add to
the total
Litter component
Cibuco I
Cibuco II
Paso del Indio
Recently fallen leaves
Old leaves
Fragmented leaves
Total leaves
Wood
Miscellaneous
2 (0.8)
5 (2)
38 (8)
45
321 (94)
36 (25)
3 (1)
9 (2)
159 (27)
171
364 (61)
20 (10)
2 (0.5)
11 (3)
114 (66)
127
334 (97)
9 (5)
Total
403 (90)
556 (77)
469 (125)
richness and the dominant species had a higher importance value among large trees
(Fig. 3.3a), than among small trees (Fig. 3.3b), or understory plants (Fig. 3.3c). In all
stands, the IV of the highest ranked large trees was uniformly higher than 95%. For
smaller trees, the number of species and IV of species increased and were reduced,
respectively, relative to large trees. The understory had the highest species richness
and lower IV for similarly ranked species when compared to small or large trees.
3.4.3 Loose Litter and Soil Compartments
Most of the loose litter in the three alluvial stands was woody litter (Table 3.6).
Leaves, mostly fragmented, were a smaller component of litter in all stands. The
stock of elements in loose litter varied among stands in proportion with the total
mass of loose litter (Table 3.7). Some of the differences in element stocks were due
to differences in element concentrations and proportions of litter components in total
loose litter. For example, leaf litter had greater element concentrations than wood
and even though leaf litter was a small fraction of the total loose litter, it contained
disproportionately high stocks of elements. Soil had much higher quantities of elements than loose litter and there was an insignificant difference in element stocks
between the two depths studied (Table 3.7). We did not find a developed humus
layer atop these soils perhaps because of insufficient time for its development, or
because periodic flooding prevents the accumulation of humus.
3.5 Ecological Comparisons and Implications
3.5.1 Structure and Species Dynamics
Our data shows that even by age 25–40 years, alluvial forests dominated by S. campanulata develop structural complexity similar to remnant mature stands of natural
alluvial forests (Table 3.8). For example, S. campanulata forests had a higher basal
13.3
0.0
0.1
13.5
10.8
0.7
24.9
Fragmented leaves
Recently fallen leaves
Old leaves
Total leaves
Wood
Miscellaneous
Total
108.1
44.6
1.0
2.3
48.0
55.0
5.1
52.5
36.5
19.8
0.0
0.2
20.0
15.6
0.9
28.1
6.5
2.2
0.2
0.3
2.7
3.6
0.3
6.2
0.6
0.2
0.1
0.9
2.5
2.8
11.3
5.4
0.1
0.3
5.8
5.1
0.4
8.3
1.5
0.1
0.2
1.9
6.1
0.3
0.4
0.3
0.1
0.0
0.0
0.1
0.2
0.0
0.9
0.5
0.0
0.0
0.5
0.4
0.0
0.5
0.2
0.0
0.0
0.2
0.2
0.0
Cibuco II
0.7
0.1
0.0
0.0
0.2
0.6
0.0
Cibuco 1
1.0
19.8
5.4
0.0
0.3
5.7
22.3
0.1
9.3
Total
10.4
0.6
1.6
12.6
38.6
1.3
4.2
4.0
0.0
0.2
4.2
15.5
0.1
36.6
0.2
0.0
0.0
0.2
0.3
0.0
Fragmented leaves
Recently fallen leaves
Old leaves
Total leaves
Wood
Miscellaneous
85.6
4.6
0.1
0.3
5.0
4.1
0.2
25.0
1.2
0.1
0.2
1.5
2.6
0.1
Na
Total
20.6
0.1
1.1
21.7
14.0
0.8
Mn
0.5
0.0
0.0
0.5
0.4
0.0
41.3
0.1
2.9
44.3
39.8
1.5
Mg
13.9
0.1
0.8
14.7
9.7
0.6
K
Fragmented leaves
Recently fallen leaves
Old leaves
Total leaves
Wood
Miscellaneous
Fe
Paso del Indio
Ca
Loose litter
Al
5.9
2.2
0.0
0.1
2.4
3.2
0.3
2.4
0.4
0.0
0.0
0.5
1.7
0.2
3.5
1.3
0.0
0.1
1.5
2.0
0.0
P
2,580
650
14
46
710
1,771
99
1,791
128
9
21
159
1,448
184
2031
329
9
44
382
1614
35
C
10.8
4.4
0.2
0.3
4.9
5.3
0.6
5.0
0.8
0.1
0.2
1.1
3.5
0.4
5.1
1.8
0.1
0.2
2.1
2.9
0.1
S
432
5
11
446
409
27
480
12
92
371
24
563
Ash
80.7
883
(continued)
35.0
0.4
1.9
37.3
37.5
5.9
33.1
6.8
0.3
0.9
8.0
20.9
4.1
36.4
14.8
0.3
1.6
16.6
19.0
0.7
N
Table 3.7 Element content of litter (kg/ha) and soil (Mg/ha) compartments of Spathodea campanulata alluvial forests in northern Puerto Rico. Data are rounded
to the nearest kg or tenth of kg and thus total might not add. Empty cells mean no data available
3 Post Sugar Cane Succession in Moist Alluvial Sites in Puerto Rico
83
30.6
60.6
23.2
34.5
57.7
15–30
Total (0–30)
0–15
15–30
Total (0–30)
11.0
6.2
4.8
14.3
6.7
7.5
84.5
50.2
34.3
89.5
44.7
44.8
1.5
0.9
0.7
1.5
0.7
0.8
1.1
14.4
8.5
5.9
15.8
7.8
8.0
11.1
0.3
0.2
0.2
0.2
2.1
1.3
0.8
0.3
0.2
0.1
Cibuco II
2.3
1.2
1.2
Cibuco I
1.5
0.1
0.1
30.0
63.6
0.7
0–15
27.7
5.6
5.5
41.2
0.5
0.6
Total (0–30)
32.3
31.3
Na
0.7
12.2
15.5
Mn
20.3
Mg
20.9
K
15–30
Fe
0–15
Ca
Paso del Indio
Al
Soil depth (cm)
Table 3.7 (continued)
1.2
0.7
0.6
1.3
0.6
0.7
1.1
0.5
0.6
P
256.0
142.1
113.9
288.5
141.1
147.4
237.4
110.5
126.8
C
37.9
18.7
19.2
43.0
20.0
23.0
47.5
18.7
28.8
S
0.5
0.2
0.2
0.5
0.2
0.3
0.4
0.2
0.2
N
4.2
2.2
2.0
4.1
1.9
2.1
3.9
1.6
2.3
Ash
84
O. Abelleira Martínez, A. E. Lugo
3 Post Sugar Cane Succession in Moist Alluvial Sites in Puerto Rico
85
Table 3.8 Species richness and structural parameters of lowland alluvial moist forests in Puerto
Rico (Lugo 2005). Minimum diameter was 2.5 cm. The Complexity Index (CI) is the product of
number of species, stem density, basal area, and height divided by 1,000. The number of species,
density and basal area are per 0.1 ha. Standard error is in parenthesis. For Pterocarpus, n was 10.
Manilkara data (n = 9) are from Gould et al. (2006). Spathodea campanulata data are from this
study (n = 3)
Area
sampled (ha)
Number of
tree species
Stem density
(stems/ha)
Pterocarpus*
0.1
Syzygium
Manilkara
0.1
0.09
8
(2.2)
21
38
1,705
(208)
1,880
4,389
(165)
2,527
(391)
Forest type
S.campanulata
10
(2.7)
Basal area
(m2 /ha)
42.6
(3.7)
41.6
50.0
(11.1)
91.1
(8.2)
Height (m)
18.6
(1.8)
19.7
26
(2.8)
17.5**
CI
118
(37)
423
2,168
402
*Two Pterocarpus stands on alluvial soils several km east of our sites had basal areas of 73.5 and
68.9 m2 /ha (Eusse and Aide 1999)
**Mean of tallest three trees
area than any other forest on alluvial soil, but they had lower species richness and
intermediate tree densities. As a result, the Complexity Index is higher than those
of Pterocarpus wetlands, comparable to introduced Syzygium forest, and lower than
native Manilkara forest. The Manilkara forest, a mature native stand, had taller
trees, higher tree density, and higher species richness than the S. campanulata forest.
These differences between the S. campanulata and Manilkara forest demonstrates
the floristic distance between S. campanulata and native vegetation on alluvial soils
as pointed out by Chinea (2002). However, the age differential between these two
forest types is considerable as the Manilkara forest is probably over 200 years old
compared to a maximum age of 40 years for S. campanulata. In spite of this difference in age, the high productivity of S. campanulata is evident. If we use the mean
basal area in Table 3.8, a wood density of 0.26 g/cm3 (Francis 2000), a mean stand
age of 30 years, and a bole height of 10 m, we obtain a net rate of wood biomass
accumulation of 8 Mg/ha.yr, a wood biomass productivity value higher than all but
one of the values reported by Brown and Lugo (1982) in a review of tropical forest
productivity.
Age plays a role in the development of community structure and composition
among S. campanulata stands. The older Cibuco II stand had higher understory
species richness and plant density, and larger basal area, than the other and younger
stands (Table 3.5). Growth is so rapid in these forests that a decade difference in age
represents significant opportunity for development. Aide et al. (2000) reported that
S. campanulata stands on karst and volcanic parent materials are replaced by native
species after 40 years. Our 40-year old stand at Cibuco II showed signs of S. campanulata being replaced in the near-term. Spathodea campanulata had the lowest
IV in the understory of Cibuco II (Table 3.4). The reason why S. campanulata might
still prevail after 40 year on alluvial soils might be the result of different conditions
in the Aide et al. (2000) study. We have observed that on karstic and volcanic
soils S. campanulata stands accumulate more species than stands under alluvial
86
O. Abelleira Martínez, A. E. Lugo
conditions. This is consistent with the mature stands studied by Aide et al. (2000).
They found more species than we did (20 species of large trees per 0.1 ha), the basal
areas were lower (<50 m2 /ha), and the density of S. campanulata was lower (<400
individual/ha) than in our alluvial stands (Table 3.5). It appears that on alluvial soils
following sugar cane cultivation, flooding and absence of propagules slow forest
succession. We suggest that the causal factors for this slowing of succession are
flooding, which stresses trees and selects for a few species that can tolerate floods,
and the history of land use, including the length of time and extensiveness of cultivation. The initial deforestation and conversion of the alluvial valleys to agricultural
use affected large land areas and kept them free of trees for centuries. Such land use
must have reduced the tree species pool available in the alluvial valleys as well as
limiting the availability of seed sources for re-colonization of abandoned fields.
Clearly, the length and intensity of sugar cane cultivation do not appear to curtail
the primary productivity of the study sites (discussed below) but cultivation does
appear to have an effect on the species composition of stands. Our results reflect a
high proportion of introduced species (35%) in these forests (Tables 3.3 and 3.4).
Many of these species are indicators of human use such as for fruit consumption,
ornamental use, or use as shade trees (Psidium guajava, Delonix regia, and Erythrina poeppigiana, respectively). The presence of these species in the understory
suggests that they are persistent in these altered landscapes and are not likely to
disappear as succession proceeds to mature states (Lugo and Helmer 2004, Lugo
2004). However, in spite of the persistence of introduced species and low species
richness in S. campanulata alluvial forests, the species list we found suggests steady
enrichment with native species, particularly the 13 species found in the understory
but not yet in the small tree category.
In a study of rain forest fragments in India, Muthuramkumar et al. (2006) found
results similar to ours in the sense that past land use was reflected in the species
composition of the fragments. They found that S. campanulata was most abundant
(43% of trees present) in smaller and previously disturbed fragments. Other introduced and pioneer species were associated with S. campanulata in these disturbed
fragments. However, the presence of the introduced species did not hinder the development of stands with high species richness of native species. In abandoned wet
pastures on volcanic-derived soils at 420 m elevation in Puerto Rico, S. campanulata did not appear as a component of the early succession (Myster 2003). Instead,
the introduced Syzygium jambos dominates the tree layer in these successions but
they lag behind herbaceous grasses, forbs, and ferns. Myster (2004) showed through
experiments that seed predation and disease limited tree establishment in these pastures and not seed rain.
In the moist alluvial old fields, the early stages of succession appear to be dominated by grasses including the grass Panicum laxum, which Britton and Wilson
(1923) identified as a grass typical of moist and wet places. Other herbaceous plants
growing on those areas were Commelina diffusa (native and pantropical) at Cibuco I,
and Cyperus odoratus (native and pantropical) at Cibuco II, but both were present at
much lower densities than P. laxum. An understory herbaceous plant in the alluvial
forest was Petiveria alliacea (native to the Americas including Puerto Rico) found
at Cibuco I at low densities. Syngonium podophyllum (introduced from continental
3 Post Sugar Cane Succession in Moist Alluvial Sites in Puerto Rico
87
America) was present and abundant as herbaceous ground cover mostly in gaps or
near-gaps and grew as a climber in nearly all S. campanulata stems within all sites.
Dalbergia monetaria (native to the Americas including Puerto Rico) was another
climbing plant that was found at all sites although at much lower densities than S.
podophyllum.
Spathodea campanulata is able to grow through the grass cover just described
and eventually shade them out. Other fast-growing species such as S. jambos and
the native C. schreberiana germinate in the understory and within gaps of S. campanulata forests, respectively. Apparently the filtering mechanisms (sensu Myster
2004) that operate in wet pasture succession at higher elevation are not as effective
on moist alluvial pastures.
We anticipate a continuing floristic enrichment of S. campanulata stands and
eventual enrichment of the canopy as other species increase in importance and S.
campanulata declines. But the increase should not be to the levels measured in
other locations without floods (Aide et al. 2000). Although S. campanulata and
other introduced species will not grow under deep shade (S. jambos excepted), it
is possible that these species will not disappear from these stands for at least two
reasons. First, these alluvial sites flood periodically and the flooding is likely to limit
the richness of species that can grown and persist under flooded conditions. At Paso
del Indio, S. campanulata floods to depths of almost 2 meters, and these periodic
floods are likely to select for low tree density, low species richness, but high basal
area stands of flood-adapted species such as S. campanulata, Pterocarpus (not found
in our study) or Calophyllum calaba. The endemic palm Roystonea borinquena also
tolerates some flooding, as does the introduced S. jambos . If the species richness
remains low, competition with S. campanulata also decreases relative to that in other
soil types and it is more likely it will persist.
A second likely factor that might favor the persistence of introduced species in
these stands is wind. Periodic windstorms and hurricanes will continue to open
the canopy and favor the rapid growth of S. campanulata in canopy clearings. In
a closed canopy mature forest at the Luquillo Experimental Forest, Thompson et al.
(in press) recorded an increase in the density of S. campanulata and S. jambos after
hurricane Georges. This mechanism might be in effect at Cibuco I, which is the
only site exposed eastward to oncoming winds and which appears to have more
gaps in relation to the other sites. This can be the reason for the presence of shade
intolerant colonizers Cassia siamea and Zanthoxylum martinicense amongst small
trees at Cibuco I and their absence at the other sites. Normally the colonizer role
is played in moist and wet forests in Puerto Rico by the native Cecropia schreberiana, but Cecropia cannot invade abandoned old fields or pastures (Silander
1979). The absence of this native species in old field succession is probably one
reason why S. campanulata has been as successful as a colonizer of abandoned
sugar cane fields (its wind-dispersed seeds also help). But we have seen Cecropia
trees growing under the shade of S. campanulata in alluvial stands near highway
PR 2, an event that raises interesting future possibilities as continuous presence
of forest conditions on alluvial soils allow this native species to again establish
large populations in places where it was extirpated as a result of past cultivation
of sugar cane.
88
O. Abelleira Martínez, A. E. Lugo
3.5.2 Nutrient Dynamics
The nutrient stocks of alluvial soils of the study sites compared favorably with soils
from native forests in areas with different climates and geology (Table 3.9). Taking
into consideration differences in sampling depth, alluvial soils were comparatively
lower in soil organic matter and N and higher in P, K, Ca, and Mg than soils of
natural stands on different parent materials. Perhaps due to the annual fertilization
of sugar cane fields or the continuous input of silt from the uplands, the heavily
used sugar cane fields maintained soil fertility in contrast to the reductions of P
shown in Table 3.9 in dry karst forests used for various other purposes. Deep alluvial soils, when allowed to be flooded periodically are thus resilient to long-term
cultivation, as it appears to be the case in our study sites. Thus, a high nutrient
demanding tree species like S. campanulata can sustain its rapid growth and high
primary productivity in these fertile abandoned sugar cane fields. This is reflected in
a comparison of nutrient stocks in S. campanulata loose litter with litter from other
forest types in Puerto Rico (Table 3.10). In spite of the lower mass of accumulated
loose litter in alluvial stands, the stock of P, Ca, and Mg is disproportionately high
in S. campanulata.
Table 3.11 shows that S. campanulata leaf litter is high in N and P concentration
relative to leaf litter of mahogany plantations and mature dry karst forest. However,
alluvial soil was not as high in N concentration as the soils of those forest types,
although alluvial soil was higher in P, K, and Mg. The ecological implications of
these data are that the organic material produced by S. campanulata can circulate
rapidly in the forest ecosystem given its high nutritional quality for decomposers.
The presence of high densities of earthworm populations in these forests (Lugo et al.
Table 3.9 Comparison of mass and total nutrient content in soil of various forest types with contrasting use history in Puerto Rico. Soil organic matter is SOM in Mg/ha and nutrients are in
kg/ha to the specified depth. The last two rows correspond to this study of Spathodea campanulata
alluvial forests. Natural forest data are from a summary in Lugo (2005) and the rest is from Molina
Colón (1998). Standard error of the mean and n are in parenthesis for SOM. Empty cells mean
there are no data
Forest type/depth (cm)
Use History SOM
N
P
K
Ca
Mg
Moist alluvial (30)
Natural
Moist to wet ultramaphic
(30)
Moist volcanic (30)
Dry karst (15)
Dry karst (10)
Dry karst (10)
Natural
Dry karst (10)
Dry karst (10)
Dry karst (10)
Moist alluvial (15)
Moist alluvial (30)
Natural
Natural
Natural
Baseball
Park
Charcoal
pits
Farmlands
Houses
Sugar cane
Sugar cane
319
(63, 4)
879
(197,6)
594
238
80
71
12,550
5,000
4,700
655
569
6
540
115
274
6,500
4,800
8,400
1,220
263
659
85
5,300
9
203
5,600
293
73
98
129
261
5,000
6,000
2,146
4,062
12
21
603
1,196
361
410
681
1,393
6,300
6,700
9,265
17,639
460
333
6,469
13,788
3 Post Sugar Cane Succession in Moist Alluvial Sites in Puerto Rico
89
Table 3.10 Comparison of mass (Mg/ha) and nutrient content (kg/ha) in fine litter of various forest
types with contrasting use history in Puerto Rico. The rows labeled alluvial correspond to this study
of Spathodea campanulata alluvial forests on abandoned sugar cane fields. All other forest stands
were on land previously deforested and either abandoned (dry karst and wet volcanic) or planted
(mahogany plantations). Data are from Lugo and Fu (2003) and Lugo and Murphy (1986). Empty
cells mean there are no data. Standard error of the mean and n are given in parenthesis followed
by n, when sample size is > 3. For alluvial data, n = 3
Forest type
Mass
N
P
3.7
1.9
4.1 (1.1, 4)
3.4
1.5 (0.5)
2.7
2.4
5.1 (1.0, 4)
4.7
3.9 (1.0)
K
Ca
Mg
17
9.1
27 (15, 4)
20
1.7 (0.5)
64
524
193
35 (11)
14
32
22
4 (1.2)
8
11.3
33 (16, 4)
32
5.6 (0.7)
78
611
235
82 (16)
11
40
30
10 (0.9)
Leaf litter
Dry karst
Wet volcanic
Mahogany
Mahogany
Alluvial
12.3
1.1 (0.4)
165
70
132 (39, 4)
101
20 (9)
Wet volcanic
Wet volcanic
Mahogany
Mahogany
Alluvial
7.4
7.3
12.6 (1.7, 5)
11.7
4.8 (0.4)
87
83
163 (39, 4)
130
50 (15)
9.4 (1.8, 5)
Total litter
in press) also suggests that organic matter and nutrient turnover in these alluvial
stands are high. This strategy of rapid nutrient flux contrasts with that of plantation
forests such as mahogany and Pinus caribaea (Table 3.10, Lugo 1992), which tend
to accumulate nutrients on sites.
In summary, our study shows that abandoned sugar cane plantations on alluvial
soils favor the rapid establishment and growth of forests of S. campanulata, an
introduced wind-dispersed species. In spite of centuries of cultivation, these sites
maintain high soil fertility (Tables 3.9 and 3.11) and support a rapid rate of growth
of S. campanulata. In 25–40 years, closed canopy stands develop a complex forest structure characterized by high basal area (Table 3.5), but low species richness
(Tables 3.2 and 3.3 and Fig. 3.2). Initially S. campanulata forms almost pure stands
(Fig. 3.3a), but these stands do not remain monodominant because other species
(both native and introduced) establish populations and contribute to the diversification of the stand (Fig. 3.3b), particularly the understory (Fig. 3.3c). Because alluvial
Table 3.11 Weighted nutrient concentration (mg/g) of leaf litter and soil of various forest types in
Puerto Rico. Data are derived from Tables 3.9 and 3.10. Empty cells mean there are no data
Forest type
N
P
K
Ca
Mg
Leaf litter
Mature dry karst
Mahogany plantation
Alluvial Spathodea campanulata
13.4
14.0
18.2
0.3
0.4
1.4
Mature dry karst
Dry karst farmlands
Alluvial Spathodea campanulata
52.7
68.5
15.6
2.8
0.2
4.6
1.4
2.9
1.5
56
32
3.4
3.6
27
86
68
5.1
6.3
52.8
Soil
2.3
4.9
5.3
90
O. Abelleira Martínez, A. E. Lugo
sites flood, it is possible that the species list in these locations will be shorter than
observed in other types of abandoned agricultural fields. For example, abandoned
coffee plantations, pastures on volcanic soils, or abandoned agriculture on dry or
moist karst, all result in new forest with higher species counts than S. campanulata
forest on alluvial soils (Aide et al. 2000, Lugo and Brandeis 2005). However, forests
following these other types of abandoned agriculture do not develop the structural
complexity that S. campanulata achieves on alluvial soils. High primary productivity and apparent high turnover of nutrient-rich litter suggest that S. campanulata
forests on alluvial soils function at rapid rates commensurate with the fertility of the
soil. Soil fertility and abundant water supplies implies that abandoned sugar cane
fields on alluvial soils can be quickly reforested regardless of the length of time
of cultivation. However, because of the long history of cultivation, the restoration
of a full complement of native species on these sites appears problematic. We saw
indications of a return of Cecropia schreberiana and we found two endemic and
20 native species in these forests, both positive developments, but we believe that
restoring the original forest will be retarded by disturbances that favor the presence
of introduced species such as S. campanulata.
Acknowledgment This study was conducted in collaboration with the University of Puerto Rico.
The study is part of a larger study of S. campanulata forests in Puerto Rico. We thank the following
people for making the study possible: Lucy Gaspar, Lydia Sáez, and Josefina Arce officials of
the ALACIMA program of the National Science Foundation; Iván Vicéns, Cariel Arocho, Palma
Ripole, Mark McGarrigal, Brett Wolfe, Helena Román, Arnaldo Torres, Dixon Irizarry, and Angelo
Rodríguez for their field assistance; Edwin López, Mary Jeane Sánchez and Olga Ramos for laboratory assistance; Frank Axelrod for taxonomic assistance; Carlos Rodríguez for GPS support;
and Lía Sánchez and Mildred Alayón for their help with the manuscript. The following colleagues
reviewed the manuscript: Julie Hernández, D.J. Lodge, E. Medina, R. Myster, and F.H. Wadsworth.
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Chapter 4
Soil Fungi and Macrofauna
in the Neotropics
Yiqing Li and Grizelle González
4.1 Introduction
Decomposition is a critical ecosystem function that decomposes dead organic materials, removes wastes, recycles nutrients and renews soil fertility. In natural ecosystems most nitrogen (N) and phosphorous (P) required for plant growth are supplied
through the decomposition of detritus, relying therefore on the activities of soil
microbes and macrofauna. Decomposition is a consequence of interacting physical and chemical processes occurring inside and outside of living organisms in an
ecosystem (Chapin III et al. 2002). The two major processes include fragmentation,
breaking larger pieces of organic mater into smaller ones, and chemical alteration,
the change of chemical elements of dead organic mater into others. Fragmentation is
largely a consequence of feeding activity of soil macrofauna and chemical alteration
is primarily associated with the activity of bacteria and fungi.
Land conversion from forests to pasture or agricultural land alters the belowground ecosystem and decomposition process, and depletion of soil organic C
(Doran and Zeiss 2000). Degradation processes such as losses in soil C, nutrient
depletion and reduced water holding capacity often lead to considerable change in
soil microbial community and macrofuana (Bever et al. 1996, Picone 2000, Thomas
et al. 2004). In the Neotropics, land conversion from tropical forest to agricultural
land and then to pasture is the most common kind of land use practices (Fearnside
1993, Thomlinson 1996, Cleveland et al. 2003). Studies of the dynamics of microbes
and macrofauna following land conversion are thus particularly of importance for
establishing effective strategies of post-agriculture recovery and forest ecosystem
restoration because soil microbial community and soil macrofauna play a key role
in successional pathways.
4.2 Fungi
Together with bacteria, fungi are the main initial decomposers of dead plant and
animal materials and account for 80–95% of the total decomposer biomass and heterotrophic soil respiration (Macfadyen 1963, Coleman 1985, Chapin III et al. 2002).
R. W. Myster, Post-Agricultural Succession in the Neotropics.
C Springer 2008
93
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Y. Li, G. González
Fungi alone constitute a major portion of the soil biomass, with fungal biomass carbon (C) equal to or exceeding root biomass C (Paul and Clark 1996). Their biomass
in soils ranges from 1,000 to 15,000 kg/ha in the top 15 cm (Brady and Weil 2002).
Typically, fungi produce slender filaments or hyphae with a dense concentration of cytoplasm, either branched or un-branched, and commonly multinucleate.
The hyphae constitute the soma, mycelium and thallus and contain more vacuoles
when resources are scare, which enables fungi to grow into new areas to explore
for substrate. Fungi secrete enzymes capable of breaking down complex cell wall
components, such as lignin, enabling them to penetrate the cuticle of dead leaves
or the exterior of roots to gain access to more labile compounds that occur within
the cell. Reproduction in fungi is sexual, asexual or both. Asexual spores are produced at hyphal tips in structures and sexual spores are produced following hyphal
conjugations. In asexual reproduction, the offspring are genetically identical to the
“parent” organism (clones). During sexual reproduction, a mixing of genetic material occurs so that the offspring exhibit traits of both parents. Many species can use
both strategies at different times or different conditions, while others are apparently
strictly sexual or strictly asexual. As heterotrophs, fungi depend on leaving or dead
organic materials for their energy and C sources.
As major decomposers of organic materials in soil, fungi comprise an extremely
diverse group of microorganisms. Thousands of species have been identified in
soils, representing some 170 genera, and at least 1 million of fungal species in soils
are waiting for discovery (Brady and Weil 2002). Two groups of fungi common to
soils are the single-celled organisms called yeasts and the multicellular filamentous
organisms, such as molds, mildews, smuts, and rusts. Generally, filamentous fungi
are abundant in well-aerated environments, whereas yeasts inhabit in poor-aerated
soils. Most filamentous fungi lack a capacity for anaerobic metabolism and therefore
often absent from anaerobic soils. Mushrooms are the true reproductive structures
of filamentous fungi. The aboveground fruiting structure of most mushrooms is
only a small part of the total body. An extensive hyphae network of mushrooms
permeates below ground. Mushroom fungi are important wood rots that attack
standing and felled timber, stumps, slash and wooden structures, given appropriate
moisture contents. Molds are microscopic and semi-macroscopic fungi that play a
much more important role in soil organic matter decomposition than the mushroom
fungi. The four most common genera of molds in soils are Penicillium, Mucor,
Fusarium and Aspergillus.
In many terrestrial ecosystems, fungi are the most abundant decomposers which
are more effective than bacteria. Fungi are better equipped for carrying out the
decaying of the dead organic materials since their hyphae networks can penetrate
new substrates and proliferate both within and between dead plant cells. Fungi
also have an ability to transport nutrients through their hyphal networks to zones
of exploitation. Fungi have enzyme systems capable of breaking down all types of
plant compounds. All these advantages of fungi over bacteria made fungi dominant
in most ecosystems. Up to 50% of the substances decomposed by fungi are used
for fungal tissue, compared to around 20% for bacteria (Brady and Weil, 2002).
Fungi continue to decompose various complex organic materials which bacteria and
actimycetes cannot decompose.
4 Soil Fungi and Macrofauna in the Neotropics
95
Fungi play a critical role in the mineralization of nutrients. Fungi, together with
bacteria and other microbes, break down soluble and insoluble organic matter and
covert it into inorganic forms which can used by plants. The majority of N is contained in dead organic matter which depends upon the decomposition by microbes
for the N availability of plants. Nitrogen fixation by microbes is another route
through which plants obtain N. These microbes form nodules with the plant root
where they gain carbohydrate from the plants and in turn they provide the plants
with amino acids. Although bacteria play a key role in N fixation through symbiotic
association with plants (legumes), the primary route by which N enters this soil is
through the break down of legume leaf litter inputs that are enriched with N (Read
and Perez-Moreno 2003, Bardgett 2005). The breaking down of these leaf litters
are mainly completed by fungi. Phosphorous availability for plants is determined
by geochemical process rather than biological process, soil microbes however are
actively involved in the cycling of P (Sanders and Tinker 1971, Dodd et al. 1987,
Jayachandran et al. 1992, Salas et al. 2003, Joner and Jakobsen 2005).
4.3 Soil Macrofauna
Soil macrofauna play a very important role in improving structure, organic matter
content and distribution patterns of nutrient elements. Soil macrofauna may affect
soil function and processes in a variety of ways, and could be used as indicators of
nutrient status of soils (Vanlauwe et al. 1997, Rao et al. 1998, Martius et al. 2001).
Members of the macro-fauna group of soil animals include: arthropods (isopods,
amphipods, millipedes, termites), annelidans (earthworms) and molluscs (slugs and
snails). These animals are mostly responsible for the initial shredding of plant
remains and its redistribution within decomposer habitats (Swift et al. 1979) as
their body sizes are large enough to disrupt the structure of mineral and organic
soil horizons through their feeding and burrowing activities (Anderson 1988). The
macro-fauna have an important role in the maintenance of soil structural stability
and fertility in many natural and human-modified habitats (Gillison et al. 2003).
Not surprisingly then, the soil macro-fauna are considered important ecosystem
“engineers”; particularly in tropical ecosystems where their bioturbation actions
are major determinant of soil processes (Lavelle 1997), climatic conditions rarely
limit their activities, and their feeding strategies produce biogenic structures that can
determine the activities of microorganisms and other smaller invertebrates (Lavelle
et al. 2001).
The function of soil decomposer organisms may be categorized by body size, or
by various physiological aspects of trophic function (Swift et al. 1979). In terms
of body size, soil fauna were traditionally subdivided on the basis of body length
(Wallwork 1970). However, the use of meshed litterbags to separate the effects of
particular groups of the fauna on litter decomposition led to the classification of
soil fauna based on their body diameter classes (Swift et al. 1979). Thus, soil invertebrates can be classified by body width into micro- (< 100 μm), meso- (< 2 mm)
and macro-fauna (> 2 mm). This latter classification being more of a functional
96
Y. Li, G. González
classification with respect to litter breakdown and decomposition processes
(Fig. 4.1, Swift et al. 1979). These size categories are somewhat arbitrary because
juveniles of some species may be classified differently as adults, and also ignores
the range of trophic diversity of animals in the soil (Anderson 1988). Yet, body
width rather than body length broadly defines the extent to which the feeding and
burrowing activities of soil animals are constrained by, or modify, the structure of
soil and litter habitats (Anderson 1988).
The size of an organism is also an important determinant of its distribution (Swift
et al. 1979). It has been argued that micro-fauna species tend to have large geographic distribution patterns as a result of the size and ubiquity of the microsites
they occupy, whereas the macro-fauna show a greater discontinuity in distribution
(Fig. 4.2, Anderson 1977). The microorganisms and the micro-fauna may be widely
dispersed by abiotic and biotic agents, whereas most of the larger sapotrophic animals are sedentary and rarely dispersed except by their locomotive efforts (Swift
et al. 1979). The relative contribution of different functional groups of the soil fauna
to soil / ecosystem processes is dependant on the confounding effects of their size,
abundance, diversity and functionality (e.g., Hansen 1999, Heneghan et al. 1999,
Irmler 2000, González and Seastedt 2001) but also by the landscape in which they
Mesofauna
100 um
Microflora and microfauna
Macro and megafauna
2 mm
20 mm
Bacteria
Fungi
Nematoda
Protozoa
Rotifera
Acari
Collembola
Protura
Diplura
Symphyla
Enchytraeidae
Chelonethi
Isoptera
Opiliones
Isopoda
Amphipoda
Chilopoda
Diplopoda
Megadrili (earthworms)
Coleoptera
Araneida
Mollusca
1
2
4
8
16
32
um
64
128
256 512
1024
2
4
8
16
32
64
mm
Fig. 4.1 Size classification of organisms in decomposer food webs by body width (from Swift
et al. 1979)
4 Soil Fungi and Macrofauna in the Neotropics
97
Macrofauna
Size of organisms
Greater discontinuity
in distribution
Mesofauna
Co
sm
op
ol
ita
n
Microfauna
Microflora
Greater continuity
in distribution
Order
Family
Genus
Species
Size of taxa
Fig. 4.2 Biogeographical distribution of soil animals in relation to body size and taxonomic level
(modified from Anderson 1977)
occur (i.e., the degree of fragmentation, state of plant community succession and /
or habitat alteration) as it can affect their distribution patterns.
Conversely, soil organisms can have substantial ecosystem effects on natural and
human-altered habitats. This is important to consider given that (1) more than 95%
of the earth’s land surface is affected by human activities, including agriculture and
urbanization (Moguel and Toledo 1999) and (2) the majority of animals in terrestrial
habitats are invertebrate members of the decomposer community (Hansen 2000).
Thus, soil macro-fauna might be more prone than smaller invertebrates to show
an even greater discontinuity in distribution and more patchy effect on ecosystem
processes in the context of habitat alteration and as compared to their expected
distribution given size. In addition, land use intensification and fragmentation in
the Neotropics can significantly modify the relative contribution of native and
exotic elements within the macro-fauna; further challenging the incorporation of
soil organismal effects on ecosystem processes into sound management plans and
practices.
Ecosystem engineers—like the macro-fauna—are less able to withstand high levels of natural and anthropogenic environmental constraints and lower order groups
may become predominant with significant differences in the function of the ecosystem (Lavelle 1997). Consequently, modifications of soil fauna communities may
lead to loss of diversity and soil functions (Lavelle 1997). In the following section
of this chapter, we will discuss the effects that soil macro-fauna have on soils and
ecosystem functioning under different post agricultural management practices; and
highlight gaps in knowledge that might stimulate further studying of soil macrofauna effects in the Neotropics.
98
Y. Li, G. González
4.4 Discussion
4.4.1 Fungi
4.4.1.1 General Patterns in Tropical Forests and Pastures
Since the Industrial Revolution, first temperate and then tropical land-use change has
significantly affected biogeochemical and biophysical processes of various ecosystems at multiple scales. Typically, when tropical forests are cut down the land is
converted for cattle ranching or agricultural use. Human alteration of natural forest
ecosystems for pastures is an important component of tropical land use change, and
has the potential to alter soil nutrient and C cycles (Buschbacher et al. 1988, Neill
et al. 1997, Holl 1999, Nepstad et al. 1999, Post and Kwon 2000, Amundson 2001,
Townsend et al. 2002, Myster 2004). The nature and extent of such changes of nutrient and C cycles vary widely across gradients in climate, soil type and management
strategies (Spaans et al. 1989, Groffman et al. 2001). Although the changes of biogeochemical patterns following land conversion are becoming well documented, the
major mechanisms driving the variation of such changes are still poorly understood.
Most previous studies have focused on environmental or plant-driven factors over
biogeochemical changes following deforestation. However, soil microbial community, one of the possible mechanical controls over the biogeochemical changes, has
been less documented. Land conversion disrupts the structure and activity of soil
microbial community (Borneman and Triplett 1997, Nüsslein and Tiedje 1999) and
microbes in turn affect the biogeochemical variables. Fungi play a significant role in
regulating populations of other soil organisms and ecosystem processes (Fitter and
Garbaye 1994). Thus, it is essential to understanding the relationship between patterns and activities of microbial community and biogeochemical changes following
land conversion in the tropical ecosystems.
A main component of the tropical soil microbial community is fungi. Fungi
are primarily responsible for mineral nutrient recycling through decomposition
of organic matter and the transfer of nutrients into plants by mycorrhizal fungi
(Lodge 1993, Myster 2006). The conversion of forest to pasture causes significant changes to soil environment, leading to changes in fungi community
(Nüsslein and Tiedje 1999, Cleveland et al. 2003). Studies suggested the shift
in overall microbial composition could be up to 50%. The change of fungal
diversity following land conversion is closely related to initial fungal species
in different regions even with similar climates. For example, the Neotropics
have a very high fungal diversity comparing to African and Asia in the tropics (Lodge and Cantrell 1995). In addition, microbial biomass C also declines
when from forest to pasture. For example, in an oxisol pasture in Costa Rica,
microbial biomass (main component being fungi) is 40–60% lower than values
in the forest (Cleveland et al. 2003). The decrease in microbial biomass may
not entirely driven by soil organic matter because the decrease of soil organic
matter in oxisol pasture was around 20% lower in pasture than in forest. Generally, the impacts of land use change on soil fungi community is lager in
4 Soil Fungi and Macrofauna in the Neotropics
99
nutrient-poor sites and relatively modest in more fertile sites (Luizao et al. 1992,
Henrot and Robertson 1994). Groffman (2001) found that land conversion did
not decrease microbial biomass and activity in a relatively nutrient-rich tropical
ecosystem. Fungal biomass and total biomass of all microorganisms contain
significant fractions of labile nutrients in forest floor of tropical ecosystems.
During the conversion from forest to pasture most of these microbial related
labile nutrients are lost. Based on litter and basidiomycete nutrient concentration, 22% of litter P and 4% of K could have been immobilized in fungal
biomass (Lodge 1993). Phosphorus immobilization by fungi in the Luquillo
Experimental Forest is considerable because leaf decomposer fungi maintained
P concentration (Lodge 1987). In pastures, the total amount of immobilized P
in litter significantly decreased because of less fungal biomass in the pasture
litters.
Mycorrhizal associations have been intensively studied over the past several
decades and increased understanding of the important role of this symbiosis in the
functioning and performance of plants in tropical ecosystems. Besides mycorrhizal
fungi providing phosphorus from host plants and reciprocal C provision from host
to fungus, additional effects of mycorrhizal fungi on the functioning of their host
plants include increased disease resistance, improved water relations, alternation in
other soil properties, and acquisition of soil nutrients. Mycorrhizal fungi are of particularly importance in the tropical ecosystems because of their widely distribution
with a higher proportion of total fungal biomass. In nutrient-poor soils of humid
tropics, a number of late-successional woody species are obligately dependent on
arbuscular mycorrhizae (Picone 2000). Previous studies have suggested that plant
succession, restoration or reforestation could be inhibited by either the abundance
or the diversity of arbuscular mycorrhizal fungi in old pastures or agricultural lands
(Jenos 1988, van der Heijden et al. 1998). Because disturbances to the host plant
community alter the mycorrhizal fungi community (Bever et al. 1996) and host
plant diversity is much lower in pastures than in forests, mycorrhizal fungal diversity
dramatically declines. Thus, unlike in forests, pastures are usually dominated by a
few mycorrhizal species with each species produces more spores than the species
in forests. Several hypotheses have been suggested to explain why most species of
arbuscular mycorrhizal fungi produce more spores in pasture than in native forests
(Picone 2000). For example, it is suggested by these hypotheses that higher fine root
density and turnover rate, higher pH value, and more host specific sporulation in pasture favor producing more spores. Species richness of arbuscular mycorrhizal fungi
also decreases in agroecosystems comparing to natural ecosystems. Frequently irrigated and fertilized agricultural lands reduce more arbuscular mycorrhizal fungi
than un-intensive managed agricultural lands since high nutrient inputs diminish the
soil colonization potential and the dependency of plants on mycorrhization (Höflich
and Metz 1997). Phosphorous limitation to fungal processes has been demonstrated
in forest ecosystems and such constraints are also present in the pastures. However,
since conversion from forest to pasture dramatically reduces the abundance of spores
and the diversity of arbuscular mycorrhizal fungi, forests tend to conserve limiting
nutrients, such as P, more tightly than pastures because of the nutrient immobilization by fungi which maintain limiting nutrient concentration.
100
Y. Li, G. González
4.4.1.2 A Case Study: Fungal Biomass Dynamics Following Reforestation
in Puerto Rico
Fungal biomass contains significant fractions of labile nutrients in the forest floors
and surface soils of some tropical ecosystems (Yang and Insam 1991, Li et al.
2005). Nutrients can be immobilized and conserved by fungi and other microbes
in their biomass during periods of high precipitation that cause leaching (Behera
et al. 1991, Yang and Insam 1991). Therefore, fungal biomass and total biomass
of all microorganisms are important indicators for the nutrient dynamics and availability. However, relatively few publications contain measurement of fungal or total
microbial biomass in tropical primary forests and other forest ecosystems following
reforestation, such secondary forests and plantations (Frankland et al. 1990, Korf
1997, Tufekcioglu et al. 2001, Myster and Schaefer 2003, Li et al. 2004). Secondary
forests account for 40% of the total area of tropical forests and this percentage is still
increasing throughout the tropics (Brown and Lugo 1990). The success of management of tropical forests in the future might well depend upon the adequacy of our
ecological understanding of secondary forests (Cheng 1993). Tree plantations, covering 11 million ha of the tropics (Lanly 1982), are rapidly increasing in area in the
tropics as a reforestation approach. Reforestation through plantation on abandoned
and degraded agricultural lands in the tropics has been proposed as an effective C
management approach (Montagnini and Porras 1998).
In this case study, total fungal biomass and active fungal biomass were measured in surface soil in subtropical wet forest zone at Guzmán site in Puerto Rico.
Using a 20-year old pine plantation and a secondary forest originating from the
same abandoned farmland in the wet tropics in Puerto Rico, we examined microbial
biomass after 7-year litter exclusion, root exclusion, and both litter and root exclusion treatments in the plantation and the secondary forest. The study was conducted
on two sites that were within 100 m of each other. One site was in a Pinus caribaea
(Morelet)-dominated plantation and the other was in a secondary forest. Both sites
were located in the Guzmán sector of the Luquillo Experimental Forest in northeastern Puerto Rico (18◦18‘N, 65◦ 50‘W). The plantation and the secondary forests
originated from the same abandoned agricultural land with the same cropping system and management/disturbance history (Lugo 1992). The sites were characterized
by a wet tropical climate with mean annual precipitation of 3,920 mm and mean
annual air temperature of 22.3◦C. The temperature was mild and stable with diurnal
and seasonal temperature ranges of 3–4◦C. Soils were classified as mixed isothermic
tropohumult in both the plantation and the secondary forest. The sites were relatively
flat with a slope of <5 degrees and an elevation of about 400 m above sea level.
The tree plantation was established on the abandoned cropland in 1976 as part
of a reforestation program of the United States Forest Service (Lugo 1992). The
secondary forest had naturally developed also on the same abandoned cropland
since 1976. The plantation was dominated by Pinus caribaea (Morelet) with small
trees and grass species underneath the canopy. When our study started in 1996, the
average tree height was about 15 m and the average diameter at breast height (DBH)
was 22 cm in the plantation. The secondary forest was characterized by a sparse
4 Soil Fungi and Macrofauna in the Neotropics
101
overstory and a dense understory with abundant shrubs and grasses comparing to the
plantation but there were still a lot of spaces among the vegetations. The dominant
canopy species in the secondary forest include Myrcia splendens (Sw.), Miconia
prasina (Sw.) and Casearia arborea (L C Richard). The major understory species
include Casearia sylvestris (Sw.), Miconia mirabilis (Aubl.) and Tabebuia heterophylla (DC.). Total microbial biomass was measured using a fumigation-incubation
procedure (Jenkinson and Powlson 1975) in August 1996 and March 1997, which
represented a wet and a dry season, respectively. Biomass of active and total fungi
was estimated using the agar film techniques (Lodge and Ingham 1991). Biomass of
active and total bacteria was obtained using fluorescein isothiocyanate techniques
(Zou and Bashkin 1998).
In this study, we found total soil microbial biomass in the litter-and-root exclusion plots decreased the most among all the treatments in both forests. Specifically,
total fungal biomass was reduced 84.6 and 87.5% and total bacterial biomass was
reduced 62.1 and 56.9%, respectively, in the plantation and the secondary forest.
We found that microbial biomass demonstrated considerable seasonal variation in
both the plantation and the secondary forest. Total microbial biomass, in general,
was greater in the wet season than in the dry season with exceptions for the root
exclusion and the litter-and-root exclusion plots in the secondary forest. Total and
active fungal biomass was significantly (P ≤ 0.05) higher in the plantation than in
the secondary forest in all treatments during both the wet and dry seasons (Figs 4.3
and 4.4). Total and active fungal biomass was significantly greater in the wet season
than in the dry season in both forests.
Microbial biomass was also significantly different among different treatments.
The litter exclusion and the litter-and-root exclusion significantly reduced total fungal, total bacterial, active fungal, active bacterial biomass except for the active bacteria in the secondary forest during the wet season. In most plots, the litter-and-root
exclusion reduced microbial biomass the most in comparison with the control and
the other treatments. While root exclusion significantly reduced total fungal biomass
in both forests (Fig. 4.3 a, b). Root exclusion plots, in general, had the least reduction
in fungal biomass among all the treatment plots.
We also measured soil CO2 efflux in the same site in another study (Li et al.
2005) and found microbial activity in the plantation was more sensitive to the litter
input than the secondary forest. Soil CO2 efflux had higher correlation with total
fungal biomass but lower correlation with total bacterial biomass in the plantation
than in the secondary forest, suggesting that the fungal communities may dominate
soil heterotrophic respiration in the plantation, while bacterial communities may
take the control in the secondary forest. By separating the total fungal biomass into
active and inactive components, we found that active fungal biomass demonstrated
higher correlations with soil CO2 efflux than their corresponding total fungal and
bacterial biomass.
In conclusion, fungal biomass and total microbial biomass are important variables in the reforestation processes. Although the plantations and the secondary
forests originated from the same abandoned farm lands but they have different proportions of fungal communities and therefore might have different nutrient cycle
102
Y. Li, G. González
1000
A
Control
L-exc
R-exc
L&R-exc
800
Total fungi biomass (mg kg-1 soil)
600
400
200
0
1000
B
Control
L-exc
R-exc
L&R-exc
800
600
400
200
0
Wet
(mg C kg−1
Dry
Fig. 4.3 Active fungal biomass
soil) in the plantation (A) and the secondary forest (B)
and active bacterial biomass (mg C kg−1 soil) in the plantation (C) and the secondary forest (D)
in the wet (August) and dry (March) seasons. Common letters on the bars indicate no significant
difference between the treatments at 95% confidence level
4 Soil Fungi and Macrofauna in the Neotropics
103
120
100
A
Control
L-exc
R-exc
L&R-exc
Active fungi biomass (mg kg-1 soil)
80
60
40
20
0
70
60
Control
L-exc
R-exc
L&R-exc
B
50
40
30
20
10
0
Wet
Dry
Fig. 4.4 Total fungal biomass (mg C kg−1 soil) in the plantation (A) and the secondary forest (B)
and total bacterial biomass (mg C kg−1 soil) in the plantation (C) and the secondary forest (D)
in the wet (August) and dry (March) seasons. Common letters on the bars indicate no significant
difference between the treatments 95% confidence level
104
Y. Li, G. González
pathways. In addition, separation of active fungal biomass and total fungal biomass
may enhance our understanding on the interactions between microbial communities
and soil nutrient and organic C cycles.
4.4.2 Macrofauna
4.4.2.1 General Patterns of Macro-Fauna Composition Under Natural
and Post-Agricultural Management Practices in the Neotropics
Earthworms, termites, ants, coleopterans, and myriapods constitute the largest component of the macro-fauna. Lavelle and Fragoso (1999) compared 12 communities
from tropical rain forests, and estimated that earthworms and termites are the most
important macro-fauna in these soils. Lavelle et al. (1994) analyzed the composition
of 73 macro-invertebrate communities from 29 sites in the humid tropics and found
termites, earthworms and litter arthropods (mostly millipedes and coleopterans) to
account for the dominant proportion of the biomass. Further, Lavelle et al. (1997)
argued that earthworms and termites are the most important engineers in terrestrial ecosystems based on the influence they exert on the diversity and activity of
biota in subordinate trophic levels and the creation of organo-mineral structures
that can affect soil physical and chemical properties. In general, earthworms are
best represented in grasslands in humid areas as their abundance decreases towards
humid forests and dry forests and grasslands (Lavelle et al. 1994). Termites are well
adapted to dry environments where earthworms are not found (Menaut et al. 1985).
Millipedes are important in litter dominated ecosystems; where their feeding activity is specialized on dead organic matter, or on saprophytic organism consumption
(Bertrand and Lumaret 1992).
Disturbances linked to land use practices seem to severely affect the species
richness and abundance of soil invertebrate communities world-wide (Lavelle et al.
1997). There are clear differences in soil macro-fauna between forested and grassland systems and other types of land use (Giller 1996). In the Neotropics, annual
crops on sites where the natural vegetation has been cleared have depleted macroinvertebrate communities in terms of abundance and biomass (Lavelle et al. 1994);
usually irrespective of the annual crop species but mostly associated to soil disturbance and the absence of a permanent soil cover (Barros et al. 2005. Termites
seem thrived better than earthworms in annual crops and fallows and as compared
to pastures, forests and agroforestry systems (e.g., Lavelle and Pashanasi 1989,
Barros et al. 2002). Also for example, Lavelle and Pashanasi (1989) described the
soil macrofauna communities of Yurimaguas (Peruvian Amazonia) and found the
primary forest to have diverse and abundant fauna with a population density of
4,303 ind. m−2 and biomass of 53.9 g. m−2 —2 to 3 times higher than in similar
environments in Mexico. Yet, clearing and cropping soon destroyed most of the
fauna where half of the taxonomic units disappeared (see Table 4.1, Lavelle and
Pashanasi 1989); where termites and ants become the dominant in terms of abundance. They also found pastures to have the highest biomass (83.3–159.2 g. m−2 )
4 Soil Fungi and Macrofauna in the Neotropics
105
but the lowest population densities (922–2,347 ind. m−2 ) and taxonomic richness
(20–27) due to the dominant development of populations of Pontoscolex corethrurus
(Oligochaeta: Glossoscolecidae) which represent 82.3–95% of the biomass (also
see Chapter 5 on more information in earthworm dominance in pastures). Further, secondary forests had a soil invertebrate community with population densities on the same order (4,032 ind. m−2 ) of the primary forest but biomass and
taxonomic units were lower in the secondary forest as compared to the latter (see
Table 4.1).
4.4.3 Crop Systems
Leaving soils to “rest” or “fallow” is a traditional management practice throughout
the tropics for restoration of soil fertility during cropping (Sánchez 1995). Successful restoration of soil fertility normally requires a long fallow period for sufficient regeneration of the native vegetation and establishment of tree species (Young
1997). In this context, Thomas et al. (2004) studied the biodiversity of soil macrofauna along a chronosequence of abandoned rice paddy fields and plots without
agricultural disturbance in Northern Argentina. They found that species diversity
varied strongly along the chronosequence; it was highest in the natural grassland,
lowest in the 2 years- fallow and increased progressively through the chronosequence, but was still lower in the fallow after 15 years of abandonment as compared
to the natural grassland. Yet while working on Andean agroecosystems in Colombia,
Barrios et al. (2005) found that improved fallows (as an alternative to the natural regeneration of the native flora) could help regenerate soil fertility in degraded
volcanic-ash soils via changes in the soil macrofauna diversity. Similarly, in the
Table 4.1 Taxonomic richness, mean abundance and biomass of macroinvertebrates in different
types of land-use at Yurimaguas, Peru. Modified from Lavelle and Pashanasi, 1989
Type of land use
Forests
Primary forest
Secondary forest (15 years)
Crops (1st crop)
High input maize
Low input rice
Cassava
Pastures
Improved (Bracharia-Desmodium)
Moist
Dry
Fallows
6-mo. Centrosema after pasture
6-mo. Kudzu after high input crop
Peach-palm + Kudzu cover
Taxonomic
units (no.)
Density
(ind. m−2 ) ± SE
Biomass
(g. m−2 ) ± SE
41
27
4,304 ± 933
4,099 ± 1,828
53.9 ± 8.54
24.1 ± 5.85
20
24
18
730 ± 221
3,683 ± 1,059
1,197 ± 283
3.1 ± 1.04
8.5 ± 2.75
7.6 ± 3.26
27
23
20
922 ± 81
1,768 ± 308
2,347 ± 453
159.2 ± 16.4
121.1 ± 20.8
82.3 ± 15.5
22
23
32
1,546 ± 253
2,214 ± 969
1,858 ± 380
111.9 ± 13.2
15.5 ± 6.2
93.9 ± 25.2
106
Y. Li, G. González
above mentioned study from the Peruvian Amazonia, Lavelle and Pashanasi (1989)
found that recent fallows (6 mo.) of Kudzu (Leguminosea) and Centrosema retained
characteristics of the system from which they were originally started; continued
with low taxonomic richness but then showed increases in the density of the macrofaunal population. Hence, they argued these fallows were on a slight development
towards the original forest situation (Lavelle and Pashanasi 1989). Meanwhile, they
found the association of palm-trees and Kudzu cover had macro-faunal elements of
the primary forest and new species introduced after disturbance and the land was
cleared. Thus, it seems that invertebrate communities are best conserved when the
derived system has a structure similar to that of the original system, such as pastures
planted in savanna areas and tree-based systems in forested areas (Decaëns et al.
1994, 2004, Fragoso et al. 1997, Barros 1999 and as presented by Barros et al.
2002; see Fig. 4.5). In Western Brazilian Amazonia (Barros et al. 2005) found that
macrofauna communities in agroforestry systems were similar to the undisturbed
Forest
Dramatic effects on:
Microclimate
OM inputs
Litter availability
Pasture
Decrease macrofaunal
diversity & biomass
Change sp composition
Pasture sustainability
+
Maintain macrofaunal
Diversity & composition
Increase biomass
Pasture
Savanna
Slights effects on
OM inputs
Litter availability:
Fig. 4.5 A model of the response of macroinvertebrate communities to pasture establishment and
the potential effects on agro-ecosystem sustainability (modified from Decaëns et al. 2004)
4 Soil Fungi and Macrofauna in the Neotropics
107
lowland rain forest, in spite of higher densities of social insects in the former. Yet
in terms of abundance, they showed that land use practices (pastures, fallows, agroforestry and annual crop) were able to sustain sizeable macro-faunal communities
similar to those of a disturbed forest.
4.4.4 Agroforestry
In general, pastures have been found to enhance earthworm abundance but reduce
the overall macro-fauna diversity, whereas tree-based systems seem to be the best for
conserving diversity and sustaining high levels of abundance (Mboukou-Kimbatsa
et al. 1998, Barros et al. 2002). Alteration of the land use history plays a major
role in determining the abundance and community structure of earthworms and
the establishment of exotic earthworms in areas previously inhabited by worms
(González et al. 2006). For example, in the tropics, conversion of forest to pastures
has been associated to significant decreases in soil macro-invertebrate diversity
(Lavelle and Pashanasi 1989) and the dominance of a few exotic earthworm species
that can persist along different stages of plant succession after disturbance (e.g., Zou
and González 1997, Sánchez-de León et al. 2003). Yet, considering the vast area
under pastures and the different edapho-climatic and management parameters in the
Neotropics, it is not surprising that the results on biological, physical and chemical
properties may sometimes appear contradictory (Brossard et al. 2004). For example,
Brown et al. (2004) studied soil macro-faunal communities in native and introduced
pastures in Southeastern Mexico and suggested that if nutrient-poor savannas are
converted to native pastures, these can maintain higher soil macrofauna populations
and earthworm diversity than introduced pastures. Yet, Jiménez et al. (1998) studied
earthworm communities in native savannas and man-made pastures of the Eastern
Plains of Colombia and found a positive response of earthworm communities
(higher density and biomass) to improved pastures; which is a type of land use being
increasingly adopted in most Neotropical savannas. Further, Benito et al. (2004)
compared macro-fauna communities from the native vegetation of a Brazilian
Cerrado with renewed pastures and showed that renewed pastures had less diversity,
population density and biomass of macro-fauna than the deforested Cerrado.
With the vast amount of abandoned tropical land due to non-sustainable farming practices, tropical tree plantations become an effective means in restoring soil
productivity and preserving ecosystem biodiversity (Zou and González 2001). Tree
planting accelerates the process of tree invasions and establishment (Myster 1993)
in abandoned agricultural fields and it has been shown to mitigate barriers to secondary succession by re-establishing nutrient dynamics and improving the microclimate for plant understory species (Hagger et al. 1997, Binkley and Resh 1998).
Forest management practices can influence earthworm communities through changing physical and chemical properties of soil, net primary productivity, or plant litter chemistry (Fig. 4.6, González et al. 1996, Zou and González 2001). Yet the
response of the macro-faunal community to the forest management practices might
differ depending among the various groups of fauna given their functionality. For
108
Y. Li, G. González
NATIVE FOREST
SECONDARY SUCCESSION
FARMLAND
Species selection
Site preparation
Fertilization
Harvesting
TREE PLANTATION
Single species
Mixed Species
Agroforestry
WATER
pH
SOIL PROPERTIES
LITTER INPUTS
ORGANIC MATTER
QUANTITTY
C/N, C/P
SECONDARY CHEMISTRY
OTHERS
MACROFAUNA ABUNDANCE
SPECIES COMPOSITION
Fig. 4.6 Forest management practices (species selection, site preparation, fertilization and harvesting) that affect earthworm abundance and community structure in tropical tree plantations. Native
forest refers to forests that are not disturbed by human activities. Farmland includes crop fields and
pastures (modified from González et al. 1996)
example, Warren and Zou (2002) studied soil macro-fauna and litter inputs in three
tropical plantations started from a disturbed site in Puerto Rico, and found that earthworm biomass did not differ among plantation treatments. However, they found the
millipede biomass and abundance were significantly higher in Leucaena than in
Casuarina and Eucalyptus plantations. Further, the millipede biomass was highly
correlated to the N concentration and the C/N ratio in the incompletely decomposed
organic layer (Oi ). Warren and Zou (2002) suggested that in this study, millipedes
may be influenced by litter chemistry at fairly small spatial scales, and earthworms
are probably more affected by higher order controls such as soil physical conditions.
In the Neotropics and the world in general, soil fauna are important to consider if we are to sustain and maintain, functioning and healthy soils. As forests
are converted to pastures, the effects of heavy machinery and trampling by cattle
as a result from techniques used for deforestation and pasture management can
have widespread and long lasting effects on the reduced abundance and diversity of
macro-faunal communities in the newly created pastures (Chauvel et al. 1999); further instigating a cascade of deleterious effects which are linked to ecosystem health
(see Chapter 5 for detailed examples on the effects of exotic earthworms on ecosystem processes). Land use alteration in the tropics has historically been dominated
by a shift from forest to agriculture, but there are also trends towards increasing
urbanization and reforestation. The role of exotic earthworms (the dominant group
4 Soil Fungi and Macrofauna in the Neotropics
109
in tropical pastures) includes influence on current pasture ecosystems, their potential
invasion into surrounding forests and consequent ecosystem effects, their influence
on regenerating secondary forests in abandoned pasture, and their overall effects on
tropical biodiversity. Many of these potential effects are yet unstudied and unknown
(González et al. 2006).
In conclusion, the study of fungi and macrofauna, the major players in the
decomposition processes of ecosystem functioning, is of particular importance since
knowing the relationship between patterns and activities of fungal and macrofauna
communities and biogeochemical changes following land conversion could greatly
improve our understanding of the post-agricultural recovery mechanisms. In this
chapter, we discussed general patterns of fungi and macrofauna in forests, pastures
and agroforestry systems. In general, fungal and total microbial biomass significantly reduced following the conversion from forests to pastures or agricultural
lands. In the case study of fungal biomass dynamics following reforestation in
Puerto Rico, we found the plantations and the secondary forests have different proportions of fungal communities although they were developed originally from the
same abandoned farm lands and therefore might have different nutrient cycle pathways. Soil macrofauna play a very important role in improving structure, organic
matter content and distribution patterns of nutrient elements. Soil macrofauna may
affect soil function and processes in a variety of ways, and could be used as indicators of nutrient status of soils. In general, earthworms are best represented in grasslands in humid areas as their abundance decreases towards humid forests and dry
forests and grasslands. Termites are well adapted to dry environments where earthworms are not found. Millipedes are important in litter dominated ecosystems where
their feeding activity is specialized on dead organic matter, or on saprophytic organism consumption. Fungal and macrofauna communities might be the most important
variables controlling the biogeochemical changes following land conversion. However, most previous studies have focused on environmental or plant-driven factors
over these biogeochemical changes. Future studies on post-agricultural recovery
mechanisms need to pay more attention to the forms of trophic interactions among
the functional groups, such as fungi and macrofauna, in soil.
Acknowledgment While writing this manuscript, Grizelle González received support of grant
DEB-0218039 from the National Science Foundation to the Institute of Tropical Ecosystem Studies, University of Puerto Rico, and the USDA Forest Service, International Institute of Tropical
Forestry as part of the Long-Term Ecological Research Program in the Luquillo Experimental
Forest. Additional support was provided by the Forest Service (U.S. Department of Agriculture),
and the University of Puerto Rico. María M. Rivera helped with graph reproduction. We are grateful
to Dr. William A. Gould for valuable comments on early versions of the manuscript.
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Chapter 5
Earthworms and Post-Agricultural Succession
Grizelle González, Ching-Yu Huang and Shu-Chun Chuang
5.1 Introduction
Earthworms are the best known and often the most important animals influencing
the functioning of soil ecosystems (Hendrix and Bohlen 2002). In mature forests of
the Neotropics, earthworms usually dominate the soil food web in terms of biomass
(e.g., Odum and Pigeon 1970, Lee 1985, Lavelle, Brussaard and Hendrix 1999).
Consequently, they have significant influence on soil structure, nutrient cycling and
crop productivity as their activities can increase soil porosity, stimulate microbial
activities, and accelerate litter decay and the release of nutrients into the soil (Lee
1985, Lavelle et al. 1999, González and Seastedt 2001, González 2002, Liu and Zou
2002).
Earthworms are classified into endogeic, anecic, and epigeic species to represent soil, soil and litter, and litter feeders, respectively (Bouché 1977). Earthworms
can alter soil physical properties and biogeochemical processes (e.g., Edwards and
Bohlen 1996) according to their functionality. Through borrowing activities, earthworm may alter soil structure and soil water infiltration rates, thus alter the environmental conditions for microbial growth and biogeochemical fluxes (González
et al. 2006). Earthworms also act as inoculators of soil microbes for the freshly
fallen plant materials, and consequently accelerate the decomposition of plant
materials (González 2002). Endogeic earthworms alter soil properties primarily
through changing soil physical and chemical properties and epigeic earthworms
mainly affect microbial inoculation and the associated decomposition processes.
Anecic earthworm influence soils through both borrowing and inoculation activities
(González et al. 2006).
Plant community succession alters the quantity and chemistry of organic inputs
to soils (Zou and González 1997). These differences in the organic input influence
the patterns of response of the soil biota as do tree species during the process of
change from pastures to post-agricultural forests. In addition, land use history and
invasions by earthworm exotic species can affect the relative abundance and species
composition of a local earthworm fauna distinctly from soil, climate, vegetation, and
topography (Hendrix and Bohlen 2002). Regularly in the Neotropics, cosmopolitan
earthworm species dominate the pastures at the expense of native ones (Table 5.1).
With the conversion of pastures to secondary forests, the richness of earthworm
R. W. Myster, Post-Agricultural Succession in the Neotropics.
C Springer 2008
115
Mexico
Tabasco
Abandoned plantation of rubber trees
Brazil
Manaus
Brachiaria decumbens pasture
Cynodon plectostachyus pasture
Successional forest
Secondary forest
Primary forest
Pattern
Location
None
None
None
Andiorrhinus sp. 1
An. amazonicus
An. venezuelanus tarumanis
Rhinodrilus contornus
R. priollii
R. brasiliensis
Tuiba dianaea
Data not available
Data not available
Native earthworm
P. corethrurus
P. corethrurus
P. corethrurus
None
None
Pontoscolex corethrurus
Dichogaster bolaui
Exotic earthworm
Table 5.1 Changes of native and exotic earthworm communities along post-agricultural successional stages in the Neotropics
(continued)
Geissen and Gúzman (2006)
Römbke et al. (1999)
Reference
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G. González et al.
Sabana, Luquillo Mt.
None
P. spiralis
E. montana
E. gatesi
None
None
E. gatesi
E. gatesi
Swietenia macrophylla plantation
Secondary forest
Forest
Active pasture
Grass-vine-fern
Shrub-small tree
None
Pinus caribaea plantation
None
Borgesia sedecimesetae
Estherella sp.
Onychochaeta borincana
Neotrigaster rufa
Trigaster longissimus
Secondary forest
Mature forest
Luquillo Mt.
None
Abandon pasture
Puerto Rico
Cayey Mt.
Native earthworm
Pattern
Table 5.1 (continued)
Location
P. corethrurus
P. corethrurus
P. corethrurus
A. rodericensis
P. corethrurus
A. rodericensis
P. corethrurus
A. rodericensis
P. corethrurus
P. corethrurus,
A. rodericensis
P. corethrurus
Amynthas gracilis
P. corethrurus
P. corethrurus
Exotic earthworm
Zou and González (1997)
González et al. (1996)
Sánchez-de León et al. (2003)
Reference
5 Earthworms and Post-Agricultural Succession
117
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G. González et al.
species increases due to the presence of native species (Zou and González 1997).
Yet, the exotic earthworms are still dominant in terms of biomass and density (Zou
and González 1997, Sánchez-de León et al. 2003). In Puerto Rico, González et al.
(1996) found that Pontoscolex corethrurus (an exotic earthworm) dominated both
plantations and secondary forests but native earthworms occurred only in secondary
forests; suggesting that naturally regenerated secondary forests are preferable to
plantations for maintaining a high level of earthworm density, biomass and native
species in post-agricultural forests. The restoration of secondary forests through natural regeneration on abandoned pastures can promote the recovery of both anecic
earthworms and native species (González et al. 1996, Sánchez-de León et al. 2003).
Within this context then, in this chapter, we discuss (1) factors that affect the composition of earthworms (soil physico-chemical properties, soil nutrient status, and
biotic interactions), (2) habitat disturbance and the invasions of exotic earthworm
species, and (3) the physiological mechanisms by which exotic earthworms can
successfully invade pastures and disturbed environments. Finally, we highlight gaps
in knowledge that might stimulate further studying of native and non native earthworms in natural, agricultural and post-agricultural sites in the Neotropics.
5.2 Discussion
5.2.1 Factors that Affect the Composition of Earthworms
Earthworm communities can be described in terms of numbers, biomass, composition, and diversity of species. These components of earthworm communities
are regulated by a combination of environmental factors and biotic interactions.
Beyond landscape and regional scales, phylogenetic constraints and geological history determine the distribution and composition of earthworm assemblages. In the
North Neotropical region, earthworm communities are mainly constituted of Megascolecidae (Acanthodrilinae and Dichogastrini), then followed by Ocnerodrilidae and
some Glossoscolecidae (Fragoso et al. 1995). Glossoscolecidae becomes dominant
at South America, especially in Brazil, Perú and Colombia (Fragoso et al. 1999a).
Earthworm biogeography is also related to past geological history. For example,
plate tectonic movements in Caribbean islands might partly explain the differences
in earthworm communities and abundance among the Greater Antilles, the Lesser
Antilles, and the Bahamas archipelago (Fragoso et al. 1995, James 2004). At the
local scale, earthworm community structure may be controlled by a variety of factors such as climate, vegetation, edaphic properties and the biotic interactions in the
soil. It has been suggested that these factors affect earthworm communities hierarchically; climate (air temperature) being the predominant factor explaining the
variation of earthworm communities, followed by soil nutrient status and seasonal
fluctuations of humidity (Fragoso et al. 1999a). However, the predominant factors
determining earthworm communities can vary depending on local habitat characteristics. Also, it is difficult to separate the independent effects of environmental
5 Earthworms and Post-Agricultural Succession
119
factors as they are often interrelated (e.g., soil temperature and moisture retention)
(Lee 1985). Biotic factors may become determinant for earthworm communities,
especially in areas invaded by exotic species. Predatory, competitive, or facilitative
interactions between native and exotic earthworm species may change the dynamics
and composition of the original earthworm community in invaded areas. The community structure we observe in the field is a result of the combined effects from the
aforementioned and other factors. We can divide the regulating factors of earthworm
communities into two main categories, abiotic and biotic; and the following sections
discuss the potential impacts of each on earthworm communities.
5.2.1.1 Abiotic Factors
Soil temperature and moisture. – Edaphic properties have strong controls on earthworm community structure (Fragoso and Lavelle 1992), primarily soil temperature
and moisture content, which govern basic requirements for maintaining earthworm
activities. Few field studies describe the preferred and/or optimum soil temperatures
in relation to earthworm populations due to the difficulty of standardization and
concurrently measuring soil temperatures and earthworm activities in natural settings. Generally, the temperature tolerance of earthworms varies depending on the
species, collecting site, and the observed conditions (laboratory vs field observation, and methods of experimentation employed) (Lee 1985). However, it has been
determined that the optimum temperature range for tropical earthworm growth and
activity is between 20–30◦C at the surface or sub-surface of the soil in the field (Lee
1985).
Among all environmental factors within the Neotropics, moisture seems to be
the principal regulator of earthworm dynamics. Soil moisture content also regulates soil water tension, gas transfer, and other soil fauna activities (i.e. microbial
populations), which indirectly affect earthworm communities (Araujo and LópezHernández 1999). In general, total earthworm numbers and biomass increase with
an increase in soil moisture content (Garnsey 1994, Araujo and López-Hernández
1999, Schmidt and Curry 2001). Fragoso and Lavelle (1992) found maximum earthworm density and biomass in wet tropical forests with 2,000–4,000 mm annual
rainfall. Beyond this range, the soils were either too wet or too dry for sustaining an optimum earthworm community. Nevertheless, once soil moisture content meets the basic requirement for earthworm activities, this factor may have
less influence on earthworm dynamics. In Puerto Rico, González et al. (1999)
found that soil water content, which ranged 40 to over 100%, in wet (tabonuco)
forests had no influence on earthworm communities and Pontoscolex corethrurus
densities.
Soil moisture content usually closely relates to soil temperature fluctuations. The
rise of soil temperature due to removal of litter floor or loss of canopy may also
result in an increase in soil evapotranspiration rates; which in turn decrease the
soil water content. Both soil moisture and temperature fluctuations cause spatial
and temporal (seasonal and annual) heterogeneity in earthworm communities at a
variety of ecosystems within the Neotropics. Araujo and López-Hernández (1999)
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found that soil temperature and moisture were positively correlated to earthworm
densities in a natural savanna of Venezuela. In the east Mexican tropics, soil temperature and moisture significantly affected the relative dominance of endogeic and
epigeic groups of earthworm communities in different types of forests (Fragoso et al.
1995). In the tropics, earthworms can either become inactive or migrate into deeper
soil when confronting extremely low soil moisture conditions like during droughts
(Lavelle 1988). In Colombian pastures and savannas, earthworms showed uniform
vertical distribution in the top 20 cm of the soil profile during the rainy season; while
during the dry season, most worms moved to deeper soils responding to the decline
of soil moisture at the top 20 cm of soil (Jiménez and Decaëns 2000). Seasonal or
frequent droughts may convert earthworm communities from mixed species with
different ecological strategies to the dominance of some specific species, which
are more likely to be opportunistic poly-humic species (Lavelle 1988). Earthworm
species richness is negatively correlated to the quiescent period of earthworm population in tropical regions (Lavelle 1988). However, other edaphic characteristics
and biotic regulating factors may become more predominant than soil temperature
and moisture when the climate factors remain relatively constant, such as in some
tropical wet and rain forests.
Soil pH and texture. – Soil properties may be useful predictive variables of earthworm communities, particularly soil pH and texture. Some studies have showed a
strong correlation between soil texture and earthworm communities in the Neotropics. Hubers et al. (2003) suggested that “severe” soil physicochemical conditions
in Nipe soils of Maricao State Forest, Puerto Rico, resulted in lower diversity and
abundance of earthworm communities as compared to Mexico, Venezuela, and other
sites in Puerto Rico. These factors included lower soil pH, lower organic matter,
and higher exchangeable iron (Fe), magnesium (Mg) and aluminum (Al) concentrations in Nipe soils. However, in the Luquillo Mountains of Puerto Rico, soil
pH alone could not explain the distribution and biomass of earthworm communities (González et al. 1996, González et al. 1999, González and Zou, 1999b). The
correlation between soil texture and earthworm communities should be explained
with caution, since the casting and burrowing activities of earthworms result in the
modification of soil texture (silt, clay, and sand contents), soil pH, and bulk density
(Lee 1985, Edwards 2004).
Soil nutrients. – Depending on the feeding strategies of earthworms (epigeic,
endogeic and anecic groups), earthworms can utilize litter, soil organic matter, plant
roots, root exudates and microbes as food resources. Litterfall input (quantity) and
quality are more important determinants of the presence and abundance of epigeic
and anecic than endogeic earthworm species because of direct consumption of
litter by the former groups. Fragoso and Lavelle (1992) showed that earthworm
biomass increased with an increase in litter quantity among 14 localities of tropical rain forests. Further, the reduction of litter inputs in most agroecosystems via
forest clearing and aboveground crop harvest, which is detrimental to epigeic and
anecic species, have resulted in the dominance of endogeic earthworm species in the
tropics (Fragoso et al. 1999a). Earthworm abundance is often positively correlated
to litter quality (such as litter nitrogen, or phosphorus) and is negatively related
to litter polyphenol content (Lee 1985). In the Luquillo Experimental Forest of
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Puerto Rico, González and Zou (1999) described a correlation between litter N and
phosphorus (P) contents and the biomass of anecic earthworms in both Dacryodes
and Heliconia tree communities. Fragoso and Lavelle (1987) found that litter quality
and soil organic matter explained the aggregated distribution of earthworm communities in a Mexican tropical rainforest.
Soil organic matter is considered an important food source for endogeic earthworm species (Lavelle 1988). Earthworms are always found in soils with higher soil
organic matter (Lee 1985, Hubers et al. 2003). Hubers et al. (2003) found the density
of P. corethrurus was higher in sites with soil organic matter above 6.5%. Fragoso
and Lavelle (1992) showed that the soil nutrients status, including N, calcium (Ca),
and Mg, accounts for 28% of the variation of earthworm communities, which is the
second most important factor after soil moisture (45%), in tropical rainforests of
Central America, South America, South Africa and Asia. The improved soil fertility status in the soil, either due to increased litterfall inputs and/or the applications
of organic and inorganic fertilizers, can enhance earthworm biomass, density and
diversity (Zou and González 1997, López-Hernández et al. 2004). Plant living and
dead roots, root exudates and associated microbes which form the “rhizosphere”,
provide alternative food resources for earthworms (Lee 1985, Brown et al. 2000).
Aggregated earthworm communities are usually found near the rhizosphere and
dominated by endogeic species, because of higher soil carbon input in this zone
as compared to the surrounding soil. For example, direct and indirect evidence
showed that endogeic earthworm, P. corethrurus, fed in and aggregated around the
rhizosphere (Brown et al. 2000, Sánchez-de León et al. 2003). P. corethrurus was
shown to assimilate root-derived carbon of maize plants and sugar cane by directly
observing the 13 C shift in worm tissues from an originally lower 13 C value (from C3
vegetation) to a higher 13 C content (from C4 plants) (Brown et al. 2000). Indirectly,
Sánchez-de León et al. (2003) suggested that grass roots can be an important food
resources for endogeic earthworms, P. corethrurus, because root biomass affected
the earthworm community and distribution in a chronosequence of pastures in the
Cayey Mountains of Puerto Rico. Moreover, Zou and González (1997) suggested
root quality as an important factor explaining the difference on earthworm communities between pastures and woody plantations in Puerto Rico. Still, few studies deal
with the potential effects that root exudates or microbial-earthworm interactions
in the rhizosphere might have on earthworm community patterns. More research
is needed in this area; as both microorganisms and root exudates are believed to
be favorable food for earthworms. Generally, energy (food quality) and nutrient
availability in the soil are good indicators of earthworm communities and their
distributions.
5.2.1.2 Biotic Factors
Decaëns and Rossi (2001) investigated the spatio-temporal structure of earthworm
communities and its relationship with soil heterogeneity in Colombian pastures; and
found that the spatial and temporal variations of soil heterogeneity at regional scales
could not completely explain the variability of earthworm distributions (Decaëns
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and Rossi 2001). Further, they argued that soil properties explained a small portion
of the spatial and temporal structure of earthworm distribution, suggesting that
earthworm communities themselves were at a non-equilibrium state, which was
possibly caused by biotic interaction, such as competition exclusion (Decaëns and
Rossi 2001) or predatory pressure.
Predatory pressure. – The earthworm tissue is enriched with proteins and thus
earthworms are favored food for many animal species, which include birds, mammals, amphibians, reptiles, beetles, flatworms and some other invertebrates (Macdonald 1983, Lee 1985, Judas 1989, Blackshaw 1995). Most studies on earthworm
predation focused either on the effects of predators on a single earthworm species
or on the predator behavior. Lee (1985) argued that few studies have gathered data
on amounts of earthworms captured or eaten by the predators, how predators affect
earthworm populations, or the significance of earthworms in the energy flow of the
food web. Yet, some studies had been done on the potential impacts of predatory
pressure (mostly birds) on the structure of earthworm communities in temperate
forests (Bengtson et al. 1976, Barnard and Thompson 1985, Judas 1989, Blackshaw
1995). For example, Bengtson et al. (1976) found twice the abundance and biomass
of earthworms in protected plots than in the plots exposed to predation of golden
plovers (Pluvialis apricaria) in an Icelandic hayfield. Judas (1989) found the average mortality of earthworms that were exposed to the predators (chilopod, Strigamia
acuminata) was two times higher (64%) than the control (chilopod-excluded; 31%)
in a 2-month field microcosm experiment. However, he did not find any significant
effects of predatory pressure from macro-predators (birds, carabids, rodents, and
shrews) on earthworm populations and vertical distribution in exclusion and control
plots in another 7.5 mo field exclusion experiment in a beech forest (Judas 1989).
The predation pressure may not always play a significant role on earthworm communities, and its inconsistent influence may depend on the seasons and certain situations (such as changes in vegetation and invasion of new predators). Shifts of vegetation types due to anthropogenic factors can result in the presence and/or absence
of litter floor and canopy, which in turn may change the intensity of predatory pressure on earthworm communities though predatory efficiency and/or as related to
different assemblages of earthworm predators. Blackshaw (1995) investigated the
population sizes of invasive predatory flatworm, Artioposthia triangulata, and its
impacts on earthworm communities in grasslands of Ireland. With the presence of
predatory flatworms, he found a reduction in earthworm species richness (Blackshaw 1995). However, potential relationships between earthworms and their predators due to vegetation changes have not been intensively explored. In the Neotropics,
many tropical pastures have been gradually converted to grasslands or secondary
forests following the abandonment of agricultural practices due to soil degradation
and changes in economic strategies (e.g., Brazil and Puerto Rico, see Buschbacher
1986, Helmer 2004). This post-agricultural succession provides great opportunities
to investigate the dynamics of earthworms and their potential predators along the
successional stages.
Intra- and inter-populations interactions. – Three ecological groups of earthworms (epigeic, endogeic, and anecic groups) represent different abilities to utilize
resources (niche partitioning), including space and nutrients (Bouché 1977). Epigeic
5 Earthworms and Post-Agricultural Succession
123
species are litter-dwelling and mostly feed on surface litter, while endogeic species
live in the soil and consume organic residues and soil. Anecic species, which live in
permanent vertical burrows, eat and bury surface litter (Lee 1985). When the niche
of different earthworm species overlap, significant competition would be expected
due to the use of similar resources. Since 1980s, several researchers have been
interested in the competitive interaction among earthworms. Competition intensities
are found to vary among different earthworm species from many laboratory studies
(Abbott 1980, Hamilton et al. 1988, Butt 1998, Dalby et al. 1998, Baker et al. 2002,
Lowe and Butt 2002). For example, Abbott (1980) showed persistent competition
for food between Eisenia foetida and Microscolex dubius, but not between E. foetida
and Allolobophora trapezoides. In a laboratory study by Butt (1998), the decrease
of growth rates and cocoon production of Allolobophora longa resulted from both
intra-specific competition and inter-specific competition with Lumbricus terrestris.
Winsome et al. (2006) described that Argilophilus marmoratus had lower growth
and reproduction rates in the presence of A. trapezoides in different habitats in a
California grassland. They attributed this to the competition for exploiting similar
food resources (particularly microbes) (Winsome et al. 2006). In contrast to the
abovementioned studies from temperate ecosystems, few studies have focused on
the competitive relationships between native and exotic earthworm species in the
tropics (Hendrix et al. 1999, Lachnicht et al. 2002). Hendrix et al. (1999) examined earthworm resource utilization under field conditions in three different forest
ecosystems (elfin forest, tabonuco forest, and a pasture) in Puerto Rico by using
the natural abundance of 13 C and 15 N. The differences in 15 N fractions within
native and exotic earthworm species when they co-exist in the soil indicated that
there might be resource competition between native and exotic species (Hendrix
et al. 1999). Also in Puerto Rico, Lachnicht et al. (2002) set up a 19-day incubation experiment to evaluate the interactive effects of P. corethrurus (an exotic
earthworm) with native Estherella sp. by using stable isotopes. P. corethrurus and
Estherella sp. were found to inhabit different soil layers when incubated together.
In addition, P. corethrurus assimilated different nitrogen resources in the presence of Estherella sp. (Lachnicht et al. 2002). In the field, habitat disturbance
and invasion of exotic earthworm species usually trigger or enhance competitive
exclusion within earthworm communities due to the shortage of food resources.
Nevertheless, it is still not clear how competitive exclusion and its intensity affect
the structure of earthworm communities in different ecosystems; especially in the
tropics.
Although environmental (abiotic) and biotic factors are determinants on earthworm communities, earthworm activities can reversibly affect their environments.
Earthworms, as ecological engineers, have significant impacts on the soil abiotic and
biotic environment and can regulate the availability of resources for other species
by directly or indirectly modifying their habitats (Lee 1985, Edwards 2004). Soil
moisture, soil texture, organic matter distribution, and soil nutrient cycling can be
modified by earthworm activities, e.g. casting and building burrows (Lee 1985,
Edwards 2004). Hence, unique earthworm assemblages are generally believed to
reflect not only their specific adaptations or preferences to the local surroundings,
but also the habitat they build.
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5.2.2 Habitat Disturbance and the Invasion of Exotic
Earthworm Species
Habitat disturbance can modify biotic and abiotic factors, which in turn can have
significant effects on the composition of earthworm communities as discussed earlier. Natural disturbances, such as flooding and landslides caused by hurricanes, can
result in temporal changes of earthworm communities through the alteration of litter
inputs and soil properties in localized areas. However, anthropogenic disturbances,
like forest clearing, human inhabitation and land-use management, can occur more
frequently than natural disturbances and have more permanent and devastating influences on earthworm communities. For instance, large areas of tropical moist forests
had been converted into savannas and pastures in Asia, Africa, South America,
and Central America (Buschbacher 1986). Tropical trees in the original forests of
Central America have been replaced with shrubs, annual/perennial crops or grasses
(Nations and Komer 1983). This transformation of vegetation types causes a reduction of litter quality/quantity, alteration of root distribution and availability, and
modification of soil properties. As a result, these abrupt fluctuations on soil physical/chemical conditions and food resource availability from habitat disturbance
can modify the original composition and distribution of earthworms taxonomically
and functionally (Abbott 1985, González et al. 1996, Sánchez-de León et al. 2003,
Sánchez-de León and Zou 2004). Often, there is a significant decline of earthworm
diversity, with endogeic earthworm species becoming dominant after vegetation
conversion from forests to agroecosystems in most tropical countries (Blanchart
and Julka 1997, Fragoso et al. 1999b, González et al. 2006). However, the alteration
of earthworm communities will vary with the nature of land-use management and
as it relates to the intensity and frequency of the disturbance (Abbott 1985, Fragoso
et al. 1999a). For example, the dominance of earthworm ecological groups can be
modified from the endogeic species in natural savannas (Glossodrilus n. sp.; 80%)
to anecic species (88%) in the man-made pastures in Eastern Colombia (Jiménez
et al. 1998).
Recently, it has been found that anthropogenic habitat disturbance can enhance
the introduction of exotic earthworm species (Fragoso and Lavelle 1992, González
et al. 2006). The invasion of exotic earthworm species can alter local native earthworm communities (Kalisz 1993, Kalisz and Wood 1995). The success for exotic
earthworms to invade novel areas or their potential impacts on native earthworm
communities depends greatly on population characteristics of both exotic and native
earthworms and the local environmental and soil properties. Hendrix et al. (2006)
illustrated the possible sequences of invasion depending on degree of habitat disturbance and invasion success by exotic earthworms invading ecosystems inhabited by
native earthworms (Figure 5.1). They hypothesized that the intensity of habitat disturbance (severe, moderate or minimal) would influence native earthworm communities in the natural ecosystems differentially. Exotic species can occupy disturbed
areas, exclusively or by coexisting with native earthworm species where competition
displacement could happen in the later case (Hendrix et al. 2006). Hendrix et al.
(2006) also proposed three main mechanisms (propagule pressure, habitat match
5 Earthworms and Post-Agricultural Succession
INITIAL
STATE
Disturbance
A
Severe
INTERMEDIATE
STATE
125
Exotic Invasion
Native Worms
Eliminated
Successful
CURRENT
STATE
Exotic Worms
Exclusively
Successful
'Pristine' System
(Native Worms
Exclusively)
B
Moderate
Native Worms
Diminished
1
2
Native Worms
Exclusively
1
2 Unsuccessful
Successful ?
Native and Exotic
Worms
Coexisting
Successful ?
C
Minimal
Native Worms
Exclusively
Fig. 5.1 Hypothesized sequences of invasion depending on degree of habitat disturbance and invasion success by exotic earthworms invading ecosystems inhabited by native earthworms (reproduced with kind permission of Springer Science and Business Media from Fig. 1 of Hendrix et al.
2006)
and biotic resistance) to determine invasion success or failure of exotic earthworm
species during the invasion processes. Propagule pressure relates to the magnitude
of the exotic earthworm introduction into the invaded soils, in terms of the frequency
of the introduction and size of the introduced population (Hendrix et al. 2006).
Repeated introductions and more individuals involved during each introduction can
enhance the success of exotic earthworm populations to establish, proliferate and
disperse. Hence, population characteristics (such as life history and reproduction
strategy) of the invasive species become determinant elements at this initial stage
of invasion. For example, high fecundity and parthenogenesis can boost certain
exotic earthworm species to build their populations in a relative short time (Hendrix
et al. 2006). Nevertheless, even with the advantage of the propagule pressure, exotic
earthworm species still have to adapt to the local environment that they attempt
to invade, which may encompass different plant communities, food resource availability, and soil properties. As noted earlier, these abiotic environmental factors are
basic requirements for earthworm populations and significantly affect earthworm
communities and distributions. Hendrix et al. (2006) called this “adaptation” step
to the invaded environment as the “habitat matching” mechanism, which is relevant
for distribution, growth and reproduction of exotic earthworms. The characteristics
of higher tolerance and flexible plasticity to the environment can make exotic earthworm species more competitive than other earthworm species, especially when they
invade into the disturbed areas. Superior adaptation ability to new environments and
reproduction biology (e.g., parthenogenesis and r-selected strategy) are generally
attributed to the success of exotic earthworm species, e.g. Pontoscolex corethrurus, a
pan-tropical distributed invasive earthworm species (Fragoso et al. 1999a, González
et al. 2006, Hendrix 2006). The third mechanism, “biotic resistance”, hypothesized
the potentially resistant forces from local native communities to impede the invasion of exotic earthworms. Predatory pressure, parasitism, and competition with
indigenous earthworm species for resources can retard the establishment of exotic
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G. González et al.
earthworms even when they are able to overcome the challenges of propagule pressure and habitat matching (Hendrix et al. 2006). Biotic resistance gives a reasonable explanation for the lack of invasion of exotic earthworms in some undisturbed
forests adjacent to areas they have invaded (Lavelle and Pashanasi 1989, Kalisz
1993). In addition, recognition systems on exotic earthworms by new predators and
parasites may not develop in a relative short time at the beginning of the invasion.
Competition relationships with local native earthworm communities may play a
more important role for invasion success of exotic earthworm at the early stage of
invasion. The resistance to the invasion from the native earthworm communities can
be weakened by habitat disturbance. The shortage of food availability and unsuitable
soil conditions from habitat disturbance are usually detrimental to the local native
earthworms. As a result, habitat disturbances devastate the integrity and equilibrium of local earthworm communities and can curtail their resistance strengths to
impede the invasion of exotic species. In disturbed areas, the mechanisms for exotic
earthworms to become the dominant species include not only the two advantages
discussed above (adaptation plasticity and reproduction biology), but also the lack of
competition stress from local native earthworm communities (Fragoso et al. 1999a,
González et al. 2006).
Consistent with Hendrix et al. (2006), González et al. (2006) showed different
pathways by which exotic earthworms can establish populations in natural and disturbed ecosystems (Figure 5.2). In natural ecosystems with no or minor disturbance,
Invasion of exotic
earthworms
Native ecosystem
Native species
only
Disturbed ecosystem
Habitat
disturbance
Secondary
succession
Competitive
exclusion of exotics
Natives only
Reduction of
native species
Reduced competition
from natives
Exotics and natives
co-exist
Exotics and natives
co-exist
Extirpation of
native species
No
competition
with natives
Exotics only
Natives only
Increase gradient of disturbance
Decrease in native species and increase in exotic species
Fig. 5.2 Model illustrating the paths by which invasion of exotic earthworms affect native earthworm species in undisturbed and disturbed ecosystems (reproduced with kind permission of
Springer Science and Business Media from Fig. 1 of González et al. 2006)
5 Earthworms and Post-Agricultural Succession
127
the integrity of local native earthworm communities gives few chances for exotic
earthworm species to invade. Once native earthworm communities are excluded
and/or disrupted by habitat disturbance in the disturbed areas, exotic earthworm
species can occupy the empty niches released by native earthworms to build their
populations. The intensity and frequency of disturbance will significantly affect the
completeness of native earthworm communities (biotic resistance), therefore the
invasion success for exotic earthworm species (González et al. 2006). In reality,
these three mechanisms (propagule pressure, habitat matching, and biotic resistance) correlate to one another when explaining the invasion success of exotic earthworms. More efforts are needed to test and clarify each of these mechanisms. The
relative importance of each mechanism may be diverse in the different ecosystems.
5.2.3 The Physiology of Exotic Earthworms: Mechanisms
for Successful Introductions in Pastures and Disturbed
Environments
The reproductive biology of exotic species is an important characteristic to deem
in the context of invasion (González et al. 2006). Tropical peregrine earthworms
(e.g., P. corethrurus, Perionex excavatus, Dichogaster modigliani and Polypheretima elongata) are often considered to be continuous breeders with high fecundity
(Bhattacharjee and Chaudhuri 2002). So that on the basis of response to selection
pressure, high fecundity, short incubation period with high hatching success are
probably adaptative strategies of r-selected organisms that enables them to survive drastic environmental changes, especially heat, drought and predation in the
soil (Pianka 1970, Bhattacharjee and Chaudhuri 2002). Up until now, endogeic
earthworms are more frequent invaders of disturbed tropical pastures than epigeic
species. Interestingly, exotic endogeic species (e.g., P. corethrurus, P. elongata and
Drawida nepalis) have been shown to increase their rate of cocoon production and
incubation period with increased temperature (Figure 5.3) while epigeics decreased
their reproductive capabilities (Bhattacharjee and Chaudhuri 2002). Thus, the interactions between the characteristics of the exotic earthworm species—particularly
their functionality and activities—could determine their potential for establishment
(González et al. 2006).
Animals have physiological adaptations to the environment, and earthworms
are no exception. Efficient adaptations to the environment by earthworms ensure
the survival of species, particularly that of non-natives. How can P. corethrurus
adapt and distribute widely in tropical areas around the world? The physiological
adaptations of this species may be very important to consider when answering
this question. For example, some earthworm species, such as Lumbricus terrestris, Amynthas gracilis, A. robustus, and Metaphire schmardae, crawl out
of the soil after a heavy rain exposing themselves to predation (Chuang et al.
2004, Darwin 1881, Tsai 1964). Other earthworm species do not. In tropical
and subtropical Taiwan, P. corethrurus is a common exotic worm; where it has
128
70
(a)
60
Incubation period (days)
Fig. 5.3 Relationship between temperature
and incubation period in different earthworm
species (a) Polypheretima elongata,
(b) Drawida nepalensis, (c) Pontoscolex
corethrurus , (d) Dichogaster modiglianii,
and (e) Peryonyx excavatus. Endogeic species
are represented with open symbols and
epigeic species are represented with filled
symbols (modified from Bhattacharjee and
Chaudhuri 2002)
G. González et al.
50
(b)
40
(c)
30
20
10
0
27
28
29
30
31
(d)
(e)
32
33
Temperature (ºC)
not been observed to crawl out of the soil after a heavy rain. To understand why
some earthworms crawl out of the soil, Chuang et al. (2004) and Chuang and Chen
(in press) investigated the respiration physiology of A. gracilis (a native earthworm
to Taiwan) and P. corethrurus by placing earthworms in open and sealed bottles
filled with air-saturated water. They found the mean survival time of A. gracilis in
water was 5.4 ± 1.34 h in a sealed bottle and 13.4 ± 6.95 h in an open bottle. A
control earthworm remained alive for more than 72 h (end of testing) in a humidity box. However, P. corethrurus survived five to ten times longer than A. gracilis
under the same conditions, surviving in water for 45.4 ± 3.58 h in a sealed bottle
and to the end of the test (76 h) in an open bottle and for at least 96 h (end of
test) in a humidity box (Figure 5.4). When A. gracilis was classified as dead, the
residual oxygen concentration in the sealed and open bottle was 1.87 or 1.5 μg/ml,
respectively. However, P. corethrurus could survive until the oxygen concentration
in the sealed and open bottles fell to 0.73 or 0.64 μg/ml, respectively (Figure 5.5).
Therefore, it is reasonable to infer that P. corethrurus might use anaerobic respiration in oxygen-shortage conditions. Amynthas gracilis was found to have a diurnal
rhythm of oxygen consumption, using more oxygen at night than during the day.
However, P. corethrurus lacked a diurnal rhythm (Figure 5.6). P. corethrurus consumed similar amounts of oxygen at different time periods and temperatures. In
addition, the cocoon and juveniles of P. corethrurus are easily found on the soil surface throughout the entire year unless the soil is extremely dry (Chuang et al. 2004).
If the soil humidity is too low, then P. corethrurus will go into torpor. They twist
their body into a coil and secrete mucus to maintain moisture (Chuang et al. 2004).
This phenomenon was also described for Glossoscolex paulistus (Abe 1985), and
they are both classified as Glossoscolecidae. Thus, P. corethrurus has lower oxygen
consumption, tolerates poor oxygen concentrations in the environment and adapts
to a wide range of soil temperatures. These physiological adaptations could well
5 Earthworms and Post-Agricultural Succession
129
Fig. 5.4 Survival times of submerged earthworms. Earthworms were kept in an open or sealed
bottle filled with artificial spring water (ASW) at 25◦ C. Different letters indicate significant differences. The data are the mean±SD (n = 20) for the survival time (Chuang et al. in press).
Fig. 5.5 Oxygen remaining in artificial spring water-containing bottle at 25◦ C. The value was measured immediately after the worm was found to be dead. Different letters for different treatments
indicate significant differences. The data are the mean±SD (n = 20) for the residual oxygen content
(Chuang et al. in press)
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G. González et al.
A. gracilis
Oxygen consumption (ug/g)
180
160
140
120
100
80
60
40
20
0
140
120
100
P. corethrurus
A) 20°C
*
*
B) 25°C
*
*
80
60
40
20
0
120
100
80
C) 30°C
*
*
*
60
40
20
0
4:00-6:00
11:00-13:00
19:00-21:00
Time (min)
Fig. 5.6 Total oxygen consumption of two species of earthworms at different temperatures. The
oxygen consumption of A. gracilis was significantly higher than P. corethrurus, irregardless of
being at the different temperature (P < 0.001). The asterisks indicate significant difference (from
Chuang et al. 2004)
explain why P. corethrurus is widely distributed in the tropics around the world.
On the other hand, the high oxygen demand of A. gracilis (especially at night),
causes it to expose its body on the soil surface to exchange for air, consequently
making them more susceptible to predation. This behavior is not observed in P.
corethrurus (as it is an endogeic species), which might be another explanation as to
why P. corethrurus can be a successful exotic earthworm (Chuang et al. 2004) in
pastures and disturbed sites with a non continuous canopy cover. Another characteristic that makes P. corethrurus a hard-hitting invader is the ability that the juveniles
5 Earthworms and Post-Agricultural Succession
131
have to enter diapause and regenerate after amputation independent of soil moisture
(Fragoso and Lozano 1992). This phenomenon and the fact that parthenogenesis
is common for most tropical exotics (Fragoso et al. 1999b) could well be adaptive
strategies of tropical exotic earthworms to avoid predation pressure in a climatically
harsh environment; making them without doubt strong invaders and competitors
(González et al. 2006).
In addition to the respiration adaptation to water-saturated environments,
P. corethrurus exhibits other traits that help them survive in severe environments.
P. corethrurus has high tolerance to ultraviolet radiation (UV) (Chuang et al. 2006).
Chuang et al. (2006) tested UV effects on the behavior and physiology of three
species of earthworms: A gracilis, M. posthuma and P. corethrurus. They found
that P. corethrurus would crawl slowly after UV (Figure 5.7), but surprisingly,
they would not die or have tissue damage even after 24 h of ultraviolet radiation
(Figure 5.8). The influence of UV to P. corethrurus may be temporary, as UV
influences their crawling activity, but these earthworms may have some mechanism
to protect themselves. It is known that the development of pigment is an important
(A)
Delaying crawling activity
3.5
*
UV-A
3.0
2.5
2.0
1.5
1.0
0.5
0.0
control
500
1000
1500
Dose (J/m2)
(B)
Delaying crawling activity
4.5
UV-B
4.0
*
3.5
*
3.0
2.5
2.0
1.5
1.0
0.5
0.0
control
500
1000
1500
Dose (J/m2)
Fig. 5.7 Relative crawling time of earthworms after exposure to either UV-A and UV-B. A) The
crawling time of Pontoscolex corethrurus exposed to 1,500·J/m−−2 of UV-A was significantly
increased compared to controls (ANOVA, N=6, *P<0.01) B) After UV-B exposure, P. corethrurus
show a significantly increased crawling time compared to controls (ANOVA, N=6, *P<0.01) (from
Chuang et al. 2006)
132
(A)
G. González et al.
(B)
Fig. 5.8 Cross section photographs of Pontoscolex corethrurus tissue A) with no UV radiation
(control) and B) as exposed to UV radiation at a dose of 1,500 J/m2 after 24 h. No changes in the
epidermis were observed after 2, 12 and 24 h (as shown in B) of UV radiation. Also note the many
granular cells in the epidermis by which P. corethrurus can secrete more mucus than A. gracilis or
M. posthuma (photographs taken by Shu-Chun Chuang)
adaptation against solar radiation. However, P. corethrurus does not show much
pigmentation in the epidermis. Yet, the mucus and epidermis of P. corethrurus is
thicker than other species of earthworms (Chuang, personal observation). In addition, P. corethrurus mucus may contain chemicals that help absorb UV and protect
their skin. Experimentation with earthworm mucus is problematical given difficulties in the extraction of mucus from the earthworm skin surface (Chuang, personal
communication). Accordingly, Chuang et al. (2006) proposed that P. corethrurus
might utilize substances such as mycosporine-like amino acid (MAA) or flavonol
quercetin, and/or some enzymes like photolyase to protect itself. Chuang et al.
(unpublished data, not shown), showed that P. corethrurus has a high activity of
Protein kinase before and after UV radiation. Protein kinase activity is related to the
release of acetylcholine, which is an important hormone in neuromuscular junction.
Earthworms can have better movement behavior with higher acetylcholine activity.
On the other hand, oxidative stress and lipid peroxidation after ultraviolet radiation
may be lethal for earthworms because the final product of lipid peroxidation is
lipid peroxide (LOOH). LOOH will lead to the accumulation of malondialdehye
(MDA) in the earthworm muscle/skin tissue. When MDA reacts with thiobarbituric
acid (TBA), it forms TBA-reactive substances (TBARS). Therefore, we can extract
MDA from earthworm skin/muscle tissue to react with TBA and test the level of
lipid peroxide. The higher TBARS means the higher LOOH. Therefore, TBARS
can be good indicators of lipid peroxidation. For example, when P. corethrurus is
exposed to 1,000 J/m2 UV-B, the concentration of TBARS increased after 2 hr,
however, at other doses of UV-B radiation, the concentration of TBARS did not
change; meaning that P. corethrurus did not have much lipid peroxidation after
UV-B exposure (Table 5.2, unpublished data).
Chuang et al. (unpublished data) also found that the anti-oxidant enzyme activity
of other earthworm species decreased after UV radiation, but in P. corethrurus,
it remained high after UV radiation (Table 5.2, unpublished data). Girotti (2001)
classified the photo-oxidative damage into four degrees and according to these
5 Earthworms and Post-Agricultural Succession
133
Table 5.2 MDA level and catalase activity in muscle/skin tissue (Chuang et al. unpublished data)
Dosage (J/m2 )
Control
500
1,000
1,500
3,000
Time (min)
TBARS (mmol/g)1
Catalase activity (mU/mg)2
15
120
240
15
120
240
15
120
240
15
120
240
15.029±1.91
17.57±1.38
15.07±3.92
14.88±2.24
20.12±2.85
23.45±1.22 (*3 )
14.53±1.01
22.12±3.97
23.31±3.58
17.74±2.68
16.06±4.27
21.40±2.52
14.78±1.12
4.67±0.94
5.62±0.36
5.52±0.62
5.04±0.52
5.65±0.38
5.07±1.13
4.94±0.24
5.95±0.27
4.79±0.87
5.30±0.37
5.69±0.38
5.53±0.38
4.87±0.21
1 The MDA level of muscle/skin tissue depended on minutes expose and doses of UV-B radiation.
The concentration of MDA was used to estimate the extent of lipid peroxidation.
2 Effect of UV-B radiation on catalase activity in muscle/skin tissue. The activity meant the ability
of anti-oxidative stress.
3 Asterisk (*) meant significant differences compare to control treatment (Two way ANOVA test,
p<0.05).
results from P. corethrurus; it should be classified into secondary degree. It means
that epidermal cells of P. corethrurus have been influenced slightly after ultraviolet
radiation, but the increase of lipid peroxidation also turned on the anti-oxidative
mechanism and further protected the epidermal cells. In the aforementioned study,
the anti-oxidative enzyme of P. corethrurus was also found to be kept at a high
activity after UV radiation and the reasons for this are unknown. Therefore,
P. corethrurus still had enough protection from UV and did not die. The oxidative
stress relates not only to the effects of ultraviolet radiation but also to other stresses
such as heavy metal pollution and an adversely compromised physiology. Therefore,
if P. corethrurus has high tolerance to oxidative stress, they would better survive
in a harsh environment. Certainly, we also argue that this mechanism can help
P. corethrurus invade successfully disturbed sites throughout in the tropics.
5.3 Summary
In the Neotropics, large areas of pasture have been gradually converted to grasslands or secondary forests following the abandonment of agricultural practices.
These areas covered with the secondary succession are expected to expand in the
near future because of the adjustment of economic strategies and soil degradation
problems in these tropical countries (e.g., Brazil and Puerto Rico, Buschbacher
1986, Birdsey and Weaver 1987, Helmer 2004). For now at least, these postagricultural succession processes boost the recovery of natural landscapes and vegetations and help the restoration of natural ecosystems. Natural succession (no
attempted management) evolved from these post-agricultural lands provides an
134
G. González et al.
opportunity to observe the dynamics of vegetation communities, the re-colonization
sequence of soil fauna, and further the interactive relationships among plants, soil
fauna, and microbes within considerable areas along the continuum of successional
forests. These areas also represent an excellent opportunity to explore the linkage between aboveground communities (forest types, vegetation cover) and belowground soil fauna (especially earthworm communities) in post-agricultural lands. In
addition, the variety of exotic and native earthworm communities observed in postagricultural fields provide for good opportunities to test the mechanisms of invasion
biology. The case study from post-agricultural pastures and secondary forests of
Puerto Rico showed that native earthworm communities are gradually re-colonizing
the forests as aboveground plant communities mature. We can test the three mechanisms (propagule pressure, habitat matching, and biotic resistance) proposed by
Hendrix et al. (2006) by monitoring dynamics of exotic and native earthworm populations and/or conducting field experiments along post-agricultural/succession gradients. Comparisons between different ecosystems and geographic regions are also
possible by using earthworms as an experimental model for invasion biology.
5.4 Future Research Directions
The history of the introductions of non-native flora and fauna is much more complex in the Neotropics than in temperate North America; as it is related to the
complex human history of migration and use of the landscape, water barriers and
island ecosystems. An interdisciplinary approach (i.e., history and ecology) can
help elucidate the spreading of non-native species in the tropics and the development of sustainable land management practices (González et al. 2006). Landscape scale dynamics of earthworm distributions within changing landscapes; and
physiological and other mechanisms controlling community dynamics are important areas of future study on earthworm ecology in the Neotropics. As discussed
above, P. corethrurus shows better physiological adaptations than other earthworms
in terms of oxygen utilization, oxygen consumption, ultraviolet tolerance and high
anti-oxidative enzyme activities. These physiological traits can be powerful and
effective competitive adaptations when non native earthworms invade new areas.
Yet, many questions related to the physiological adaptation of exotic earthworms
like P. corethrurus remained unanswered. For example, why can P. corethrurus survive in oxygen-shortage conditions? Why does P. corethrurus have high tolerance
to ultraviolet radiation? Why does it have higher anti-oxidative stress in its body?
The hemoglobin of earthworms (mega-hemoglobin) is different from vertebrates.
Thus, is the high oxygen binding affinity of P. corethrurus hemoglobin that can help
worms survive in anoxia condition? Also, although we do know that P. corethrurus
is tolerant to UV; we still do not understand the mechanism by which it can protect
itself from the radiation. P. corethrurus mucus is thick and sticky and we infer it
has some protective chemicals. MAA is a chemical which can absorb ultraviolet
in sea hare (Carefoot et al. 1998) and photolyase is an enzyme known to repair
DNA damage in frogs (Blaustein et al. 1998). To date, it is unknown whether or not
5 Earthworms and Post-Agricultural Succession
135
earthworms have these chemicals, but these topics merit study. On the other hand,
heavy metals, pesticides or chemical fertilizers are also harmful to earthworms, but
their toxicity to P. corethrurus is not well understood. Exactly how species such as
P. corethrurus can survive in pastures and / or agricultural fields requires further
investigation.
Acknowledgment While writing this manuscript, Grizelle González received support of grant
DEB-0218039 from the National Science Foundation to the Institute of Tropical Ecosystem Studies, University of Puerto Rico, and the USDA Forest Service, International Institute of Tropical
Forestry as part of the Long-Term Ecological Research Program in the Luquillo Experimental Forest. Additional support was provided by the Forest Service (U.S. Department of Agriculture), and
the University of Puerto Rico. María M. Rivera helped with graph reproduction. We are grateful to
Drs. William A. Gould, Timothy R. Seastedt, and Ariel E. Lugo for kindly commenting on earlier
versions of the manuscript.
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Part II
Mechanisms and Tolerances that Cause
Plant-Plant Replacements, Producing Old
Field Vegetation Patterns
Chapter 6
Factors Affecting the Species Richness
and Composition of Neotropical Secondary
Succession: A Case Study of Abandoned
Agricultural Land in Panama
Elaine R. Hooper
6.1 Neotropical Secondary Succession
Deforestation of extensive areas throughout the tropics and consequent biodiversity
loss is of worldwide concern. Often, unsustainable land use follows deforestation,
producing vast areas of degraded land in various stages of secondary succession
(Brown and Lugo 1990, FAO 2001). Increasing fragmentation of mature tropical
forests is resulting in landscapes where primary forest appears as islands in a
sea of secondary vegetation and anthropogenic grasslands (Uhl 1987). While
species richness may recover rapidly during secondary forest succession, species
composition of secondary forest remains very different from that of mature forest
for many decades (Brown and Lugo 1990, Finegan 1996). The processes determining the species composition during neotropical secondary succession remain
poorly understood (Wijdeven and Kuzee 2000). Successional research in temperate
areas has been used as the basis for modeling neotropical secondary succession,
however, temperate-based models do not include stochastic factors such as seed
availability and dispersal (Finegan 1984, McCook 1994). In tropical areas, most
tree species are animal-dispersed (Howe and Smallwood 1982) suggesting potential limitations in characterizing tropical succession using temperate models. For
example, some of the major seed dispersers in the tropics (frugivorous primates,
large frugivorous birds and bats) are absent from temperate areas (Chapman and
Chapman 1999).
The classical description of neotropical secondary forest succession involves four
phases, each characterized by a different suite of species. Changes in environmental
conditions and differences among species’ growth rates, shade-tolerance, longevity
and size at maturity are thought to drive the sequential physiognomic dominance of
the site by various suites of species as succession proceeds (Finegan 1984, 1996).
The first phase of succession is dominated by herbs, shrubs and climbers. These
disappear under the shade of short-lived pioneers trees, which rapidly develop a
closed canopy and dominate the second phase for 10–30 years. Long-lived pioneers
predominate in the third phase of succession, which can last for 75–150 years.
R. W. Myster, Post-Agricultural Succession in the Neotropics.
C Springer 2008
141
142
E. R. Hooper
Colonization by shade-tolerant, mature forest species occurs continuously, and
eventual dominance by these species characterizes the final, mature forest stage
(Finegan 1996, Guariguata and Ostertag 2001). While life history variation and
competition among species are thought to drive successional change, factors like
initial site conditions and seed dispersal also play a role, and must be factored in for
a more complete description of succession (Finegan 1996).
6.2 Factors Constraining Secondary Succession
in the Neotropics
The successional sequence following deforestation, agricultural use and subsequent
land abandonment often follows less predictable successional trajectories (Guariguata and Ostertag 2001) because land degradation can delay forest regeneration
in proportion to the intensity of prior land use (Uhl et al. 1988, Aide et al. 1995).
Site-specific factors also add to the unpredictability of successional species composition. Interactions between site-specific factors and land-use may facilitate, inhibit
(sensu Connell and Slatyer 1977), or even divert successional sequences predicted
by the general successional model, depending upon the type and intensity of land
use before abandonment (Mesquita et al. 2001). While old-growth forests are high
priorities for conservation because of their complex structure and diverse floristic
composition, the conservation value of secondary forests is becoming increasingly
important as mature forests are rapidly lost through logging and conversion to agriculture (DeWalt et al. 2003). Successional processes may restore the diversity and
species composition to that of old-growth forests, however research throughout the
neotropics has shown that a number of factors, often interacting, may impede tropical forest regeneration at deforested, abandoned sites. These include biotic factors
such as seed dispersal limitations and seed predation, and abiotic factors including
grass competition, fire, drought and low soil nutrients.
6.2.1 Biotic Factors
When intensive land use, often coupled with fire eliminates tree regeneration from
the seed bank, root-sprouts, and advance regeneration (Uhl et al. 1988, Nepstad
et al. 1990), the recovery of mature-forest species composition can be very slow
(Guariguata and Ostertag 2001) and dependent on seed dispersal as the critical
mechanism for determining the template of species recruitment. When seed dispersal is the only possibility for regeneration, succession can potentially be constrained.
While over 80% of tropical tree species have animal-dispersed fruits (Howe and
Smallwood 1982), most animals will not venture into abandoned old fields because
of limited caloric and structural resources, such as perching sites (Cardosa da Silva
et al. 1996) and they often do not venture far from forest edge (Aide and Cavelier
1994). Large-seeded species are often the most severely dispersal-limited because
the predominant frugivores in abandoned areas (small birds and bats) carry small
seeds (Nepstad et al. 1996). Consequently, insufficient seed dispersal is a major
6 A Case Study of Abandoned Agricultural Land in Panama
143
limitation to natural regeneration throughout the neotropics (Janzen 1988, Uhl et al.
1988, Aide and Cavelier 1994, Thomlinson et al. 1996, Holl 1999, Holl et al. 2000,
Wijdeven and Kuzee 2000, Zimmerman et al. 2000). In extremely degraded and
fragmented landscapes, such as the state of São Paulo, Brazil, where remaining
forested areas are estimated to be only 9% of the original native vegetation, even
areas adjacent to remaining forest fragments have limited floristic development
because of the small fraction of regenerating species originating from dispersed
seeds (Maluf de Souza and Batista 2004).
The effects of remnant, or residual vegetation (Chazdon 2003) can also play a
critical role in structuring species richness and composition following disturbance.
For example remnant trees (Guevara et al. 1986, Guevara and Laborde 1993),
shrubs (Vieira et al. 1994, Holl 2002, but see Zahawi and Augspurger 1999) and
large-leaved monocots (Duncan and Chapman 1999) can attract seed dispersers
and facilitate native rainforest regeneration under their canopies. They may also
attract seed predators and herbivores, whose activity can modify the original seed
dispersal shadows centered around this remnant vegetation (Myster and Pickett
1992). The ability of some tree and shrub species to resprout following disturbance
is another important successional mechanism. For example, trees damaged during
hurricanes can resprout, promoting a rapid return of the species composition of
mature forest by circumventing secondary succession (Yih et al. 1991). Species
differ in their ability to resprout following cutting and fire. In areas where fire is a
recurrent disturbance phenomena, this differential resprouting ability can have an
unparalleled impact on forest regeneration, extinguishing those species which can
not resprout and consequently impoverishing and altering successional trajectories
(Cochrane and Schulze 1999).
Post-dispersal biotic factors such as seed and seedling predation may play a significant role in determining the species richness and composition of regenerating
communities, as well as convergence characteristics between secondary and primary
forest. Research in primary forest suggests that post-dispersal factors are a major
factor increasing diversity from dispersed seeds to recruits (Harms et al. 2000, Wills
et al. 2006). Whether these post-dispersal factors contribute to the diversity and
species composition of forests regenerating after agriculture deserves research attention. In temperate areas, differential palatability to seed and seedling predators can
influence species composition during succession by modifying seed and seedling
distributions characteristic of dispersal shadows (leptokurtic seeding distributions
centered around adult trees) (Myster and Pickett 1992). In the neotropics, research
suggests that seed predation has the potential to make significant impacts upon the
diversity of floristic recovery, for example losses due to seed predators can be the
predominant mechanism of seed mortality in some neotropical old-fields (Nepstad
et al. 1996, Myster 2003a, 2004). Seed predation may play a pivotal role in structuring species composition during succession if seed removal is species specific, as has
been found in old fields in Bolivia (Peña-Claros and De Boo 2002) and Puerto Rico
(Myster 2003b). At one site in Costa Rica, dominance of the species Pentaclethra
macroloba in mature-forest canopy is mirrored by dominance of the same species
in the seedling communities in secondary forests 16–18 years after pasture abandonment. In contrast, other mature-forest trees in the region, for example species of
Dipteryx, Tetragastris and Pouteria that are present as remnant overstory individuals
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in secondary stands have low regeneration. The authors (Guariguata et al. 1997)
attribute this dominance by Pentaclethra to the fact that its seedlings lack significant predation, while Dipteryx, Tetragastris and Pouteria have minimum juvenile
recruitment because their fleshy fruits are prone to heavy seed predation.
Abandoned agricultural lands are often invaded by exotic grasses, which can
constrain or arrest succession of tree species. Grass invasions limit tree regeneration at many ecological levels (D’Antonio and Vitousek 1992). At the population
level, alien grasses successfully compete with tree seedlings for water and nutrients
because the grass has dense, shallow root systems (Nepstad 1989, Nepstad et al.
1996, Gonzales-Montagut 1996). Grass competition can affect plant community
composition by preventing many tree species from regenerating, especially smallseeded species (Nepstad et al. 1990, 1996, Hooper et al. 2002).
6.2.2 Abiotic Factors
The harsh physical conditions resulting from the presence of grasses including a
high irradiance regime, desiccating microclimate, and exposure to drought and fire
may also prevent many tree species from regenerating (Nepstad et al. 1990). At the
ecosystem level, grasses are more flammable and able to recover more rapidly after
fire in comparison to trees. Fire arrests natural forest regeneration in abandoned,
deforested lands and encroaches on uncut forest (Janzen 1988, Nepstad et al. 1991).
Fire can also impoverish soils, consequently reducing seedling growth (Aide and
Cavelier 1994), and exacerbating the problem that total nutrient stocks (especially
N) decline with increasing land-use intensity prior to abandonment (Buschbacher
et al. 1988, Reiners et al. 1994). At some sites, soil nutrient losses following tropical
deforestation can limit forest regeneration (Buschbacher et al. 1988). For example,
in abandoned Amazonian pastures, total nutrient stocks and forest recovery were
negatively related to land use intensity prior to abandonment; lightly-used sites were
N depleted while heavily-used sites were N, K and Mg depleted (Buschbacher et al.
1988). At other sites, research suggests that a lack of nutrients reduces seedling
growth in abandoned agricultural lands (Aide and Cavelier 1994). This combination
of factors can make these exotic grassland systems too extreme for successful establishment of many rain forest tree species throughout the tropics (Ashton et al. 2001).
Those species that do dominate extreme post-agricultural landscapes are often a fire
resistant, drought resistant subset of the original native forest species complement
with fast growth in full sunlight and the ability to resprout repeatedly when burned
or cut with a machete (Boucher et al. 2001).
6.3 Measuring Floristic Recovery During Succession
Forest recovery during succession can be measured using a number of variables,
including structural (basal area, biomass, tree height or stem density), functional
(soil variables including soil carbon, nitrogen mineralization, soil water availability,
6 A Case Study of Abandoned Agricultural Land in Panama
145
cation exchange capacity) or floristic. Floristic variables, including species density,
richness and composition have been used by researchers to assess forest recovery
and are the subject of this comparison as they have a high relevance to conservation and restoration efforts. Research throughout the neotropics suggests that
species richness and composition vary independently (Finegan 1996). Secondary
forest succession can recover species richness rapidly, however virtually all studies
of post-agricultural succession suggest that the recovery of mature-forest species
composition is comparatively slow (Brown and Lugo 1990, Chazdon 2003, Howorth
and Pendry 2006). Chronosequence studies suggest that while species richness may
recover to mature-forest conditions over time in neotropical forests, species composition may not converge. Instead, initial site conditions, proximity to mature forest,
type and degree of disturbance may interact to prevent successional forests from
achieving the species composition of the original forest (Capers et al. 2005). Yet,
more information is needed on plant community floristics during secondary succession to assess their potential for maintenance of plant species diversity, stand
compositional dynamics, and to delineate guidelines for their management and conservation (Guariguata et al. 1997).
6.3.1 Multivariate Techniques- Indirect Gradient Analysis
Multivariate techniques are effective at summarizing data that relates changes in
species composition over time, and these methods have advantages over traditional
descriptive studies (Austin 1977). Multivariate analyses have been successfully
applied to analyze successional data and have allowed researchers to dissect temporal trends from environmental patterns and randomness and to understand successional trajectories of species composition. While both direct and indirect methods
of gradient analysis have been used, most studies have utilized indirect gradient
analysis, which allows a description of changes in vegetation structure in relation
to differences at sites. Indirect gradient analysis has been used to confirm the general successional model (Peña-Claros 2003) proposed by Finegan (1984, 1996), to
show a floristic separation between secondary forest regenerating following different land-use histories (Mesquita et al. 2001, Healey and Gara 2003), hurricane
compared to agricultural disturbance (Boucher et al. 2000), on land with different
soil and topography (Finegan and Delgado 2000), as well as a chronosequence of
sites with different times since disturbance (Knight 1975, Sheil 1999, Aide et al.
2000, Howorth and Pendry 2006).
6.3.2 Multivariate Techniques- Direct Gradient Analysis
Direct gradient analysis is advantageous because it can be used to determine the
mechanisms driving successional changes. Direct gradient analysis can be used with
chronosequence data to determine how much of the floristic variation is explained
by successional trend and to illustrate changing patterns of species composition
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in relation to time since disturbance. Additionally, it can show the mechanisms
underlying successional change (ex Purata 1986, Aide et al. 1996), perhaps
contributing to a recent increase in the use of these direct gradient multivariate
analyses by scientists throughout the neotropics to test hypotheses regarding the
effect of different variables on vegetation patterns. For example, direct gradient
analysis has been used to examine how patterns of species composition are affected
by factors thought to affect floristic development such as land use history, soils
and elevation (ex. Aide et al. 1996, Pascarella et al. 2000). Purata (1986) used
direct gradient analysis to show how the slow recovery of species composition
in secondary forest can be attributed to inadequate seed dispersal and the extent
of site degradation. Throughout the neotropics, chronosequence analysis using
multivariate analysis illustrates some of the complexities and barriers to achieving
community compositional convergence. Multivariate techniques are advantageous
in that they can show the interaction of time since disturbance (succession) with
mechanisms that may alter successional pathways, and allow a quantification of the
amount of floristic variation that is explained by these various interacting factors.
In the following overview, I will present results from a range of floristic studies that
use multivariate analyses to characterize forest recovery following deforestation
and agricultural land use, beginning with sites that present a minimum of barriers to
succession (nutrient rich sites close to seed sources) and progressing through those
where barriers to forest regeneration can be severe.
6.3.3 Lightly-Disturbed Sites
A recent study in Costa Rica (Capers et al. 2005) on lightly-used sites suggests that
the species richness of secondary forest seedling communities 13–26 years following pasture abandonment was similar to that of mature forest. However, community
composition of chronosequence plots showed no evidence of increasing similarity
over time. Detrended Correspondence analysis (DCA) showed an initial site distinctness in seedling community composition; these sites remained distinct over the
period of study (5 years) and did not show any trend of convergence (Capers et al.
2005). This result is underscored by examining the DCA biplot; the primary axis
of variation, time since disturbance explained over 50% of the variation in floristic composition, whereas the secondary axis explained only approximately 10% of
the variation. Another study in Costa Rica of secondary stands 16–18 years after
relatively light pasture use (Guariguata et al. 1997) showed that tree and treelet
species richness was lower in secondary stands than in mature forest stands, yet
these differences were much less pronounced in seedling and sapling communities.
Sapling community composition in these secondary forests was floristically more
similar to mature forest stands than was the overstory, mainly because of dominance
of Pentaclethra macroloba in both (Guariguata et al. 1997). The ubiquity of Pentaclethra macroloba in the study area and its ability to survive seed predation and
shade conditions may allow for a greater convergence in floristic composition than
at most other neotropical sites where species dominance is much less pronounced.
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A chronosequence study in Panama following abandonment of lightly-used sites
near mature forest suggests that secondary succession on such sites can proceed
rapidly, with species richness approaching that of mature forest within 20 years
of regrowth (Denslow and Guzman 2000, DeWalt et al. 2003). However, even on
these sites where successional processes are uninterrupted (DeWalt et al. 2003),
floristic differences were pronounced 100 years following abandonment. Secondary
forest stands differed in tree and liana species composition among themselves, and
with old-growth stands. While similarity to old-growth composition increased with
secondary forest age, 100 year old secondary stands and old growth forest had only
a 41% similarity (Sorenson’s index) in tree species composition, suggesting that
species composition converges on old-growth more slowly than does forest structure
(DeWalt et al. 2003).
In Northeast Puerto Rico, conditions for forest recovery are also considered
favourable because of small pasture size, short distances to nearby forest, lack of
fire, and nutrient rich soils. In these areas, direct gradient analysis suggests that successional trend is not deflected by factors related to anthropogenic use, but is related
to natural site factors, specifically elevation (Aide et al. 1996). Species distributions
were significantly affected by age since abandonment and elevation; these effects
were nearly perpendicular on the ordination and explained 14% of the variation.
No effect of distance to the forest was detected, suggesting that the relatively short
distances to forest in this landscape do not limit forest regeneration.
6.3.4 Moderately-Disturbed Sites
In a subsequent study in Puerto Rico at sites with more severe disturbance, direct
gradient analysis similarly identified the importance of age since abandonment and
elevation in structuring species community composition, and additionally showed
that substrate type and bulldozing significantly contributed to species composition
(Chinea 2002). In total 26% of the variance in species community composition was
explained, with over half the variance explained by the two environmental variables (elevation, substrate), while slightly less variation was explained by land-use
history variables (site age, bulldozing). Very little variance was shared by the two
sets of variables; this effect is underscored by the fact that their effect on species
composition was perpendicular in ordination space. The first ordination axis was
positively correlated with bulldozing and negatively correlated with age since abandonment, suggesting that bulldozed sites have a species composition similar to
early-successional communities. Site elevation was positively correlated with the
second axis, a result similar to sites studied by Aide et al. (1996).
6.3.5 Heavily-Disturbed Sites
At sites where previous land use has been more severe, multivariate analyses suggest that not only does time since disturbance play a primary role in structuring
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species composition (as was shown for lightly-used sites), but land-use type use
also explains much of the floristic variation, and these two factors interact. Correspondence analysis (CA) of a chronosequence (11–50 years after abandonment)
in Northwest Argentinian neotropical montane forest comparing abandoned herbaceous crops and abandoned citrus orchard shows that time since abandonment and
previous land use were both important influences on species composition (Grau
et al. 1997). The first axis of the CA accounted for 41% of the floristic variation,
and was related to chronosequence age, while the second CA axis also explained a
significant proportion of the variation (23%) and was related to the difference in land
use (herbaceous vs. citrus). The floristic trend suggests that dispersal mechanisms
mediated by crop structure play an important role in structuring species composition; animal-dispersed species predominated in the abandoned citrus orchards,
while wind-dispersed species dominated early successional stages. Species richness at these sites reached similar values to old-growth forest within 50 years
(Grau et al. 1997).
In abandoned fields in the Mexican humid tropics, research also suggests the
importance of land use and seed dispersal; Purata (1986) studied how intensity
of land-use following abandonment and proximity to forest affect chronosequence
results. She used a direct gradient analysis (Canonical Correspondence Analysis –
CCA) to illustrate how regrowth age (successional trend) interacts with length of
cropping period and proximity to forest to determine floristic composition. The two
main CCA axes of variation explained a high proportion of the floristic variation
(Axis 1–50%; Axis 2–22%) and interestingly all three factors were significantly
related to the first axis, while the second axis was significantly related to regrowth
age and forested perimeter. In contrast to previously-discussed results from lightlyused sites (ex Capers et al. 2005) where the major axis of floristic variation is
explained solely be time since disturbance, at these more heavily-used sites, time
since disturbance no longer primarily determines floristic development. Instead it
interacts with intensity of land use and proximity to forest showing how these factors
can permanently deflect succession and prevent convergence. Floristic development
was positively related to regrowth age and amount of forested perimeter and negatively related to length of cropping period. The latter result suggests that intense
land use inhibits succession; at sites where the length of the cropping period was
low, diversity was higher and floristic composition richer than when it was high.
The positive effect of forested perimeter on floristic development stresses the importance of nearby forest propagules to community recovery; without the addition of
forest species floristic trend is deflected and does not converge with mature forest
composition (Purata 1986).
While species composition may not follow predictable trajectories because of
interacting factors such as unique site and disturbance characteristics, predictable
trends in life-form composition of the seedling community with time have been
shown at a number of sites. In Costa Rica, canopy palms, understory palms and
the proportion of canopy tree seedlings increased in abundance and species richness
over time, while shrub and liana relative abundance declined; these trends correlated to declining light levels with succession (Capers et al. 2005). The proportion of rare species also increased over time because of the increasing recruitment
6 A Case Study of Abandoned Agricultural Land in Panama
149
of canopy tree species whose propagules originated outside the study plots. The
authors (Capers et al. 2005) suggest that these rare species have large propagules
that are animal-dispersed and therefore seed dispersal by animals may be a driver
of this diversity-generating process. In Panama, palm importance values similarly
increased with stand age (DeWalt et al. 2003). Understory shrub assemblages in
Costa Rican premontane tropical wet forest (La Selva Biological station) were more
diverse, richer and had a higher density in secondary forest than in old-growth forest
(Laska 1997), suggesting an inverse correlation between diversity trends in trees and
shrubs. While these predictable trends in life form may suggest that species composition will return to original conditions through the process of secondary succession,
recent research in Puerto Rico shows irreversible changes in species composition
due to human activity (Lugo and Helmer 2004). These changes include persistence
of alien species in mature phases of forest development and a low complement of
endemic species compared to native forests.
6.4 Species Richness and Composition on Sites with Heavy
and Prolonged Disturbance: a Case History from Panama
While previous floristic studies have shown the relationship between time since
disturbance, anthropogenic and landscape factors which can affect species composition on lightly to heavily disturbed sites, no multivariate studies have shown how
species composition can be modified on heavily disturbed sites invaded by exotic
grasses and modified repeatedly by fire. In central Panama, many abandoned sites
are invaded by an exotic grass species, Saccharum spontaneum L. ssp. spontaneum
, that grows in tall (2.5 m) dense, impenetrable, fire-prone stands. In the following
case study, I discuss results from experiments (Hooper et al. 2002, 2004, 2005)
designed to investigate factors affecting the floristics of early stages of secondary
forest succession on abandoned agricultural lands in Panama invaded by Saccharum
spontaneum. At five study sites located near the Panama Canal (9◦06 N, 79◦53 W),
4 km southwest of the Barro Colorado Nature Monument (BCI) I used a multifactorial experimental design to test the effect of limited seed dispersal, Saccharum
competition, fire, and remnant vegetation (trees, large-leaved monocots and shrubs)
on the diversity and floristic composition of regenerating tree and shrub species
[see Hooper et al. (2004, 2005) for a detailed description of the experiments]. The
effect of distance from the forest edge (10, 35, and 85 m) was assessed to evaluate
the role of potential limitations to seed dispersal. Competition with Saccharum was
studied by comparing mown (once and thrice mown yearly) and unmown treatments. The effect of fire was investigated by performing a prescribed burn. Soil
nutrients were compared in Saccharum-dominated sites and adjacent forest. Tree
and shrub sprouts, seedlings and saplings were monitored over a 1-year period
beginning in August, 1996 to assess the effects of these factors on natural regeneration. To determine the effect of trees, shrubs and large-leaved monocots (Musa
and Heliconia spp.) on natural tree regeneration I documented their location and
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measured their proximity to each regenerating tree and shrub seedling. In another
experiment (Hooper et al. 2002), I planted 15,000 seeds of 20 native tree species
varying in seed size and shade tolerance characteristics under combinations of Saccharum shading and mowing treatments and followed their germination and survival
to further assess the effect of Saccharum competiton on tree regeneration. ANOVA,
linear and nonlinear regression analyses and various multivariate analyses including
distance-based redundancy analysis (Legendre and Anderson 1999) forward selection using RDA (following appropriate transformation) and fourth-corner analysis
(Legendre et al. 1997) were used to evaluate the effect of the experimental factors.
In total, I documented 4,984 individuals of 80 species (mean diameter 6 mm)
naturally regenerating in the Saccharum spontaneum. Seedling density was surprisingly high underneath the dense Saccharum, and depended on the season (dry season: 29 ±6 10 m−2 ; wet season: 35 ± 8 10 m−2 ). The most common species (total
number of individuals) were Gustavia superba (1,060), Piper marginatum (522),
Cochlospermum vitifolium (400), Spondias mombin (290) and Cordia alliodora
(241); a full species list (Appendix C) is included in Hooper et al. (2004). Direct
gradient analysis results show that Saccharum competition (P = 0.001), distance
from the forest (P = 0.038) and time since fire (P = 0.016), significantly affect
species community composition of regenerating tree and shrub seedlings (Hooper
et al. 2004). Fire and distance from the forest significantly affect species diversity.
Remnant vegetation in the Saccharum also significantly affects the density, species
richness, and species composition of forest regeneration.
6.4.1 Saccharum competition
Results from the experiment where I planted 15,000 seeds of 20 native tree species
in four Saccharum mowing and shading treatments and unshaded, unmown controls (Hooper et al. 2002) shows that the invasive grass, Saccharum spontaneum,
negatively affects germination, survival, and growth of native tree seedlings in abandoned Panamanian farmland. Most tree species had higher germination, survival and
growth when the Saccharum was shaded in contrast to unshaded control conditions.
As shading treatments essentially eliminated the Saccharum, consequently reducing below-ground competition, while understory light levels were similar when the
Saccharum was 95% shaded as compared to the control, I attribute high seedling
performance in shaded treatments to a decrease in below-ground competition with
the Saccharum. The enhancement of seedling performance in shaded treatments as
compared to the Saccharum control was pronounced in the dry season, leading to a
further conclusion that dry-season water stress limits the regeneration in the Saccharum. This conclusion concurs with results in Amazonian pastures, where seedling
establishment and survival was similarly limited by high temperatures and lower
moisture availability in tropical pastures as compared to forest because the dense
grass root mass in the upper 50 cm of soil significantly decreases soil moisture
availability (Uhl 1987, Nepstad et al. 1990).
Species responded differently to shading and mowing conditions, suggesting that
species community composition can be affected by conditions in the Saccharum.
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151
These species differences in response to experimental treatments were consistently
predicted by their seed size and shade tolerance characteristics. Small-seeded, lightdependent species (Trema, Annona, Vochysia and Jacaranda spp.) had the lowest
overall germination and survival in the Saccharum control and their performance
in the treatments suggest that they can tolerate neither above nor below-ground
constraints imposed by the Saccharum. Light-demanding gap-colonists (Byrsonima, Spondias and Sterculia spp.) with larger seeds had high performance in the
thrice mown Saccharum treatment and low performance in the Saccharum control,
suggesting that these species will colonize after a fire, but will not persist once
overtopped by the Saccharum. Small-seeded, shade-tolerant species (Posoqueria,
Genipa, and Heisteri spp.) had high performance in shaded treatments, but low
performance in the mown and control Saccharum, suggesting that these species
could colonize sites where the Saccharum is shaded out, such as below remnant
vegetation but will not successfully regenerate in open conditions after fire, or in
dense Saccharum. Large- to very large-seeded (2.9 to 50.4g), shade-tolerant species
(Dipteryx, Calophyllum, Carapa, and Virola) characteristic of mature-forest had the
highest performance in the Saccharum control, combining high germination and
high survival.
These results suggest that the potential for succession to be constrained in
Saccharum-dominated sites is high. At sites in close proximity to ours (< 4 km)
on Barro Colorado Island (BCI) that are not invaded by Saccharum spontaneum,
successional trajectories follow Finegan’s (1996) general neotropical successional
model. For example, Kenoyer (1929) reports a four-stage, post-disturbance succession on BCI: (1) grasses and weedy plants, large monocots (Heliconia spp.), and
shrubs of the genus Piper; (2) dominance by short-lived pioneers; (3) mixed secondary forest of longer lived pioneers; (4) climax forest stage. In contrast, experimental results at Saccharum-dominated sites (Hooper et al. 2002) suggest that
this successional sequence may be constrained. Small-seeded, light-demanding tree
species of the second phase of succession (pioneers) can not invade because they do
not successfully compete with the Saccharum. The larger-seeded, light-dependent
species which would form the third successional phase (long-lived pioneers) are
also precluded because they need more light. Their high performance in mown
Saccharum suggests that these species may potentially regenerate directly after fire
before the Saccharum has re-established. Small-seeded, shade-tolerant species of
the mature-forest which would continuously colonize the understory in the typical
successional sequence (Finegan 1996) are also absent because they can not compete with the Saccharum. In contrast, large-seeded, shade-tolerant species typical of
mature forest may regenerate in the Saccharum as they have the highest potential
performance in the Saccharum controls, and were the only species able to resprout
after recurrent fires.
6.4.2 Seed Dispersal
A comparison of the above results from experimentally-dispersed species to those
naturally-regenerating in the Saccharum suggests insufficient seed dispersal is a
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major barrier to forest regeneration. When seed input was equalized, large-seeded
shade-tolerant species had the highest germination and survival but most individuals
naturally regenerating were small-seeded gap species (Hooper et al. 2005). The high
germination and survival of the large-seeded shade-tolerant species suggest that if
they were dispersed into the Saccharum, they would be found in the highest proportions. Yet, they are rarely found naturally regenerating. These findings suggest similar regeneration constraints in Panamanian and Amazonian secondary succession.
In Amazonian pastures, the small-seeded bat- bird-dispersed species most likely to
be dispersed into abandoned Amazonian pastures had the lowest survival whereas
large-seeded species were constrained by the fact that their mammalian dispersers
rarely moved into abandoned pastures (Nepstad et al. 1990, 1996). My research
results suggest that these dispersal limitations depend on distance from the forest,
distance from remnant vegetation, and time since fire.
6.4.3 Distance from the Forest
Distance from the forest significantly affected species richness (ANOVA; P = 0.036)
and species community composition (db-RDA; P = 0.001). Species richness was
almost threefold lower at 35 m as compared to 10 m from the forest. Similar declines
in species richness away from forest edge throughout the neotropics (Uhl et al. 1988,
Aide and Cavelier 1994, Holl 1999, Cubiña and Aide 2001, Rodrigues et al. 2004)
suggest that if distance to the nearest seed source is large, this will limit the speed,
diversity, and species composition of natural forest regeneration. The effect of distance from the forest on species composition depended on time since fire (db-RDA;
effect of site × distance interaction on community composition: P = 0.001, 12.2%
of the variance explained) (Hooper et al. 2004). Sites with early post-fire communities (1–2 years post-fire) had significant differences in community composition
as a result of distance from the forest, while sites surveyed 3 and 4 years after fire
showed no significant effect (P = 0.119, P = 0.116) of distance from the forest on
community composition. Purata (1986) similarly reports that the interaction of time
since disturbance and distance to the forest was important in determining floristic
composition in abandoned Mexican fields. Finegan (1996) also reports how initial
site conditions affect propagule dispersal, which determines species richness and
composition during neotropical rainforest succession.
At Saccharum-dominated sites, community composition of tree seedling regeneration 1 year after fire is dominated by light-demanding, wind-dispersed species,
with some small-bird-dispersed species present; their abundance declines exponentially with increasing distance from the forest. This decline is steep enough that few
individuals are found 35 m from the forest. In contrast to sites close to the forest,
the density of the bat-dispersed (Piper marginatum) increases with distance from the
forest, peaking at 35 m and dominating the community there. At other neotropical
sites, bats have a similar effect on tree regeneration because they defecate in flight
(Charles-Dominique 1986) and thus disperse more seeds to locations farther from
perches than birds (Gorchov et al. 1993, Cardosa da Silva et al. 1996, Medellin and
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Gaona 1999). At distances furthest from the forest, few individuals are found except
for species that can recruit from seed after fire or are able to reproduce in these
Saccharum grasslands. No large monocots, trees, saplings, or shrubs are present
1 year after fire. In contrast, as time since fire increases, naturally-regenerating
wind-dispersed species become progressively less abundant, large-animal-dispersed
species become more abundant, remnant vegetation become established and reaches
a height great enough to overtop the Saccharum, and the importance of distance
from the forest on seedling community composition declines. Distance from remnant vegetation becomes more important with time. Two years post-fire, significant
differences in species composition result from the interaction of distance to remnant vegetation and forest. Three and 4 years post-fire, no significant differences
in community composition are found because all sites are dominated by species
regenerating close to the remnant vegetation (Hooper et al. 2004).
6.4.4 Remnant Vegetation
I utilized non-linear regression analysis to model the effect of distance from the forest and distance from various types of remnant vegetation (isolated trees, shrubs, and
large-leaved monocots) on seedling density and species richness and used a backwards selection procedure to determine the relative importance of each (Table 6.1).
Results suggest that remnant vegetation, specifically large-leaved monocots is the
single most important factor affecting seedling density. Distance from the large
monocots accounts for 56.6% of the variation in seedling density, with a fourfold
increase in seedling density in plots with large monocots present as compared to
those with no large monocots. Three of the most common species (accounting for
42% of the total density of all naturally-regenerating individuals) had densities that
declined exponentially with distance from these monocots including a bat-dispersed
species, an agouti-dispersed species and a species dispersed by both birds and bats
(Dalling et al. 1998b). Other remnant vegetation in these Panamanian old-fields also
influenced different species of tree regeneration. Isolated trees and shrubs had more
small-bird-dispersed species, and more Byrsonima crassifolia, a species dispersed
by large birds and mammals. These results concur with research throughout the
tropics, where trees (Guevara et al. 1986, Guevara and Laborde 1993, Cardosa da
Silva et al. 1996, Duncan and Chapman 1999, Toh et al. 1999, Wijdeven and Kuzee
2000), shrubs (Vieira et al. 1994, Holl 2002, but see Zahawi and Augspurger 1999)
and large monocots (Duncan and Chapman 1999) act as recruitment foci for tropical
tree seedlings at abandoned sites because they attract seed dispersers. Differences in
the relative density of recruitment below different types of remnant vegetation may
result from differential attraction of frugivores.
In contrast, the density of naturally-regenerating trees and shrubs did not correlate with distance to the forest. Duncan and Duncan (2000) and Slocum and Horvitz
(2000) similarly report that in areas containing remnant vegetation, seedling density does not correlate with distance from the forest, and Guevara et al. (1986)
and Willson and Crome (1989) record little effect of distance from the forest on
–
–
–
3. Ballistic-dispersed species
Hyba pr
0.113−
0.091−
0.133−
–
0.611−
–
–
–
–
–
–
Ex
Linc
Forest
1.Wind-dispersed species
(not reproducing)
Tric ga
Lueh se
Cord al
2.Wind-dispersed species
(reproducing)
Coch vi
Total number of species
Total density of all individuals
Speciesb
0.335−
–
–
–
–
0.662−
0.566−
Ex
Monocot
–
–
–
–
–
–
–
Lind
–
–
–
–
–
–
–
Ex
Tree
0.155−
–
–
–
–
–
–
Line
–
–
–
–
–
–
–
Ex
Shrub
–
–
–
–
–
–
–
Linf
–
Dead 0.120−
Fire 0.111−
–
–
–
–
–
–
linear
Fire 0.108−
–
Fire 0.164+
–
–
linear
Habitata
0.464
0.000
0.091
0.133
0.000
0.679
0.566
partial
Biotic
0.000
0.171
0.108
0.000
0.164
0.000
0.000
partial
Abiotic
0.464
0.171
0.161
0.133
0.164
0.679
0.566
Total
F5,85 = 15.85
F3,87 = 4.49
F3,87 = 6.15
F2,88 = 6.80
F1,89 = 17.47
F4,86 = 7.24
F3,87 = 22.65
F
<0.0001
0.0139
0.0032
0.0107
<0.0001
0.0002
<0.0001
P
Table 6.1 Summary of results of nonlinear regression analyses, modeling the effect of distance from various types of old-field vegetation (biotic) and abiotic factors
on the density of naturally-regenerating tree and shrub species per 10 m2 . Only the variables that were significant (P < 0.05) following backward elimination are
presented. For old-field vegetation types, both the negative exponential (Ex) and linear functions (Lin) were used in the model. All reported values in the 10
left-hand columns are R2 statistics resulting from regression against a single explanatory variable of distance from the various old-field vegetation types or
abiotic factors. Positive exponent: increasing density with distance from source; negative exponent: decreasing density with distance from source. Source includes:
distance from the forest (Forest), distance from the nearest large-leaved monocot (Monocot), distance from the nearest old-field tree larger than 5 m in height
(Tree), and distance from the nearest shrub or sapling 2.5–5 m in height (Shrub). Habitat: see notea . Biotic = partial R2 for all biotic variables. Abiotic = partial R2
for all abiotic variables. Total = total R2 for the (multiple) nonlinear regression model. F = F-statistic; P = P-value; n = 90 (reported in Hooper et al. 2004; digital
Appendix D)
154
E. R. Hooper
–
–
–
–
–
–
–
–
–
–
–
–
–
0.449−
0.096−
–
–
0.250+
0.223−
–
–
–
–
–
–
–
–
–
–
–
–
–
–
–
–
–
–
–
–
–
–
–
–
–
–
0.168+
–
–
–
–
–
–
–
–
–
–
–
–
–
–
–
–
–
–
–
0.112−
–
–
–
–
–
–
–
–
–
Dead 0.283+
–
–
–
Dead 0.284+
–
–
–
–
Mid 0.136+
–
–
–
–
–
–
–
–
–
–
–
–
Low 0.091−
–
–
–
0.577
0.173
0.440
0.000
0.000
0.147
0.283
0.000
0.000
0.000
0.284
0.000
0.000
0.000
0.000
0.149
0.000
0.000
0.000
sec−1 ),
0.449
0.096
0.000
0.112
0.250
0.223
0.000
sec−1 ), Low = Light (PAR) at 0.1 m (μ moles m−2
0.173−
–
–
–
0.147−
(μ moles m−2
–
0.577−
–
–
–
–
–
–
–
–
0.312−
–
–
–
–
–
–
–
–
–
–
–
–
–
0.0368
0.0252
0.0024
0.0318
F2,88 = 4.50
F2,88 = 5.18
F3,87 = 6.48
F2,88 = 4.76
0.479 F4,86 = 17.73 <0.0001
0.096
0.000
0.112
0.317
0.223
0.000
0.577 F3,87 = 47.97 <0.0001
0.173 F2,88 = 9.21 0.0002
0.487 F6,84 = 12.03 <0.0001
0.000
0.000
0.147 F2,88 = 7.59 0.0009
include: Mid = Light (PAR) at 0.5 m
Live = biomass of live Saccharum
(g), Dead = biomass of Saccharum litter (g), Fire = time since wildfire (months), Slope = slope (angle above horizontal) . All abiotic variables were ln (x + 1)
transformed, except fire and slope. Slope was square root (x + 0.1) transformed; time since the last wildfire was untransformed
b Species: Bana gu = Banara guianensis, Byrs cr = Byrsonima crassifolia, Cecr in = Cecropia insignis, Coch vi = Cochlospermum vitifolium, Cord al = Cordia
alliodora, Gust su = Gustavia superba, Hyba pr = Hybanthus prunifolius, Inga ve = Inga vera, Lueh se = Luehea seemannii, Mico ar = Miconia argentea, Pipe ma
= Piper marginatum, Psyc gr = Psychotria grandis, Quas am = Quassia amara, Spon mo = Spondias mombin, Swar s2 = Swartzia simplex var. ochnacea , Thev
ah = Thevetia ahouai, Trem mi = Trema micrantha, Tric ga = Trichospermum galeotti
c Square root of distance from the forest
d Distance from the nearest large-leaved monocot (Musa sp., Heliconia sp.)
e Square root of distance from the nearest isolated old fields tree larger than 5 m in height
f Ln (distance from the nearest shrub or sapling between 2.5 and 5 m in height)
a Abiotic variables
4. Small-bird-dispersed species
Mico ar
Cecr in
Bana gu
Trem mi
Psyc gr
5. Bat-dispersed species
Pipe ma
6. Large-arboreal-animal-dispersed species
Thev ah
Quas am
Byrs cr
Swar s2
Inga ve
Spon mo
7. Agouti-dispersed species
Gust su
6 A Case Study of Abandoned Agricultural Land in Panama
155
156
E. R. Hooper
animal-dispersed seed rain near isolated trees and shrubs present in old-fields.
Slocum and Horvitz (2000) conclude that distance to the forest is most important
when remnant vegetation is not found in degraded lands, like very early in succession. Seedling distributions at my Panamanian sites support this hypothesis. While
I attribute these results to the effect of different seed dispersal vectors, it is possible that post-dispersal seed predation modifies the original leptokurtic distributions
of seeds dispersed near forest. For example, Myster and Pickett (1992) similarly
found that the negative exponential decline in density of dispersed species from
vegetation became weaker as succession proceeded, and attribute this result to herbivore pressure. My results can not distinguish between the effects of seed dispersal
and post-dispersal factors as I measured the distributions of seedlings, however the
impact of seed predators and herbivores deserves research attention.
Species richness was significantly affected by both distance from the forest and
distance from the large-leaved monocots, accounting for 68.1% of the variation
(Table 6.1). Many species were significantly affected by proximity to the forest,
with the density of larger-arboreal-animal-dispersed, wind-dispersed, and ballisticdispersed species declining linearly with increasing distance from the forest. Holl
(1999) and Aide and Cavelier (1994) record similar declines in seedling species
richness with distance from the forest edge at other neotropical sites, which may
result from dispersal limitations. For example, Zimmerman et al. (2000) document
a decline in the species richness of seed rain with increasing distance from the forest.
In contrast to these biotic factors, abiotic factors in Panamanian old-fields did not
significantly affect the overall density or species richness of naturally-regenerating
trees and shrubs. Similarly, soil nutrient levels were not a significant factor in influencing natural regeneration (Hooper et al. 2005). These results suggest that biotic
factors are much more important than abiotic post-dispersal factors in affecting
species richness and composition. However, plots with remnant vegetation, and
those close to the forest may also provide superior abiotic conditions for growth
and survival compared to grass-dominated sites (Guevara et al. 1992, Vieira et al.
1994, Nepstad et al. 1996, Holl 2002, Hooper et al. 2002) because of lowered grass
competition and a more favourable microclimate (Vieira et al. 1994) These results
underscore the importance of remaining forest and remnant vegetation in attracting seed dispersers (and potentially seed predators) and consequently increasing
diversity and driving succession in these Saccharum-dominated old-fields. These
results suggest that in their absence, forest regeneration will be slow and limited
to locations close to pre-existing forest. Proximity to forest and different remnant
vegetation in the grasslands differentially affects seed dispersal vectors, resulting in
spatially distinct variations in community composition in relation to this vegetation.
6.4.5 Fire
Fires burn yearly in the dry season in these Saccharum-dominated abandoned pastures. My data show that these wildfires are a major barrier to tree regeneration
because they decrease species richness and change community composition. Fire
6 A Case Study of Abandoned Agricultural Land in Panama
157
significantly (P = 0.025) decreased the number of species from 6 ± 0.1 10 m−2
pre-fire to 3 ± 0.6 10 m−2 post-fire. In contrast, the density of regenerating tree
and shrub species was not significantly affected by burning because a subset of
the original species resprouted in high densities after fire. Species that were able to
sprout after fire formed a large part of the initial floristic recovery. The percentage of
seedlings originating as rootsprouts significantly (ANOVA; P = 0.036) increased in
burned plots (unburned: 31 ± 7 %, burned: 52 ± 10 %) (Hooper et al. 2005). Species
differed in the effect that fire had on the amount of regeneration from resprouts and
recruits (from seed); these differences significantly impacted upon community composition (db-RDA; P = 0.001) (Hooper et al. 2004). Many more species showed a
decline in one or both mechanisms of regeneration compared to those that increased
in response to fire, leading to a low-diversity community following fire, as has been
similarly documented at other tropical sites (Uhl et al. 1988, Nepstad et al. 1990).
For example, 43 species had decreased recruitment from seed, while only 11 had
enhanced recruitment after fire (Hooper et al. 2004). Half the species able to resprout
had lowered or no resprouting following fire. Other authors similarly found that
burning reduces the number of species resprouting at neotropical sites (Uhl et al.
1990, Sampaio et al. 1993). The few species that had increased resprouting after
fire, especially Gustavia superba, Cochlospermum vitifolium, and Cordia alliodora
dominated post-fire community composition.
6.4.6 Succession Following Fire
To characterize species differences in relation to time since fire, I used a series of
4th corner correlations (Legendre et al. 1997) relating time since fire with species’
dispersal mechanism (Table 6.2), species’ regeneration characteristics (shadetolerance, seed size, resprouting ability) (Table 6.3) and life-form (Table 6.4). The
species characteristics negatively correlating with time since fire include wind
dispersal and ability to reproduce in Saccharum-dominated old-fields (Table 6.2),
shade-intolerance (Table 6.3), and mid-sized tree life-form (Table 6.4). In contrast,
the species characteristics positively correlating with time since fire include largeanimal-dispersal (Table 6.2), regeneration from root sprouts (Tables 6.2 and 6.3),
Table 6.2 Dispersal vector / mechanism (see text) compared to time since fire. The table reports
correlation fourth-corner statistics r(i,j), and probabilities (Prob) tested using 999 permutations followed by adjustment using Holm’s procedure (reported in Hooper et al. 2004; digital Appendix E)
r(i,j)
Prob
1. Wind, not 2. Wind,
reproducing reproducing 3. Ballistic
–0.0435
–0.1329
0.0806
0.5400
0.0080*
0.1200
4. Small
birds
–0.0390
0.5400
5. Bats
–0.0380
0.5400
6. Large
arboreal
animals
0.1231
0.0080*
7. Agouti
–0.0364
0.5400
Resprout
indexa
0.1382
0.0080*
*P < 0.01
a Resprout index, calculated as the number of individuals originating from root sprouts divided by
the total number of individuals, indicates the ability of each species to resprout
158
E. R. Hooper
Table 6.3 Species characteristics including light index, seed size and resprout index compared
to abiotic variables. Only those abiotic variables that significantly correlated with species characteristics are presented. The table reports correlation fourth-corner statistics r(i,j), and probabilites
(Prob) tested using 999 permutations (reported in Hooper et al. 2004; digital Appendix F)
Time since fire (months)
r(i,j)
Prob
Light indexa
Seed size (mm)
Resprout indexb
–0.15507
0.00100*
0.04682
0.08200
0.16238
0.00100*
*P < 0.01
a Light index was calculated in Condit et al. (1996). Species with a high light index ( > 3.0) are
pioneers and light demanding, those with a low light index (< 0) are shade tolerant, and those
between 0 and 3.0 are gap-dependent
b Resprout index (calculated as the number of individuals originating from root sprouts divided by
the total number of individuals) indicates the ability of each species to resprout
shade-tolerance (Table 6.3), and shrub life-form (Table 6.4). These results suggest
a very general characterization of succession following fire begins with lightdependent species regenerating from small-seeded, wind-dispersed seeds; these
mid-sized tree species such as Cochlospermum vitifolium are able to reproduce
quickly within the old-fields. As succession proceeds, larger-seeded, shade-tolerant
species, especially shrubs but also large trees (result marginally non-significant; P =
0.07) that can resprout and that are dispersed by large arboreal animals predominate.
This data generally supports predictions by the general successional model
of Finegan (1996) that species composition will change from small-seeded pioneer species dominating early in succession, to more shade-tolerant, mature forest
species dominating late-successional communities. However, the chronosequence
at my Panamanian sites differs fundamentally from Finegan’s successional model
(Hooper et al. 2004). While I similarly record a dominance of the site by grasses
(Saccharum) as a first successional stage, the second successional stage of shortlived, light-demanding species is virtually absent. Short-lived pioneers of the genera
Cecropia, Trema and Ochroma predominate the 3rd to 15th years following disturbance at nearby sites that are not invaded by Saccharum (ex: Kenoyer 1929). In
contrast, light-demanding species significantly declined in abundance after only a
few years following fire, suggesting that pioneer species do not persist in Saccharum grasslands. Saccharum sites lack dominance by core pioneers (especially of the
genus Cecropia) that dominate the seed bank at unburned sites less than 4 km away
(Dalling et al. 1998a). Research at Amazonian sites also demonstrates that Cecropia
Table 6.4 Adult size of individuals, as defined in Condit et al. (1995), regenerating in the old
fields, related to time since fire (months). The table reports correlation fourth-corner correlation
statistics r(i,j), and probabilities tested using 999 permutations (reported in Hooper et al. 2004;
digital Appendix G)
r(i,j)
Probability
*P < 0.05;
**P < 0.01
Large tree
Mid-sized tree
Treelet
Shrub
0.0711
0.0700
–0.1842
0.0040**
0.0441
0.1560
0.1201
0.0150*
6 A Case Study of Abandoned Agricultural Land in Panama
159
species are extinguished as a component of the regeneration (Didham and Lawton
1999, Mesquita et al. 2001) when fire is present. The small-seeded pioneers that do
survive in the seed bank and germinate following fire have very low performance
in the Saccharum because they can tolerate neither above-ground nor below-ground
Saccharum competition (Hooper et al. 2002).
Species of the long-lived pioneer genera Luehea, Apeiba, Miconia, and Cordia,
characteristic of the third phase of the general neotropical successional model are
the major component of the regeneration at Saccharum-dominated sites. Results
from the planting experiment (Hooper et al. 2002) suggest that these species can
tolerate below-ground Saccharum competition because of their larger seed size.
Their abundance supports the hypothesis of Finegan (1996) that regeneration of the
long-lived pioneers is abundant in the first years of succession. Interestingly, many
species regenerating in the Saccharum are characteristic of mature-forest (Hooper
et al. 2004); for example 54 of the 80 species regenerating in Saccharum are also
present in mature-phase forest on Barro Colorado Island nearby (Condit et al. 1996).
The spatial distribution of these species is positively affected by proximity to the
forest and remnant vegetation, while species composition of these communities is
significantly affected by the identity of the vegetation (ex forest, large monocot)
found in close proximity; these results follow the nucleation model of Yarranton
and Morrison (1974). Therefore, while species composition is similar to the third
and fourth stages of the general successional model (Finegan 1996), their spatial
distribution follows the nucleation model of succession, highlighting the importance of seed dispersal and initial site conditions on propagule dispersal, seedling
establishment, and survival. Results at Saccharum-dominated sites suggest that
both temporal and spatial factors must be considered for a complete description of
succession.
In the continual presence of fire, however, this successional sequence does not
have a chance to proceed, and a Saccharum-dominated grassland will continue to
effectively preclude woody species regeneration. Authors throughout the neotropics
report how grass dominance on repeatedly-burned lands can suppress tree seedling
recruitment for years (Cochrane and Schulze 1998) or permanently deflect succession to a savanna and scrub vegetation (D’Antonio and Vitousek 1992). This could
lead to impoverishment of vast expanses of tropical lands as fire becomes more
widespread (Uhl 1998). This irreversible degradation is extremely likely in the presence of invasive exotic grasses, as demonstrated in much of the Old World tropics
(Richards 1964, Otsamo et al. 1995). At neotropical sites invaded by exotic grasses,
results in Panama and Brazil suggest that the inability to resprout can drive local
species extinction, lower biodiversity and modifying species composition if seed
dispersal is limited, and the seed bank becomes degraded (Miller and Kauffmann
1998, Hooper et al. 2004).
In conclusion, results from heavily-disturbed sites in Panama concur with
research at other moderately- to heavily-disturbed neotropical sites (Purata 1986,
Grau et al. 1997, Chinea 2002) that illustrate how floristic composition following
land abandonment differs from that predicted by the general successional model
because of the interaction of previous land use and the spatial dynamics of seed dispersal. In abandoned, Saccharum-dominated sites in Panama, as at other neotropical
160
E. R. Hooper
locations (Cochrane and Schulze 1999), fire is a recurrent disturbance phenomena.
Fire permanently deflects succession and lowers biodiversity by extinguishing many
species from the seed bank and favouring a limited subset of fire-adapted species
that can resprout. Competition with the Saccharum spontaneum further modifies
predicted successional trajectories by preventing small-seeded species from regenerating, in effect eliminating the pioneer phase of succession. Larger-seeded species
are able to survive Saccharum competition, however these species are mostly
animal-dispersed and consequently dispersal-limited, a result consistent at many
neotropical sites (Janzen 1988, Uhl et al. 1988, Aide and Cavelier 1994, Cardosa da
Silva et al. 1996, Thomlinson et al. 1996, Holl 1999, Holl et al. 2000, Wijdeven and
Kuzee 2000, Zimmerman et al. 2000). The density and spatial distribution of these
vertebrate-dispersed species are influenced by the location of pre-existing forest
and also remnant vegetation, such as trees, shrubs, or large-leaved monocots which
attract seed dispersal vectors and ameliorate conditions for seedling regeneration.
Direct gradient analysis confirms that at neotropical sites with moderate to heavy
disturbance histories (Purata 1986, Grau et al. 1997), and Panamanian Saccharumdominated sites, distance from this vegetation explains much of the variation in
species composition, and is consequently a major driver of the floristics during
succession. These results concur with McCook (1994) and Finegan (1984) who
suggest that spatial and stochastic factors, such as seed availability and dispersal
should be included for a more complete description of neotropical succession.
Acknowledgment This work was supported by a grant from the Fundación AVINA. Richard
Condit and Pierre Legendre provided invaluable support throughout. The Smithsonian Tropical
Research Institute in Panama provided generous logistical assistance. I thank members of the Center for Tropical Forest Science, including Jorge Aranda, Salomón Aguilar and Rolando Pérez for
tree seedling identification, Jorge Aranda for help with vegetation mapping, Suzanne Loo de Lao
for help with initial database management, and Lidia Valencia for logistical support. I am grateful
to M.S. Ashton for valuable comments on later versions of this Manuscript. Special thanks to the
people of Las Pavas for their dedicated field assistance and Alejandro Sanchez for sharing his
knowledge of the local flora.
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Chapter 7
Seed Dispersal by Cattle: Natural History
and Applications to Neotropical Forest
Restoration and Agroforestry
Clara Luz Miceli-Méndez, Bruce G. Ferguson and Neptalí Ramírez-Marcial
7.1 Introduction
Agricultural expansion in Latin America is intimately associated with tropical deforestation and biodiversity loss (e.g. Angelsen and Kaimowitz 2001). Although agricultural activities are frequently abandoned (Aide and Grau 2004), their effects on
forest regeneration processes may be much more persistent (e.g. Ferguson et al.
2003). On land subjected to prolonged and/or heavy use, potential for forest recovery is limited by the dearth of tree seedlings, resprouts and seed banks (Uhl et al.
1990). At such sites, seed dispersal from off-site is the principal colonization route
for woody plants. However, distance to seed sources, lack of seed disperser activity,
seed predation, herbivory, poor soils, high light intensity, frequent fires and competition with herbaceous ground cover impede forest regeneration from seed rain (Holl
1998b, 1999, Meli 2003).
Among the most common of these barriers is the lack of animals capable of
transporting seed between wooded and open areas (Holl 1999, Guariguata and
Ostertag 2002, Meli 2003). Animals disperse as many as 66% of tree species in
Neotropical forests, principally by swallowing and defecating their seed (Howe
and Smallwood 1982, Wunderle Jr. 1997). The major dispersers are birds and
bats, although primates, rodents and other mammals may play a complementary
part. Many wild seed dispersers, particularly forest birds, do not cross open areas
(Wunderle Jr. 1997). Frugivorous birds that do enter open areas generally limit their
activities to patches of woody vegetation within the agricultural landscape, and their
movements depend upon plant phenology, particularly fruiting period. The result is
a scanty seed rain that is heterogeneous in both space and time (da Silva et al. 1996,
Wunderle Jr. 1997).
Restoration ecologists have proposed techniques to overcome or bypass this barrier to tropical forest regeneration. Such techniques include the manual collection
and planting of tree seeds and/or seedlings and the artificial establishment of structural diversity to attract dispersers to open areas (McClanahan and Wolfe 1993, Holl
1998a, Holl and Kappelle 1999, Parrotta and Knowles 1999). While these efforts
have met with some success, they can be costly to implement at large scale (Lamb
et al. 2005).
R. W. Myster, Post-Agricultural Succession in the Neotropics.
C Springer 2008
165
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C. L. Miceli-Méndez et al.
Another option for reestablishing tree cover over large areas arises from the
observations of Janzen and Martin (1982), who argued that numerous Central American trees are adapted for dispersal by large herbivores. When much of the Pleistocene megafauna went extinct around 10,000 years ago, they maintain, numerous
tree species were left without efficient dispersers. Since colonial times, seeds of
many such trees have been dispersed by introduced livestock, particularly cattle and
horses.
Under the right circumstances, cattle, a salient feature of many Neotropical
landscapes, are indeed important seed dispersal agents. Where management permits, cattle cross easily between secondary forests, younger fallows, and pastures
(Hernández and Benavides 1995), and may also be rotated through fields to graze on
crop residues (Jiménez Ferrer et al. 2003). They eat the fruit of a wide variety of trees
and shrubs (Moreno Casasola 1996), and their digestive processes are mild enough
to leave many seeds intact (Razanamandranto et al. 2004, Pascacio Damián 2006).
Furthermore, light grazing can diminish the competitive effect of grasses on woody
colonizers (Posada et al. 2000) another important barrier to tree establishment (Holl
et al. 2000, Posada et al. 2000).
Although many instances of seed dispersal by cattle (hereafter referred to as
“bovinochory”) are known, systematic studies of the relationship are lacking. We
review ecological and agronomic literature to typify bovinochorous trees and shrubs
and describe the methods used to study the phenomenon. We describe fruits, seeds
and habitat of the species involved, quantify reports of bovinochory by plant family
and summarize the authors’ conclusions with regard to the impact of bovinochory.
Next, taking the ecologically diverse state of Chiapas, Mexico as a case study,
we analyze the management potential of bovinochory as a tool for reestablishing
woody plant cover under both active cattle production and ecological restoration
scenarios. We present preliminary interview and sampling data identifying woody
bovinochores in Chiapas, then combine information from interviews and literature
review to describe the distribution, physiological requirements, uses and aptitude
as successional nuclei of these species. Finally, we propose a research agenda to
support full realization of bovinochory’s management potential.
7.2 Methods
7.2.1 Literature Review
Between September, 2004 and June, 2006, we consulted databases including
Science Citation Index expanded (ISI web of science, 1951–2005), Google
Scholar and the Livestock, Environment and Development Digital Library
(lead.virtualcentre.org). Search terms included combinations of “seed dispersal,” or
“germination,” and “livestock,” “cattle,” or “bovine,” in both English and Spanish.
We also searched for articles citing Janzen and Martin (1982). We carried out additional searches to develop descriptions of the characteristics of each species relevant
to their dispersal, habitats, uses and potential function as successional nuclei.
7 Seed Dispersal by Cattle
167
7.2.2 Chiapas Field Studies
During 2005 and 2006 we conducted interviews on 60 ranches to identify
bovinochorous, woody species in both the wet and dry lowlands of Chiapas. During
both the wet and dry seasons, we surveyed dung pats on the same ranches for
germinating seedlings. We also collected dung and sieved a portion for seeds while
spreading the rest over soil-filled trays in the greenhouse to facilitate germination.
This work is ongoing and the preliminary results are reported here as a case study
of the management potential of bovinochory. Details of the methodology will be
provided elsewhere (Miceli Méndez In prep.).
7.3 Results
7.3.1 Studies Identified and Their Methodologies
Our searches identified 38 articles that mention 24 bovinochorous, woody
species (cited here by the scientific names utilized by the Missouri Botanical
on mobot.mobot.org; Table 7.1). We classify these according to the methodology
they employed to study bovinochory as experimental, descriptive and/or modeling.
The experimental studies include germination trials of seed banks in pasture soil and
dung (Gutiérrez and Armesto 1981, Janzen et al. 1985, Somarriba 1986, Peinetti
et al. 1993, Campos and Ojeda 1997, Brown and Archer 1999, Doucette et al.
2001, Villagra et al. 2002, Kneuper et al. 2003, Razanamandranto et al. 2004) as
well as trials of pregermination treatments (McCully 1951, Argaw et al. 1999,
Rubio-Delgado et al. 2002). Other experimental trials compared the effectiveness
of simulated bovine digestion with those of other scarification treatments (Janzen
et al. 1985, Somarriba 1986). Studies described as observational/descriptive mainly
documented foraging behaviour, presence of viable seed in dung and the distribution
of seedlings. Their data come from both field observations and interviews. The
modeling studies simulated the population dynamics of bovinochorous species
under differing management scenarios (Radford et al. 2001, Tews et al. 2004).
7.3.2 Characteristics of Bovinochores
Sixteen of the 24 bovinochorous species identified belong to the Fabaceae family,
two species are Rosaceae, and 6 other families, Boraginaceae, Caprifoliaceae, Myrtaceae, Rutaceae, Sterculiaceae, and Tiliaceae, are represented by a single species
each (Fig. 7.1). Two of these families are in the same order (Malvales) but no other
relationships are apparent at this taxonomic level. Of the Fabaceae, 14 species are
in the Mimosoideae subfamily and two are Ceaesalpinoideae.
Type of study
E
E
D
Species
Acacia caven
(Mol.) Molina
Mimosoideae Fabaceae
Acacia dudgeoni
Craib ex Hall
Mimosoideae
Acacia seyal
Del. Mimosoideae
Burkea africana
Hook. f.
Caesalpinioideae
Prosopis africana
(Guill., Perrott. & Rich.)
Taub.
Mimosoideae
Fabaceae
Acacia farnesiana
(L.) Willd.
Mimosoideae
Fabaceae
Country
Chile
Burkina Faso
USA
México
Findings
• Germination of seeds extracted from
dung was 5 times that of seed extracted
directly from pods.
• Due to the larger diameter of their
digestive systems, cattle affect viability
and germination of these four species
less than sheep do.
• The moist, acid environment of dung
may accelerate germination.
• Large herbivores may plan a major
role in long-distance seed dispersal.
• Colonization of pastures and other
disturbed areas by these species results
from bovinochory (a).
• Cattle can disperse large amounts of
viable seed (b)
Table 7.1 Cattle-dispersed woody plants reported in the literature and further findings regarding bovinochory
Study
a) (Parrotta 1992)
b) (Pascacio Damián 2006)
(Razanamandranto et al. 2004)
(Gutiérrez and Armesto 1981)
168
C. L. Miceli-Méndez et al.
Acacia pennatula
Schldl. & Cham.
Benth.
Mimosoideae
Fabaceae
Acacia nilotica L.
Willd. ex Del.
Mimosoideae
Fabaceae
E(d)
D (a–c)
M
D
• The range of this species has increased
due to human activity (d).
• Cattle disperse its seeds (a–e).
• Seedlings grow directly from cattle
dung (b, c)
• Germination is greatest in the
shade (d).
• The size distribution varies in response
to canopy cover and disturbance
intensity (d).
• This species seems to facilitate
colonization by other woody plants (d).
• This is an invasive plant in Australian
rangelands.
• Its invasive power is a product of
prolific seed output and efficient
dispersal by cattle (80% of seeds
defecated intact)
• Only 1% of seed in sheep dung is
intact.
• Models predict patchy invasion related
to microclimate and water sources for
livestock.
Mexico (a, d, e)
Nicaragua (b)
Costa Rica (c)
Australia
(continued)
a) (Greenberg et al. 1997)
b) (Nieto et al. 2001)
c) (Ibrahim and Camargo 2001)
d) (Rubio-Delgado 2001)
e) (Pascacio Damián 2006)
(Radford et al. 2001)
7 Seed Dispersal by Cattle
169
D
E
D
D (a, c)
E (b)
Citrus limetta
RissoRutaceae
Cordia curassavica
(Jacq.) Roem. &
Schult.
Boraginaceae
Dichrostachys
cinerea
(L.) Wight &
Arn.Fabaceae
(see also Acacia
tortilis)
Type of study
Acacia tortilis
(Forsk.)
Mimosoideae
Fabaceae
Dichrostachys
cinerea
Mimosoideae
Fabaceae
Table 7.1 (continued)
Species
Country
Ethiopia
Nicaragua
Cuba (c),
Etiopía (a–b)
Findings
• A majority of seeds of both species,
consumed in their pods by cattle, were
still viable upon defecation (while
seeds of Acacia senegal, A. seyal, and
Balanites aegyptiaca were not).
• Seed of all 5 species collected from the
soil, under tree canopies, and from
barns germinated in the laboratory.
• The abundance and breadth of seed
distribution of A. tortilis and D.
cinerea is a result of bovinochory.
• Seeds of both species were most
numerous in barns.
• Farmers report bovinochory.
• Seedlings sprout from cattle dung.
• This treelet, introduced to Cuba from
Madagascar, is considered an invasive
pest although it has multiple domestic
and industrial uses (a, c).
• Cattle disperse this species (c,b) and
are the principal cause of its invasion
of Cuban pastures (a).
a) (Mabberley 1997)
b) (Argaw et al. 1999)
c) (Méndez Santos and Ramos
Jalil 2004)
(BGF pers. obs.)
(Argaw et al. 1999)
Study
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C. L. Miceli-Méndez et al.
M
D (a–c)
Guazuma ulmifolia
Lam.
Sterculiaceae
E (b)
D (a, c)
Grewia flava
Tiliaceae
Enterolobium
cyclocarpum (Jacq.)
Griseb.
Mimosoideae
Fabaceae
South Africa
Guatemala (b)
Costa Rica (b, c)
• Cattle consume and disperse the seeds,
favoring establishment on degraded
lands (a, b).
• Bovinochory favors germination (a).
• This species was likely dispersed by
extinct Pleistocene megafauna (a).
• This species favors colonization by
other woody plants (c).
Costa Rica (a–c)
• This is an invasive species in
rangelands.
• Cattle distribute the seed into open
grassland.
• Simulations predict that bovinochory
promotes invasion.
• Cattle and horses consume the seeds,
favoring their germination from dung
and soil (a, c).
• In vitro simulations of bovine and
equine digestion found that seeds
defecated within a day of ingestion
survive at rates of 11 and 9%
respectively. Longer retention reduces
viability (b).
(continued)
a) (Janzen 1982b)
b) (Ferguson 2001)
c) (Ibrahim and Camargo 2001)
(Tews et al. 2004)
a) (Janzen and Martin 1982)
b) (Janzen et al. 1985)
c) (Ibrahim and Camargo 2001)
7 Seed Dispersal by Cattle
171
D, E
E
Prosopis caldenia
Burkart
Mimosoideae
Fabaceae
Proposis flexuosa
DC
Mimosoideae
Fabaceae
Type of study
D (a, b)
Pithecellobium
saman
Benth
Mimosoideae
Fabaceae
Table 7.1 (continued)
Species
Country
Australia (a)
Costa Rica (b)
Argentina
Argentina
Findings
• The species is bovinochorous (a, b).
• This is a multipurpose species suitable
for silvopastoral systems (a).
• The fruit is valuable forage.
• The gradual increase in this species’
range is due to bovinochory.
• The species can become a pest as a
result of bovinochory.
• Excreted seed exhibited lower viability
and higher germinability than
uningested seed.
• Bovinochory favors germination more
than dispersal by other wild and
domestic animals (a, b).
• However, seeds obtained directly from
the tree exhibit greater viability than
those defecated by cattle (a).
• Both cattle and Patagonian maras
(Dolichotis patagonum) separate seeds
from their endocarp, favoring
germination (a, b).
a) (Campos and Ojeda 1997)
b) (Villagra et al. 2002)
(Peinetti et al. 1993)
a) (Durr 2001)
b) (Ibrahim and Camargo 2001)
Study
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C. L. Miceli-Méndez et al.
D (a, b, d, e)
Psidium guajava
L. Myrtaceae
E (c)
E (a, b)
D (b, c)
Prosopis juliflora
(Sw.) DC
(synonym of
Prosopis glandulosa
Torr. var.
Glandulosa)
Mimosoideae
Fabaceae
USA (a, b)
Ethiopia (c)
Costa Rica (a–d)
Colombia (e)
• Cattle defecate viable seed (b, c).
• However, seed consumed by sheep and
goats are less viable (b).
• Cattle may accelerate invasion of
rangelands by this species (a).
• Invasion is homogeneous and
independent of ground cover (a).
• Cattle disperse the seeds, favoring
seedling establishment (d,e).
• During peak fruiting, cattle may
consume 49,500 seeds/day (a,b).
• The number of seeds dispersed
diminishes with increasing fruit size
(a,b).
• Simulated bovine digestion did not
affect seed germination (c).
• This species can become a pest in
tropical rangelands as a result of
bovinochory (d).
(continued)
a) (Somarriba 1985a)
b) (Somarriba 1985b)
c) (Somarriba 1986)
d) (Somarriba 1995)
e) (Esquivel Sheik and Calle
Díaz 2002)
a) (Brown and Archer 1999)
b) (Kneuper et al. 2003)
c) (Shiferaw et al. 2004)
7 Seed Dispersal by Cattle
173
E
Rosa woodsii Lindl
Rosaceae
Symphoricarpos
albus (L.)
BlakeCaprifoliaceae
Senna atomaria (L.)
H.S.Irwin &
BarnebyFabaceae
D, E (a, b)
Type of study
Rosa bracteata
Wendl. Rosaceae
Table 7.1 (continued)
Species
USA
Canada
• Cattle defecate viable seed of this
species (a, b).
• Approximately 80% of seeds
consumed were recovered between the
second and third days following
ingestion (a).
• Eight % of defecated seeds were
physically damaged (a).
• Bovinochory favors germination (a).
• Seeds that remain in dung maintain
their viability (a).
• The species is an invasive exotic in SE
USA (a, b).
• Viability following excretion by cattle
was 77.4 and 69.3% respectively.
• Cattle could be managed to
disseminate these species.
• R. woodsii invades rangelands.
• Cattle can disperse large amounts of
viable seed.
Country
Findings
(Pascacio Damián 2006)
(Doucette et al. 2001)
a) (McCully 1951)
b) (Douce et al.)
Study
174
C. L. Miceli-Méndez et al.
7 Seed Dispersal by Cattle
175
16
14
Number of species
12
10
8
6
4
2
0
Boraginaceae
Lamiales
Caprifoliaceae
Dipsacales
Fabaceae
Fabales
Myrtaceae
Myrtales
Rosaceae
Rosales
Rutaceae
Sapindales
Sterculiaceae
Malvales
Tiliaceae
Malvales
Plant family, order
Fig. 7.1 Cattle-dispersed woody plants reported in the literature, by botanical family and order
The fruits of these species are variously dry and indehiscent (n=16), dry and
dehiscent (n=1) or fleshy (n=7); drab (n=13), yellow (n=4), red-orange (n=4) or
white (n=1); and may be fragrant or without odor. The mean fruit length for species
for which data could be found is 10.3 ± 6.6 (standard deviation) cm. The typical
bovinochorous fruit is a legume with dry, elongated, indehiscent, brown-black pods.
Seeds of bovinochorous are also highly variable. Their mean length is
7.0 ± 4.2 mm, with a range of 2–23 mm. Seed weight averages 83 ± 154 mg across
species, ranging from 4 to 700 mg. Appendix 1 provides detailed descriptions of
bovinochorous seeds and fruits.
7.3.3 Where does Bovinochory Occur?
The studies we identify report bovinochory of trees and shrubs in North, Central and
South America, the Caribbean, Africa and Australia. They describe bovinochory
in tropical and temperate savannas and, less frequently, in areas where the native
vegetation is deciduous or semi-deciduous tropical dry forest, tropical rain forest or
pine-oak forest (Table 7.1).
7.3.4 Effects of Bovinochory on Seed Fate
Cattle digestive processes often have minimal effect on seed viability and germination (Peinetti et al. 1993, Razanamandranto et al. 2004). However scarification
176
C. L. Miceli-Méndez et al.
in the bovine digestive tract may soften the seed coat, both diminishing the proportion of viable seed and accelerating germination of surviving seed (Janzen et al.
1985, Doucette et al. 2001). For several species (Acacia dudgeoni, Acacia seyal,
Rosa woodsii, Symphoricarpos albus, Burkea africana, Prosopis africana), viability varies inversely with duration of retention in the digestive tract (Doucette
et al. 2001, Razanamandranto et al. 2004). Small, smooth seeds tend to pass more
rapidly and suffer less damage (Razanamandranto et al. 2004). Cattle digestion augments germination rate more than other physical and chemical treatments for some
species, including Enterolobium cyclocarpum, Psidium guajava, Rosa bracteata,
Rosa woodsii, and Symphoricarpos albus (McCully 1951, Janzen et al. 1985,
Somarriba 1986, Doucette et al. 2001).
The molar mill may also play a role in scarification, as in the case of the Acrocomia aculeata (Jacq.) Lodd. ex Mart. In a series of scarification trials, this palm
germinated only after being chewed and spit out by cattle (Orantes García 1999).
Cattle dung can be a propitious microsite for germination and establishment
of some plants. Dung provides both an acid environment in which further scarification takes place, and abundant nutrients for seedling growth (Somarriba 1986,
Razanamandranto et al. 2004). Species exhibiting increased establishment in dung
microsites include Acacia dudgeoni, Acacia seyal, Burkea africana, Prosopis
africana, Prosopis flexuosa (Villagra et al. 2002, Razanamandranto et al. 2004,
Loth et al. 2005).
7.3.5 Bovinochory in Chiapas
The state of Chiapas, Mexico is an appealing context in which to explore the ecology
and management of bovinochory for several reasons: (1) it affords a broad range of
topographic and climatic conditions, resulting in a correspondingly broad variety
of ecosystems (González-Espinosa et al. 2005); (2) these ecosystems host extraordinary biological diversity, including 9 of the 24 bovinochorous species reported
in Table 7.1; (3) cattle ranching is of economic and social importance throughout
the state, and is practiced under a variety of management scenarios. In this section, we report preliminary results of extensive surveys designed to identify further
bovinochorous species in lowland Chiapas.
The ranchers surveyed identified a total of 17 woody species that germinate from
cattle dung, including representatives of the Anacardiaceae, Burseraceae, Bignoniaceae, Cactaceae, Fabaceae, Malpighiaceae, Moraceae, Myrtaceae, Rutaceae,
Sapotaceae, Sterculiaceae, Ulmaceae and Tiliaceae. Of these, we have thus far
confirmed the presence of viable seed of 12 tree species and one shrub species in
cattle dung. Eight of the trees have been previously reported as bovinochorous.
While our surveys focused on lowland Chiapas, our literature and database
searches found that the 13 confirmed bovinochorous species are distributed across
a broad range of ecosystems (Miranda 1998, Pennington and Sarukhán 1998). Most
are not frost-resistant, but at least three species reach altitudes of 2,000 m or more.
7 Seed Dispersal by Cattle
177
All 13 grow in the dry lowlands and five are also present in the humid tropics.
This suite of species also grows across a spectrum of soil types, from rich soils of
volcanic origin to calcareous soils, to highly-weathered, acidic soils.
7.4 Discussion
Notably, most of the studies we identify in our literature review treat one or a few
bovinochorous species, often as invasive plants that present a challenge to rangeland management (Brown and Acher 1989). Some of these cases present dramatic
demonstrations of the influence of bovinochory on the composition and structure of
plant communities. However, few studies (e.g. Argaw et al. 1999) attempt to identify
the suite of bovinochorous plants in a given system and we found no systematic
studies designed to identify bovinochorous species in the Neotropics. Nonetheless,
our review and findings from Chiapas corroborate Janzen and Martin’s (1982) generalizations regarding the importance of livestock as dispersal agents for neotropical
trees and shrubs. We are able to draw some useful conclusions regarding characteristics of bovinochorous species, when and how bovinochory occurs, and how the
phenomenon might be exploited for both enhancement of tropical rangelands and
for catalyzing ecological restoration.
7.4.1 The Ecology of Bovinochory
As Janzen and Martin (1982) predicted, bovinochorous fruits are highly variable
in size, color, morphology, odor, nutrient composition, seed size, etc., defying definition of a consistent bovinochorous dispersal syndrome (Howe 1985). Indeed, as
Bruun and Poschlod (2006) note, cattle and other ungulates may be important vectors for many species without obvious dispersal adaptations. One generalization that
does seem to hold true is that the seeds are hard and smooth, protected within a hard,
smooth nut, or else very small (Janzen and Martin 1982).
While legumes with indehiscent pods are disproportionately favored by cattle,
bovinochory manifests itself across numerous plant families (Fig. 7.1). Janzen and
Martin (1982) mention species from 14 plant families in Pacific lowland, deciduous
forest of Costa Rica that are dispersed by livestock and/or may have been dispersed
by extinct megafauna. Eleven of these families (Anacardiaceae, Arecaceae, Bignoniaceae, Bromeliaceae, Fabaceae, Malpighiaceae, Moraceae, Myrtaceae, Sapotaceae,
Sterculiaceae, Tiliaceae) along with three additional families (Burseraceae, Cactaceae, Rutaceae) are represented among the species we have thus far identified in
Chiapas as bovinochorous.
Both the literature and our preliminary data suggest that bovinochory is more
prevalent in dry forest areas than in the humid tropics. This observation is consistent
with the hypothesis that bovinochory is a characteristic inherited from coevolution
178
C. L. Miceli-Méndez et al.
with large, extinct herbivores that maintained, savanna-like vegetation (Janzen and
Martin 1982). Such plants must have been able to survive in the relatively dry microclimate of these open areas and could be expected to be most abundant in today’s
dry forests.
Bovinochory has been reasonably well documented in the range management
context (e.g. McCully 1951, Brown and Acher 1989, Cox et al. 1993, Auman et al.
1998, Doucette et al. 2001, Bruun and Fritzboger 2002, Edwards and Younger
2006). However, its influence upon successional patterns in abandoned neotropical pastures seems to have been largely disregarded by ecologists since Janzen and
Martin (1982), perhaps because of the more obvious association between cattle and
deforestation (Hecht 1993).
Intriguingly, we and the ranchers we have interviewed have observed latency
in seeds of cattle-dispersed taxa, including Acacia, spp. and Guazuma ulmifolia,
species that are often important early-successional colonizers (e.g. Ferguson et al.
2003). Thus these species may germinate not from post-abandonment seed rain, but
from a seed bank that accumulates in pastures during years of use. Draft animals and
livestock grazing on crop stubble may also contribute to seed banks and successional
patterns in croplands.
Furthermore, the structural diversity and resources provided by bovinochores that
establish in active pasture doubtless also play a role in post-abandonment succession. In open vegetation, structural diversity is vital to maintaining seed dispersal
by wild animals, and isolated trees offer key microsites for establishment of other
woody species (Guevara et al. 1986, Vieira et al. 1994, da Silva et al. 1996, Ferguson
2001). In the Central Valleys of Chiapas, bovinochorous trees including Enterolobium cyclocarpum, Pithecelobium dulce, Guazuma ulmifolia, Acacia pennatula,
and Acacia farnesiana are salient features of the landscape. These trees can offer
important habitat for birds (e.g. Greenberg et al. 1997), contributing to successional
dynamics.
7.4.2 Management Potential of Bovinochory in the Neotropics
Strategies that contribute simultaneously to rural livelihoods and biological conservation are urgently needed throughout the Neotropics (Sánchez et al. 2000,
Angelsen and Kaimowitz 2001). Our findings with respect to the diversity of
Neotropical bovinochorous trees and shrubs, the range of ecosystems in which
they grow and the apparent importance of bovinochory in structuring vegetation
in grazing land in the dry tropics all highlight the potential for management of this
relationship. Establishment of trees via bovinochory holds particular promise for
both silvopastoral systems and ecological restoration. This strategy could be used to
plant trees at low cost by exploiting relationships and resources already familiar to
rural people.
7 Seed Dispersal by Cattle
179
7.4.2.1 Silvopastoral Systems
Silvopastoral systems rely on trees, particularly multiple-use trees to improve the
productivity of grazing land (Sánchez 1998). Our searches found multiple uses for
all of the bovinochorous trees and shrubs we report from Chiapas. In addition to
forage, these plants offer wood for diverse applications, extracts for tanning, shade
for coffee or cattle, medicine, perfume, soil retention, wind breaks and live fences
among other products and services (Miranda 1998).
Availability of quality forage in the dry season is a major determinant of the productivity of Neotropical rangelands (Botero and Russo 1998). Most bovinochorous
species fruit during the dry season when grasses are less abundant and cattle seek
alternative forage. Their fruits and seeds generally possess high concentrations of
lipids, proteins and soluble carbohydrates that both attract and nourish livestock
(Chargoy Zamora 1988, Razanamandranto et al. 2004). We have identified several
ranchers in the Central Valleys who use trees dispersed by their cattle as key dry
season forage. One of these ranchers relies on bovinochory to establish Acacia
pennatula, thinning the trees as necessary to maintain a favorable balance between
trees and grass. Two more ranchers use cattle-dispersed Guazuma ulmifolia, and
one of them spread G. ulmifolia seed across his ranch intentionally by feeding
its fruits to his cows. These ranchers report that within 3 months of germination,
G. ulmifolia can withstand frequent browsing. In addition to high quality forage,
G. ulmifolia provides firewood, wood, and medicine (CATIE 1991, Miranda 1998,
Pennington and Sarukhán 1998). Other bovinochorous species that may represent
considerable forage resources in Chiapas include Acacia farnesiana, Enterolobium cyclocarpum, Manilkara achras, Nopalea dejecta, Parmentiera aculeata,
Pithecellobium spp., Prosopis juliflora, Psidium guajava, Senna atomaria y
Senna spectabilis.
7.4.2.2 Restoration
Where ecological restoration is the primary management objective, bovinochory
may catalyze successional processes by increasing woody plant seed rain in open
areas and diversifying microsites for establishment. Bovinochorous colonizers help
re-establish the vegetation structure and food resources that attract wild seed vectors, which in turn will diversify the seed rain. Seeds deposited by wild dispersers
will find improved microsites for germination and recruitment under bovinochorous
plants (e.g. Ferguson 2001). Thus two of the most frequent barriers to forest regeneration in tropical pastures, lack of seed rain and of safe sites for establishment of
woody plants (Holl et al. 2000), may be overcome by bovinochory.
Several bovinochores hold considerable potential as successional nuclei. All colonize open areas and most tolerate poor soils (CATIE 1991, Miranda 1998, Pennington and Sarukhán 1998). Many such early colonizers improve soil conditions
or microclimate or attract seed-dispersing wildlife (Vieira et al. 1994, Meli 2003).
Some, such as A. pennatula, G. ulmifolia and P. dulce are of particular importance
180
C. L. Miceli-Méndez et al.
for wildlife conservation (Greenberg et al. 1997, Leon-Cortes et al. 2004, B.G.F.
pers ob.). However some bovinochorous species, including Acacia spp., D. cinerea,
Prosopis sp. and P. guajava form dense thickets that could impede colonization by
other species or restrict cattle movement and the productivity of rangelands (Peinetti
et al. 1993, Somarriba 1995, Méndez Santos and Ramos Jalil 2004).
7.4.3 Research Priorities
Despite the success of some of the ranchers we interviewed at managing
bovinochory, broad and efficient exploitation of this phenomenon will require
deeper understanding of its ecology and natural history. Here we outline some
pertinent areas of research.
The range of plant species susceptible to bovinochory has yet to be studied in
most places. Interviews with ranchers, observations of germination in dung pats,
controlled germination of seed from dung and simple experiments involving the
addition of seeds or fruits to livestock diets may all quickly yield this kind of
information.
Differences among livestock species and even breeds with respect to their capacity for seed dispersal should also be explored. We have focused on cattle as the most
ubiquitous livestock in neotropical landscapes. However horses are also effective
seed dispersers and they use different habitat and graze different species than do
cattle (Cosyns et al. 2005). Sheep, while less likely than cattle to excrete whole
seeds (Razanamandranto et al. 2004), are locally important in tropical highlands
(Jiménez Ferrer et al. 2003), while tropical sheep are increasingly grazed in the
lowlands (Nuncio-Ochoa et al. 2001).
The distribution and characteristics of various microsites arising from bovinochory influence seed and seedling fate. Patterns of dung deposition correspond to
herbivores’ habitat preferences and can be influenced by management (Wiegand
et al. 1999, Cosyns et al. 2005). Dung pats can create short-lived gaps in dominant
vegetation (Cosyns et al. 2006, Kohler et al. 2006), offer an acid environment that
scarifies seed, and provide nutrients that favor seedling establishment (Razanamandranto et al. 2004). Dung beetles may act as secondary dispersers, diversifying the
range of microsites encountered by bovinochores, but may also bury seed well below
the soil surface where establishment is unlikely (Díaz Valdiviezo 2005). Microsite
conditions arising from bovinochory and secondary dispersal are likely to favor
some species in some environments, while filtering out others.
Changes in management of grazing lands bovinochory can alter the significance
of bovinochory over a range of spatial scales. In northern Europe, for example,
a reduction in the number of free-ranging livestock, increased fencing of pastures and less common long-distance seasonal movement of livestock have all
contributed to a decline in seed dispersal in modern times (Bruun and Fritzboger
2002). Recent changes in Mexico’s land reform law permitting division and sale
of ejido land may similarly alter patterns bovinochory in that country. At the
same time, enforcement of bans on burning seems to be favoring increased tree
7 Seed Dispersal by Cattle
181
cover, particularly of bovinochores, in some ranching-dominated landscapes of
the Central Valleys of Chiapas.
Grazing pressure likely interacts with bovinochory to determine germination and
establishment rates. Intriguingly, because cattle generally prefer herbaceous forage, woody plant establishment can be greater at low livestock densities than in
the absence of grazing where grasses are unchecked (Posada et al. 2000, Tobler
et al. 2003). Thus cattle may simultaneously increase seed rain and lessen competition faced by establishing seedlings. However the form of the relationship between
grazing pressure and woody plant establishment is little known as is the effect of
timing of grazing in relation to germination. Further complicating this relationship,
well-timed grazing may reduce fuel loads and thus the frequency and severity of fire
during early succession (Fonseca and Beita In press).
For purposes of both restoration and silvopastoral systems, we need a better
understanding of the interactions of bovinochorous shrubs and trees with other
species. Which bovinochores permit or even favor grass growth beneath their
crowns, and which grasses are most compatible with them (Hernández-Daumas
2000)? Conversely, which create shade levels or other microsite conditions that favor
establishment of more woody plants (Vieira et al. 1994)? Which create habitat for
wildlife species of interest (e.g. Greenberg et al. 1997), particularly for wild seed
dispersers (Vieira et al. 1994)? Which species, under what conditions, are likely
to form dense stands, and which maintain wider spacing that permits continued
grazing and/or establishment of other species (Peinetti et al. 1993)? Does early
dominance by bovinochores cause a “founder effect” that influences successional
trajectory over the long term (Drake 1991)?
7.5 Conclusions
Cattle ranching is justifiably viewed as a major agent of neotropical deforestation
(e.g. Hecht 1993). However under some conditions, seed dispersal and grazing by
cattle can overcome major barriers to reestablishment of woody cover in open areas,
catalyzing successional processes. Exploitation of this relationship holds promise
as a cheap, efficient tool for both silvopastoral systems and ecological restoration.
Further research will help realize the full potential of this strategy across a range of
tropical ecosystems.
Acknowledgment We thank Mario Alberto López Miceli, Ivonne Anahí López Miceli, María
Guadalupe Pascacio Damián, María Lesvia Dominguez Morales, Henry Castañeda Ocaña, Miguel
Martínez Icó and Juan López Méndez for valuable support in the field, laboratory, greenhouse
and herbarium. We are also grateful to Lorena Soto Pinto and René Pinto Ruíz for comments
on an early draft. This research was supported by the Fondo sectorial CONACyT-SEMARNAP
through the project “Uso sustentable de los Recursos Naturales en la Frontera sur de México”,
SEMARNAT-2002-C01-1109.
Dry indehiscent,
sessile or on
short stem,
cylindricalfusiform
Dry indehiscent,
woody
Acacia farnesiana (L.)
Willd. Tree/Shrub
Acacia nilotica L.
Willd. ex Del. Tree
Dry indehiscent, Brown
oblong, flattened,
glabrous
Acacia dudgeoni
Craib ex Hall
Tree/Shrub
Elliptical,
reniform
Ellipsoid,
smooth,
compressed
Brown to 12.5 × 1.8
black
NA
Compressed,
smooth,
ellispsoid
Shape
Brown to 5.5 × 1.3
black
5.5 × 2
Brown to 5 × 2
black,
shiny.
Dry indehiscent,
thick, oblong
Acacia caven (Mol.)
Molina Shrub/Tree
Color
Type
(Moreno 1984)
Species and growth
form
Size
Length ×
width ×
thickness (cm)
95.2
83
8.54
115.1
Mass (mg)
6.5 × 4.8
6.6 × 4.7
8.2 × 8.2 × 1.8
7.5
Size Length ×
width ×
thickness (mm)
Appendix 1 Characteristics of fruits and seeds of woody plants cited in the literature as cattle-dispersed
Fruit
Seed
Appendix
Source
Mar.–June
Jan.–May
Dec.
(NewCROP, Sánchez de
Lorenzo-Cáceres,
Winrock International,
Radford et al. 2001)
(PROCYMAF, Sánchez
de Lorenzo-Cáceres,
Standley et al.
1946–1975, Parrotta
1992, Vázquez-Yanes
et al. 1999, Cordero and
Boshier 2003, Pascacio
Damián 2006)
(CIAT/FAO,
Razanamandranto et al.
2004).
Dec.–April (CIAT/FAO, Sánchez de
Lorenzo-Cáceres,
Piaggio 2003)
Fruiting
season
182
C. L. Miceli-Méndez et al.
Dry indehiscent,
woody, sessile,
turgid
Dry indehiscent,
glabrous or
pubescent
Dry dehiscente,
falcada, glabra
Dry indehiscent,
pubescent,
elliptical
Globose with
fleshy pericarp
Fleshy, exocarp
thin and smooth
Fleshy
Acacia pennatula
Schldl. & Cham. Benth.
Tree
Acacia tortilis (Forsk.).
Tree/Shrub
Acacia seyal Del.
Tree
Burkea africana
Hook. F.Tree
Brosimum alicastrum
SwTree
Citrus limetta
Risso. Tree
Cordia curassavica
Shrub
NA
6 × 2.5
170
NA
Ovoid
1750
10
3.83
Jan.
Mar.–May
Apr.–Jun.
Apr.–Nov.
11.8 × 15 × 15
11.6 × 5.8 × 4.4
3.7 × 4.2 × 5.9
April
Oct.–Dec.
Nov.–Jan.
8.1 × 6.1 × 2.5
6.6 × 4.2 × 3.1
33.33 5.5
60.60 7 × 4.3 × 3.5
Ovoid
Spherical with
flattened tips,
papery testa
Elliptical,
compressed,
wrinkled,
funiculate,
elongate
NA
10 × 0.7
13.5 (length)
× 0.7(thickness)
Globose,
elliptical, hard,
smooth
12.5 × 2.1 ×
0.8
Yellowish 2.3
green,
orangish
red
Pale
5.8 × 6.1
yellow
Red
3.7 × 4.2 ×
5.9
NA
Brown
Brown
Brown
(continued)
(Miranda 1998, Miceli
Méndez In prep.)
(Missouri Botanical
Garden)
(Pennington and
Sarukhán 1998) Orantes
et al. 2007
(CIAT/FAO,
Razanamandranto et al.
2004)
(Sánchez de
Lorenzo-Cáceres,
Razanamandranto et al.
2004)
(FAO, Sánchez de
Lorenzo-Cáceres,
Winrock International,
Duke 1983, Argaw et al.
1999, Loth et al. 2005)
(Standley et al.
1946–1975, Missouri
Botanical Garden,
Martínez 1992, Miranda
1998, Rubio-Delgado
2001, Pascacio Damián
2006)
7 Seed Dispersal by Cattle
183
Fleshy
Dry indehiscent,
woody, ellipsoid
Guazuma ulmifolia
Lam. Tree
NA
5.5 × 7.1 ×
1.2
6 × 1.1
Size
Length ×
width ×
thickness (cm)
Black
2.2 × 1.8
with
small
protuberances,
hard
NA
Brown
Dry indehiscent,
woody, curved
Grewia flava
Shrub
Brown
Dry indehiscent,
glabrous, oblong
Dichrostachys cinerea
(L.)
Wight & Arn.
Shrub/Tree
Enterolobium
cyclocarpum (Jacq.)
Griseb. Tree
Color
Type
(Moreno 1984)
Fruit
Species and growth
form
Appendix 1 (continued)
Spheroid
NA
Biconvex,
obovate,
elliptical,
subcircular
Ovoid, flattened
and hard
Shape
Seed
6.1
NA
700
29.41
1.8 (diameter)
NA
23 × 15
4×3
Size Length ×
width ×
Mass(mg) thickness (mm)
End of
rainy
season
Sep.– Apr.
Mar.–May
Aug.
Fruiting
season
Source
(Missouri Botanical
Garden, Janzen 1982b,
Miranda 1998,
Pennington and
Sarukhán 1998,
Villatoro 1998,
Vázquez-Yanes et al.
1999)
(Tews et al. 2004)
(Standley et al.
1946–1975, Janzen
1982a, Janzen et al.
1985, Miranda 1998,
Pennington and
Sarukhán 1998)
(FAO Feed Resources
Group, ISSG)
184
C. L. Miceli-Méndez et al.
Dry indehiscent,
soft exocarp, tic
mesocarp
Dry indehiscent,
elongate,
compressed,
glabrous
Proposis flexuosa
DC.Tree
Prosopis juliflora (Sw.)
DC
Tree
Dark
15
yellow
with
violet
spots
Straw16.5 × 1.0
colored
with
violet to
black
splotches
Yellow to 18.5 × 1.3
violet
Dry indehiscent
Round to oval
Ovoid
Ovoid
NA
15 × 3
Dark red
Oblong
Thin, hard
17.5 × 1.2
Dry indehiscent,
cylindrical
Brown
Prosopis africana
(Guill.,
Perrott. & Rich.)
Taub. Tree
Prosopis caldenia
Burkart. Tree
Yellowish 15 × 6.5
green
Fleshy, elongate
with longitudinal
furrows
Dry indehiscent,
linear,
compressed
Parmentiera aculeata
(Kunth)
Seemann/Tree
Pithecellobium saman
Benth. Tree
NA
36.36
32.35
24.09
15.71
Feb.–Mar.
Dec.–Mar.
Dec.–Jan.
Jan.–May
6.1 × 3.7 × 2.2
6.8 × 4.5 × 2.3
7.5 × 5.3 × 3
Feb.–May
Jan.–Dec.
8.7 × 5.8 × 4.0
175.43 6.5
NA
(continued)
(CIAT/FAO, Martínez
1992, Moreno Casasola
1996, Miranda 1998,
Shiferaw et al. 2004)
(CIAT/FAO, Campos
and Ojeda 1997,
Villagra et al. 2002,
Cosiansi et al. 2003)
(CIAT/FAO, Peinetti
et al. 1993, Cosiansi
et al. 2003)
(PROCYMAF, Miranda
1998, Pennington and
Sarukhán 1998)
(Standley et al.
1946–1975) (Sánchez
de Lorenzo-Cáceres,
Miranda 1998, Zamora
et al. 2001)
(CIAT/FAO,
Razanamandranto et al.
2004)
7 Seed Dispersal by Cattle
185
Color
Size
Length ×
width ×
thickness (cm)
Symphoricarpus albus
(L.) Blake
Shrub
Fleshy, globose
White
NA
Yellowish 8 (diameter)
cream to
pink,
fragrant,
sweetsour
Rosa bracteata
Fleshy, globose
Orange
NA
Wendl. Shrub
to red
Rosa woodsii Lindl
Fleshy, globose
Red
0.9
Shrub
to elipsoid
(diameter)
Senna atomaria (L.) H. Dry indehiscent, Dark
28.5 × 1
S. Irwin & BarnebyTree linear-cylindrical brown
Type
(Moreno 1984)
Fruit
Psidium guajava L.Tree Fleshy, globose
to ovoid
Species and growth
form
Appendix 1 (continued)
13.6
36.36
Oval, hard
Hard, smooth
4.6
NA
NA
NA
NA
NA
Aug.
Feb.–Jun.
NA
4.5 × 2.5
5.2 × 2.8
3 × 1.2
Jul.–Feb.
Fruiting
season
NA
4
Size Length ×
width ×
Mass(mg) thickness (mm)
Round to
triangular
Shape
Seed
(Douce, Moorhead and
Bargeron)
(Johnson and Hoagland,
Doucette et al. 2001)
(Palma and Román
2000, Cordero and
Boshier 2003, Pascacio
Damián 2006)
(USDA, Doucette et al.
2001)
(Somarriba 1985b,
1986, Moreno Casasola
1996, Pennington and
Sarukhán 1998)
Source
186
C. L. Miceli-Méndez et al.
7 Seed Dispersal by Cattle
187
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Chapter 8
The Roles of Disperser Behavior
and Physical Habitat Structure in Regeneration
of Post-Agricultural Fields
K. Greg Murray, Kathy Winnett-Murray, Jason Roberts,
Katherine Horjus, William A. Haber, Willow Zuchowski,
Mark Kuhlmann and Tammy M. Long-Robinson
8.1 Introduction
Since the 1980s, the plight of tropical forests and the factors that drive their conversion to human-dominated uses have received widespread attention in both the
scientific and popular media. Worldwide, conversion to croplands and logging for
timber and fuelwood have led the forces driving deforestation, but in Latin America
creation of cattle pastures has been the major cause (Myers 1980). In Amazonia
alone, more than 20 million hectares of forest had been converted to cattle pasture
by the 1990s (Nepstad et al. 1996).
Despite their prevalence in many parts of the neotropics and their importance
in local economies, productivity of cattle pastures often degrades rapidly through
a combination of poor management, high pest pressures, and the loss of nutrients
from what are often highly weathered soils (Serrão and Toledo 1990). The ultimate
fate of most degraded pastures is abandonment, after which regeneration to forest
via natural succession is possible. Forest regeneration on such abandoned pastures
and other agricultural lands has received considerable attention (e.g., Uhl et al.1988
Nepstad et al. 1991, 1996, Buschbacher et al. 1992, Lugo 1992, Gerhardt 1993,
Aide and Cavelier 1994, Brown and Lugo 1994, Kuusipalo et al. 1995, Sarmiento
1997a, Duncan and Chapman 1999, 2002 Holl 1999, Wijdeven and Kuzee 2000,
Myster 2004b), much of it centered on the factors that limit the rate of regeneration
and determine which forest species colonize. Soil degradation, competition with
pasture grasses, seasonal drought, and herbivory on seeds and seedlings have been
implicated as significant limiting factors in a number of studies (e.g., Uhl 1987, Holl
et al. 2000, Zimmerman et al. 2000, Myster 2004a). Even more frequently, regeneration is thought to be limited by the availability of seeds dispersed into abandoned
agricultural lands from the surrounding forest. Since most woody tropical species
are adapted for dispersal by vertebrates (Howe and Smallwood 1982; Murray et al.
2000), many investigators have focused on the roles of seed dispersing animals and
on those of remnant trees (i.e., those left standing when the pasture was created) as
192
R. W. Myster, Post-Agricultural Succession in the Neotropics.
C Springer 2008
8 Roles of Disperser Behavior and Physical Habitat Structure
193
foci for the recruitment of forest plants (e.g., Guevara et al. 1986, 1992, Janzen 1988,
Nepstad et al. 1990, Uhl et al 1990, Guevara and Laborde 1993, Vieira et al. 1994,
Green 1995, Cardoso da Silva et al. 1996, Sarmiento 1997b, Wunderle 1997, Holl
1998, Harvey and Haber 1999, Otero-Arnaiz et al. 1999, Toh et al. 1999, Zahawi
and Augspurger 1999, Harvey 2000, Harvey et al. 2000, 2007, Holl et al. 2000,
Slocum 2001). Most studies have found that remnant trees or even artificial structures (e.g., Holl 1998) attract greater numbers of animal-dispersed seeds than adjacent sites without overhanging vegetation, presumably because fruit-eating animals
are attracted to them. Similar effects of natural and artificial structures have been
demonstrated at temperate latitudes as well (McDonnell and Stiles 1983, McClanahan and Wolfe 1993, Robinson and Handel 1993, 2000).
Given that secondary forests derived from former agricultural lands are likely to
be increasingly important repositories of tropical biodiversity, understanding how
recruitment foci operate will be a useful aid to understanding and directing the rate
and trajectory of forest regeneration on such sites.
Despite the considerable number of recent studies on tropical pasture regeneration, almost none have followed the process over an extended period of time (but
see Zahawi and Augspurger 1999, Aide et al. 2000, Finegan and Delgado 2000).
In this chapter we address the role of physical habitat structure and seed dispersal
in the regeneration process, specifically the role of remnant trees as recruitment
foci. Using data from a montane site in northwestern Costa Rica, we first test
the hypothesis that colonization of abandoned pastures is enhanced by the presence of remnant trees, and that the effect acts through influencing the behavior
of fruit-eating animals rather than through shading effects. Second, we test the
hypothesis that remnant trees that themselves produce fleshy fruits eaten by animals are more effective as recruitment foci for animal-dispersed plants than are
trees that produce wind-dispersed seeds. Finally, we examine post-abandonment
regeneration patterns over a period of 30 years to determine whether the effects
of remnant trees on density and species composition in a regenerating forest persist
over time.
8.2 Study Site and Methods
The Monteverde area (Provincia de Puntarenas) lies on a gently sloping plateau
in the Cordillera de Tilarán in northwestern Costa Rica (10◦18 N, 84◦ 48 W).
Vegetation in the area is loosely termed tropical montane cloud forest, but includes
life zones (Holdridge life zone classification system, Holdridge 1967) from tropical
moist forest on the Pacific slope through lower montane rainforest along the continental divide, and down to tropical wet forest on the Caribbean slope. Climate and
vegetation of the area are treated in detail by Clark et al. (2000) and Haber (2000),
respectively.
Settled by Costa Rican farmers in the 1920s and 1930s, much of the area was substantially deforested for pasture and subsistence farming by the 1950s. Large-scale
194
K. G. Murray et al.
dairy farming began in 1951 after the arrival of North American settlers. Despite
the continuing importance of dairy farming in the area, many pastures have been
abandoned since the 1960s, and second-growth forests now occupy several areas
that were active pastures as late as the 1950s and 1960s.
In order to examine the influence of remnant trees on patterns of pasture regeneration, we collected data on woody plant recruitment into two abandoned pastures
on the Pacific slope, both of which were created by clearing of lower montane
wet forest. Pasture A was 1.1 ha in area and was located at 1,525 meters elevation
approximately 0.5 km west of the entrance to the Monteverde Cloud Forest Reserve
(Fig. 8.1). It was surrounded on two sides by primary forest and on one side by
approximately 40-year-old secondary forest. The remaining side was bordered by a
dirt road and another pasture of similar age and structure. It was abandoned (cattle
removed permanently) in 1984 after approximately 32 years of use. We censused
woody recruits into the pasture in plots beneath remnant trees and in plots with no
overhanging vegetation in 1987 and again in 1998 (hence, 3 years and 14 years
post-abandonment).
Pasture B was located at approximately 1,500 meters elevation just within the
Monteverde Cloud Forest Reserve, approximately 500 meters from pasture A. It was
abandoned in approximately 1968 after approximately 20 years of use, and had a
closed canopy dominated by the early-successional species Conostegia oerstediana
(Melastomataceae) and Hampea appendiculata (Malvaceae) when it was censused
in 1998 (hence, approximately 30 years post-abandonment). Including Pasture B in
our design allowed us to determine whether the successional trajectories suggested
by the first 14 years of regeneration are likely to persist over even longer periods
of time. Although this design potentially confounds site effects with temporal ones,
we feel that the proximity of the two pastures (0.5 km) and the similarity in species
composition of forest surrounding them minimizes any site differences.
Census plots in pasture A centered on the boles of 77 remnant trees and 7 dead
snags of canopy tree size, and on 35 sites without overhanging vegetation. Remnant
trees and dead snags were those that had been left standing when the pasture was
created, and the sites without overhanging vegetation (“open-sky” plots) were scattered systematically throughout the remainder of the pasture. In the 1987 census,
sample plots were circular, with 2 m radii ( = 12.57 m2 ). In 1998, we censused only
one randomly chosen quadrant of each such plot ( = 3.14 m2 ) due to higher overall
recruit densities. Censuses consisted simply of counting and identifying all woody
seedlings and saplings within the plot.
Pasture B was dominated by a closed canopy of even-aged early-successional
species (as noted above) when we censused it in 1998, but we were able to locate and
identify the trees that were remnants of the original forest at the time of abandonment as the only large late-successional species within the area. “Open sky” plots
for this pasture were paired with each of the remnant trees sampled, and located 15
meters from the bole of the remnant tree in a randomly chosen compass direction.
Plots were established and censused in this pasture in the same way as in pasture A
in 1998.
Univariate statistical analyses reported here follow Sokal and Rohlf (1995) and
were performed with SYSTAT 8.0.3 (Systat software, Inc., 1735 Technology Drive,
8 Roles of Disperser Behavior and Physical Habitat Structure
195
Fig. 8.1 Map of the Monteverde area, showing the locations of pastures A and B
Suite 430, San Jose, CA 95110). Ordination analyses (detrended correspondence
analysis; Gauch 1982) were performed with DECORANA (Microcomputer Power,
111 Clover Lane, Ithaca, NY 14850 USA).
196
K. G. Murray et al.
8.3 Results
8.3.1 Remnant Tree Species Present in Pastures
Of the 77 remnant trees in pasture A in 1987, 73 were identified at least to genus,
and represented 26 species in 19 families (Table 8.1). Of these, 62 produce fruits
adapted for dispersal by animals and 11 are adapted for dispersal by wind. Between
the 1987 and 1998 censuses, some of the remnant trees in pasture A fell or were
removed by the landowner. In addition, some of the original open sky sites were
destroyed by the construction of pathways, such that we were able to include only
26 remnant tree and 16 open sky sites in the 1998 census.
We located and identified 11 remnant trees (all animal-dispersed) in pasture B,
representing 7 species in 7 families (Table 8.1).
Table 8.1 Tree species present as pasture remnants. All remnants were sampled in pasture A,
but in pasture B only those that were much larger than post-abandonment recruits were sampled.
Names and dispersal modes follow Haber (2000); bd = bird, bt = bat, am = arboreal mammal, and
tm = terrestrial mammal, wd = wind
Dispersal
Family
Species
mode
# in pasture A # in pasture B
Annonaceae
Bombacaceae
Boraginaceae
Celastraceae
Clusiaceae
Euphorbiaceae
Guatteria oliviformis
Quararibea costaricensis
Cordia cymosa
Maytenus reconditus
Chrysochlamys psychotriifolia
Sapium glandulosum
Hyeronima oblonga
Flacourtiaceae Xylosma oligandra
Hasseltia floribunda
Lauraceae
Beilschmiedia costaricensis
Ocotea tonduzii
Ocotea whitei
Nectandra smithii
Malvaceae
Hampea appendiculata
Meliaceae
Guarea tonduzii
Cedrela tonduzii
Trichilia havanensis
Moraceae
Ficus sp.
Pseudolmedia mollis
Myrsinaceae
Ardisia palmana
Myrtaceae
Eugenia sp.
Proteaceae
Roupala glaberrima
Rhizophoraceae Cassipourea elliptica
Rubiaceae
Chione sylvicola
Rutaceae
Stauranthus perforatus
Zanthoxylum sp.
Sabiaceae
Meliosma idiopoda
Sapotaceae
Pouteria sp.
Tiliaceae
Mortoniodendron guatemalense
Verbenaceae
Citharexylum caudatum
–
Unidentified species / dead snags
bd
am
bd
bd
bd
bdx
bd
bd
bd
bd
bd
bd
bd
bd
bd
wd
bd
bd/bt/am
bd
bd
bd or bt
wd
bd
bd
bd
bd
bd
am, tm
bd
bd
–
1
1
2
1
5
1
1
1
3
6
4
1
2
2
1
3
1
3
1
1
2
1
1
1
11
1
2
8
1
4
5
1
2
11
8 Roles of Disperser Behavior and Physical Habitat Structure
197
8.3.2 Remnant Trees as Early Recruitment Foci
Remnant trees were exceedingly important as foci for early recruitment of woody
plants. Three years after abandonment of pasture A, we found a total of 2,039
allospecific recruits (i.e., of species other than the overhanging remnant tree), representing 27 species and at least 19 families, within the 73 identified remnant tree plots
(Fig. 8.2; Table 8.2). The 10 most common recruit taxa were Eugenia guatemalensis
and Myrcianthes fragrans (we were unable to distinguish reliably between seedlings
of these in 1987), Conostegia oerstediana, Sapium glandulosum, Ocotea whitei,
Hampea appendiculata, Viburnum costaricanum, Stauranthus perforatus, Ocotea
tonduzii, Myrsine coriacea, and Citharexylum caudatum. In contrast, we found only
34 woody recruits, representing 11 species in 8 families, in the 35 plots without
overhanging vegetation (Fig. 8.2). The difference in mean density of recruits in
the two plot types is highly significant (Mann-Whitney U = 59.5; p < 0.0001). The
great majority of recruits were of species dispersed by birds or bats, although two
species (Persea americana and Pouteria fossicola) dispersed primarily by arboreal
mammals (e.g., monkeys, Kinkajous and Olingos) were also well-represented. Few
recruits of wind-dispersed species were present, despite the fact that the single most
common remnant tree species in pasture A (Roupala glaberrima) produces winddispersed seeds.
It might be argued that recruits would occur most commonly beneath remnant
trees without any dispersal at all, since many seeds simply fall beneath parent plants.
This source of bias was not important in our study, however: 12 of the 24 recruit
species identified 3 years post-abandonment were not even present as remnant trees
in the pasture, and therefore must have come from the surrounding forest. Moreover, data reported above do not include 344 recruits that were censused beneath
remnant trees of the same species. Our approach therefore conservatively estimates
the effects of remnant trees on recruitment patterns.
Clumping of recruits beneath remnant trees might also result simply from higher
survival of seeds or seedlings there, e.g., due to favorable microclimate conditions. While we could not address that possibility directly, we did so indirectly
Fig. 8.2 Mean density of
allospecific wind- and
animal-dispersed recruits in
73 remnant tree and 35 “open
sky” plots in pasture A, 3
years post-abandonment.
Error bars represent one
standard error of the mean
Recruits / m2
3
2
1
DISPERSAL MODE
animal
wind
0
open
remnant
Plot Type
Actinidiaceae
Apocynaceae
Aquifoliaceae
Araliaceae
Malpighiaceae
Lauraceae
Fabaceae
Flacourtiaceae
Cunoniaceae
Elaeocarpaceae
Euphorbiaceae
Capparidaceae
Caprifoliaceae
Celastraceae
Asteraceae
Species
Saurauia montana
Tabernaemontana longipes
Ilex lamprophylla
Schefflera rodrigueziana
Oreopanax oerstedianus
Oreopanax xalapensis
Neomirandea angularis
Koanophyllon pittieri
Capparis discolor
Viburnum costaricanum
Maytenus reconditus
Perrottetia longistylis
Weinmannia wercklei
Sloanea brenesii
Alchornea latifolia
Sapium glandulosum
Lonchocarpus sp. (Haber 11102)
Casearia tacanensis
Hasseltia floribunda
Xylosma oligandra
Beilschmiedia costaricensis
Cinnamomum “paraplitrinerve”
Ocotea meziana
Ocotea tonduzii
Ocotea whitei
Nectandra smithii
Pleurothyrium palmanum
Persea americana
Bunchosia veluticarpa
Family
bd, am
bd
bd
bd
bd
bd
wd
wd
am
bd
bd
bd
wd
bd
bd
bd
wd
am
bd
bd
bd
bd
bd
bd
bd
bd
bd
am
bd
Dispersal mode
early
late
late
early
late
early
early
late
late
early
late
late
late
late
early
early
late
late
late
late
late
late
late
late
late
late
late
late
late
Successional status
x
x
x
x
x
x
x
x
x
Pasture A, 3 years
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
Pasture A, 14 years
x
x
x
x
x
x
x
x
x
x
x
x
Pasture B, 30 years
Table 8.2 Woody recruits in pasture A (3 and 14 years post-abandonment) and pasture B (30 years) in 1987 and 1998. Dispersal mode and successional status
(early vs. late) is based on personal observation and Haber (2000)
198
K. G. Murray et al.
Sapindaceae
Sapotaceae
Sabiaceae
Rutaceae
Rubiaceae
Oleaceae
Proteaceae
Rhizophoraceae
Rosaceae
Myrtaceae
Myrsinaceae
Moraceae
Melastomataceae
Meliaceae
Malvaceae
Hampea appendiculata
Wercklea insignis
Conostegia oerstediana
Trichilia havanensis
Cedrela tonduzii
Guarea rhopalocarpa
Guarea tonduzii
Ficus hartwegii
Ficus pertusa
Ardisia palmana
Ardisia compressa
Myrsine coriacea
Eugenia guatemalensis + Myrcianthes fragrans*
Eugenia sp.
Calyptranthes pittieri
Myrcia fallax
Psidium guajava
Forestiera cartaginense
Roupala glaberrima
Cassipourea elliptica
Prunus skutchii
Prunus brachybotria
Guettarda poasana
Elaeagia auriculata
Chione sylvicola
Stauranthus perforatus
Zanthoxylum sp.
Meliosma idiopoda
Meliosma subcordata
Matayba oppositifolia
Pouteria fossicola
bd
wd
bd, bt
bd
wd
bd
bd
bd
bd
bd
bd
bd
bd or bt
bd
bd
bd
tm, bt
bd
wd
bd
bd
bd
bd
wd
bd
bd
bd
bd
am
bd
am
late
late
late
late
early
early
early
late
late
late
late
late
late
late
late
early
late
late
late
late
early
late
late
late
late
late
early
late
late
late
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
(continued)
x
x
x
x
x
x
x
x
x
x
8 Roles of Disperser Behavior and Physical Habitat Structure
199
am
bd
bd
wd
bd
bd
–
Dispersal mode
late
late
early
early
late
early
Successional status
*We were unable to distinguish between seedlings of these species in 1987
Verbenaceae
Symplocaceae
Thymelaeaceae
Tiliaceae
Pouteria sp.
Symplocos brenesii
Daphnopsis americana
Heliocarpus americanus
Mortoniodendron guatemalense
Citharexylum caudatum
unknown species
total species
total families
Table 8.2 (continued)
Family
Species
x
3
27
19
x
x
Pasture A, 3 years
x
9
54
29
x
x
Pasture A, 14 years
x
x
–
25
17
x
Pasture B, 30 years
200
K. G. Murray et al.
8 Roles of Disperser Behavior and Physical Habitat Structure
201
by comparing numbers of wind-dispersed recruits in the two plot types. If seeds
and seedlings simply survive better under the shade of remnant trees, we would
expect the same distribution patterns among wind-dispersed and animal-dispersed
species. Data on the few wind-dispersed recruits show clearly that this was not
the case; wind-dispersed recruits were no more common beneath remnant trees
than in plots without overhanging vegetation (Fig. 8.2; Mann-Whitney U = 1,270.5;
p > 0.9), while animal-dispersed recruits were (Fig. 8.2; Mann-Whitney U = 55.5;
p < 0.0001).
8.3.3 Effect of Distance to Nearest Possible Seed Source
To determine whether the number of recruits per plot might be a function of proximity to seed sources, we regressed the number of allospecific recruits of each
species in each plot on two independent estimators of proximity to seed sources:
distance to nearest conspecific remnant within the pasture, and distance to the nearest pasture/forest boundary. Distance to the nearest conspecific tree was a significant
predictor of the number of recruits per plot for only two of the 7 taxa for which
we had adequate sample sizes of conspecific remnants: Ocotea whitei and Eugenia
guatemalensis + Myrcianthes fragrans. (Table 8.3).
Table 8.3 Results of multiple regression analyses of numbers of allospecific recruits on distance to
nearest conspecific remnant and distance to pasture edge. Cells with dashes indicate species with
insufficient numbers of conspecific remnants for analysis
Species
Viburnum costaricanum
Sapium glandulosum
Stauranthus perforatus
Beilschmeidia costaricensis
Ocotea tonduzii
Ocotea whitei
Persea americana
Trichilia havanensis
Hampea appendiculata
Conostegia oerstediana
Myrsine coriacea
Eugenia guatemalensis +
Myrcianthes fragrans
Roupala glaberrima
Prunus sp.
Pouteria exfoliata
Pouteria fossicola
Heliocarpus americanus
Citharexylum caudatum
Distance to nearest
conspecific remnant
Distance to
pasture edge
–
ns
ns
–
ns
***
–
–
–
–
–
*
ns
ns
ns
ns
ns
ns
ns
ns
ns
ns
ns
ns
ns
–
ns
–
–
–
ns
ns
ns
*a
ns
ns
*p < 0.05 ***p < 0.001 ns p > 0.05
a regression coefficient significant, but in the opposite direction of that predicted.
202
K. G. Murray et al.
The number of allospecific recruits was dependent upon distance to pasture edge
for only one species, Pouteria exfoliata. In this case, however, the effect was in the
direction opposite of that predicted: trees farther from the forest edge tended to have
more Pouteria seedlings than those nearer to it.
8.3.4 Effect of Dispersal Mode of Remnant Tree
As a conservative test of the hypothesis that remnant trees are particularly effective as recruitment foci for species with similar fruit types, we compared the mean
numbers of animal-dispersed recruits beneath animal-dispersed and wind-dispersed
remnants. Surprisingly, the mean number of animal-dispersed recruits was actually somewhat higher beneath wind-dispersed remnants (Fig. 8.3), although the
difference was not significant (Mann-Whitney U = 247; p = 0.147). In fact, even
dead snags without foliage were as effective as animal-dispersed remnant trees as
recruitment foci for animal-dispersed species (Fig. 8.3; Mann-Whitney U = 210;
p = 0.889). Similarly, species richness of allospecific animal-dispersed recruits did
not differ significantly among the two remnant tree types and dead snags (Fig. 8.3;
Kruskal-Wallis one-way ANOVA statistic = 0.936, p = 0.626). Thus, remnant trees
do not seem to attract animal-dispersed recruits only if they also serve as fruit
sources for foraging animals. Rather, trees act as foci for seed input simply because
they provide perch sites or sources of insect prey also eaten by frugivorous birds and
mammals.
5
Recruits / m2
4
3
2
1
Fig. 8.3 Mean density and species richness of
allospecific animal-dispersed recruits in plots beneath 11
wind-dispersed and 62 animal-dispersed remnant trees,
and beneath 7 dead snags in pasture A, 3 years
post-abandonment. Error bars represent one standard
error of the mean
Species per Plot
0
8
6
4
2
0
animal snag wind
Dispersal Mode of Remnant
8 Roles of Disperser Behavior and Physical Habitat Structure
203
8.3.5 Persistent Effects of Recruitment Foci on Regeneration
Recruit density increased in both plot types between 3 and 14 years postabandonment in pasture A, but remnant tree plots still had significantly higher stem
densities than open sky plots (Fig. 8.4). By 30 years post-abandonment at pasture
B, however, densities in both plot types had declined to levels below that typical of
remnant plots at 3 years post-abandonment, presumably due to self-thinning. More
importantly, recruit density in Pasture B (30 years post-abandoment) was similar in
the two plot types.
Temporal trends in species richness of recruits paralleled those in density—
species richness increased in both plot types from 3 to 14 years post-abandonment,
with richness significantly higher in remnant tree plots. Richness then declined by
30 years, and was similar in both plot types.
8.3.6 Effects of Remnant Trees on Species Composition of Recruits
We investigated the long-term impact of remnant trees on regeneration of a
species-rich assemblage of late-successional species in two ways. First, we
compared density and species richness of early- vs. late-successional recruits in
remnant tree and open sky plots as pastures aged. At 3 years post-abandonment,
*
12
Recruits / m2
10
8
6
4
2
*
Fig. 8.4 Mean density and species richness of all
woody recruits in remnant tree and “open sky” plots
vs. time since pasture abandonment. Error bars
represent one standard error of the mean, and
asterisks denote significant differences at p < 0.001
(Mann-Whitney U test). Note that the 3 year
post-abandonment data are based on 12.6 m2
samples, while the 14 and 30 year data are based on
3.14 m2 samples. Comparisons of species richness
between 3 years post-abandonment and that for
either the 14 or 30 year censuses must be made with
caution
Species per Plot
0
10
*
8
6
*
Plot Type
open sky
remnant
4
2
0
3
14
30
Time Since Abandonment (yrs)
204
K. G. Murray et al.
Fig. 8.5 Mean density of recruits of early- and
late-successional species (see text) in both plot
types vs. time since pasture abandonment. Error
bars represent one standard error of the mean,
and asterisks denote significant differences at p <
0.001 (Mann-Whitney U test)
3
Open Sky Plots
*
0
Recruits / m2
12
Remnant Tree Plots
*
9
Recruit Type
pioneer
primary
6
3
*
*
0
3
14
30
Time Since Abandonment (yrs)
early-successional species slightly outnumbered late-successional ones in remnant
tree plots and approximately equaled them in open sky plots (Fig. 8.5). By 14
years, however, late-successional species far outnumbered early-successional ones
in remnant tree plots. In open sky plots, early- and late-successional species were
still similar in density. By 30 years post abandonment, early-successional species
were significantly more dense in both plot types.
We also investigated the similarity in species composition among plot types and
over time using detrended correspondence analysis (Gauch 1982). For purposes of
our analysis, we included all remnant tree and open sky sites censused in the two
pastures in 1987 and 1998. To retain the plots with no woody recruits in the analysis,
we created a dummy species and included it as the only recruit in otherwise empty
plots. In order to focus only on species composition rather than the differences in
abundance discussed above, we normalized the overall density in each plot to 1.0;
each species thus contributed a proportion to that overall density.
Species composition differed substantially between the two plot types at 3 years
post-abandonment (Fig. 8.6A), largely because most open sky sites had few or no
recruits. Open sky sites were more heterogeneous for the same reason. At 14 years
post-abandonment, species composition was both more homogeneous within plot
types and more similar between them (Fig. 8.6B), probably because by that time
the original open sky sites had been colonized by recruits deposited by frugivores
perched in new overhanging vegetation (i.e., from the original recruits that by this
time were 3–5 meters tall) that didn’t exist at the time of pasture abandonment. By
30 years, species composition was more heterogeneous in both plot types (Fig. 8.6C)
probably because self-thinning reduced stem density and thus gave slight compositional differences a larger effect in the ordination. Direct comparisons between the
30-year data and that for 3 and 14 years must be done with caution, since the former
was done in pasture B and the latter were done in pasture A half a kilometer away.
8 Roles of Disperser Behavior and Physical Habitat Structure
400
A
310
AX4
Fig. 8.6 Results of detrended correspondence analyses
(DECORANA) on woody recruits in remnant tree
(filled symbols) and “open sky” (open symbols) plots
vs. time since pasture abandonment. (A) pasture A, 3
years post-abandonment. The tightly clustered open
symbols represent open sky plots with no recruits. (B)
pasture A, 14 years post-abandonment. (C) pasture B,
30 years post-abandonment. Confidence ellipses are
centered on the bivariate means, and their sizes are
such that 68.27% (1 standard deviation) of the data
points are included within. Note that axes 3 and 4 are
shown here for each ordination. Axes 1 and 2 separated
plots more on the basis of the composition of a few
idiosyncratic plots
205
220
130
40
-50
0
300
600
900
600
900
600
900
AX3
400
B
AX4
310
220
130
40
-50
0
300
AX3
400
C
AX4
310
220
130
40
-50
0
300
AX3
8.4 Discussion
Data presented here show clearly that forest regeneration on abandoned pastures can
occur very rapidly at Monteverde, and that the physical structure of the pasture plays
a key role in determining the rate and successional trajectory of that regeneration.
The fact that a species-rich assemblage of plants appears beneath remnant trees so
soon after cattle are removed from a pasture suggests strongly that seeds are being
deposited continuously even in active pastures, but that the cattle simply eat the
resultant seedlings soon after germination. Studies of the seed rain in pastures verifies this pattern at Monteverde (Harvey 2000, R. Boyce and K. G. Murray, unpublished data) and elsewhere (e.g., Guevara and Laborde 1993, Vieira et al. 1994, Holl
1998).
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8.4.1 Remnant Trees as Recruitment Foci
In Monteverde, seed dispersal by fruit-eating birds and mammals was largely
responsible for the rapid colonization of pastures by woody plant species; of the
67 plant species identified as recruits in the study plots in our two pastures, all but 9
have seeds adapted for dispersal by animals.
Patterns of early recruitment were decidedly non-random with respect to physical
habitat structure: significant recruitment of woody species occurred only beneath
the crowns of remnant trees in pasture A. The reason for this pattern is obvious:
fruit-eating animals forage in trees and shrubs, preferentially regurgitate or defecate
while perched rather than in flight, and therefore deposit seeds beneath overhanging vegetation rather than in open pasture. Previous work at many other sites has
also demonstrated this pattern (e.g., Debussche et al. 1982, McDonnell and Stiles
1983, Guevara et al. 1986, McDonnell 1986, McClanahan and Wolfe 1987, 1993,
Bass 1990, Crome 1990, Uhl et al. 1991, Guevara 1992, Guevara and Laborde
1993, DeBussche and Isenmann 1994, Vieira et al. 1994, Toh et al. 1999, Zahawi
and Augspurger 1999, Slocum 2001). This pattern also coincides with anecdotal
observations of the behavior of fruit-eating birds (e.g., Cardoso da Silva et al. 1996,
K. G. Murray, personal observation).
In contrast to Harvey (2000), we found no evidence that bat-dispersed species
are more likely to colonize open pasture than are bird-dispersed ones: recruits of
both Conostegia oerstediana and Eugenia guatemalensis + Myrcianthes fragrans
were just as heavily concentrated beneath remnant trees as those of bird-dispersed
species. Some authors have explained the deposition of bat-dispersed seeds in open
pastures or gap interiors on the basis of bats’ propensity to defecate in flight (e.g.,
Charles-Dominique 1986, Harvey 2000). The lack of such a pattern in our study
may result from several factors. First, both bat-dispersed taxa are dispersed by birds
as well as bats (Haber 2000). In fact, C. oerstediana is dispersed primarily by birds:
only a small proportion of bat fecal samples collected by Dinerstein (1983) contained seeds of this species, while Wheelwright et al. (1984) found it to be one of
the most commonly eaten species by frugivorous birds at Monteverde. Second, E.
guatemalensis has large seeds that are never ingested by bats. Rather, they are discarded after the pulp is removed, which probably occurs at a feeding roost. Finally,
bats may deposit the majority of the seeds they carry beneath feeding roosts despite
the fact that they also defecate in flight, simply because they spend more time roosting than flying.
We found no evidence that the shade produced by remnant trees contributed to
higher recruitment beneath them, as was suggested by Nepstad et al. (1991) for
abandoned Amazonian pastures and by Holl et al. (2000) for a montane site in Costa
Rica. This effect, which presumably results from amelioration of water stress during
the dry season or inhibition of pasture grasses, is undoubtedly important at some
locations—particularly at lower elevations where the dry season is more severe. Our
study was conducted at an elevation where blowing mist usually prevents serious
water stress even during the height of the dry season. Working at the same elevation
8 Roles of Disperser Behavior and Physical Habitat Structure
207
in southern Costa Rica, Holl (1999) found no evidence of water stress among forest
tree seedlings planted in abandoned pastures.
The particular effectiveness of animal-dispersed remnant trees as recruitment
foci is widely cited (e.g., Janzen 1988), though few studies explicitly present
data to support this conclusion. Groom (2000) found higher recruitment of woody
plants beneath three species of animal-dispersed remnants than beneath one winddispersed species, and the pattern was especially pronounced for recruits in the
Lauraceae, a family that depends disproportionately on large fruit-eating birds like
Resplendant Quetzals and Three-wattled Bellbirds for dispersal. Similarly, Slocum
and Horvitz (2001) found higher seed input (in terms of both density and species
richness) beneath animal-dispersed remnants than beneath those with wind- or
ballistically-dispersed seeds, but input beneath both types of remnants exceeded that
at sites without overhanging vegetation. Censuses of established recruits showed
similar patterns, but the differences between focal tree types were less pronounced
(Slocum 2001). In contrast, Guevara et al. (1986) found that species richness was no
higher beneath the five animal-dispersed remnants they censused (mean 11.2, s.d.
4.38) than beneath two wind-dispersed ones (mean 10.5, s.d. 3.54). They did not
report data on the density of recruits. Similarly, Otero-Arnaiz et al. (1999) found no
relationship between the species composition of recruits beneath pasture trees and
the identity or dispersal mode of that tree. On abandoned farmlands in Queensland,
Toh et al. (1999) reported that the presence of a fruit reward was less important
than the structure of a remnant tree and its suitability as a perch in determining its
effectiveness as a recruitment focus.
We too found no evidence that remnant trees that themselves produce fleshy
fruits are more effective than other trees as recruitment foci. Instead, we found
similar densities of animal-dispersed recruits beneath both animal-dispersed and
wind-dispersed remnants, and even beneath dead snags. These findings and those
reported above suggest that dispersers do not visit isolated pasture trees just for the
fruit crops they contain, but perhaps as sources of cover (i.e., concealment from
predators), as sites for courtship or territorial displays (e.g., singing perches), or as
sources of other food, e.g. arthropods. Many of the bird species that frequent the
pasture include substantial numbers of insects in their diets in addition to fruits,
most notably many of the tanagers and thrushes. Cardoso da Silva et al. (1996)
also found that tanagers in particular were the primary dispersers of seeds beneath
isolated trees and shrubs in recently abandoned pastures in Brazil.
The finding that even artificial perches enhance recruitment (e.g., McDonnell and
Stiles 1983, McClanahan and Wolfe 1993, Robinson and Handel 1993, 2000, Holl
1998) also suggests that the effectiveness of elevated structures as recruitment foci
is largely due to their use as perches. Holl (1998) found that birds (and deposited
seeds) were attracted by leafless tree branches placed upright in a southern Costa
Rican pasture, but that they were no more effective when she added banana to
them. And the greater seed input beneath animal-dispersed remnants in Slocum and
Horvitz’ (2000) study occurred even during times of the year when the remnant trees
were not producing fruits. The importance of a remnant tree’s own dispersal mode
in determining its effectiveness as a recruitment focus thus remains ambiguous.
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Virtually all remnant trees and other elevated structures facilitate secondary succession on post-agricultural lands by attracting animals, and sometimes (but not
always) the effect is enhanced by the presence of fruits in the remnants themselves.
8.4.2 Distance to Seed Sources
We found little evidence that distance to nearest possible seed source had any effect
on the density of most species of recruits in pasture A. Apparently the distances
involved were small enough that most seed dispersers readily visited even the trees
in the most central portions of the pasture. In contrast, Holl (1999) found almost no
seed deposition beyond 5 meters from the edge of an abandoned montane pasture in
southern Costa Rica, but this finding was more likely due to the fact that seed traps
were placed in locations without overhanging vegetation. The only taxa in our study
for which density was positively correlated with proximity to either pasture edge
or nearest conspecific remnant (Ocotea whitei and Eugenia guatemalensis + Myrcianthes fragrans) both produce fruits with single large seeds that are dispersed by
a few species of large frugivores (e.g., Resplendant Quetzals, Black Guans, Emerald Toucanets, and Three-wattled Bellbirds; Wheelwright et al. 1984). Such large
seeds are voided rapidly by regurgitation (Wheelwright 1991), and are thus often
not dispersed far from the parent plant (cf. Wenny and Levey 1998).
We do not suggest that the lack of a relationship between recruitment and distance to seed sources in this study is a general one, however. Pasture A is fairly
typical of those in the Monteverde region in that it is relatively small and has forest
in close proximity. Pastures are apparently small in much of Puerto Rico also, and
like us Aide et al. (1996) found no evidence for a distance effect on pasture regeneration there. These conditions are not typical of pastures in much of South America,
however, nor of those at lower elevations in much of Central America, and we would
expect a much stronger effect of distance to pasture edge on recruitment beneath
remnant trees in them. In large Amazonian pastures, for example, even tanagers
characteristic of the forest/pastures boundary rarely venture more than 150 meters
from the forest edge (Cardoso da Silva et al. 1996).
8.4.3 Successional Trends
Early-successional species were particularly well-represented among the early
colonists of both remnant tree and open sky plots, perhaps because their small
seed size permits dispersal by a greater number of bird species than do the larger
seeds of late-successional species (e.g., Wheelwright 1985). Species with minute
seeds (< 1mm, e.g., Conostegia oerstediana) in particular are often dispersed by
small tanagers, which were abundant in the study area and readily forage within
pastures at Monteverde (K. G. Murray, personal observation) as elsewhere (e.g.,
Cardoso da Silva et al. 1996). Tanagers fail to disperse seeds of species with larger
8 Roles of Disperser Behavior and Physical Habitat Structure
209
seeds because they routinely separate seeds from the pulp before swallowing it
(Levey 1986, Murray 1988). Most other studies have documented dominance by
early-successional species in the early stages of post-agricultural regeneration as
well (e.g., Aide et al. 1996, 2000, Myster 2004b).
By 14 years post-abandonment, remnant tree plots in our study were dominated
by late-successional species, while open sky plots were still dominated by earlysuccessional ones. While we did not address the reason for this trend directly, it
probably derives from the combination of reduced germination of early-successional
species beneath the increasingly shaded (by previous recruits) remnants and the
simple accumulation of recruits from a larger pool of available primary species. By
30 years post-abandonment, both plot types were dominated by late-successional
species to about the same degree.
Both remnant tree and open sky plots continued to accumulate species richness and total stem density for at least several years after pasture abandonment.
Indeed, by 30 years no difference in either stem density or species richness was
apparent between the two plot types. The reason for continued accumulation is
obvious at least for remnant tree plots—fruit-eating animals continued to deposit
seeds beneath perch sites. For open sky plots, however, the reason for eventual
accumulation of recruits is less obvious, since they failed utterly to attract significant numbers of recruits soon after pasture abandonment. The probable answer is
that rapid recruitment beneath the crowns of remnant trees leads to the formation
of “vegetation islands” that expand in size and coalesce with one another and with
regeneration proceeding into the pasture from the edge. At our study site, the density
of remnant trees was high enough (and the pastures small enough) that most of
the original open sky plots came to be beneath the crowns of trees recruited into
the pasture by 14 years post-abandonment. As soon as these plots had overhanging vegetation, they began to attract recruits for the same reasons as the original
remnant tree plots. By 30 years post-abandonment, the entire pasture had been
covered for many years by the crowns of trees recruited after pasture abandonment. The elimination of the original “open sky” plots via the growth of vegetation islands, together with the self-thinning of recruits as light intensity near the
ground decreased over time, led to the loss of any recognizable difference in either
species richness or total stem density between the two plot types by 30 years post
abandonment.
This relatively rapid convergence of density and species composition in remnant
tree and open sky plots should not be taken as evidence that remnant trees have
only transient effects on pasture regeneration. Indeed, the rapid convergence only
occurred because the pastures we studied were relatively small and had large numbers of remnants (which were therefore close together). The most direct measure of
the importance of remnant trees would require comparison of our data with that from
a large pasture completely lacking remnant trees—a comparison we were unable to
make. Working in Brazil, however, Chris Uhl and his colleagues have been able
to study very large pastures (> 10 ha) in which any original remnant trees have
been killed by fires. They estimate that without human intervention regeneration
of species-rich forests on such pastures may take hundreds of years (Buschbacher
et al. 1992).
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Aide et al. (2000) also found that species richness (as well as density, basal area,
and aboveground biomass) in abandoned pastures in Puerto Rico converged on that
typical of old-growth forest by 40 years post-abandonment, but that the species composition of the secondary forest remained quite different from old growth. Clearly,
we need more long-term studies of the trajectory of succession in post-agricultural
landscapes to better understand whether, and how rapidly, regenerating communities converge on the species richness and composition of the communities that
surround them.
8.4.4 Other Habitat Structural Elements
Other structural habitat features influence the rate and trajectory of succession on
post-agricultural lands as well. Several investigators have documented facilitation
of recruitment of woody forest species on rotting logs, (e.g., Peterson and Haines
2000, Slocum 2000). Such logs are often remnants of the original forest, and they
may enhance recruitment by acting as perch sites for frugivores, because cattle can’t
trample seedlings that colonize them, because they are devoid of the pasture grasses
that compete with tree seedlings, or because they have great water-holding capacity
after they partially decompose.
Patches of ferns or shrubs also facilitate colonization by forest species, in some
cases because they are avoided by cattle (e.g., Slocum 2000), and in others because
they inhibit the herbaceous vegetation that competitively exclude woody forest
species (e.g., Posada et al. 2000). As importantly, shrubs, like trees, provide elevated perch sites that attract foraging birds and mammals, and thus enhance seed
deposition. Cardoso da Silva et al. (1996), for example, found that birds moving
from forest into Amazonian pastures spent much of their time in shrubs, even when
larger trees (e.g., Cecropia sp.) were present. It is possible, of course, that dense
shrub patches in post-agricultural fields inhibit the growth rates of woody recruits
that do establish via the shade they produce (e.g., Zahawi and Augspurger 1999). By
enhancing seed deposition and inhibiting herbaceous vegetation, however, their role
overall is more likely to be one of facilitation (e.g., Holl et al. 2000 and references
cited therein).
8.4.5 Implications for Management
Secondary forests on previously abandoned tropical pastures are likely to become
increasingly important repositories of tropical biodiversity and providers of other
ecosystem services like erosion control and flood amelioration, even if they are ultimately converted again to some other land use type. Understanding the key roles
played by frugivores and remnant trees can give those responsible for managing
tropical landscapes an important incentive for conserving frugivore populations and
8 Roles of Disperser Behavior and Physical Habitat Structure
211
seed sources, as well as allow them to manage such lands for maximum speed and
species richness of recovery.
Results presented here and elsewhere suggest management practices that would
speed succession to a species-rich forest on abandoned tropical pastures. Key among
the factors that affect the rate and trajectory of regeneration is the physical structure
of the pasture: all else being equal, small pastures with remnant trees are likely
regenerate most rapidly. And while characteristics that would speed recovery after
abandonment might best be addressed at the time of conversion from forest, farmers
are more likely to have a greater incentive to make pastures large and to leave only
enough trees to provide their cattle with a few sources of shade. After abandonment,
however, some relatively simple manipulations might greatly accelerate seed input
even in large barren pastures. The finding that remnant trees enhance recruitment
primarily through a “perch effect” suggests that the rate of forest regeneration would
be maximized by planting the most rapidly growing native trees locally available,
regardless of the dispersal mode for which their own seeds are adapted. And since
many seed dispersers will travel among forest patches through even narrow shelterbelts of trees (e.g., Harvey 2000), managers might speed recovery by connecting the
areas bordering abandoned pastures with other forest patches via the construction
of shelterbelts of rapidly growing species.
We acknowledge the danger of drawing conclusions about temporal trends from
data on spatial ones (i.e., comparison of pasture B at 30 years post-abandonment
with pasture A at 3 and 14 years post-abandonment), and of drawing conclusions
about pasture regeneration from a study of just two pastures in the study described
here. Because the pastures were only 0.5 km apart, however, and because the species
composition of forest surrounding them is so similar we are confident that the patterns seen in the two pastures reflect a successional trajectory common to both. We
also remain confident that our results contain important insights about the time scale
on which the interaction between initial habitat structure and successional trajectory
occurs, and that our results reflect processes that operate widely. Like several of
the other studies cited above, which were conducted on four continents, this study
demonstrates a profound effect of physical habitat structure on successional trajectory. Few studies of tropical pasture regeneration have followed the process for more
than a few years post-abandonment, however, and we believe that future investigations will also find that some effects of initial habitat structure weaken with time
(e.g., recovery of density and species richness concentrated beneath remnant trees),
while others persist for much longer (e.g., species composition).
Finally, we acknowledge that drawing conclusions about secondary succession
on post-agricultural lands in general from studies concentrated on abandoned pastures can lead to oversimplification. Like the work reported here, most of the literature on regeneration of former agricultural lands concerns abandoned cattle pastures rather than croplands. That bias is partly geographical; most studies are from
the Neotropics, where conversion to cattle pasture is the single greatest driver of
deforestation (Myers 1980). It may also reflect greater permanence of other types
of agricultural land use. Different agricultural uses result in profound differences
in soil characteristics, disperser, herbivore and pathogen communities, and physical
habitat structure at the time of abandonment, however (e.g., Myster 2004b), and
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K. G. Murray et al.
given that all of these factors have been shown to influence the rate and trajectory
of regeneration we should expect details of the process to reflect the unique history
and characteristics of each site.
It would be naïve to suggest that the pristine tropical forests remaining today can
be protected permanently from conversion to agriculture, and there is widespread
agreement that maintaining or restoring biodiversity over wide areas will require
regeneration of lands already deforested. Over the long term, sustainable use of
many tropical lands will involve repeated cycles of conversion to agriculture and
regeneration, and learning to manage regeneration should remain a high priority.
Acknowledgment We are especially grateful to Alan Pounds and Marlene Brenes for allowing
us to collect data in pasture A. The Tropical Science Center, and especially the former director of
the Monteverde Cloud Forest Preserve, Robert Carlson, greatly facilitated the project by providing
laboratory space and giving us access to pasture B. Nalini Nadkarni and Rodrigo Solano kindly
shared previous data on the identities and locations of remnant trees at the pasture B site. The
work was funded by Faculty Development Grants from Hope College and by NSF grants to Hope
College (BBS-8712566 and DBI-9322220) and to K. G. Murray (DEB-9208229). Christopher
Lepczyk, Dan Wenny, and Chris J. Peterson provided useful comments on an earlier version of
the manuscript. Finally, KGM thanks Chris Uhl for many stimulating discussions about pasture
regeneration, and for the opportunity to see it in Amazonia.
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and Conservation of a Tropical Cloud Forest, eds. N. M. Nadkarni and N. T. Wheelwright,
pp. 448–450. New York, NY: Oxford University Press.
Otero-Arnaiz, A., Castillo, S., Meave, J., and Ibarra-Manríquez, G. 1999. Isolated pasture trees
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243–254.
Peterson, C. J., and Haines, B. L. 2000. Early successional patterns and potential facilitation of
woody plant colonization by rotting logs in premontane Costa Rican pastures. Restoration
Ecology 8:361–369.
Posada, J. M., Aide, T. M., and Cavelier, J. 2000. Cattle and weedy shrubs as restoration tools of
tropical montane rainforest. Restoration Ecology 8:370–379.
Robinson, G. R., and Handel, S. N. 1993. Forest restoration on a closed landfill: rapid addition of
new species by bird dispersal. Conservation Biology 7:271–278.
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Journal of Ecology 75:377–407.
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degradadas. Ciência Hoje 13:22–31.
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Chapter 9
Neotropic Post-Dispersal Seed Predation
Randall W. Myster
9.1 Introduction
The dynamics of seeds and seedlings determine much of plant community development after disturbance (Grubb 1977). For old field succession, whether in the
temperate zone or in the Neotropics, such regeneration comes overwhelmingly
from outside the field (Purata 1986, Uhl, Buschbacher and Serrao 1988, Myster
1993, Medellin and Gaona 1999) because few species sprout from remnant trees
or leftover rootstocks (although some shrubs can recruit asexually: Miconia spp. in
Neotropical fields [Myster 2004b] and Rhus spp. in temperate fields [Myster 1993]).
Since most species depend on dispersed seeds for their recruitment, dispersal and
the fate of seeds once they have landed in fields are of primary concern.
As previous chapters have shown, large numbers of seeds disperse into Neotropic
pastures and old fields which, when combined with the low seedling emergence
from Neotropic seed banks (Vazquez-Yanes and Orozco-Segovia 1993), points to a
critical mechanism that removes large amounts of seeds from old fields after dispersal. Numerous studies in both temperate and tropical America have shown that seed
predators can eat up to 100% of seeds presented to them in old field experiments
(reviewed in Myster 1993, 2004b), implying that seed predation is one of those key
mechanisms.
In temperate old fields, small mammals account for most of those losses (Gill
and Marks 1991, Myster and Pickett 1993). As a field ages and becomes more complex structurally and as more mammal refuges become available, there occurs an
increase in predation levels. In general, predators in old fields prefer woody patches
to herbaceous patches (Webb and Willson 1985) and avoid open areas (Burton
1989). Also because of woody cover predators choose seeds that are closer to the
forest, which becomes less true over time as fields develop patches of tall vegetation
(Gill and Marks 1991). There are other sources of variation that affect the behavior
of seed predators. For example, additions of litter have been shown to decrease
predation (Myster and Pickett 1993) while spring rain have been shown to increase
it (DeSteven 1991). Furthermore, studies have generally shown a strong ranking of
species-specific predator preferences, which may correlate more with tree taxa (i.e.,
conifers are the least desired) than with common seed sizes (DeSteven 1991) which
change as a field ages (Myster and Pickett 1993).
216
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9 Neotropic Post-Dispersal Seed Predation
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The investigation of seed predation in Neotropic old fields has built on the results
from temperate old fields to further examine how the interactions among various
factors affect rates of seed removal, with the aim of more precisely quantifying
the seed predation niche for Neotropic old field plant species. Thus, the study of
predator preferences in the Neotropics has been a priority, but also litter interactions, density of seed, age of field, distance to forest edge, field-type, patch-type, and
season.
9.2 Results from Past Experiments in Puerto Rican
and Ecuadorian Old Fields
A number of seed predation experiments were performed in (1) pastures, Banana
plantations, Sugarcane plantations, and a Coffee plantation in the Luquillo Experimental Forest of Puerto Rico (Myster 2003a, 2003b, 2004a) and in (2) pastures,
Banana plantations, and Sugarcane plantations in the Maquipucuna Reserve in
Ecuador (Myster 2004a, 2007: see Chapter 1 for a description of these study sites).
At both locations, seeds were put out in 9cm diameter plastic Petri dishes (except
for the Coffee plantation where a plastic grid was used). After 2 weeks, seed loss
was scored in each dish (or grid cube) and assumed to be due to predation, not
wind or rain, since plastic seed mimics of similar size, shape, and weight were
not lost.
The analysis of data from the Puerto Rican pastures indicated that most seeds
were lost to predators with significant differences in species (Syzgium jambos had
50% losses, Solanum tovum had 65% losses, Miconia prasina had 76% losses, and
Piper aduncum had 79% losses), in density (78% losses at 5 seeds per dish, 69%
losses at 10 seeds per dish, 51% losses at 20 seeds per dish), and in patch type (grass
patches had 69% losses, Miconia spp. patches had 65% losses). In addition, there
were significant interaction effects (1) between species and density for P. aduncum
and S. tovum (88% losses at low density vs. 85% losses at medium density vs. 56%
losses at high density, 83% losses at low density vs. 70% losses at medium density
vs. 27% losses at high density, respectively) and (2) between species and patch-type
for M. prasina (93% losses in grass patches vs. 55% losses in Miconia spp. patches).
Another experiment in the same pastures and in a Puerto Rican Coffee plantation
showed large species differences (Guarea guidonia had 34% losses, Gonzalugunia
spicata had 55% losses, Miconia racemosa had 66% losses, and Casearia arborea
had 72% losses), with predators taking seeds at spatial scales of 2–11cm. Puerto
Rican Banana plantations and Sugarcane plantations also demonstrated significant
differences in species (Gonzalagunia spicata had 80% losses, Palicourea riparia
had 63% losses, Piper aduncum had 74% losses, and Miconia pracina had 86%
losses).
Experiments in Ecuadorian pastures, Banana plantations, and Sugarcane plantations yielded similar results: most seeds were lost to predators while there
were significant differences in species (Cecropia monostachya had 76% losses,
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R. W. Myster
Solanum ovlifolium had 82% losses, Piper aduncum had 57% losses, and Cestrom
megalothyllum had 80% losses). There were certain interactions between species
and predator taxa. C. monostachya had 86% of its predation from insects rather
than mammals, S. ovlifolium had 48%, Piper aduncum had 56% with most of the
predation occurring in the forest, and C. megalothyllum had 40%. More seeds were
eaten in the pasture (97% losses) than in Sugarcane (76% losses) and Banana fields
(43% losses).
9.3 Discussion
The results suggest that those aspects of plant niches that deal with seed predation
are very important in determining plant-plant replacements during old field succession. Therefore, seed predation is critical in creating invasion and establishment
patterns. Studies on seed predation that were conducted in different Neotropic areas
(Puerto Rico, Ecuador) and that involved different species, field-types, densities,
and patch-types all report significant differences in the way predators eat seeds of
different species. Due to this dominant species effect, other sources of variation in
the working of seed predation may best be seen as the results of interaction with
the significant species effects which are always present. In addition to the species
effect, there has been found a general negative correspondence between rates of seed
predation and seed size (as in Uhl 1987, Nepstad et al. 1996, alternatively see Holl
and Lulow 1997). Other Neotropic studies also indicated that seed losses were large
and dominated by species differences (Table 9.1).
The effect of woody patches on seed predation is variable. Whereas in the present
series of studies woody patches decreased predation levels (patch-type effects in
Table 9.1 Other seed predation studies of Neotropic old fields and pastures. All studies were
done in lowland wet rainforest, with the exception of Sarmiento which was done in montane wet
rainforest at the Maquipucuna Reserve. For dry forest seed predation results see Chapter 11.
% lost
time period
field
variation
country
reference
10–100%
19 days
pasture
species field age
Colombia
up to 98%
7–100%
2 months
24 days
field
pasture
Mexico
Costa Rica
20–100%
30 days
pasture
Costa Rica
Holl and Lulow
1997
15–100%
200 days
pasture
species field age
species patch-type
year
species forest
distance
patch-type
species
Aide and
Cavelier 1994
Hammond 1995
Holl 2002
Brazil
4–100%
28 weeks
field
species
Peru
90%
35–100%
5–100%
2 weeks
4 months
6–8 months
pasture
field
pasture
species
species field age
species
Ecuador
Brazil
Brazil
Nepstad et al.
1996
Notman and
Gorchov 2001
Sarmiento 1997
Uhl 1987
Vieira and
Scariot 2006
9 Neotropic Post-Dispersal Seed Predation
219
Holl and Lulow 1997) due to increased predation on mammals by birds in those
patches, other studies showed more predation in woody patches and in the forest
(Nepstad et al. 1996, Aide and Cavelier 1994, Holl 2002) perhaps due to the greater
litter accumulation there. As in temperate fields, the large seed losses caused by
predation and the highly specific preferences of predators (Holl and Lulow 1997)
suggest that the issues of variation in predator density and distribution in old fields
and in predator foraging behavior, in particular how these aspects of their biology
and ecology are affected by the availability and structure of plant cover, should be
an essential component of research into seed predation.
Holl and Lulow (1997), Norman and Gorchov (2001), and Vieira and Scariot (2006) used enclosures to implicate small mammals as seed predators on the
Neotropic mainland. The presented case studies implicated both small mammals
and insects (mainly ants) in Ecuador, but only insects (again ants) in Puerto Rico,
as the major seed predator taxa. Nepstad et al. (1993) reported similar results
that implicated both ants and small mammals as seed predators while suggesting
that seeds larger than 0.4g would be taken by mammals and smaller than 0.4g by
insects. Researchers have shown both a decrease (Uhl 1987, Hammond 1995) and
an increase (Aide and Cavilier 1994) in predation ever successional time, and effects
of forest edge on predation are also not conclusive (Holl and Ludlow 1997).
A comparison of the results from Neotemperate and Neotropical studies brings
forward the following points. Old fields in both places show high predation rates, up
to 100% losses for some species, and more predation both under shrubs and at low
density (Myster and Pickett 1993). Seasonal effects, however, are more pronounced
in temperate studies (reviewed in Myster 1993, 2003b). Reduced rates of predation
on seeds under litter and distance away from the forest (Gill and Marks 1991) have
not yet been examined in Neotropical old fields.
All studies point to seed predation as a critical mechanism that determines old
field invasion patterns. Results show that, first, predators prefer some species over
others. This predilection gives the less preferred species an advantage in that they
can establish themselves both before and more extensively than the preferred, more
eaten species. Second, predation may be reduced under “nurse” trees, especially at
the higher seed densities that are naturally found under shrubs and trees (satiation
hypothesis: Janzen 1971, Myster 2006), which allows regeneration to radiate from
these trees. Finally, patch-types with less predation may be able to facilitate more
regeneration while field-types with more predation, such as pastures, may lag behind
developmentally for years.
References
Aide, T. M., and Cavelier, J. 1994. Barriers to lowland tropical forest restoration in the Sierra
Nevada de Santa Marta, Colombia. Restoration Ecology 2:219–229.
Burton, P. J. 1989. Constraints to tree invasion on a nutrient rich site during old-field succession.
Ph.D dissertation. University of Illinois, Urbana.
DeSteven, D. 1991. Experiments on mechanisms of tree establishment in old field succession:
seedling emergence. Ecology 72:1066–1075.
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Gill, D. S., and Marks, P. L. 1991. Tree and shrub seedling colonization of old fields in central
New York. Ecological Monographs 61:183–205.
Grubb, P. J. 1977. The maintenance of species-richness in plant communities: the important of the
regeneration niche. Biological review of the Cambridge Philosophical Society 52:107–145.
Hammond, D. S. 1995. Post-dispersal seed and seedling mortally of tropical dry forest trees after
shifting agriculture, Chiapas, Mexiso. Journal of Tropical Ecology 11:295–313.
Holl, K. D. 2002. Effect of shrubs on tree seedling establishment in an abandoned tropical pasture.
Journal of Ecology 90:179–187.
Holl, K. D., and Lulow, M. E. 1997. Effects of species, habitat and distance from edge on postdispersal seed predation in a tropical rainforest. Biotropica 29:459–468.
Janzen, D. H. 1971. Seed predation by animals. Annual Review of Ecology and Systematics
2:465–492.
Medellin, R. A., and Gaona, O. 1999. Seed dispersal by bats and birds in forest and disturbed
habitats of Chiapas, Mexico. Biotropica 31:478–485.
Myster, R. W. 1993. Tree invasion and establishment in old fields at Hutcheson Memorial Forest.
The Botanical Review 59:251–272.
Myster, R. W. 2003a. Seed regeneration mechanisms over fine spatial scales on recovering Coffee
plantation and pasture in Puerto Rico. Plant Ecology 166:199–205.
Myster, R. W. 2003b. Effects of species, density, patch-type and season on post-dispersal seed
predation in a Puerto Rican pasture. Biotropica 35:542–546.
Myster, R. W. 2004a. Regeneration filters in post-agricultural fields of Puerto Rico and Ecuador.
Plant Ecology 172:199–209.
Myster, R. W. 2004b. Post-agricultural invasion, establishment, and growth of Neotropical trees.
The Botanical Review 70:381–402.
Myster, R. W. 2006. Shrub vs. grass patch effects on the seed rain and the seed bank of a five-year
pasture in Puerto Rico. Ecotropica 12:12–19.
Myster, R. W. 2007. Early successional pattern and process after Banana, Sugarcane and Pasture
in Ecuador. The New Zealand Journal of Botany 45:35–44.
Myster, R. W., and Pickett, S. T. A. 1993. Effects of litter, distance, density and vegetation patch
type in postdispersal tree seed predation in old fields. Oikos 66:381–388.
Nepstad, D. C., Uhl, C., Pereira, C. A., and Cardoso-Da Silva, J. M. 1996. A comparative study
of tree establishment in abandoned pasture and mature forest of eastern Amazonia. Oikos
76:25–39.
Notman, E., and Gorchov, D. L. 2001. Variation in post-dispersal seed predation in mature Peruvian
low-land tropical forest and fallow agricultural sites. Biotropica 33:621–636.
Purata, S. E. 1986. Floristic and structural changes during old field succession in the Mexican tropics in relation to site history and species availability. Journal of Tropical Ecology 2:257–276.
Sarmiento, F. O. 1997. Landscape regeneration by seeds and successional pathways to restore
fragile Tropandean slopes. Mountain Research and Development 17:239–252.
Uhl, C. 1987. Factors controlling succession following slash-and-burn agriculture in Amazonia.
Journal of Ecology 75:377–407.
Uhl, C., Buschbacher, R., and Serrao, E. A. S. 1988. Abandoned pastures in eastern Amazonia.
I. Patterns of plant succession. Journal of Ecology 73:663–681.
Vazquez-Yanes, C., and Orozco-Segovia, A. 1993. Patterns of seed longevity and germination in
the tropical rainforest. Annual Review of Ecology and Systematics 24:69–87.
Vieira, D. L. M., and Scariot, A. 2006. Effects of logging, liana tangles and pasture on seed fate of
dry forest tree species in Central Brazil. Forest Ecology and Management 230:197–205.
Webb, S. L., and Willson, M. F. 1985. Spatial heterogeneity in post-dispersal predation on Prunus
and Uvula seeds. Oecologia 57:150–153.
Chapter 10
Facilitation Versus Competition in Neotropical
Old-Fields: A Case Study After Pinus taeda
Cultivation in Brazil
Gislene Ganade, Lessandra Zanini and Ingo Hübel
10.1 Facilitation and Competition in Neotropical Old-Fields
Plantations of exotic tree monocultures for wood production are extremely
widespread in the Neotropics (Richardson 1998). In Brazil, Pinus trees introduced
from Europe and the US have been planted and commercialized for more than half
a century. Most of these monoculture plantations have short clear-cut cycles that
range from 7 to 10 years. Shortly after monocultures are removed, open fields are
re-invaded by Pinus seedlings that regenerate from the seed bank or arrive in the site
by dispersion from nearby parent trees (Higgins and Richardson 1998, Nathan et al.
2002). After clear-cut, a new tree monoculture can be naturally re-established in the
site over a period of 10–15 years (Chilvers and Burdon 1983, Rouget et al. 2001).
Therefore, old field succession after Pinus plantation suffers from the inevitable
threat of aggressive exotic tree invasion (Kulmatiski 2006). The present chapter will
discuss how the early successional vegetation established in these old-field sites
could influence either the likelihood of Pinus invasion or the chances of native tree
species to establish in those sites.
Two mechanisms of plant- plant interactions are considered to be crucial to define
the structure and successional dynamics of old field communities: competition and
facilitation (Callaway and Walker 1997). Competition is the mechanism by which
plants are suppressed by neighbors due to resource depletion while facilitation is the
mechanism by which plants are benefited by the presence of neighbor vegetation
due to improvements of microclimatic and soil conditions (Callaway and Walker
1997). In old fields areas competition seems to be a much more frequent process
influencing succession (Loik and Holl 2001, Myster 2004) while facilitation seems
to occur more frequently in cases in which disturbance creates harsh conditions that
are ameliorated by neighbors (Wilson 1999). Additionally, as old field succession
proceeds, facilitation is suggested to become rarer while competition starts to dominate the plant interaction scenario (Wilson 1999).
The “stress gradient hypothesis” states that facilitation is more expressive in
harsh environments where plants can experience a great deal of stress, e.g. deserts or
alpine areas, while competition would be more important in low stress environments
(Bertness and Callaway 1994). This hypothesis has been confirmed by works based
R. W. Myster, Post-Agricultural Succession in the Neotropics.
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221
222
G. Ganade et al.
on meta-analysis which revisited studies performed in arid and semi arid environments (Gomez-Aparicio et al. 2004, Lortie and Callaway 2006). Nevertheless, productive rainforest areas would be expected by this hypothesis to hold competition as
the most important process driving the interplay of species interactions during succession. Experimental studies, however, have shown that facilitation may be much
more frequent than expected in neotropical old-fields (Vieira et al. 1994, Ganade
and Brown 2002, Zanini and Ganade 2005, Zanini et al. 2006).
Switches from facilitation to competition in a given study site are quite frequently
registered in the literature (Callaway and Walker 1997, Rousset and Lepart 2000,
Alados et al. 2006). The occurrence of these switches may depend on levels of
resource availability and / or environmental stress (Callaway et al. 2002). However, the degree in which a plant suffers from stress does not depend exclusively on
resource conditions. Different life stages of a single plant may respond to environmental stress in distinct ways. For example, old fields can be extremely harsh environments for seedlings but not for adult trees. In old fields of Brazilian Amazonia
seedlings of rain forest species rely on their plant neighbors to improve microclimatic conditions for germination, nevertheless, saplings have their growth reduced
by light competition with neighbor vegetation (Ganade and Brown 2002, Ganade
2007). This same pattern has been registered in old-fields of temperate and sub
tropical areas (De Steven 1991a, 1991b, Gill and Marks 1991, Hastwell and Facelli
2003). It is possible, therefore, that the facilitation process may be quite common in
Neotropical old-fields during seedling establishment phase but may become rarer as
seedlings grow through sapling and adult stages.
We exemplify this argument by a graphic model which shows how the outcome
of facilitation and competition can change depending on the interplay between plant
life stage and environmental stress (Fig. 10.1). For all life stages, facilitation is
more pronounced when environmental stress increases. However, switches from
competition to facilitation along a stress gradient differ for the distinct plant life
stages. Seedlings experience facilitation by neighboring vegetation at a lower level
Fig. 10.1 Graphical model showing the outcome of facilitation and competition correlated with
the interplay between plant life stages and environmental stress
10 Facilitation Versus Competition in Neotropical Old-Fields
223
of environmental stress (Fig. 10.1a) than saplings and adults trees (Fig. 10.1b, c).
In the next section, we show the results of different experiments conducted in a
Neotropical old field developed after the removal of a Pinus taeda plantation. These
results confirm that neighbor vegetation quite frequently improves seedling performance in Neotropical old fields.
10.2 Old Field Succession After Pinus Plantation:
A Study Case in Southern Brazil
We performed various field experiments in an old field area located at the National
Forest of São Francisco de Paula (29◦23‘S; 50◦ 23‘W), RS, southern Brazil
(Fig. 10.2). The reserve covers 1,600 ha and is mostly dominated by native Araucaria forest interspaced with plantations of exotic tree species of the genera Pinus
and Eucalyptus as well as plantations of Araucaria angustifolia (Araucariaceae)
the brazillian pine. The Araucaria forest is characterized by the canopy prevalence
of A. angustifolia. The study site encompasses a 2 ha area that has been formally
occupied by a Pinus taeda plantation for 30 years before clear-cut. In January
2001 the pine trees were cut at soil level and removed from the site for wood
commercialization. After tree removal, the area was characterized by a mosaic of
native early successional grasses and herbs as well as a small number of shrubs.
Tree saplings of various species, including A. angustifolia and P. taeda in which
Fig. 10.2 Map showing the location of the study area in Rio Grande do Sul state, southern Brazil
224
G. Ganade et al.
the height could reach up to approximately 0.5 m were also present at the site. The
edge of the nearest native forest was located at about 50 m from the study area.
10.3 Methods
We present here results of four experimental studies on seedling establishment that
were carried out during the first years of old field succession. The studies aimed
to test the effect of neighbor vegetation on establishment and performance of colonizing woody seedlings. The presence and absence of vegetation was assigned in
3 × 3m plots arranged in a randomized block design using 10 replicated blocks, 20
plots in total. Blocks were located at least 20 m apart. Pairs of plots within each
block were randomly assigned to have vegetation removed or left intact. The vegetation removal treatment comprised the removal of all above ground plants, surface
litter, and all non woody plants that became established throughout the experiment.
Experiment 1: To test the influence of neighbor vegetation on exotic Pinus invasion, we conducted an experiment where seeds of P. taeda were introduced in oldfield plots where early successional vegetation was removed or left intact. A group
of 20 seeds, 5 cm apart, were placed in each treatment and the number of seedlings
established was monitored over an 8 months period.
Experiment 2: To investigate how spontaneous establishment of the woody
species community would be affected by the presence of neighbour vegetation, we
monitored the number and richness of woody seedlings naturally established in plots
were vegetation was removed or left intact. Established seedlings were monitored
within 1 m2 subplots placed inside each 3 × 3m plot. Seedlings were marked counted
and morphotyped at monthly intervals over a 1 year period.
Experiment 3: To investigate which process competition or facilitation with
neighbour vegetation could influence the survival and growth of a late successional
tree we transplanted seedlings of Inga virescens (Mimosaceae) in all 3 × 3 m plots.
Five seedlings averaged 15 cm in height were transplanted in each plot, at least 50
cm apart. The number of seedlings that survived and their heights were registered
monthly during 16 months. Seedling growth was calculated by subtracting the registered seedling height (cm) by its height measured at the beginning of the experiment.
Experiment 4: To investigate how availability of soil nutrients could influence
neighbour effects (competition or facilitation) seedlings of Araucaria angustifolia
(Araucariaceae) were transplanted in the following four treatment combinations:
(1) presence of neighbour vegetation without nutrients (control), (2) presence of
neighbour vegetation with nutrients, (3) absence of neighbour vegetation without
nutrients and (4) absence of neighbour vegetation with nutrients. Treatments were
assigned in 3 × 3 m plots, in a randomized block design using 10 replicated blocks,
40 plots in total. Five seedlings averaged 30 cm in height were transplanted at least
50 cm apart in each plot. Seedling height was registered monthly during 22 months.
Seedling growth was again estimated by the difference between a given seedling
height and its initial height.
10 Facilitation Versus Competition in Neotropical Old-Fields
225
10.4 Results
Seedling establishment
Old-field invasion by the exotic species Pinus taeda was facilitated by neighbour
vegetation (Fig. 10.3). Seedling establishment in the first 2 months did not differ
from vegetation treatments, however, by the third month the number of seedlings
established was significantly enhanced in plots where vegetation was left intact
(significant interaction: time x vegetation treatment, F7,63 = 5.4; P < 0.001, repeated
measures ANOVA).
The establishment of other woody species was also improved by neighbour vegetation. The presence of vegetation increased abundance (F1,38 = 11.4; P < 0.01,
Fig. 10.4) and richness (F1,38 = 8.9; P < 0.01, Fig. 10.5) of native woody seedlings
compared to plots where vegetation was removed.
Survival of I. virescens seedlings was improved by the presence of early successional vegetation. By the end of the experiment, seedling survival was 6 fold higher
in intact vegetation plots in relation to plots where vegetation was removed (significant interaction: time × vegetation treatment F9,162 = 7.7; P < 0.001, repeated measures ANOVA, Fig. 10.6). Growth of I. virescens seedlings was also enhanced by the
presence of neighbour vegetation. Seedling growth was similar between vegetation
treatments in the first 6 months of the experiment, but, by its end, seedling growth
increased 10 fold in plots where vegetation was left intact (significant interaction:
time x vegetation treatment, F15,135 = 12.9; P < 0.001, repeated measures ANOVA,
Fig. 10.7).
Nutrient availability was able to modulate the influence of early successional
vegetation on growth of A. angustifolia seedlings. A significant third order interaction between time, vegetation and nutrient treatments confirmed this results
(F21,105 = 2.295; P < 0.01). Without nutrient application, seedling growth was
higher when vegetation was removed, indicating competition between colonizing
seedlings and neighbour vegetation (Fig. 10.8a). Nevertheless, when nutrients were
applied, seedling growth was higher when vegetation was left intact (Fig. 10.8b),
indicating that once competition for nutrients is overcome due to increases in
6
Pinus taeda
5
4
Presence of
neighbour
Absence of
neighbour
3
2
1
0
F
F
M M A M
Time
J
J
A
S
Fig. 10.3 Mean number of Pinus taeda seedlings established in the presence (vegetation intact,
black symbols) or absence (vegetation removed, white symbols) of neighbour vegetation over a 8
months period. The study was undertaken in a neotropical old field created after a Pinus plantation
clear-cut. Error bars represent ± 1 standard error
G. Ganade et al.
Seedling abundance/m2
226
4
3
2
1
0
Presence of neighbour
Absence of neighbour
m2
Fig. 10.4 Number of native woody seedlings per
established at the presence (vegetation intact)
and absence (vegetation removed) of neighbour vegetation. The study was undertaken in a neotropical old field created after a Pinus plantation clear-cut. Error bars represent ± 1 standard error
Seedling richness/m2
3
2
1
0
Presence of neighbour
Absence of neighbour
Fig. 10.5 Richness of native woody seedling morphotypes per m2 established at the presence
(vegetation intact) and absence (vegetation removed) of neighbour vegetation. The study was
undertaken in a neotropical old field created after a Pinus plantation clear-cut. Error bars represent
± 1 standard error
Inga virescens
Seedling survival
100
80
60
Presence of
neighbour
Absence of
neighbour
40
20
0
M J J A S O N D J F M A M J J O
2002
2003
2004
Time
Fig. 10.6 Mean percentage of survival of Inga virescens seedlings transplanted in the presence
(vegetation intact, black symbols) or absence (vegetation removed, white symbols) of neighbour
vegetation over a 16 months period. The study was undertaken in a neotropical old field created
after a Pinus plantation clear-cut. Error bars represent ± 1 standard error
10 Facilitation Versus Competition in Neotropical Old-Fields
227
Inga virescens
Seedling growth
15
10
Presence of
neighbour
Absence of
neighbour
5
0
-5
J J A S O N D J F M A M J J O M
2002
2003
2004
Time
Fig. 10.7 Mean seedling growth (cm) of Inga virescens seedlings transplanted in the presence
(vegetation intact, black symbols) or absence (vegetation removed, white symbols) of neighbour
vegetation over a 16 months period. The study was undertaken in a neotropical old field created
after a Pinus plantation clear-cut. Error bars represent ± 1 standard error
nutrient available, the presence of vegetation may improve seedling growth in
relation to plots where vegetation is absent.
10.5 Discussion
Facilitation was the most common process occurring in the old field studied, because
early successional herbaceous vegetation, in most cases, improved establishment,
survival and growth of colonizing woody species. Facilitation might have occurred
because vegetation had positive effects on local microclimate, protecting establishing seedlings from low winter temperatures, summer drought and even seed
predators (Zanini and Ganade 2005, Zanini et al. 2006). In this study, however,
facilitation occurred during the growth phase which is a rather rare phenomenon.
Improvements of local microclimatic conditions provided by the pioneer vegetation could be responsible for the facilitation process registered for I. virescens. In
this case, vegetation provided protection against consecutive frosts during winter
months, which directly avoided leaf loss (Zanini et al. 2006).
Facilitation of woody seedling establishment by the herbaceous community has
been reported in old fields of temperate areas (De Steven 1991a, 1991b, Gill and
Marks 1991). However, various studies in Neotropical old fields report that competition with grasses is one of the main factors that restrain seedling establishment,
slowing down forest regeneration (Nepstad et al. 1991, Aide and Cavelier 1994).
This pattern may be due to the pasture fields studied being dominated by highly
aggressive exotic grasses. These grasses have fast growth rates and use a great
deal of water and soil resources that would otherwise be available for seedlings.
Additionally its thick root and litter layer may restrict the availability of microsites
228
G. Ganade et al.
30
Without Nutrients
A
Seedling
25
Sapling
Growth
20
15
Presence of
neighbour
10
Absence of
neighbour
5
0
-5
N D J F M A M J J A S O N D J F M A M J J O
2002
2003
2001
Time
30
With Nutrients
B
Seedling
Sapling
25
Growth
20
15
Presence of
neighbour
Absence of
neighbour
10
5
0
-5
N D J F M A M J J A S O N D J F M A M J J O
2001
2002
2003
Time
Fig. 10.8 Mean seedling growth (cm) of Araucaria angustifolia seedlings transplanted in the
presence (vegetation intact, black symbols) or absence (vegetation removed, white symbols) of
neighbour vegetation over a 22 months period. Graph A represents plots without nutrients (soil
fertilizer not added) and graph B represents plots with nutrients (soil fertilizer added). The first
year of measurement represent seedling phase and the second year represents the sapling phase.
Error bars represent ± 1 standard error
for seedling establishment. In our study, we do not know which particular species
improved seedling performance (Myster and Pickett 1992). We, nevertheless, know
that the average effect of grasses and herbs on woody seedlings was positive (diffuse
facilitation). This pattern may have occurred because the herbaceous vegetation of
this study was composed by less aggressive species which are native grasses and
herbs endemic from highland fields. Experimental works performed in this area have
confirmed that these native species do not show strong competitive responses such
as fast biomass accumulation after nutrient addition (Zanini and Ganade 2005).
As explained by the graph model proposed in Fig. 10.1, seedlings were indeed
more likely to have their performance improved by neighbor vegetation. This is so
because tree seedlings of mid and late successional species have evolved to establish
under the shade of a tree canopy. Thus, although resources such as light may be
10 Facilitation Versus Competition in Neotropical Old-Fields
229
more available during early old-field succession, the benefits of light may be outweighed by the stress of environmental exposure. Therefore, the seedling stage may
be in greater need of a neighbor to improve its microclimatic conditions for survival
than other more advanced developmental stages such as saplings and adults. In this
study, the only case in which competition was revealed was related to the growth
of A. angustifolia saplings. Our results showed that competition did not occur when
transplants were still small seedlings but competition started to occur when transplants achieved a larger size during sapling phase (Fig. 10.8a). However, when
fertilizer was applied competition switched to facilitation (Fig. 10.8b). The results
indicate that nutrient competition by neighbour vegetation starts to occur only when
seedlings achieve a larger size. However, when nutrients are added to the soil and
competition is overcome, the presence of vegetation still improves seedling growth.
Invasion of the exotic pine has also been improved by the presence of early successional vegetation. Indeed, pine invasion in the studied old-field was quite aggressive over time. When naturally established pine were counted 5 years after clear-cut
we found 1,224 pine individuals established at the site, almost half of them were
already in sub-adult stage. By that time the site has also been dominated by shrubs
and small early successional trees. The pairwise effects of these shrubs and trees
on colonization of late successional species are still unknown, however, ongoing
studies have shown that the spatial distribution of Pinus species is negatively correlated with the spatial distribution of early successional trees of the genus Vernonia
(Asteraceae). Future experimental studies may look at possible inhibition patterns.
10.6 Concluding remarks
This chapter provides evidences that mechanisms of facilitation may occur quite frequently in neotropical old-fields. Most of the reasons underlying this pattern relates
to seedlings being more sensitive to environmental stress than other life stages such
us saplings and adult trees (Fig. 10.1). Moreover, early-successional vegetation not
only improves the survival of native tree seedlings but it also improves the survival
of undesirable exotic Pinus seedlings that reinvade old field areas after the Pinus
plantation is clear-cut. These results have important consequences for programs of
restoration and management of such areas. Removal of early successional vegetation
may be applied in situations where desirable species have already established and
achieved sapling phase. Further investigations on species by species interactions are
needed to provide specific information on species that might inhibit exotic Pinus
growth while not affecting native seedlings performance (Myster and Pickett 1992).
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Bertness, M. D., and Callaway, R. M. 1994. Positive interactions in communities. Trends in Ecology and Evolution 9:191–193.
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Ganade, G., and Brown, V. K. 2002. Succession in old pastures of central Amazônia: role of soil
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Hastwell, G. T., and Facelli, J. M. 2003. Differing effects of shade-induced facilitation on growth
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Higgins, S. I., and Richardson, D. M. 1998. Pine invasions in the southern hemisphere: modeling
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Kulmatiski, A. 2006. Exotic plants establish persistent communities. Plant Ecology 187:261–275.
Loik, M. E., and Holl, K. D. 2001. Photosynthetic responses of tree seedlings in grass and under
shrubs in early-successional tropical old fields, Costa Rica. Oecologia 127:40–50.
Lortie, C. J., and Callaway, R. M. 2006. Re-analysis of meta-analysis: support for the stressgradient hypothesis. Journal of Ecology 93:7–16.
Myster, R. W. 2004. Post-agricultural invasion, establishment, and growth of neotropical trees. The
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Chapter 11
Recruitment of Dry Forest Tree Species
in Central Brazil Pastures
Aldicir Scariot, Daniel L. M. Vieira, Alexandre B. Sampaio,
Ernestino Guarino and Anderson Sevilha
11.1 Introduction
Despite its importance as a wealthy source of unique biodiversity (Janzen 1988)
and that it covers 42% of the tropical vegetation worldwide (Murphy and Lugo
1995), the dry forest is the most endangered terrestrial ecosystem and one of the
least protected (Scariot and Sevilha 2005). In Central America for example, less than
1% of the original 500,000 km2 is preserved in conservation units (Janzen 1988,
Sanchez-Azofeifa et al. 2003). In Brazil, the seasonal deciduous forests—a type of
dry forest—covers 27,367,815 ha (3.21%) of the territory, from which 1,072,946 ha
(3.9%) are protected by conservation units. Among conserved areas of deciduous
dry forest, only 117,980 ha (0.43%) are in conservation units of restricted use as
national parks and biological reserves. Most of the conservation units (402,456 ha
or 1.47%) allow limited economic use of the land and natural resources. There is
no information of the conservation unit assigned in official database for 552,509 ha
(2.02%) of dry forests (Sevilha et al. 2004).
Dry forests occur in many soil classes. In Brazil they occur in at least 13 soil
classes of the Brazilian Soil Classification (Scariot and Sevilha 2005), but deciduous
dry forests occur predominantly in rich soils. There are many valuable trees used for
timber. These forests are the main targets of farmers, who log the trees and sell the
timber to acquire financial resources to establish farms. In Central Brazil, where
some of the most important remnants of dry forests of the country occur (Fig. 11.1),
most of the area originally occupied by dry forests has been replaced by pasture
composed by exotic grass in extensive farmlands on rich soils.
Recommendations have been made for the creation of new conservation units, but
given the high degree of landscape fragmentation any new conservation unit will
contain significant proportion of its area covered by pasture and forest fragments
with different levels of disturbance caused by logging, cattle grazing, fire and invasive species. Thus, in this scenario it is fundamental to understand the mechanisms
of tree regeneration in pasture fields to address vegetation recovery and use this
information to facilitate forest regeneration.
There are four to five times more studies on the natural regeneration of rainforests
than on dry forests, and even less on restoration (Vieira and Scariot 2006a). Taking
into account the threat level, the lack of conservation units, the lack of information
R. W. Myster, Post-Agricultural Succession in the Neotropics.
C Springer 2008
231
232
A. Scariot et al.
Fig. 11.1 Map of Brazil showing the distribution of the deciduous forests (shaded areas) and the
location of the Paranã River Valley in Central Brazil
on regeneration mechanisms, and understanding how regeneration can be enhanced
are crucial concerns for the conservation, restoration and management of the dry
forests.
11.2 The Case of Paranã River Valley in Central Brazil
The Paranã River valley has an area of 59,403 km2 in the Cerrado biome. It is
embedded in a transition zone between the wet Amazon region and the semi-arid
Caatinga region. It has a Tropical climate with two well defined seasons (AW of
Köppen) varying to Cwa (Altitude Tropical Climate) (IBGE 1995).
This region presents a mosaic of vegetation types, and has some of the last
remnants of seasonally deciduous forests in Brazil. These occur from 400–600
m above sea level, with annual precipitation of 1,236 mm ± 255 (SD, data from
1969 to 1994). 89% of the rains fall between October and March. The mean annual
11 Recruitment of Dry Forest Tree Species in Central Brazil Pastures
233
temperature is 23◦C, fairly constant throughout the year. The landscape is flat and
consists of limestone geology, with occasional limestone outcrops containing a
moderately distinct tree flora (Silva and Scariot 2003). Nitosoil is the main soil
class upon flat to lightly rugged (Krejci et al. 1982). Rocky limestone outcrops are
common in the region.
This basin is between two well known vascular plant distribution patterns in
Brazil. One of them is formed by forest species with different levels of leaf shedding that depend mainly on the presence of fertile soils within the Cerrado domain
and occurs mainly in the northeast-southeast arch, connecting the Caatinga to the
Chaco border. The other pattern is formed by species from the Amazon and Atlantic
rainforests, crossing the Cerrado biome through riparian forests (Oliveira-Filho and
Ratter 1995).
On the flat lands, the forest canopy is 17–23m height with a basal area of 23–
28m2/ha (Scariot and Sevilha 2005). The authors found 128 tree species in 90 genera
and 41 families. Most species are commonly found in large Brazilian biomes such
as Cerrado, Caatinga, Amazonia and Atlantic Forest (Scariot and Sevilha 2005).
Cattle farming, logging and incipient agriculture are the main economic activities
of the region. Wood has been harvested mainly for corral and fence building, for
the production of woodchip, and for civil construction. The main exploited species
are Astronium fraxinifolium Schott, Myracrodruon urundeuva Fr. All., Schinopsis
brasiliensis Engl. and Tabebuia impetiginosa (Mart. ex DC.) Standl., which have
been commercialized locally, regionally and also in the Southern Brazilian markets
(IBGE 1995).
This region has been acknowledged for biodiversity conservation efforts by
Brazilian researchers and the Government (Brasil 2002). However, despite its
importance, the region is dominated by cattle farms and forest fragments rarely
exceeding 250 ha (Andahur 2001). There are three conservation units, but only the
smallest one, with less than 2,000 ha, is fully implemented.
11.3 Seed Germination in Pastures
Seed dispersal is the main barrier for plant recruitment on abandoned pastures (Holl
1999, Holl et al. 2000, Wijdeven and Kuzee 2000), but it can be overcome either by
direct seeding (Engel and Parrotta 2001, Camargo et al. 2002, Doust et al. 2006) or
by installation of bird perching (Melo et al. 1997, Holl 1998, Duncan and Chapman
1999, Melo et al. 2000). However, after sowing germination can also be highly limited by (1) seed predation (Nepstad et al. 1996, Holl and Lulow 1997, Jones et al.
2003, Myster 2003, 2004a,b); (2) competition with exotic grasses (Desteven 1991,
Holl 1999, Holl et al. 2000) and (3) water shortage (Lieberman and Li 1992, Ray
and Brown 1995), and (4) pathogens (Myster 1997, 2004a). The relative importance
of each factor varies widely among species and studied areas (Holl 1999).
Although studies are scarce, seed predation has been found to have high impact
on seed germination of dry forest species in pastures in Central Brazil (Vieira and
Scariot 2006b), similarly to findings from other tropical regions (Nepstad et al. 1996,
234
A. Scariot et al.
Camargo et al. 2002, Myster 2004a,b). The intensity of seed predation is highly
dependent on the seed species (Osunkoya 1994, Holl and Lullow 1997). Medium
sized seeds (0.2–4 g) are usually more susceptible to rodents, since they are easier
to find and they are easier to manipulate than larger seeds (Price and Jenkins 1986,
Nepstad et al. 1990, Osunkoya 1994, but see Hammond 1995, Holl and Lulow 1997
for different results). Less rodent abundance in pastures may result in less mediumsize seed predation than in forests (Jones et al. 2003, Meiners and LoGiudice 2003).
Small seeds are eaten mainly by insects, such as ants (Nepstad et al. 1990, 1996,
Osunkoya 1994), resulting in higher predation of small seeds in pastures (Nepstad
et al. 1990 1996, Duncan and Duncan 2000, Jones et al. 2003). From 6 tree species
studied in active pastures of Central Brazil dry forests, the small seeded Tabebuia
impetiginosa (0.11g) and Astronium fraxinifolium Schott ex Spreng. (0.02g) were
highly preyed or removed, mostly by ants (Tabebuia: 50% and Astronium: 28%),
when compared to seeds dispersed in forest remnants. In recently abandoned pastures, small seeds are preyed and removed at rates varying from 46–73% under
grass cover, although one species, Guazuma ulmifolia experiences only 1% of seed
loss (Guarino 2004). The large seeded (>7.5g) Cavanillesia arborea (Willdenow)
K. Schum. and Swartzia multijuga Vogel, had over 90% predation by cattle in pastures (Vieira and Scariot 2006b).
Harsh climatic conditions during the dry season and dry spells in the wet season
can be critical for seed survival in dry forest regions (Gerhardt 1994, Ray and Brown
1995, McLaren and McDonald 2003). Desiccation killed ca. 75% of the thin-coated
and high water-containing seeds of Swartzia multijuga and Eugenia dysenterica DC.
in the pasture (Vieira and Scariot 2006b). A small amount of shade in the forests,
however, was enough to prevent seed desiccation. Small seeds with low water content and small mass can survive the harsh environmental conditions of pastures
(Vieira and Scariot 2006b). In the wet season, grasses maintain soil moisture to the
same levels found under deciduous dry forests canopy (22–23%) thus contributing
to a significant increase in seed germination in 6 out of 12 tree species in Central
Brazil pastures (Guarino 2004). Three species (Cedrela fissilis, Amburana cearensis
and Sterculia striata) showed increases in seed germination by 2 to 10 fold under
grass cover (Fig. 11.2).
Pathogen attack was also a significant source of seed mortality, mostly for Aspidosperma pyrifolium. A. pyrifolium and Tabebuia impetiginosa experienced high
pathogen attack in greenhouse seedlings, mainly in most shaded conditions (Vieira
2006). Even though this is well known for rain forest species (e.g. Augspurguer
1983, Dalling et al. 1998), seed mortality by pathogens has been recently reported
for dry forest species as well (Grogan and Galvão 2006, Vieira and Scariot 2006b).
Despite the lack of information, it has been reported that Fusarium and Colletotrichum fungi are the second most important seed mortality agents in abandoned
pastures and early succession areas in Puerto Rico and Ecuador (Myster 1997,
2004a). However, fungi from Phoma and Phyllosticta genera and Colletotrichum
gloeosporioides although abundant on Inga and Cecropia schreberiana respectively,
resulted in modest seedling leaf loss (Myster 2002).
The use of direct seeding in pastures must consider that adequate species choice
and seeding time are fundamental for restoration success. In the Paranã River Valley
GC
40
20
0
40
20
0
GTh
60
60
80
GC
GTh
Triplaris gardnieriana
0
0
100
0
100
Aspidosperma pyrifolium
20
20
20
80
40
40
40
0
20
40
60
80
100
60
80
60
Schinopsis brasiliensis
40
60
80
0
100
0
100
Machaerium scleroxylon
0
100
20
20
80
20
40
40
60
60
80
100
60
Cedrela fissilis
80
100
80
Amburana cearensis
GC
GTh
Pseudobombax longiflorum
Myracrodruon urundeuva
Sterculia striata
Fig. 11.2 Seed germination (mean ( 1 standard error) of nine dry forest tree species in abandoned pastures in the Paranã River Valley, Central Brazil. Shaded bars
indicate buried seeds and opens bars unburied seeds (Grass Cover (GC), Grass thinning (GTh))
Germination (%)
100
11 Recruitment of Dry Forest Tree Species in Central Brazil Pastures
235
236
A. Scariot et al.
dry forests, priority must be given to species with seeds with small mass and low
water content that can tolerate the harsh environmental conditions found in early
successional stages and are less affected by predation and removal.
The positive effect of grass cover on seed germination must be further explored to
better understand seedling establishment and growth. Burying seeds instead of disposing them on the soil surface, seems to be a good strategy to decrease mortality by
predation (Woods and Elliott 2004) and desiccation (Negreros-Castillo et al. 2003)
in agricultural lands previously covered by dry forests and degraded soils (Doust
et al. 2006). However, Guarino (2004) found contradictory results, highlighting that
this issue deserves further investigation.
11.4 Seedling Establishment
Several factors have been indicated to restrict the establishment of forest tree
seedlings in pastures. These factors can act differently in germination and seedling
establishment processes (Schupp 1995). For instance, some species can germinate
under grass cover but will not develop due to grass competition (Zimmerman et al.
2000, Holl 2002). Herbivory by wild (Nepstad et al. 1996, Holl and Quiros-Nietzen
1999, Sweeney et al. 2002) or domestic animals (Posada et al. 2000) have been
considered to affect seedling establishment in pastures. Other factors include limited water and nutrient availability (Aide and Cavelier 1994, Nepstad et al. 1996),
competition with exotic or invasive plants (Holl 2002, Sweeney et al. 2002, Hau
and Corlett 2003), high light availability, high air and soil temperatures (Nepstad
et al. 1996, Loik and Holl 1999, Alvarez-Aquino et al. 2004) and soil compaction
(Nepstad et al. 1996). Among these, grass competition is considered to be the most
limiting, especially in tropical rain forest regions (Holl et al. 2000). Compared
to forest environments, the establishment of seedlings in pastures is considerably
reduced (Aide and Cavelier 1994, Nepstad et al. 1996, Alvarez-Aquino et al. 2004).
In two abandoned pastures originally covered by seasonal deciduous forests in
Central Brazil, tree seedling survival was low but apparently was not related to
grass competition. Despite the low survival, thinning the grass cover of Andropogon
gayanus Kunth. did not interfered on the survival of 6 out of 7 species of planted tree
seedlings after 14 months (Fig. 11.3). Some grass species which develop in tussocks,
such as A. gayanus, leave bare soil patches, covered by a canopy of grass leaves, in
between the grass tussocks. In the water deficient ecosystem of dry forests, grass
canopy improves the microclimate allowing germination and establishment of tree
seedlings in between the grass tussocks (Aide and Cavelier 1994, Guarino 2004).
Pasture temperature is higher in plots where the grass cover is cleared (Gerhardt
1996, Zimmerman et al. 2000). Therefore the cooler and moister environment created by the grass cover seems to balance the negative effects of grass competition
(Gerhardt 1996).
The benefits of grass cover might be even higher than the possible negative effects
of grass competition. One of the 7 planted species in abandoned pastures in Central
Brazil, Talisia esculenta (A. St.-Hil.) Radlk., had lower survival for grass thinning
11 Recruitment of Dry Forest Tree Species in Central Brazil Pastures
237
Fig. 11.3 Percentage survival after 14 months of planting of tree seedlings in two pastures in areas
originally covered by Dry Forests (data from both pastures aggregated). Dark bars indicate no grass
thinning (N = 19 seedlings per species) and open bars indicate grass thinning (N = 57 seedlings per
species; grass thinning once, twice or four times). The only species that showed significant effect
of grass thinning was Talisia esculenta (Pearson Chi-square = 6.4, df = 1, p = 0.01)
treatment (Fig. 11.3). Talisia esculenta seems to survive less in water stressed conditions of bare soil during the dry season and seems to be able to survive under
low light levels created by the grass cover. The soil tends to be moister under the
grass canopy (Hooper et al. 2002), allowing higher survival and growth of some
tree species than in areas with less grass cover (Aide and Cavelier 1994). In dry
forest regions water might be the most limiting resource (Gerhardt 1996, Cabin
et al. 2002, Marod et al. 2002) and the grass cover in pastures might ameliorate
the water deficit for tree regeneration (Aide and Cavelier 1994, Gerhardt 1996).
The removal of dominating shrubs in a deforested seasonal moist evergreen forest
in Uganda affected tree regeneration in a negative way, indicating possible facilitation interactions between trees and shrubs (Duncan and Chapman 2003). This was
similar to the interactions for tree seedlings survival and grass cover found in our
studies in Central Brazil. The interactions among grasses and tree regeneration in
harsh environments such as pastures in dry forests tend to be more a facilitation
relationship rather than competition (Callaway and Walker 1997).
Despite of the apparent positive effect of grass cover on tree seedling survival,
seedling growth can be reduced by grass competition (Holl 1998, Hooper et al.
2002, Schaller et al. 2003). A study carried out by Sampaio et al. 2007 in the same
region in Central Brazil dry forests that cleared out the grass cover leaving only
the naturally established tree regeneration found a higher average increase in height
compared to control plots, indicating competition between natural tree regeneration
(mostly root-suckers) and grasses. The growth of dry forest tree seedlings planted
in greenhouses under different light conditions (10, 40 and 72% of full sun photon
flux density) seems to be higher in the most elevated light levels for the majority of
tested species, however seedling survival did not follow a consistent pattern according to the light gradient (Vieira 2006). Grass root system also inhibits tree root
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A. Scariot et al.
development resulting in low growth due to grass competition (Putz and Canham
1992, Schaller et al. 2003). In areas with pronounced dry season, grass root competition may have decreased seedling survival and growth. However, clearing out the
grass cover (i.e. mowing treatment) can transform pastures in even harsher environments, as detrimental to seedling establishment as grass competition (Hooper et al.
2002).
Light seems to promote seedling growth, and grass cover might out compete tree
seedlings creating shadow, inhibiting tree root development and therefore reducing
growth. In contrast, low light levels as those created by grass cover may not promote
tree seedling survival and some species regeneration may even be facilitated by grass
cover. This may indicate water deficit reduction of bare soil of dry forest regions.
Therefore, light and water interacting with grass cover seem to be promoting tree
seedling establishment at the studied pasture sites in the dry forest regions. These
results suggest that seedlings can be introduced in pastures among grass tussocks
for pasture restoration (as in A. gayanus pastures in Central Brazil, Sampaio et al.
2007). However, grass cover should be cleared just around the seedlings to improve
tree seedling development (Vieira and Scariot 2006a). This will avoid excessive
exposure to the plants under harsh conditions.
11.5 Resprouting Ability
Root-suckers from trees, shrubs and lianas are found in recently established pastures
in tropical forest areas (Uhl et al. 1988, de Rouw 1993, Nepstad et al. 1996). The
density of these shoots can be high enough to allow forest regeneration if not eliminated. Actually, shoot elimination by clipping or by use of selective herbicides and
root collection after plowing, are common management practices in Central Brazil
dry forests region. These practices respond to a large proportion of annual money
and labor expenditures in pasture management. Even though farmers battle forest
plants exhaustively, many dry forest woody plants persist resprouting many years
after pasture establishment (Vieira et al. 2006c, Sampaio et al. in press).
The real contribution of root-suckers relative to seedling establishment for forest
regeneration in pastures along tropical forest areas is neglected. To some degree
this is due to the laborious work need to detect resprouting which involves digging
the superficial soil. The other reason is simply because root-suckers are depleted by
sequential cutting, fire, and intensive tractor use (Uhl et al. 1988, de Rouw 1993,
Sampaio et al. 1993). Much more attention has been given to evaluate limiting
factors and how to improve seed dispersal and survival and seedling survival and
growth (reviewed in Holl 2002, Myster 2004b).
Plowing the soil in the middle of the dry season and seeding grass at the beginning of the rainy season is a common practice among farmers of Central Brazil to
restore grass cover in pastures. Plowing completely eliminates plant cover, allowing
plants to start resprouting soon. Vieira et al. (2006c) investigated the regeneration
from root-suckers after plowing in a 10 year-old pasture, a 25 year-old pasture and
an early successional forest recently clear cut in Central Brazil dry forests. They
11 Recruitment of Dry Forest Tree Species in Central Brazil Pastures
239
found that species richness and density were extremely high compared to other
disturbed tropical sites with comparable history of land use (e.g. studies by Uhl
et al. 1988, Nepstad et al. 1996, Zahawi and Augspurger 1999, Holl et al. 2000).
They found 14–182 individuals per 100 m2 of 30–42 species (6, 5–17, 6 species per
100 m2 ) among the three sites, representing 80% of the species richness found in
intact forest remnants (Vieira et al. 2006c). In another investigation on tree regeneration in active pastures in dry forests of Central Brazil, stem density, regardless of
the origin, did not varied with pasture age (from ≤ 6 up to 40 years old, n=25), while
richness decreased significantly with age (Sampaio et al. in press). Despite the long
time since forest removal, pastures still maintained a significant number of dry forest
species. The 40 year-old pastures still had 12 dry forest tree species regenerating
in relatively high densities (19–58 stems/100 m2 , 2nd and 3rd quartiles). Species
composition was quite similar among pasture ages and pastures and forest remnants
(Vieira et al. 2006c, Sampaio et al. in press), indicating the potential of these areas
for dry forest return.
These optimistic results corroborate others findings that dry forest species are
strong resprouters and contribute to forest regeneration after disturbance. In other
tropical dry forests, studies show a high frequency of resprouting after disturbance
(Kennard et al. 2002, McLaren and McDonald 2003). The reasons why resprouting
is a particularly important recovery mechanism in dry forests are not certain. Some
probable reasons for this trait are (1) seeds have a lower probability of establishment,
so the survivorship by resprouting was favored in this forest ecosystem (Ewel 1980);
(2) trunk bases are less prone to decay, having more chance to resprout (Ewel 1980);
and (3) dry forest plants are adapted for above ground mortality or drought, so they
usually resprout (Sampaio et al. 1993, Bond and Midgley 2001).
Although some species are strong resprouters others are not and resprouting ability will depend on the frequency, intensity and duration of the disturbance (Sampaio
et al. 1993, Kammesheidt 1999, McLaren and McDonald 2003). Information available from actual literature does not allow one to predict what species are able to
resprout in a forest community (Bond and Midgley 2001, Vesk and Westoby 2004).
Despite this limitation on the current knowledge it is essential to understand the
traits or identities of strong resprouters to be able to (1) help to understand present
and future community composition (Kennard 2002, Saha and Howe 2003); (2) use
root and branch cuttings of these species as nurse trees in restoration plans; and (3)
invest in other restoration mechanisms for weak or non-sprouting species in early
successional forests dominated by resprouters.
In dry forests of Central Brazil, species show high density in forest remnants,
but that does not happen in ploughed sites (see above in this section; Vieira et al.
2006c). Species with low-density wood such as Cavanillesia arborea and Chorisia
pubiflora (A. St.-Hil.) G. Dawson. Spondias mombin L. and Pseudobombax tomentosum (C. Martius and Zuccarini) Robyns, and soft wood species were found only in
forest fragments and early successional sites. Species with low-density wood could
lose their resprouting ability faster and sometimes did not resprout at all because
of decay. In a chronosequence of pasture implementation, most species appear to
be able to continue resprouting for up to 11–16 years after disturbance (Sampaio
et al. in press). However, in 40 year-old pastures Eugenia dysenterica dominated the
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community, which was poorer in tree species than younger pastures. E. dysenterica
occurs in Central Brazil dry forests, but occurs in high frequency in the surrounding
savanna vegetation. Species from the Brazilian savannas tend to resprout more than
cogeneric species typical from riparian forests in the Cerrado biome, where root
systems is not as developed as well as in the dry soils of savannas (Hoffmann et al.
2003). Therefore, savanna species like E. dysenterica that develop deep root system
to survive in dry soils of savannas and dry forests might be able to survive longer and
even increase its density via vegetative reproduction in pastures. The composition
of the secondary vegetation developed in abandoned pastures will be dominated by
strong resprouters, probably hard wood and savanna-related species.
Dry forests of Central Brazil show a high degree of resilience after pasture implementation due to root resprouting. Thus, planting seedlings to enhance early succession in abandoned pastures may not be desirable, particularly if standard planting
techniques are used. Soil and vegetation disturbances, such as plowing or mechanical hole digging, reduce the density of naturally regenerating trees and consequently
slow recovery (Sampaio et al. 2007). Non-resprouting species can be efficiently
introduced by cautiously planting nursery-grown seedlings, minimizing damage to
natural regeneration. Considering that resprouting ability is positively related to
branch cutting success (Itoh et al. 2002), using branch cuttings to improve dry forest
recovery should be tested. The rooting ability of branch cuttings was tested, with
relative success, for tropical rain forests (Itoh et al. 2002, Zahawi 2005), but we
have not found published studies for dry forests. Root cuttings taken from fallen
trees from deforested areas may also be tested to enhance dry forest restoration,
since planting them is equivalent to root-suckering after plowing.
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Part III
Restoration, Management, and the Future
of Post-Agricultural Areas
Chapter 12
Microbial Biomass in Native Amazonian
Ecosystems and its Adaptation to Deforestation
and Pasture Introduction and Management
Brigitte J. Feigl, Carlos Eduardo P. Cerri, Carlos C. Cerri
and Marisa C. Piccolo
12.1 Introduction
Changes in biogeochemical cycles after land conversion from native forest to
pasture have been focused in several aspects, like soil organic matter (SOM)
quantity and quality, emissions of greenhouse gases, changes in hydrological and
limiting elements cycles (Feigl, Melillo and Cerri 1995, Matson et al. 1997, Neill
et al. 1997a,b, 2001, Garcia-Montiel et al. 2000). The decomposition of SOM and
the mineralization of nutrients are mediated by microorganisms, and the significant
environmental changes associated with land conversion are quite likely to alter
their three most critical functions, i.e. labile source and immediate sink of C,
N, P and S, and substrate transformations (Dalal 1998). Microbial biomass has
even been proposed as a sensitive indicator of soil quality and soil health (Doran
and Zeiss 2000, Schloter et al. 2003). As with any reservoir, its size is important
not only because it provides an indication of slower, less easily detectable SOM
changes, but also because it represents an important labile pool of plant available
nutrients. Temporal measures of these changes are currently scarce. Users and
policymakers require measures of microbial biomass in order to feed models to
develop sustainable cropping systems.
Assessment in soil microbial biomass (SMB) was facilitated in the last 20 years
by physiological, biochemical and chemical techniques, including chloroform fumigation incubation (CFI) (Jenkinson and Powlson 1976) and chloroform fumigation
extraction (CFE) (Brookes et al. 1985, Vance et al. 1987) among many others. As
pointed out by Dalal (1998), it is difficult to compare soil microbial biomass values
obtained by different methods in different laboratories due to variants of kEC values
(the fraction rendered extractable by fumigation, used to convert the measured data
into biomass C), soil moisture contents, incubation temperatures, SOM content, and
instrumentation and analytical techniques.
Most of the data presented in this chapter were obtained at Nova Vida Ranch
(10◦30‘S, 62◦ 30‘W) in Rondônia state, Brazil (Fig. 12.1). The local climate is
humid tropical, with an annual average temperature of 25.5◦C and precipitation of
2,200 mm (Bastos and Diniz 1982) including a dry season from May to September.
Mean daily temperature for the warmest and coolest months varies less than 5◦ C.
R. W. Myster, Post-Agricultural Succession in the Neotropics.
C Springer 2008
247
248
B. J. Feigl et al.
(10º10’05’’S, 62º49’27’’W
Brasil
Nova Vida Ranch (20.000 ha)
Rondônia State
Mean annual temperature: 25.5oC
Mean annual precipitation: 2,200 mm
Dry season: July to September
Fig. 12.1 Map of the southwestern Amazon region, indicating the localization of Nova Vida
Ranch, origin of most of the microbial biomass data discussed in this chapter
Mean annual relative humidity is 89%. The predominant soil at the ranch is an Ultisol (Moraes et al. 1995), or Argissolo in the Brazilian soil classification (EMBRAPA
1999).
Soil samples under forest and different aged pastures were collected for microbial
biomass quantification in four blocks. In most of the cases, five sub-samples were
taken from the 0–10 and 10–20 cm soil layer of each block and bulked together to
make four composite samples. Microbial biomass was estimated by CFI or CFE
methods, processed by the same lab. When necessary, values were recalculated
applying the kEC =0.30 calculated by Feigl et al. (1995b) for Amazonian soils. Main
soil characteristics are shown in Table 12.1.
Table 12.1 Soil characteristics under forest and pastures of different ages
Site
Layer
pH
Clay
d
Forest
2 month-old pasture
(4 month after burning)
3 year-old pasture
(created 1994)
10 year-old pasture
(created 1987)
15 year-old pasture
(created 1983)
(cm)
0–10
10–20
0–10
10–20
0–10
10–20
0–10
10–20
0–10
10–20
(H2 O)
4.4
4.1
5.7
4.6
6.3
6.0
6.5
6.1
6.2
6.1
(%)
22
30
34
38
34
38
29
35
24
32
cm−1 )
(g
1.3
1.5
1.1
1.3
1.2
1.2
1.4
1.4
1.6
1.6
Ctotal
------g
12.8
7.6
11.7
8.6
13.6
7.8
17.5
9.3
15.5
11.8
Ntotal
kg−1
-----1.09
0.78
0.93
0.70
1.56
0.89
1.33
0.88
1.16
1.10
12.2 The Role of Soil Microbial Biomass in Amazonian
Native Ecosystems
Over three-fourths of the Amazonian rainforest grow on red and yellow clay-like
Oxisols which are acidic and low in nutrients (Moraes et al. 1995, Bernoux et al.
2001). Most of the carbon and essential nutrients are fixed in the living vegetation.
12 Microbial Biomass in Native Amazonian Ecosystems
249
As vegetation dies, dead wood, and decaying leaves are rapidly broken down by soil
biota. Carbon availability thus often controls soil microbial growth, which positively
correlated with aboveground forest litter input. In this way nutrients almost immediately returned to the system as they are taken up by roots that actually grow out of the
ground to form a mat on the forest floor in order to more efficiently collect nutrients.
In spite of its importance, there have been few studies about litter production
and decay in Amazonian rainforests. Total aboveground biomass (excluded dead
wood and litter/rootmat) was estimated by Cummings et al (2002) as about 300 t dry
matter ha−1 . This value corresponds to the mean of 20 sites located in the southwestern Brazilian Amazon. Annual litter production of this ecosystem was measured by
Luizão and Schubart (1987), who reported a range between 6.5 and 7.4 t dry matter
ha−1 y−1 for primary forest areas located near Manaus. Higher values, up to 8.8 or
even 9.7 t dry matter ha−1 y−1 were found by Vasconcelos and Luizão (2004) and
Smith et al. (1998a), respectively, in other sites of Amazonian undisturbed forest.
The rates of decay and pathways of decomposition are determined by the quality of
the litter material, the physical environment and the abundance and composition of
decomposer organisms. Smith et al. (1998b) determined an exponential decay constant (k) of 0.605 for Amazonian forest litter. Potentially, 2–3% of total aboveground
vegetation is recycled every year by microbial biomass.
A comparison of litter amount and microbial population was done at samples
collected during the rainy season (December 1998) and the dry season (August
1999) from a forest site located at Nova Vida Ranch. The forest floor (i.e., L, F,
and H, or O1 and O2 horizons) was sampled from five randomly placed 0.3 m2
quadrates per site. The obtained data (Table 12.2) showed that relative participation
of layers, as well as C:N ratio, remained almost the same over the year. Similar
microbial biomass was supported by unit of plant dry matter over the year (5.04 ×
4.68 g Cmicr kg−1 soil, and 0.79 × 0.72 g Nmicr kg−1 soil during the rainy and the
dry season, respectively), although the difference in quantity of fallen leaves makes
the dry season a larger reservoir of microbial C and N (Table 12.2). Apparently
Table 12.2 Seasonal changes in aboveground litter stand and supported microbial biomass C and
N (kg ha−1 ) under native forest conditions
Litter layer
Litter dry matter
Microbial biomass
Weight
Total C
Total N
Cmicr
Nmicr
--------------------------- kg ha−1 ---------------------------
Ctotal :Cmicr
%
Rainy season
(50%)*
(37%)
(18%)
Total
L
F
H
2,400
1,600
1,300
5,300
1,100
700
400
2,200
38
30
18
86
8.3
12.0
6.4
26.7
1.1
2.6
0.5
4.2
0.75
1.71
1.60
1.21
13.9
23.9
8.1
45.9
2.1
4.4
0.6
7.1
0.66
1.71
1.62
1.15
Dry season
(48%)
(35%)
(17%)
Total
L
F
H
4,700
3,400
1,700
9,800
2,100
1,400
500
4,000
* Relative distribution of litter layers
71
65
22
158
250
B. J. Feigl et al.
moisture was not a limiting factor for microbial population maintenance, suggesting
that faster litter decay during the rainy period can be partially attributed to increase
in the activity of decomposers. Luizão and Schubart (1987) pointed out that the
macro-arthropods appear to be the most affected by adverse humidity conditions
during the dry season. Ruan et al. (2004) also concluded that aboveground plant litter input had no effect on soil microbial biomass or on its pattern of fluctuation, since
it is not directly regulated by soil temperature or moisture at local scales within a
tropical wet forest. The authors found asynchronous fluctuation between soil microbial biomass and plant litter fall due to regulation by competition for soil nutrients
between microorganisms and plants, and regulation by below-ground carbon inputs
associated with the annual solar and drying-rewetting cycles in tropical wet forests.
At the forest site of Nova Vida ranch, microbial biomass in litter represented
a pool of labile C and N of about 36 and 6 kg ha−1 , respectively, over the year.
However, an unknown amount of the recycled material effectively reaches the first
5 cm of soil. Tree roots actively grow into the litter layer and, associated or not to
mycorrhizae, take up the released nutrients. The poorly known root biomass itself
also plays an integral role in forest floor decomposition process and N dynamics
(Smith et al. 1998b).
Microbial biomass in soil is concentrated in the top 5 to 10 cm because substrates are more plentiful there. A survey of sites across Brazilian Amazon region
(Fig. 12.2) showed that microbial biomass C (or Cmicr ) ranged from 200 to 2,000 kg
ha−1 when the upper 10 cm soil layer is considered (Table 12.3).
The soil microbial biomass dynamics in tropical forests and savannas allows
accumulation and conservation of nutrients in biologically active forms during the
8
9
7
1
2
3
4
5
6
Fig. 12.2 Geographic location of the sites referred on Table 12.3. Samples represent major Amazonian soil types according to Moraes et al. 1995 and Koutika et al. 1999
12 Microbial Biomass in Native Amazonian Ecosystems
251
Table 12.3 Survey of soil microbial biomass C and N under native forest. Data obtained by
fumigation-extraction method were recalculated, using kEC =0.30 and kEN =0.54
Site
1
2
3
4
5
6
7
8
9
Season
Late rainy
Late rainy
Late rainy
Dry
Dry
Dry
Dry
Dry
Rainy
Dry
Rainy
Annual
Dry
Rainy
Dry
Dry
Soil
Clay
Layer
Soil microbial
biomass
Cmicr :Corg
Oxisol
Ultisol
Ultisol
Alfisol
Ultisol
Ultisol
Ultisol
Alfisol
Oxisol
Oxisol
Oxisol
Oxisol
Oxisol
Oxisol
Oxisol
Oxisol
(%)
53
67
35
20
10
18
25
23
65
65
70
70
78
35
80
80
(cm)
0–10
0–10
0–10
0–5
0–5
0–10
0–10
0–10
0–5
0–5
0–5
0–5
0–10
0–5
0–10
0–10
(mg C kg−1 )
292
579
426
293
193
407
337
500
1,388
784
1,035
1,287**
780
586
524
970
(mg N kg−1 )
41
125
86
46
30
63
50
74
36*
47*
–
–
102
–
35
111
(%)
0.7
2.3
2.2
1.7
1.4
3.2
3.3
2.5
5.0
2.8
1.4
2.5
3.1
4.5
2.1
2.9
1 Unpublished data (10◦ 05‘S–67◦ 15‘W)
2 Unpublished data (03◦ 16‘S–48◦ 26‘W)
3 Unpublished data (10◦ 14‘S–62◦ 10‘W)
4 Unpublished data (10◦ 44‘S–62◦ 13‘W)
5 Feigl et al. 1995b
6 Carvalho et al. (in prep.)
7 Pfenning et al. 1992, Luizão et al. 1992, Feigl et al. 1995b
8 Pfenning et al. 1992
9 Geraldes et al. 1995, Feigl et al. 1995b
*Determined by ninhydrin reactive compounds
**Determined by fumigation-incubation method, using kC = 0.45
dry period (high biomass, low turnover), when plant activities are low. At the beginning of the rainy period (low biomass, high turnover), nutrients are then released,
being rapidly taken up by plants (Singh et al. 1989).
Factors influencing organic matter decomposition of topsoils from the Brazilian
Amazon Basin were evaluated by several authors. Koutika et al. (1999) concluded
that, independent of geographical location and vegetation cover, C decomposition
rates are lower on topsoils of fine texture with a sand/clay ratio of less than 1,
compared to topsoils of coarse texture, with a sand/clay ration ranging from 2 to 8.
According to Telles et al. (2003) clay content exerts a major control on the amount of
slowly cycling carbon and therefore influences the storage and dynamics of carbon
in tropical forest soils.
12.3 Alteration of Soil Microbial Biomass During
Deforestation Process
The cumulative deforested area in Brazilian Amazonia is estimated to have reached
55,000 km2 by 2000 (INPE 2000). Cattle ranching is thus the leading cause of deforestation in the Brazilian Amazon, since more than 90% of this area has been used
252
B. J. Feigl et al.
for this purpose. When forest is converted to pasture/grazing areas or agriculture,
primary burning after deforestation cleans up the areas and promotes a transient
increase in nutrient availability. This tillage practice prepares the field for sowing and fertilizes the soil by adding an important amount of nutrients in the form
of ashes (Fearnside et al. 2001). In addition, burning organic matter releases the
mineral nutrients as oxides or carbonates, which have alkaline reaction and reduce
soil acidity, promoting higher phosphorus availability, depending on the type and
amount of biomass. These significant environmental changes associated with land
conversion are quite likely to alter the structure and activity of the soil microbial
community, as indicated by Cleveland et al. (2003).
The effect of deforestation on several parameters, including microbial biomass,
was measured during a field experiment realized at Nova Vida Ranch aiming a
better understanding of the early biogeochemical alterations due to land-use changes
(Neill et al. 1997b, Graça et al. 1999, Garcia-Montiel et al. 2001). The slash and burn
procedure was carried out at one hectare area by traditional Amazonian methods, i.e.
vegetation was cut down at the end of rainy season (May) and left on place till end
of the dry season (September). Fire was lit at the edges of the clearing so the flames
would meet in the middle and maximize the effect of burning (Graça et al. 1999).
Microbial biomass C and N, and other eco-physiological indicators were
measured simultaneously in the upper soil layers (0–10, 10–20 and 20–30 cm)
under intact forest, cut-down and dried plant material, and 2 days after burning
(Table 12.4). Native forest soil data had slightly lower microbial biomass C and N
than the perturbed sites, while basal respiration and metabolic quotient (or community respiration per biomass unit) decreased after disturbance. This is an indication
that more carbon and nitrogen, originated mainly by leaves, were available for
biomass production, although changes in percentage microbial carbon (Cmicr ) to
total organic carbon (Corg ) were not found.
Table 12.4 Changes in soil microbial biomass C and N (n = 4, sd in parenthesis) due to clear
cutting, drying and burning native vegetation
Depth
cm
Microbial C
μg C
g−1
soil
Microbial N
μg N g−1 soil
Native forest site
0–10
254.3
(7.0)
42.1
(2.8)
10–20
248.1
(7.8)
50.6
(1.9)
20–30
241.9
(3.5)
31.3
(2.2)
Dry cut-down vegetation site, before burning
0–10
303.9
(11.1)
63.8
(2.2)
10–20
204.7
(3.7)
37.3
(1.9)
20–30
229.5
(5.5)
54.2
(2.4)
Carbonized vegetation site, 2 days after burning
0–10
285.3
(8.9)
55.4
(1.6)
10–20
260.5
(6.7)
55.4
(1.6)
20–30
254.3
(6.2)
40.9
(1.2)
*(ng CO2 -C g−1 soil h−1 )
**(ng CO2-C ug−1 Cmicr h−1 )
C:N
*Respiration **qCO2
Cmicr :Corg
6.0
4.9
7.7
479.2
364.6
208.3
1.9
1.5
0.9
2.2
3.0
5.0
4.8
5.5
4.2
270.8
229.2
145.8
0.9
1.1
0.6
2.3
2.5
4.8
5.2
4.7
6.2
229.2
156.2
187.5
0.8
0.6
0.7
2.4
3.2
5.3
12 Microbial Biomass in Native Amazonian Ecosystems
253
Different from other studies, fire had little immediate reduction effect on microbial biomass, even in the surface 0–10 cm soil layer. A review done by GonzálezPérez et al. (2004) reports drastic decrease mainly of fungal biomass after wildfire.
Increase of available nutrients in soil, mainly in the form of water-soluble components of ash that became available to living organisms, also commented by the
authors, was not observed, since evaluation was done before any rainfall occurred.
In so far, the increase in soil pH associated to increase in exchangeable cations in
soil which affects soil microbial populations could be observed only in later measurements (Table 12.5).
Table 12.5 Changes in soil microbial biomass C and N during early pasture cultivation
Depth
FE-C
FE-N
(cm)
(μg C g−1 soil)
(μg N g(−1) soil)
C:N
*Respiration
3 months after burning, rainy season
Native forest site
0–10
274.7
(2.6)
41.1 (0.9)
6.7
257.3
10–20 285.5
(3.8)
60.6 (1.0)
4.7
157.9
20–30 209.7
(2.7)
51.9 (1.8)
4.0
297.0
Burnt site, before grass seeds were spread
0–10
498.9
(2.7)
81.1 (1.0)
6.2
480.8
10–20 320.3
(6.1)
65.0 (1.0)
4.9
112.2
20–30 224.3
(7.7)
43.7 (1.0)
5.1
111.9
8 months after burning, late rainy season
Native forest site
0–10
252.1
(3.2)
61.1 (0.8)
4.1
324.0
10–20 246.5
(7.2)
53.0 (0.8)
4.7
136.5
20–30 179.3
(2.2)
39.2 (1.2)
4.6
230.2
5 months-old pasture site
0–10
310.0
(63.2)
43.0 (17.4)
7.2
181.6
10–20 257.7
(21.5)
56.5 (6.1)
4.6
171.2
20–30 190.5
(60.3)
44.5 (5.9)
4.3
171.2
15 months after burning, rainy season
Native forest site
0–10
548.9
(5.7)
59.7 1.01
9.2
483.7
10–20 205.5
(2.9)
33.8 1.60
6.1
251.5
20–30 162.5
(2.8)
37.0 1.30
4.4
263.3
12 months-old pasture site
0–10
276.6
(31.6)
45.6 10.59
6.1
188.6
10–20 219.9
(38.2)
32.7 4.14
6.7
259.2
20–30 179.3
(33.2)
27.7 4.23
6.5
253.8
27 month after burning, rainy season
Native forest site
0–10
328. 3 (2.5)
47.5 (2.6)
6.9
140.5
10–20 208.3
(5.6)
48.8 (1.8)
4.3
61.5
20–30 169.5
(10.8)
34.7 (0.9)
4.9
63.1
24-months old pasture site
0–10
244.7
(28.1)
44.5 (4.3)
5.5
159.0
10–20 300.9
(27.1)
52.3 (7.4)
5.8
141.7
20–30 190.3
(46.9)
43.8 (4.5)
4.3
87.0
* (ng CO2 -C g−1 soil h−1 )
**(ng CO2 -C ug−1 Cmicr h−1 )
**qCO2
Cmicr :Ctotal
0.9
0.6
1.4
2.0
3.5
4.4
1.0
0.4
0.5
2.5
2.6
2.3
1.3
0.6
1.3
1.8
3.0
3.7
0.6
0.7
0.9
1.6
2.1
1.9
0.9
1.2
1.6
3.9
2.5
3.4
0.7
1.2
1.4
1.4
1.8
1.8
0.4
0.3
0.4
2.3
2.5
3.5
0.6
0.5
0.5
1.2
2.4
1.9
254
B. J. Feigl et al.
Three months after burning, during rainy season, soils from burned sites had
higher soil microbial biomass and activity than soils under forest. Microbial C and
N in the 0–10 cm layer were about 90 and 100% greater, while respiration increased
93%. However, pools of readily available substrates were rapidly depleted. Eight
months after burning, when grass plants were about 5 months old, microbial N
already decreased to levels similar to forest. Measurements done 1 and 2 years
after pasture plantation showed differences mainly in microbial C and Cmicr :Ctotal ,
indicating that the young pasture site had less C available for population growth.
12.4 Evolution of Soil Microbial Biomass Content Under
Aging Pastures
Differences in litter composition, besides environmental factors like temperature
and humidity, determine changes in microbial biomass after forest conversion to
pasture. In general, higher rates of cellulose and hemicellulose are associated with
higher, whereas phenols are associated to lower microbial biomass.
An evaluation of Cmicr and Nmicr in the litter compartment of a 10 year-old pasture showed that forest (Table 12.2) and pasture litter (Table 12.6) maintained a
similar level of Cmicr over the year. However, pastures showed greater Ctotal :Cmicr
ratio, since total litter amount, as well as total C and N content, are lower. Pasture
Nmic followed the widely observed tendency of reduced N dynamic under altered
conditions (Neill et al. 1997b, Ellingson et al. 2000, Garcia-Montiel et al. 2001).
Soil microbial biomass content showed no significant changes which could be
attributed to aging of pastures (Table 12.7), at least during wet season. However,
Cmicr was more affected by dryness in the older pasture, probably due to a less
effective protection against water evaporation provided by plant cover. As can be
verified in Table 12.8, during the dry season the soil of the 15 year-old pasture
suffered a reduction in water content of more than 70% in relation to the wet season. In comparison, soil humidity under forest decreased 55% in the 0–10 cm soil
layer. Nmicr and Pmicr were less affected by seasonality, although the amount of
Table 12.6 Changes in litter microbial biomass C and N in a 10 year-old pasture
Litter layer
Litter dry matter
Microbial biomass
Weight
Total C
Total N
Cmicr
Nmicr
---------------------------- kg ha−1 --------------------------Rainy season
(56%)
(44%)
Total
L
F+H
2,100
1,900
4,000
900
700
1,600
(53%)
(47%)
Total
F+H
4,600
4,100
8,700
1,900
1,600
3,500
10
13
23
Ctotal :Cmicr
%
12.1
13.4
25.5
1.4
1.8
3.2
1.34
1.91
3.25
19.9
27.8
47.7
2.4
3.4
5.8
1.05
1.74
2.79
Dry season
* Relative distribution of litter layers
28
29
57
12 Microbial Biomass in Native Amazonian Ecosystems
255
Table 12.7 Changes in soil microbial biomass C, N and P under forest and pastures of
different ages
Site
Forest
3 year-old pasture
10 year-old pasture
15 year-old pasture
Soil layer
Cmicr
(cm)
----------------------------------- kg ha−1 -------------------------Early wet season
Nmicr
0–10
10–20
0–10
10–20
0–10
10–20
0–10
10–20
633.7
577.9
609.7
486.5
628.3
649.1
523.0
411.8
0–10
10–20
0–10
10–20
0–10
10–20
0–10
10–20
520.4
348.3
357.5
248.8
364.5
300.2
265.0
197.4
(42.5)
(10.2)
(31.2)
(54.2)
(70.7)
(12.4)
(35.9)
(35.5)
Pmicr
133.0
101.9
58.5
57.0
137.2
65.3
74.5
52.5
(16.5)
(7.7)
(9.1)
(12.5)
(13.6)
(12.5)
(6.9)
(5.7)
22.4
15.5
19.5
12.5
19.0
13.9
15.2
11.4
(1.2)
(0.0)
(0.0)
(1.4)
(0.0)
(1.4)
(2.8)
(1.4)
107.4
95.5
52.0
40.3
99.3
52.8
74.5
49.7
(4.7)
(11.6)
(6.5)
(11.1)
(23.1)
(8.3)
(16.6)
(12.8)
20.1
14.2
18.2
12.5
16.3
8.3
11.0
7.1
(1.2)
(2.6)
(1.3)
(0.0)
(1.4)
(1.4)
(0.0)
(1.4)
Dry season
Forest
3 year-old pasture
10 year-old pasture
15 year-old pasture
(75.5)
(43.9)
(53.3)
(26.4)
(70.7)
(16.7)
(30.4)
(25.5)
immobilized N followed the general tendency of the N cycle to be lowered after
deforestation.
The replacement of the forest, which has mostly C3 photosynthetic pathway vegetation, by C4 pasture grasses results in an enrichment in 13 C of the soil organic
matter over the years, observed by several authors (Feigl et al. 1995a, Moraes et al.
1996, Bernoux et al. 1998a). Isotopic composition of microbial biomass follows this
pattern (Table 12.9), since it preferentially uses the most available organic matter
and even concentrates C derived from pasture (Fernandes et al. 2007). Changes in
13 C
micr can already be seen 3 years after pasture plantation. At this time, about 40%
of microbial C of the upper 10 cm soil layer was obtained from grass plants. Ten
years after, this food source rose to 80% and apparently stabilized in this range over
the next 5 years. 13 Cmicr in the 10–20 cm soil layer changed slower but continuously
during the considered period.
Table 12.8 Water content in soil layers under forest and pastures of different ages
Site
Forest
3 year-old pasture
10 year-old pasture
15 year-old pasture
Layer
(cm)
0–10
10–20
0–10
10–20
0–10
10–20
0–10
10–20
Wet season
Dry season
kg−1 )
---------- Humidity (g
225.8
102.6
237.7
120.1
200.2
88.3
211.3
111.2
231.4
91.6
221.1
111.1
182.8
54.9
172.7
72.9
---------
Difference
(%)
55
49
56
47
60
50
70
58
256
B. J. Feigl et al.
Table 12.9 Changes in the δ13 C of soil microbial biomass and calculated contribution of C4 plants
to its maintenance (adapted from Fernandes et al. 2007)
Site
Forest
3 year-old pasture
10 year-old pasture
15 year-old pasture
Layer
δ13 C in MB
Cmicr derived from C4 plants
(cm)
0–10
10–20
0–10
10–20
0–10
10–20
0–10
10–20
(%)
–25.9
–25.5
–21.0
–22.8
–15.8
–18.5
–15.3
–16.1
(%)
38
22
78
56
82
75
Adapted from Fernandes et al. (2007)
12.5 Pasture Soil Management Practices (Plowing,
Crop Desiccation, Fertilization) Influencing
Soil Microbial Biomass
Intensification of agriculture is emerging as a major issue in the Amazon Basin
because it is likely to have broad-scale biological and biogeochemical effects. The
focus is to make established pastures more productive through a variety of treatments including application of herbicides, disking and replanting of pasture grass.
An experiment to test the different strategies was conducted on an existing 63
ha area of pasture in process of degradation located at Nova Vida Ranch. Details
of site characterization, experimental design and biological status were given by
Cerri et al. (2004a, 2005) and Feigl et al. (2006). In short, here we considered three
pasture reformations arranged in four blocks (replicates), besides the control plots.
The four treatments were control (C), tillage (T), one cycle of no-till soybean (S)
followed by grass renovation, and selective herbicide (H) application. Treatment
C did not receive any management than mowing at 40 cm high. Treatment T was
plowed twice, fertilized with N, P and K and had Brachiaria brizantha seeded in
early rainy season, by November 2001. Treatment S involved complete desiccation
of degraded pasture vegetation, fertilizer application like T except for N, no-tillage
seeding of soybean followed by grass cultivation. Treatment H had the standing
weeds eliminated by selective herbicide and application of N, P and K.
Microbial biomass C and N was measured during pasture restoration process
by Augusti (2004). Both parameters were decreased by tillage treatment by about
25% in relation to the control during the first month, until the grass was seeded
and fertilized (Fig. 12.3). After that, Cmicr recovered faster than Nmicr , although
with no clear pattern in relation to the control. The grass dessecation (7.5 l ha−1 of
glyphosate) process itself had no significant effect on Cmicr or Nmicr of treatment S.
An increase after fertilization was observed mainly for Nmicr , even no mineral N was
added. Soybean plants maintained Nmicr about 20% higher than control during cropping period (Fig. 12.3). Selective herbicide application (4.0 l ha−1 of 2,4-D Triet.
+ picloram) sharply decreased both, Cmicr and Nmicr , an expected effect, according
12 Microbial Biomass in Native Amazonian Ecosystems
Microbial biomass (µgg C g-1 soil)
700
257
C = Control
(A)
650
T = Tillage
S = Soybean
600
H = Herbicide
550
500
450
400
350
300
250
* * ** *
*** ** ****** **
*
**
*
*
23/3/2002
9/3/2002
16/3/2002
2/3/2002
23/2/2002
9/2/2002
16/2/2002
2/2/2002
26/1/2002
19/1/2002
5/1/2002
12/1/2002
29/12/2001
22/12/2001
8/12/2001
15/12/2001
1/12/2001
24/11/2001
17/11/2001
3/11/2001
10/11/2001
27/10/2001
20/10/2001
6/10/2001
Microbial biomass (µg N g-1 soil)
100
13/10/2001
200
(B)
90
80
70
60
*
* ** * *
*
*
*
* * ** ** * *
*
*
**
50
*
*
C = Control
T = Tillage
S = Soybean
40
H = Herbicide
S1
T2
S2 H1
H2, T3, S3
T 1 = Plow
T 2 = Grass seeded, fertilization
T 3 = Fertilization
T 4 = Selective herbicide
S4 T4
S 1 = Pasture dessecation
S 2 = Soybean seeded, fertilization
S 3 = Selective herbicide
S 4 = Fertilization
S 5 = Soybean harvest
23/3/2002
16/3/2002
9/3/2002
2/3/2002
23/2/2002
16/2/2002
9/2/2002
2/2/2002
26/1/2002
19/1/2002
5/1/2002
12/1/2002
29/12/2001
22/12/2001
15/12/2001
8/12/2001
1/12/2001
24/11/2001
17/11/2001
10/11/2001
3/11/2001
27/10/2001
20/10/2001
6/10/2001
T1
13/10/2001
30
S5
H 1 = Selective herbicide
H 2 = Fertilization
Fig. 12.3 Changes in Cmicr (A) and Nmicr (B) of the 0–5 cm soil layer during pasture restoration
process
to Prado and Airoldi (2001). Subsequent fertilization, which included mineral N,
returned Cmicr to levels similar to control treatment, but Nmicr remained lover than
control over the next 3 months. It is likely that in this tropical Ultisol fertilizer and/or
legume-derived N input stimulated C and N mineralization in the surface depth by
increasing microbial activity. Similar results were obtained by Balota et al. (2004).
258
B. J. Feigl et al.
12.6 Simulating Total Soil C and Microbial Biomass C in a
Forest to Pasture Chronosequence Using the Century Model
The effects of the conversion of tropical forest to pasture on total soil C was analyzed
by Cerri et al. (2004b), using the Century ecosystem model and chronosequence
data collected from the Nova Vida ranch. First, the model was applied to estimate
equilibrium soil organic matter levels, plant productivity and residue carbon inputs
under native forest conditions. Then the model was set to simulate the deforestation
following slash and burn. Soil organic matter dynamics were simulated for pastures
established in 1989, 1987, 1983, 1979, 1972, 1951 and 1911. The Century model
predicted that forest clearance and conversion to pasture would cause an initial fall
in the stock of soil C, followed by a slow rise to levels exceeding those under native
forest (Fig. 12.4). The model predicted the longer-term changes in soil C under
pasture close to those inferred from the pasture chronosequence. Mean differences
between the simulated and observed values for the pasture chronosequence were
about 17% for the total soil C (data not shown). Although variability in soil C observations for young pastures was high (Fig. 12.4), statistical evaluations suggested by
Smith et al. (1996, 1997) show a satisfactory agreement between the simulated and
measured data.
In addition, Century accurately simulated the level and long-term dynamics of
microbial biomass C measurements for the pasture chronosequence, providing a
further test of model performance. The active SOM fraction as conceptualized in
Century includes microbes, microbial products and labile organic substrates (Parton
5000
-2
T o t a l s o il C ( g m )
4000
3000
2000
1000
0
Native
vegetation
20
40
60
80
100
Chronosequence years
Fig. 12.4 Simulated (continuous line) and observed (square symbols) total soil C in the 0–20 cm
layer from the 7 pasture sites at Nova Vida Ranch (modified from Cerri et al. (2004b)
12 Microbial Biomass in Native Amazonian Ecosystems
259
et al. 1987) and can be roughly used to represent microbial biomass and associated
metabolites.
Figure 12.5 shows simulated and measured values for soil microbial biomass C in
the 0–10 cm layer of the Nova Vida chronosequence. Century simulates soil carbon
dynamics for the 0–20 cm layer and our microbial biomass data were determined for
the 0–10 cm layer (Feigl et al. 1995b, Fernandes 1999). Therefore, we multiplied the
simulated results by 0.65, which is the percentage (65%) of the total soil C present
in the 0–10 cm layer in relation to the total soil C contained in the 0–20 cm layer, to
make the comparison.
The simulated quantities of soil microbial biomass C fit the measured data reasonably well (r = 0.84). The model under-estimated the concentration of microbial
biomass C for the youngest pastures (created in 1989 and 1987) and overestimated
values for the oldest pastures created in 1911 (Fig. 12.5), reflected by the mean
difference between observations and simulation equals to 1.28 g C m−2 . However,
Century simulated reasonably well (CD = 1.01, CRM = 0.021, and RMSE = 17.20)
the soil microbial biomass C for the other pasture sites.
The general trend over time in microbial biomass is well represented by the
model, but as with the C and N data from the chronosequences, there is considerable
variability in the observed data which is likely associated with factors such as differences in soil temperature and moisture (Steudler et al. 1996), pasture grass status
(Moraes et al. 1995) and pre-existing differences in the total amount of soil carbon
between sites at the time of clearing (Neill et al. 1997a, Bernoux et al. 1998b).
In order to evaluate soil temperature and moisture effects we compared simulated
and measured values for soil microbial biomass-C in the 0–10 cm layer of five soil
-2
Soil microbial biomass (g m )
100
80
60
40
20
0
Native
0
vegetation
20
40
60
80
100
Chronosequence years
Fig. 12.5 Simulated (continuous line) and observed (square symbols) soil microbial biomass C in
the 0–10 cm layer at Nova Vida Ranch treated as a chronosequence (compiled from Cerri et al.
(2004b)
260
B. J. Feigl et al.
types under degraded pastures within the Nova Vida ranch area in both, wet and
dry season . Measured soil microbial biomass-C contents varied greatly from 35.4
to 235 g C m−2 in the wet season. In the dry season, biomass C values showed a
much smaller variation, ranging between 17.8 and 46.5 g C m−2 (Table 12.10). Our
measured data are consistent with those reported from Amazonian soils by Feigl
et al. (1995b) and Fernandes (1999). Results showed that soils have important differences in their soil microbial biomass contents within the same season and also
between climate seasons.
The comparison of simulated (active SOM pool) and observed microbial biomass
C showed that the model tended, in general, to underestimate the values for the wet
season (excepted for the soils TP and XA) and overestimate for the dry season. Also,
Century simulated very similar absolute biomass C values for both seasons. This
is indicative of the fact that the active pool in Century is simulated as an organic
matter fraction, governed by first-order kinetics. Thus, the fact that the simulated
active pool was higher in the dry season than in the wet season is likely due to a
reduced decomposition rate during the dry season. However, the observed microbial
biomass shows a much more dynamic seasonality and likely a response to dry season
moisture stress that is not captured in the model. Based on these comparisons, we
conclude that the Century model is not totally suitable for studies of soil microbial
biomass dynamics in different weather season conditions.
Finally, pasture development has become the largest anthropogenic disturbance
of forests in the Amazon region. While our analysis represents a single ‘case study’
of pasture conversion, our results suggested that modeling approaches can help to
evaluate the magnitude of impacts on C cycling, and determine the effect of management strategies on pasture sustainability.
Table 12.10 Observed and simulated soil microbial biomass values in the 0–10 cm layer for the
five soil types within the 63 ha pasture area at Nova Vida Ranch
Soil type
Observed values
Simulated values
Difference
------------------------- (g m−2 ) --------------------------
Mean difference
------- (%) -------
Wet season
TP
TA
RA
HX
XA
35.4
124.2
215.3
235.0
64.5
(2.9) c
(18.7) b
(42.9) a
(15.3) a
(6.7) c
60.4
51.1
58.3
55.7
65.2
−25
73
157
179
−1
−71
59
73
76
−1
17.8
12.5
25.4
23.7
46.5
(1.4) c
(2.5) c*
(2.3) b*
(2.1) b*
(4.6) a
79.8
73.7
77.7
76.5
82.2
−62
−61
−52
−53
−36
−349
−488
−206
−222
−77
Dry season
TP
TA
RA
HX
XA
Values in parenthesis are standard errors.
Means within the same season followed by different letters differ significantly at P<0.05
An asterisk indicates a statistically significant difference (P<0.05) between seasons.
12 Microbial Biomass in Native Amazonian Ecosystems
261
12.7 Final Considerations
The present knowledge on microbial biomass in natural Amazonian ecosystems
and its adaptation to deforestation and pasture introduction is still not very well
understood. It would be necessary to better understand interactions between soil
microbial abundance and activity, the quality and quantity of soil organic matter, and
the effects of environment conditions. Study cases are disperses and do not cover the
large combination of soils, vegetations and present climate conditions. In order to
better explore this issue and scale it up, some aspects should be taken into consideration. For instance, tropical area could be split into homogeneous edapho-climatic
zones. Differences in methodological aspects for microbial biomass measurements
are another example that makes difficult to make extrapolations.
The situation gets even more complex once we include climate change scenarios.
In a global warming environment it could be expected that an increase in microbial
biomass activity would lead to accelerated soil organic matter decomposition and
consequently, reducing soil carbon stocks and exacerbating greenhouse gas emissions. Nevertheless, other factors like increased net primary production, extreme
events (rainfall, storms, temperature peeks), and adaptation capacity or resilience
have to be taken into account to adequately predict soil microbial community
dynamics under global change.
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Chapter 13
Management for Sustainability and Restoration
of Degraded Pastures in the Neotropics
Florencia Montagnini
13.1 Introduction
According to recent statistics, today the total agricultural land area in the Neotropics
amounts to about 500 million ha or nearly 30% of its territory, including crop and
pastures (FAO 2000). Pastures and agrosilvopastoral land represent 77% of the total
agricultural land, mostly on poor acid soils. Crops cover the remaining 23%, which
are mostly located on better quality soils. The Neotropics have a pasture/crop land
ratio of 3.4, higher than the world ratio of 2.3 (Amézquita et al. 2005).
Conversion of forests to crops and pastures has been the most important land–use
change in the Neotropics during the second half of the twentieth century (Kaimowitz
1996). Cattle and pastures replaced large amounts of tropical forests during the
decades of the 1960s, 1970s and 1980s, especially in Central America (Kaimowitz
1996, Harvey et al. 2005). A similar pattern has been reported over large areas of the
Neotropics such as in the Brazilian Amazon (Hetch 1993, Dias-Filho et al. 2001).
Among soil nutrients, phosphorus is generally most critical to pasture productivity. Soil phosphorus availability increases with each successive burning, but at lower
and lower levels. When pasture productivity declines significantly as a result of
degradation, the cattle are taken elsewhere or the activity is discontinued altogether.
It is estimated that about 30% of these pastures are degraded as a result of stocking levels being higher than carrying capacity (Szott et al. 2000). Other sources give
estimates of degradation on more than 60% of the pasture area of the Neotropics
CIAT (1976–2000). Although pasture degradation in the Amazon basin has received
the most attention, the surface area affected by pasture degradation may be much
larger in the savannas of Bolivia, Brazil, Colombia, Guyana and Venezuela than in
the Amazon rainforest (Vera et al. 1998). Locally, the extent of pasture degradation
can be even higher: for example in the northern Caribbean zone of Costa Rica,
more than 70% of the pastures were at advanced stages of degradation in the late
1990s (Jansen et al. 1997, CATIE 2002). Differences in perception of the degree and
extent of degradation by farmers and technical personnel can account for significant
differences in these estimates (Holmann et al. 2004).
Pasture development has been ranked one of the least desirable land uses of tropical forest regions because of land degradation caused by overgrazing. Formation of
gullies can be a clear symptom of overgrazing. Gullies may appear in the landscape
R. W. Myster, Post-Agricultural Succession in the Neotropics.
C Springer 2008
265
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F. Montagnini
in very humid regions following heavy rains due to lack of soil water infiltration
caused by compaction. Furthermore, land affected by gullies is very difficult to
restore.
In regions where natural grasses are used for cattle grazing, these grasses often
have low nutrient content and are not preferred by cattle. In such situations, pastures
are burnt periodically to encourage regrowth and also to increase nutrient levels in
soils. The tender regrowth of pastures is eaten by cattle, and pasture productivity
increases temporarily (Serrao and Toledo 1990, Nepstad et al. 1991a, Dias-Filho
et al. 2001).
Abandoned pastures can remain in a state of arrested succession due to a number
of factors that delay or impede their recovery (Uhl et al. 1988, Nepstad et al. 1991a,
1991b, Reiners et al. 1994, Aide et al. 1995, Dias-Filho et al. 2001). However, there
are strategies that can be used to successfully manage tropical pastures to increase
sustainability and avoid degradation. In cases where pastures have been abandoned
due to low productivity, there are also a number of strategies that can be used to
recover the land with the objectives of converting it back to forest and restoring
plant and animal diversity. This chapter examines some of the factors involved in
pasture degradation, and presents experiences used to increase pasture sustainability
and restore degraded pastures in the Neotropics.
13.2 Pasture Degradation and Regeneration
Following Abandonment
13.2.1 Factors Resulting in Pasture Degradation
The progressive displacement of cattle raising toward marginal areas of lower productive capacity is a generalized phenomenon in the Neotropics (Holmann et al.
2004). The low availability of adapted forage materials of high productivity together
with the inadequate management of pastures has led to a rapid decrease of productivity and low revenues to livestock farmers, leading to pasture abandonment and/or
conversion to other land uses.
Pasture degradation is defined as the temporary or permanent decrease of the
soil productive capacity (Stocking and Murnaghan 2001). Pasture degradation can
result from the interaction of the following factors: (1) establishment of pastures on
fragile soils, e.g. on infertile soil or on steep slopes; (2) use of natural grasses of
low productivity, or sowing of poorly adapted species; (3) soil compaction (use of
soils that are shallow and prone to compaction, excess stocking, and improper management); (4) soil erosion (as result of compaction and establishment of pastures
on steep slopes); (5) decrease in soil fertility (especially P, N), without addition of
fertilizer; (6) uncontrolled and frequent burning; (7) overgrazing during the rainy
season due to excess stocking, and use of inadequate animals or breeds; and (8)
weed invasion (as result of the previously mentioned factors). The degradation of
pastures has serious consequences for the producer: first, it reduces milk and/or
beef yields and second, it increases production costs (Holmann et al. 2004).
13 Sustainability and Restoration of Degraded Pastures
267
Even in pastures that have been established on adequate soil and with proper
management, changes in soil productivity may lead to pasture decline and eventual degradation unless other management techniques are introduced early enough
to reverse the process. Nitrogen deficiency is the first factor that destabilizes the
pasture, initiating the degradation process (Holmann et al. 2004). After nitrogen
deficiency takes place, the quality and vigor of the pasture begin to decline and
other nutrients, such as phosphorus and sulfur, can become deficient. When the
pasture begins to lose vigor, invasive weeds appear, increasing the severity of the
problem (Spain and Gualdrón 1991, Dias-Filho et al. 2001).
After a prolonged period of grazing, important changes in the physical structure
of the soil may appear, such as compaction, which increases water runoff, diminishes root development, and decreases the extraction of nutrients located deeper in
the soil. In addition, water runoff initiates the development of erosion and the pasture
enters a process of severe degradation (Hoyos et al. 1995).
13.2.2 Obstacles to Natural Forest Regeneration
in Abandoned Pastures
Following degradation and abandonment of pastures, the efforts needed to recover
the site for continued use as pasture, or to restore the previous forest cover, depend
on several factors. Some are related to specific site conditions prior to its use as pasture, and others depend on length of use as pasture and on type of pasture management. In general, the characteristics and rate of natural forest regeneration depend
on the intensity, duration and specific nature of previous land use. The factors limiting tree regeneration can include scarcity of nutrients, soil compaction, lack of or
excess of soil humidity, high solar radiation, and intra- and inter-specific competition (Uhl et al. 1988, Nepstad et al. 1991a, 1991b, Aide et al. 1995, Dias-Filho et al.
2001).
13.2.3 Seed Dispersal and Seedling Establishment
Availability or arrival of seed from trees can be a limiting factor to tree regeneration,
especially in sites whose distance to seed sources may limit propagule dispersal
(Holl 1999, 2002, Wijdeven and Kuzee 2000). In areas that are artificially seeded or
that receive a natural seed input by seed dispersal agents or from nearby forests, still
other factors can impede or delay seed germination and seedling establishment and
growth following seed arrival. Seed and seedling mortality may occur as a result of
fungal diseases or seed predation. In addition, the specific microsite where a seed
falls may not fulfill the environmental requirements for seed germination, seedling
growth, and seedling survival (Nepstad et al. 1991a, 1991b).
Competition for space, nutrient and light resources by aggressive vegetation may
severely impede the regeneration of forest seedlings in abandoned pastures. For
268
F. Montagnini
example, at La Selva Biological Station in the Caribbean lowlands of Costa Rica,
very few woody species were found in abandoned pastures left to natural regeneration, in comparison to pasture areas planted with native trees (Figs 13.1 and 13.2).
High light availability in the open pastures favored the growth of heliophillous
species, such as grasses and ferns, impeding the growth of woody species (Cusack
and Montagnini 2004). This effect was more pronounced 3–4 years following pasture abandonment (Guariguata et al. 1995), but dominance by grasses and ferns
still persisted at 10–14 years after abandonment (Carnevale and Montagnini 2002,
Cusack and Montagnini 2004, Orozco Zamora and Montagnini 2007a, 2007b). Factors that impeded the establishment of woody species in open pastures were a combination of lack of perches for seed dispersers and invasion by herbaceous vegetation
which out-competed the tree seedlings in their growth (Montagnini 2001, Cusack
and Montagnini 2004, Orozco Zamora and Montagnini 2006).
13.2.4 Intensity and Duration of Pasture Use
Intensity and duration of use are key factors influencing pasture recovery to forest.
Recovery to forest is considerably slower following disturbances that heavily impact
the soils as well as the aboveground vegetation. Activities such as bulldozing, heavy
or long-term grazing, and severe fires, often have long-lasting effects on species
Fig. 13.1 At La Selva Biological Station in the Caribbean lowlands of Costa Rica, very few woody
species were found in abandoned pastures left to natural regeneration, which were dominated by
grasses and ferns even after 15 years following pasture abandonment. A 15-year-old plantation of
Jacaranda copaia can be seen at the back. Photo: F. Montagnini
13 Sustainability and Restoration of Degraded Pastures
269
Fig. 13.2 At La Selva Biological Station, abandoned pasture areas planted with native trees developed a diverse understory including several woody species. This mixed-species plot is composed
of Jacaranda copaia, Calophyllum brasiliense and Vochysia guatemalensis. The plantation was 15
years old. Photo: F. Montagnini
composition (Chazdon 2003). Several examples illustrate this point, often comparing forest recovery following pastures and other previous land uses. For example,
Ferguson et al. (2003) compared forest recovery in El Petén, Guatemala, following
abandonment from agroforestry, shifting agriculture, and pastures. They found that
forest succession was much faster following use in agroforestry and shifting agriculture than following abandonment from pastures or from input-intensive monocultures. Their results suggest that initial ground cover by herbs inhibited recruitment
by woody colonists, a result consistent with many other studies. In Puerto Rico,
Aide et al. (1995) and Aide et al. (2000) studied changes in vegetation in chronosequences ranging from 0 to 60 years after pasture abandonment, and found that
species richness, abundance and biomass were very low during the first 10 years.
They concluded that recovery in pastures was greatly delayed in comparison with
forest recovery following other types of human and natural disturbances. The successional trajectory was quite different, with an absence of the more typical pioneer
270
F. Montagnini
species that are more commonly found in other areas. The presence of grasses and
the rapid colonization and growth of ferns and herbaceous species in the abandoned
pastures appeared to be a major factor inhibiting the establishment of secondary
forest, and imparted a selective barrier on the colonizing woody species.
In a study conducted near Paragominas in the state of Para, eastern Amazon of
Brazil, recovery to forest following pasture abandonment was examined in sites
where pasture management intensity ranged from low to high, and duration of use
was 4–10 years (Buschbacher et al. 1987, Uhl et al. 1988). In the low-intensity
management (comprising about 20% of abandoned pastures in the region) following
forest cutting and burning the area was seeded to pasture grasses, but grasses failed
to establish due to environmental and management reasons, only few cattle were
grazed, and abandonment usually occurred within 4 years of pasture formation. In
the medium-intensity management (about 70% of pastures in the region), seeded
pastures grasses established well, the pasture was weeded with machetes when
needed, and was burnt every few years to eliminate woody seedlings and favor grass
species. Burning increased soil pH and improved availability of nutrients. Grazing density was intermediate with about one animal per hectare, and abandonment
occurred 6–8 years after pasture establishment. In the high intensity management
(only about 10% of the cases), seeded pasture grasses also established well, and after
several cycles of weeding and burning, the area was cleared with heavy machinery
to remove woody species that had become established in spite of the weeding treatments. Thereafter, the sites were mowed mechanically each year and then burned.
Grazing pressure was heavy with more than two heads of cattle per hectare, and
abandonment happened after 10 or more years of use (Buschbacher et al. 1987).
Results of evaluations of plant biomass and diversity in the three categories of
abandoned pastures showed that the abandoned low-intensity pasture sites were
progressing toward forest, in contrast to the high-intensity sites. In the low-intensity
management sites, recovery of one-fourth the original biomass in 8 years suggested
that a rapid structural recovery was occurring, with most of the biomass being comprised of trees, as opposed to herbs, shrubs and grasses. In the sites of mediumintensity use, total biomass was less than in the low-intensity use site, and similarly
most of the biomass was in trees, again suggesting recovery to forest should be fast,
although not as rapid as in the low-intensity use site. In contrast, the high-intensity
management sites were almost completely dominated by shrubs, herbs and grasses,
with only 4% of total biomass in trees, therefore the potential for this site to recover
to forest appeared to be much less than in the two other cases.
13.2.5 Influence of Soil Fertility
Sites with inherently poor, acid soil may be less conducive to woody regeneration of
abandoned pastures than soils with more favorable conditions. For example, at our
study site at La Selva, Costa Rica, abandoned pastures were dominated by ferns and
grasses several years after abandonment, while in other abandoned pastures at La
Selva and nearby in locations with better microsite conditions, natural regeneration
13 Sustainability and Restoration of Degraded Pastures
271
was abundant and diverse (Chazdon et al. 2005). In other sites at La Selva of ages
12–25 years following pasture abandonment there were more than 1,000 trees larger
than 5 cm diameter at breast height per hectare, and tree species density ranged from
64–121 species per hectare (Chazdon et al. 2005).
Other studies of vegetation and soil changes following conversion of forest to
pasture show that recovery to forest may be relatively fast following abandonment,
but soil properties may take a longer time to recover (Reiners et al. 1994). These
authors examined soil and vegetation characteristics in primary forest, actively
grazed pastures (10–36 years old), abandoned pastures (4–10 years following abandonment), and 10–20 year old secondary forest in the Atlantic zone of Costa Rica.
They found decreases in soil acidity, some increases in bases, increases in soil
compaction, decreases in porosity, and lower rates of nitrogen mineralization and
nitrification in the actively grazed pastures. Their data suggest that with pasture
abandonment and succession to secondary forests these changes can be reversible,
although they may take a long time. They emphasized that some of the soil changes
were positive, such as increases in soil pH, while some were negative such as lesser
availability of soil nitrogen and lower porosity which will lead to higher erosion
(Reiners et al. 1994).
13.2.6 Position of Pastures in the Landscape
In addition, the size and position of pastures in the landscape matrix will influence the speed and characteristics of natural regeneration following abandonment.
Different types of matrices have a large influence on the ecology of recovering
forests. Proximity of pastures to remnant forest patches promotes more rapid forest
recovery, which depends heavily on seed dispersal (Chazdon 2003). In the Amazon,
forest fragments surrounded by areas of natural regeneration 5–10 m in height had
fewer micro-climate changes and lower tree mortality than did forest fragments
surrounded by pastures (Didham and Lawton 1999). The fragments surrounded by
this low-stature secondary forest also had greater bird abundance than did those
surrounded by pastures. Certain types of matrices favor the fauna of tropical forests
more than others. In the Amazon, natural regeneration areas dominated by Cecropia
spp., which tend to grow tall and with a closed crown, are utilized more by birds,
frogs, and ants than areas of more open Vismia spp.-dominated regeneration. In
general, any type of natural regeneration that borders forest fragments favors fauna
more than pastures do (Williamson et al. 1998). These patterns are found in lowland
as well as in montane forest regions of the Neotropics (Sarmiento 1997).
A condition of interrupted or arrested succession may develop when several
interacting factors leading to pasture degradation operate together, either simultaneously or in a sequence. For example, long and high intensity of use may lead to
nutrient-depleted and/or compacted soil, over which and with high solar radiation,
invasion by aggressive or weed vegetation will occur (Buschbacher et al. 1987,
Nepstad et al. 1991b, Reiners et al. 1994, Dias-Filho et al. 2001). This may be
compounded by large size and long distance from sources of seed that can impede
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F. Montagnini
or delay re-sprouting and seed arrival. In such situations, once the condition has
been identified along with its interacting causes, active measures are needed to
artificially encourage and favor forest regeneration. This may include seeding, soil
treatments, control of invasive vegetation, and planting appropriate tree species to
serve as perches for dispersal agents and to ameliorate soil and micro-environmental
conditions, as discussed in following sections.
13.3 Pasture Restoration
The purpose of pasture restoration may involve recovery of the former vegetation
(generally forest), which in turn has the compound objectives of increasing biodiversity, recovering soils, and attaining vegetation structure and functional characteristics as close as possible to the original forest. An alternative pathway, more properly
named rehabilitation, involves managing to recover productivity of degraded pastures, which in turn involves delaying degradation, increasing and sustaining yields
in the long term. The overall objective in this case is to recover the land for continued
use in pasture.
13.4 Techniques for Restoration of Degraded Pastures:
Soils, Biodiversity, Forest Structure and Function
The presence of trees, either in plantation, in groups, in lines or in isolated form may
contribute to the recovery of environmental conditions favorable to tree regeneration
processes (Guariguata et al. 1995, Parrotta 1995). The establishment of plantations
may overcome some of the barriers to natural forest regeneration by attracting seed
dispersal agents into the landscape and by ameliorating local microclimatic conditions within the area, and thereby accelerating the recovery of these lands (Parrotta
1995, Parrotta and Turnbull 1997, Powers et al. 1997, Lamb 1998). Strategies to
encourage natural regeneration of degraded pastures and fulfilling the objectives
named before (recovery of soils, biodiversity and forest structure and function) may
involve (1) reforestation, (2) care and management of remnant trees in pastures;
setting perches for birds; (3) planting individual trees/blocks/windbreaks.
13.4.1 Reforestation of Abandoned Pastures
13.4.1.1 Impacts on Soils and Nutrient Cycling
Some degraded pasture soils can be recovered through the use of fertilizers while
others that are badly eroded, or that are fully covered by invasive species such as
grasses or ferns, need more drastic rehabilitation techniques. There are also situa-
13 Sustainability and Restoration of Degraded Pastures
273
tions of extreme degradation where soils cannot be recovered at all. The recovery
of the soil’s productive capacity is frequently very expensive, thus the techniques
involved must produce financial returns to ensure their adoptability by the local
farmers (Montagnini 2002). Reforestation can be such alternative as the established
plantations not only can contribute to recover soils and the land’s productive capacity but they can also serve for cash income and long term savings to the owner.
Plantations can also serve other environmental services such as carbon sequestration
and recovery of biodiversity (Montagnini and Jordan 2005).
Abandoned pastures are often invaded by aggressive grass or ferns for the reasons discussed previously. The competitive advantage of grasses, combined with
degraded soils and lack of nutrients, often prevents the germination and initiation of
tree seedlings. These grassland areas are also often maintained by fires which inhibit
natural colonization by tree species (Montagnini and Jordan 2005). In these cases,
if the area is left for natural regeneration to take place, it may stay in an arrested
stage of succession for many years. Often the best alternative to restore these areas
and convert them back to forest is planting fast-growing trees, sometimes even using
exotic species.
In many cases it is not feasible to plant tree seedlings without first removing the
invasive vegetation. In such cases the use of herbicides, fire or mechanical weeding
may be necessary to clear the area and allow tree seedlings to establish successfully.
In some cases treatments may involve tillage, or root removal of the invasive plants
(Montagnini and Jordan 2005).
In research to develop alternatives for the rehabilitation and use of abandoned
pastures in the Neotropics, soil chemistry and nutrient cycling parameters were
measured in sites planted with indigenous tree species and in abandoned pastures.
At La Selva Biological Station in the Caribbean lowlands of Costa Rica, in just 3
years soil conditions improved in the young, native tree plantations compared to
abandoned pasture. In the top 15 cm of soil, total nitrogen and organic matter were
higher under the trees than in the degraded pastures, with values close to those found
in adjacent 20-year-old forests. The highest values for soil organic matter, total N,
Ca, and P were found under Vochysia ferruginea, a species common in mature and
secondary forests in the region (Montagnini 2001, 2002). Subsequent measurements
revealed similar trends in the soil parameters in the following years (Montagnini
2007). Based on the standards determined by the Ministry of Agriculture of Costa
Rica for soil fertility assessments, the cation levels (Ca, Mg and K) under most of
the trees were at or above the critical values needed for conventional agriculture in
the region. In contrast, the cation levels in the adjacent abandoned pasture soils were
too low for the growth of the subsistence crops preferred in the region (rice, beans).
Soil organic matter also had positive influences on soil physical properties: soil bulk
density was lower (i.e., lower compaction) while soil moisture was higher under the
trees than in abandoned pastures.
Other research at La Selva Biological Station included mixed- and pure-species
plantations of indigenous trees as strategies to recover abandoned pastures (Montagnini et al. 1995b). Early results of this research showed that 5 years after planting,
there were decreases in the content of P, K, and Ca in the soil in pure plots of
fast-growing tree species, such as Jacaranda copaia and Vochysia guatemalensis
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F. Montagnini
(Montagnini 2007). However, in other cases there were beneficial effects upon the
soil: for example, increases in Ca in the soil under Terminalia amazonia and Virola
koschnyi, both species with a high content of Ca in their foliage and elevated rates
of annual litterfall. In a similar fashion, soils under Vochysia ferruginea had greater
concentrations of Ca, Mg, and organic matter in comparison to the other species.
This result is consistent with other studies that include this species (Montagnini
2001, 2007).
Measurements over a long period are necessary to determine the effects of tree
species on soils. When the plantations were 11–12 years old, results indicated that
while many of these trends continued, some new ones were observed. For example,
the soil under Vochysia guatemalensis had higher pH, less acidity, and greater Mg
than other treatments. On the contrary, the soils under Calophyllum brasiliense had
lower values for the same parameters. These results appear to be related to the high
capacity of Vochysia guatemalensis to recycle cations, given its high quantity and
rapid decomposition of leaf litter, while the opposite happens with Calophyllum
brasiliense. The value of soil pH in Vochysia guatemalensis (5.03) was higher than
in previous measurements; the values of soil Ca, Mg, K, organic matter, and N were
similar and the values of P were less (Montagnini 2007).
The mixed plots in this research had average values for the nutrients examined,
and even improved conditions for some soil nutrients, such as P (Montagnini 2007).
In some cases there were lower values for nutrients in mixed plantations than in
pure plantations, as was the case for Ca and Mg. This suggests that in mixed plots,
soils have a more balanced nutrient status as a result of the complementary effect
on nutrient cycling of the different species participating in the mixtures. In contrast, in pure-species plots, there were some decreases or increases of soil nutrients,
depending on the species.
In similar studies conducted in Bahia, Brazil, values of at least five soil parameters under 15 out of the 20 species of the plantations were similar or higher to those
found under primary or secondary forest. Several species contributed to increased C
and N, including: Inga affinis, Parapiptadenia pterosperma, Plathymenia foliolosa
(leguminous, N-fixing species), Caesalpinia echinata, Copaifera luscens (leguminous, non-N-fixing), Eschweilera ovata, Pradosia lactescens (of other families).
Others increased soil pH and/or some cations, such as Copaifera luscens, Eschweilera ovata, Lecythis pisonis and Licania hypoleuca (Montagnini et al. 1994, 1995a).
Most of the species identified in this research for their positive influence on soil
properties are currently being used as components in productive land use systems
such as commercial plantations and agroforestry in each region (Montagnini 2001).
Exotic tree species can also play an important role in the restoration of degraded
pastures, especially when conditions of extreme degradation persist, such as
extremely eroded, denuded soil. If there is not enough time or resources to test
the performance of local species, a well-known exotic species that can thrive in
highly adverse conditions can be planted initially and later serve as nurse or facilitator for the introduction of native species. For example, in southern Costa Rica,
Carpenter et al. (2004) recommended planting Pinus tecunumanii, an exotic tree,
on the most deeply eroded, bare areas of pasture to ameliorate hostile micro-sites
quickly. Two or more years after planting the pines, seedlings of shade-loving
13 Sustainability and Restoration of Degraded Pastures
275
native species could be intermixed. In areas that have suffered less extreme erosion,
native species such as Vochysia guatemalensis, Terminalia amazonia, Calophyllum
brasiliense are preferred, according to results of long-term studies by the authors
in the region. The native trees were interplanted with cover crops or with edible
beans in controlled experiments, where these systems rapidly restored cover to
eroded pastures without fertilization (Nichols et al. 2001). These authors suggest a
system of rotations, planting pine as a rapidly growing, frequently harvested species,
and inter-planting more slow-growing but more valuable species (Carpenter et al.
2004). However, when using exotic species for restoration one has to be aware that
the qualities that make these species adaptable to harsh environmental conditions
can also allow them to become invasive. For example in Misiones, NE Argentina,
Leucaena leucocephala, a well-known N-fixing species broadly used in agroforestry
systems worldwide, has been introduced to improve degraded soils, and today this
species has expanded beyond the planted areas and in some case has become
a weed.
Reforestation can serve diverse economic, social, political and ecological functions in the Neotropics. With considerably higher yields than managed native
forests, tropical and subtropical plantations can make substantial contributions to
world timber and pulp production (Montagnini and Jordan 2005). Tree plantations
can also be a source of cash, savings and insurance for individual farmers. In combination with subsistence and commercial crops (agroforestry) or cattle (agrosilvopastoral systems), plantations have been used as tools in rural development projects
worldwide. Plantations are often seen as alternatives to deforestation as they can
provide products that otherwise would be taken from natural forests. If plantation
species and management strategies are chosen with knowledge of their nutrient-use
efficiencies and recycling capacities, they can be highly productive and also serve a
function in ecological restoration projects.
13.4.1.2 Restoration of Biodiversity
Several authors have reported on the role of tree plantations as accelerators or
“catalysts” of natural succession in tropical and subtropical sites (Parrotta 1995,
Jussi et al. 1995, Parrotta and Turnbull 1997, Keenan et al. 1999, Parrotta 1999,
Otsamo 2000). For example, in Southeast Asia, Jussi et al. (1995) reported on
the spontaneous and fast growth of indigenous tree species under plantations
of exotic trees. On the other hand, in north Queensland, Australia, a greater
diversity of species was found in the understory of plantations of native species
than in plantations of exotic species (Keenan et al.1999). In Puerto Rico, in the
understory of plantations of the exotic Albizia lebbek, 22 species of trees and
shrubs were found, in comparison with just one species in control plots without trees (Parrotta 1992). At La Selva Biological Station, Costa Rica, results
of some studies also suggest that tree plantations have a good potential for
accelerating the processes leading to recovery of biodiversity in degraded soils
(Guariguata et al. 1995, Powers et al. 1997, Carnevale and Montagnini 2002,
Cusack and Montagninin 2004, Montagnini et al. 2005). On the other hand,
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F. Montagnini
high establishment and maintenance costs are potential disadvantages of the
use of plantations for accelerating natural regeneration, given the intensive management that is needed especially during the first 2–3 years (Montagnini et al.
1995b).
Mixed plantations may offer a more favorable environment for natural regeneration than pure plantations due to their multi-strata architecture. Mixed plantations
may have a higher variety of microhabitats for seed dispersers and potentially create a greater variety of ecological niches allowing for the establishment of diverse
regeneration. At La Selva Biological Station, mixed plantations with native tree
species had relatively high abundance and high numbers of regenerating species in
their understory in comparison with pure plantations (Carnevale and Montagnini
2002, Cusack and Montagnini 2004) (Table 13.1). Higher plant species richness
accumulated under Vochysia guatemalensis, Virola koschnyi, Terminalia amazonia,
Hyeronima alchorneoides and Vochysia ferruginea, all species commonly planted
by farmers in the region. Natural regeneration was higher in understories with low
or intermediate light availability.
Results of seed rain and seed dispersal studies showed that most of the seeds
entering the open pastures were wind-dispersed, while most seeds entering the
plantations were bird or bat-dispersed (Orozco Zamora and Montagnini 2007a)
(Table 13.2). In terms of seed rain abundance, plantation plots with either pure or
mixed species had greater abundance than the abandoned pasture plots left for natural regeneration. The pure species plantations that had greater seed rain abundance
Table 13.1 Abundance of woody seedlings and saplings in three plantations with native tree
species in pure plots, mixtures of thee species, and open pastures left for natural regeneration
at La Selva Biological Station. Plantations were 7 years old at time of sampling
Treatment
No. Individuals/32 m2
No. Individuals/ha
90
78
57
88
29
28,125
24,375
17,812
27,500
8,937
38
18
88
34
14
11,875
5,625
27,500
10,625
4,375
Plantation 1
Vochysia guatemalensis
Calophyllum brasiliensis
Jacaranda copaia
Mixture of 3 species
Natural Regeneration
Plantation 2
Virola koschnyii
Dypterix panamensis
Terminalia amazonia
Mixture of 3 species
Natural Regeneration
Plantation 3
Vochysia ferruginea
18
Hyeronima alchorneoides
25
Balizia elegans
13
Mixture of 3 species
32
Natural Regeneration
2
Sources: Carnevale and Montagnini (2002); Cusack and Montagnini (2004).
5,703
7,891
4,219
10,156
703
13 Sustainability and Restoration of Degraded Pastures
277
Table 13.2 Distribution of dispersal vectors by treatment in pure plantations, mixed plantations,
and open pastures left for natural regeneration at La Selva Biological Station. Plantations were
12–13 years old at the time of the sampling
Categories of dispersal vectors
Treatment
Birds
(%)
Mammals
(%)
Wind
(%)
Mammals Mammals,
and birds birds, and wind
(%)
(%)
J. copaia
0
2
8
70
C. brasiliense
12
–
–
64
V. guatemalensis
94
–
1
–
Mixed 1
13
–
–
80
Natural Regen. 1
79
1
–
–
T. amazonia
–
–
–
99
V. koschnyi
–
–
–
98
D. panamensis
2
–
–
98
Mixed 2
5
–
3
88
Natural Regen. 2
–
–
97.7
0.3
B. elegans
–
–
–
100
H. alchorneoides
44
–
–
50
V. ferruginea
98
–
–
1
Mixed 3
18
0
0
79
Natural Regen. 3
–
–
100
–
(Source: Orozco Zamora and Montagnini 2007a,b)
14
10
3
–
10
–
1
–
2
2
–
2
–
1
–
Others
(%)
Total
(%)
6
14
2
7
10
1
1
–
2
–
–
4
1
2
–
100
100
100
100
100
100
100
100
100
100
100
100
100
100
100
were Balizia elegans, Dipteryx panamensis, and Jacaranda copaia. Apparently,
plantation trees in either mixed or pure plots are more effective in attracting birds
and other dispersal agents than the grass and other low-stature vegetation that predominates in the abandoned pastures left for natural regeneration. In terms of seed
species richness, the abandoned pastures ranked lower than both the pure and mixed
plantation plots, again suggesting that plantations provide a more attractive habitat
for seed dispersal agents than natural regeneration plots (Orozco Zamora and Montagnini 2007a, 2007b). The lower species richness of seeds found in the abandoned
pasture treatments, especially in those populated by grasses and ferns, suggests that
natural regeneration can be delayed by the absence of seeds and seed dispersal
agents. This suggests that plantations facilitate tree regeneration by attracting seeddispersing birds and bats into the area.
The different species of the plantations created different conditions of shade and
litter accumulation, which in turn affected forest regeneration (Carnevale and Montagnini 2002). Competition from grasses is a major factor influencing woody invasion under these plantations. High accumulation of litter on the plantation floor may
help diminish grass growth and thus encourage the establishment of other woody
species under tree canopies; however, canopy shading apparently had a greater
influence on tree regeneration than litter depth (Carnevale and Montagnini 2002).
In addition, results of further studies in this site suggest that there was no significant
correlation between leaf litter in the plantation floor and abundance of regeneration (Cusack and Montagnini 2004). Powers et al. (1997) and Parrotta (1995) found
litter depth to be negatively correlated with abundance and richness of understory
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F. Montagnini
PER CEN T OF TOTAL NU MB ER
OF SEED S
regeneration. Parrotta (1995) suggested that high litter biomass suppresses regeneration of small-seeded species.
Although the results of several studies suggest that plantations can accelerate
or “catalyze” secondary forest succession, it is still unknown as to what extent
they can promote the establishment of species of more advanced stages of succession. In our research at La Selva, the majority of seeds collected under the mixed
and pure plantations were from heliophillous species, which are associated with
early secondary forests. (Orozco Zamora and Montagnini 2007a, 2007b). A total
of 35% of the species of the seeds collected in this research belonged to secondary
forests (Renealmia alpinia, Psychotria bracheata, Palicourea guianensis, among
others) (Fig. 13.3). A total of 20% of the collected species of seeds belonged to
a highly varied functional stage or location in the landscape as many species can
be found in abandoned fields, advanced secondary forests, riparian zones, open
areas, minimally altered forests, and other functional stages (e.g. Ficus spp., and
species of the Araceae and Cucurbitaceae families). Approximately 15% of the
species of the collected species were associated only with abandoned fields and
pastures (e.g. grass species, Poaceae family). The rest of the species (5–10%) are
found in one to three functional stages, all of which had secondary forests in
common (Fig. 13.3).
Recent studies suggest that there may be a limit on the use of plantations to
foster biodiversity at a particular site, although such plantations often catalyze
successional development in the plantation understory (Lamb et al. 2005). Subsequent studies can verify whether the forest species composition favored by plantations approximates more mature stages of natural forest succession in the region.
50
45
40
35
30
25
20
15
10
5
0
PAF SF
SF, ASF SF,
DF
PAF
V
SF, SF, SF, MDF
DF, ASF OA
PAF
FUNCTIONAL STAGE
Fig. 13.3 Functional stage of ecological succession of seeds identified in seed rain studies in native
tree plantations at La Selva Biological Station in Costa Rica. PAF (pastures and abandoned fields),
SF (secondary forest), DF (disturbed forests), ASF (advanced secondary forest), V (various), OA
(open areas) and MDF (minimally disturbed forests) (Source: Orozco Zamora and Montagnini
2007b)
13 Sustainability and Restoration of Degraded Pastures
279
In addition to considering the ecological factors influencing tree regeneration, the
socioeconomic feasibility of these plantations as practical systems for the recovery
of biodiversity in deforested landscapes in each region should be assessed.
13.4.2 The Value of Isolated Trees in Pastures for Restoration
13.4.2.1 Role in Favoring Tree Regeneration
In many tropical pastures there are isolated trees that can be remnants of the original
forest. They are conserved by farmers for their functions as shade for cows and
humans, and as sources of firewood, timber and fruit. Often the trees are left uncut
because they are too large or the farmer may not have the tools or labor available to
remove them. For example, isolated trees of Dipteryx panamensis, which has relatively dense wood, are often present in many pastures in Central America, because
they are harder to cut than other forest trees.
Often the microenvironment below tree crowns is favorable for the establishment
of species that cannot grow in pastures. For example, in the lowlands of the Pacific
coast of Chiapas (Mexico), researchers found a total of 65 individual trees in surveys
done in a total of 15 pastures of 1 ha in size each, grazed with a cattle density of two
animals per ha (Otero Arnaiz et al. 1999). The trees belonged to 20 different species
in 11 families. The timber species were typically from the primary forests common
in the region. In addition, under the crown of the trees they found 134 species and
45 families of plants, predominantly herbs (62 species) followed by shrubs (32),
trees (31) and lianas (3). The authors recommended that future efforts to restore the
forest from isolated trees already present in the pastures should put emphasis on
maintaining a great diversity of tree species in the pasture fields.
In a similar study conducted on 24 dairy farms in Monteverde, Costa Rica, Harvey and Haber (1999) examined a total of 237 ha of pastures. They found a total
of 5,583 trees (25 trees/ ha on average). The total number of tree species was 190,
of which 57% belonged to primary forest of the region. They also examined the
farmers’ uses of the trees: 37% were used for timber, 36% for firewood, and 20%
for posts. About 90% of the trees also provided food for birds. The authors concluded that although the tree densities were apparently high in the studied pasture
fields, the species diversity would probably decrease with time, since the farmers
were harvesting trees for the uses already mentioned. There should be incentives
for planting native trees in the pastures to replace those that were harvested or died
during previous years (Harvey and Haber 1999).
13.4.2.2 Influence on Seed Dispersal
Several species of native trees growing isolated in pastures can be effective as
perches to attract birds and aid in the establishment of other species of native forest.
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F. Montagnini
In a study conducted in “Las Alturas” biological station in southern Costa Rica, Holl
(1998) compared the effectiveness of Inga trees growing naturally in pastures, with
artificial posts set as perches. The artificial perches were cross-posts 6 m in height
with two attached 2 m length crossbars installed in abandoned pastures at different
distances from the border of the forest. In addition, dead branches of Inga were
used in the same manner, separated by 10 m between them. The quantity of seeds
dispersed by birds was greater beneath the branches than beneath the crosses or in
the pasture without perches. Nevertheless the number of plants of these species was
similar below perches and in the pasture. Despite the fact that the perches increased
the dispersion of seeds, they could not compensate for other barriers to the recuperation of the tropical forest, such as the depredation of seeds, low germination rate
of seeds, and low survival of plants.
The type of seeds dispersed by the birds should also be considered. Were they
from species that could promote successional processes or that could be used by
local farmers? Plantations with tree species are the most expensive, but they can be
the most effective strategy for recuperating the forest (Holl 1998).
Crown size and shape, tree height and other characteristics may influence the
ability of trees to attract birds and facilitate tree regeneration in abandoned pastures. In a seed rain study at finca La Suerte in the Caribbean region region of
Costa Rica, Slocum and Horvitz (2000) found 25 species of seeds from trees and
shrubs under trees that were producing fruit, with larger numbers than those found
in open pastures. These authors found larger abundance and species richness of
seeds under Ficus trees than under Cordia trees. They found a positive correlation
between height of trees and seed rain abundance and diversity, with larger seed input
under the taller trees. Apparently birds prefer taller trees as perches to have a better
view of the surroundings and to avert potential predators. In our study at La Selva,
trees of Jacaranda copia and Vochysia guatemalensis, which were the tallest, had
approximately twice as much seed rain abundance and seed species richness as the
shortest trees, Callophylum brasiliense and Dipteryx panamensis (Orozco Zamora
and Montagnini 2007a).
However, other characteristics of trees apart from height may be important
in their role as recruitment foci. For example, Slocum (2001) found that the
amount of shade cast was more important than the type of fruit produced (dry
or fleshy), with lower recruitment under the shadier canopies. In the same study
site, Slocum and Horvitz (2001) found that canopy architecture did not appear
to be a significant factor affecting seed arrival with no difference in seed input
beneath trees of dense canopy (Ficus and Pentaclethra) and those of sparse
canopy (Cordia and Cecropia). On the other hand, results of a recent study
in the Pacific region of Costa Rica suggested that crown diameter was more
important than height of trees in attracting birds (Francesconi 2006). Thus, a
combination of several characteristics may determine the role of each tree species
as recruitment foci. For example, at La Selva, Guariguata et al. (1995) suggested
that although Vochysia guatemalensis trees do not fruit early, this species looks
promising as an effective restoration tool because it appears to provide perching and roosting sites for frugivorous birds and bats while rapidly suppressing
grass cover.
13 Sustainability and Restoration of Degraded Pastures
281
13.4.2.3 Influence on Soil Nutrients
In research on the role of isolated trees on soil nutrients in pastures in lower montane ecosystems in Ecuador, the density of trees was about 30 individuals/ha, of
which about 75% were species of the nitrogen-fixing Inga spp., and the rest were
guajava trees, Psidium guajaba (Rhoades et al. 1998). Below the trees there were
moderate air temperatures and intensity of light similar to that of the forest. The
concentration of nitrates in soils beneath Inga spp. trees was four times higher than
in the open pastures. There was also increased organic matter in the soil under the
Inga trees. In other work done in the same area it had been shown that the activity
of birds was 20 times higher in pastures containing trees than in pastures without
trees. As birds can be a source of seed, their activity can favor the establishment
and growth of timber species in the pastures (Rhoades et al. 1998). The number and
diversity of species of trees could be increased in order to accelerate regeneration
of forest in abandoned pastures. These results and recommendations are relevant
since the species found in these pastures are common to many tropical regions of
the Neotropics.
13.4.2.4 Use of Nurse-Trees and Artificial Protection
Addition of artificial structures can sometimes increase survival of regenerating
seedlings when used for protection against excess insolation, drought or frost. For
example, in abandoned pastures at high elevations in Hawaii, the aim of restoration
projects is to return the original mixed forest in order to serve as habitat for threatened birds. In a study conducted in the Wildlife Refuge of Hakalon, on the island
Mauna Kea at 1,840 m elevation, with 2,000 mm of rain per year, the meadows that
remain after more than 100 years of human intervention (forest clearcut, pastures)
and conversion to grassland are generally dominated by exotic grasses (Scowcroft
et al. 2000). At high elevations, planted seedlings of native species can suffer frost
damage and desiccation when planted in the open. In their study site, Scowcroft et al.
(2000) transplanted seedlings of Metrosideros polymorpha (Myrtaceae), a native
tree of Hawaii from nearby woodlands. They planted the seedlings in the open and
within a corridor of 8-year-old, 4 m tall trees of the native species Acacia koa that
were sown earlier in order to serve as protection against frost for the Metrosiderus
trees. The use of Acacia koa trees or artificial mechanisms, such as protection with
cloth on the sides, increased the successful establishment of Metrosiderus polymorpha seedlings in these degraded, abandoned rangelands. The authors recommend
Acacia koa as a nurse tree because it is endemic to Hawaii, fixes nitrogen and is
capable of rapid growth, which allow it to quickly dominate a site when planted in
groves or corridors and enhance soil N availability. Within 2–3 years the trees should
be tall enough to be out of danger from frost, and if planted at 4–5 m intervals,
crowns should close after 10 years resulting in reduced herbaceous competition in
the understory. The improved temperature, nutrient content, and herbaceous competition regimes under canopy of Acacia koa may favor establishment of other, more
sensitive native species.
282
F. Montagnini
13.4.2.5 Role of Windbreaks to Accelerate Succession in Pastures
In regions with large agricultural fields that are far from sources of propagules,
windbreaks in pastures and agricultural fields may be important reservoirs of
native tree species (Holl 1998, Harvey and Haber 1999, Harvey 2000). For
example, the effects of planted windbreaks on seed deposition patterns were
examined in dairy farms in Monteverde, Costa Rica by Harvey (2000). The
windbreaks were planted strips of trees about 5 m wide and 9 m tall and were
7–8 years old at the time of the study. Trees and rows within the windbreaks
were spaced at 1.5 m. The most common species were Cupressus lusitanica, Croton niveus, Casuarina equisetifolia (all exotic species) and Montanoa
guatemalensis (native). Windbreaks were found to receive significantly greater
densities and species richness of seeds from tree and shrub species than open
pastures. Windbreaks received an average of 39 times as many tree seeds, and
67 times as many shrub seeds as pastures. The differences in the seeds entering
windbreaks versus pastures appeared to be due almost entirely to the enhanced
activity of birds in windbreaks. Bird-dispersed seeds occurred in greater densities
(about 100 times greater) and number of bird-dispersed species was three times
greater in windbreaks than in pastures. The high densities of bird-dispersed seeds
within windbreaks suggest that windbreaks increase forest seed recruitment by
serving as habitat and/or movement corridors for seed-dispersing birds (Harvey 2000).
Windbreaks may serve as sources of woody colonists if the agricultural lands are
later abandoned. Positioning of windbreaks within the landscape may affect seed
deposition patterns by influencing the movements of seed-dispersing birds. Tree
recruitment may be higher in windbreaks that are connected to forests. Windbreaks
could be made more attractive to birds by including native, fruit-producing trees, by
increasing their species and structural complexity, and by positioning them between
forest patches to facilitate bird movement (Harvey 2000).
13.5 Alternatives to Delay Pasture Degradation and Increase
Pasture Productivity
Pasture productivity and longevity in the Neotropics seem to be closely related to
soil fertility and efficient nutrient recycling. CIAT (International Center for Tropical Agriculture) in Cali, Colombia; EMBRAPA-CPATU (Agency for Agricultural
Research of Brazil, Center for Studies of the Humid Tropics); CATIE (Tropical
Agriculture Research and Higher Education Center) in Turrialba, Costa Rica; INTA
(National Institute for Agricultural Technology) in Misiones, Argentina and other
international and local agencies in countries of the Neotropics have developed recommendations based on long-term research carried at their stations and in local
farms. These strategies are intended to increase productivity and avoid or delay
degradation, and include the following:
13 Sustainability and Restoration of Degraded Pastures
283
1) Use of adequate stocking (animal type/breeds, weight, numbers per hectare)
2) Planting good pastures (palatable, nutritious), often including exotic pasture
species
3) Use of cover crops to protect against erosion especially on slopes
4) Use of mixed pasture species, intercropping with legumes
5) Rotational pastures (including use of electric fencing)
6) Fertilization as needed (especially N, P, K)
7) Use of agrosilvopastoral systems
13.5.1 Adequate Management Techniques to Ensure
Pasture Sustainability
13.5.1.1 Method of Land Clearing
The method of land clearing can have important and long-lasting consequences
for pasture productivity and longevity. Land clearing for pasture establishment on
forested areas of the Amazon basin generally involves cutting the forest with chain
saws, generally after extraction of the commercially valuable woody species (Serrao
and Toledo 1990, Dias-Filho et al. 2001). Mechanical clearing with bulldozers is
seldom practiced because of the high costs and the detrimental effects on soil properties, as was shown by studies done in the 1970s when clearing forest for pasture
in the Amazon first began.
For example, in a study in the Amazon forest of Peru, biomass production of
the pasture grass Panicum maximum was 68% higher in a pasture on land that had
been cleared manually (slash and burn method) in comparison with a pasture that
had been established on an area cleared with bulldozer (Seubert et al. 1977 cited
by Dias-Filho et al. 2001). Further studies have confirmed this assertion. The fire
employed in the manual clearing not only contributes to remove weeds but also
improves nutrient content of the pasture soils with inputs of nutrients in the ash.
However it should be noted that the combustion of vegetation also results in large
losses of nutrients such as nitrogen, sulfur and carbon (Dias-Filho et al. 2001),
with negative consequences to long-term pasture productivity, and air pollution as
well.
In addition, the initial input of nutrients to the soil following burning is generally
short-lived. The early claims from the 1970s that conversion to cattle pastures results
in improved soil fertility of once forested soils following clearing and burning were
refuted by Fearnside (1980). Fearnside reviewed the available evidence and concluded that long-term sustainability of the pastures was not possible under those
conditions, evidence that was supported by this same author’s subsequent articles
as well as by many others. Improvement of soil pasture nutrient status must be
obtained through the use of N-fixing species and/or the addition of fertilizer (Serrao
and Toledo 1990, Dias-Filho et al. 2001).
284
F. Montagnini
13.5.1.2 Use of Adequate Pasture Species Mixtures
Alternatives to delay pasture degradation and increase productivity developed
by EMBRAPA-CPATU in Brazil, by CIAT in Colombia and by CATIE in Central America recommend intercropping leguminous herbs such as Centrosema
pubescens, C. macocarpus, C. acutifolium, Stylosantes guianensis, Arachis pintoi
and Desmodium ovalifolium with productive grasses such as Brachiaria brizantha,
B. dictyoneura, Andropugon gayanus, Panicum maximum (Serrao and Toledo 1990,
Vera et al. 1998, Bouman et al. 1999, CIAT 1997–2000, Amézquita et al. 2005).
In unfertile soils, N-fixing species are better competitors than grasses and they
tend to dominate the pasture. Thus, in these mixtures the appropriate proportion of
legumes/grasses has to be sown, and the timing of planting the legumes/grasses has
to be set to fulfill the objectives of producing a nutritious pasture and maintaining
or improving soil nutrients. To stimulate the early growth of the leguminous herbs,
they are inoculated with Rhyzobium bacteria to ensure the formation of root nodules
and increase rates of N fixation.
The species of pasture and leguminous herbs will vary according to the specific
region and the varieties and cultivars that have been developed by the corresponding institutions in each region. For example, in the Andean hillsides of
Colombia improved pastures of Brachiaria decumbens are used, while in the
Atlantic humid region of Costa Rica mixtures of Brachiaria brizantha + Arachis
pintoi are more commonly used. In drier regions of the Neotropics, such as in
the Pacific region of Costa Rica Brachiaria decumbens is more common, among
other alternatives (Bouman et al. 1999, Amézquita et al. 2005). Brachiaria decumbens is also commonly used in the savanna regions of eastern Colombia (Vera
et al. 1998).
13.5.1.3 Management of Soil Nutrients
Fertilization with P, Ca, Mg and K is often recommended for sown pastures, however depending on the region and cost of fertilizer, these recommendations are seldom followed, except for the application of P (Vera et al. 1998). Inputs of N to the
pasture are expected to be sufficient when using mixtures with leguminous herbs.
However the leguminous herbs also may need a “starter” fertilizer application with
N, P and K to stimulate their growth and to improve the efficiency of N fixation.
Soil phosphorus availability has been shown to be of particular importance in
maintaining pasture productivity, as this element is most limiting (Garcia-Montiel
et al. 2000, Dias-Filho et al. 2001). P can be in different forms in the soil and
some may be more or less available to plants. In a study describing changes in
total P and different soil P fractions that occurred after the conversion of forest to
pasture in the Brazilian Amazon state of Rondonia, Garcia-Montiel et al. (2000)
used chronosequences of forest and pasture of different ages to document patterns
of labile, occluded, and organic P pools using a sequential P-fractionation technique. These authors found that P released from the forest biomass left on the
ground substantially increased soil available P during the first 3 to 5 years following
13 Sustainability and Restoration of Degraded Pastures
285
clearing and burning. Since large portions of P are released from the burned and
decomposing biomass, management could be directed to improve the distribution of
decomposing biomass on the site, and to adding other sources of organic P once the
residual biomass is gone.
Of the approximately 200 kg of P per hectare measured in soils of one pasture
in the eastern Amazon, less than 1% was thought to be available to plants (DiasFilho et al. 2001). Since fertilization is expensive, management techniques should
be geared towards effective P recycling and conservation. Much of the P taken up
by plants is returned to soils via animal manure, however cattle tend to deposit dung
in concentrated areas rather than uniformly across the pasture. Pasture creation also
impacts carbon and nitrogen cycles by reducing standing stocks and cycling of these
elements (Dias-Filho et al. 2001). C and N stocks may increase in pasture soils, but
such increases are unpredictable and dependent on native soil fertility, fertilization,
climate, fire frequency, and grazing intensity. Even when standing stocks increase,
research from the eastern and western Amazon shows that cycling rates of these
elements consistently decrease (Dias-Filho et al. 2001).
P supply to plants in pastures is strongly influenced by type of management. In
a study done in soils of pastures in Colombia, after little P fertilization and 15 years
of continuous growth of a grass alone (Brachiaria decumbens) and a grass/legume
(B. decumbers + Pueraria phaseoloides) pasture, it was found that reserves of
labile organic P fractions associated with humic and fulvic acids was higher in the
grass/legume pasture than in the grass alone (Guggenberger et al. 2000). Earthworm
casts were highly abundant in the grass/legume pasture, and the authors concluded
that earthworm activity improves the supply of P in the soils under tropical pastures
by creating an easily available organic P pool (Guggenberger et al. 2000).
13.5.1.4 Pasture Rotations
Leaving grazed fields or paddocks to fallow is key to slowing pasture degradation.
The length of the fallow period differs according to the region and the practices
used. For example, in surveys conducted in pastures located in the eastern savannas of Colombia, a 3–4 week fallow period followed P fertilization, while a fallow
period following mechanical reclamation was 6–9 weeks depending on the specific site conditions (Vera et al. 1998). The fallow period has to be balanced by the
opportunity cost represented by the gain in weight or other productivity parameter
of the cattle that has to be taken to another grazing field. Use of electric fencing
often improves the feasibility of moving the cattle to nearby fields, especially when
relatively small grazing paddocks are used.
13.5.1.5 Animal Stocking
The choice of the right type of animal and its stocking is necessary to attain the
best productivity and sustainability of any cattle system. Much of the earlier discussion has referred to beef or dairy cattle of either European or Asian origin (Bos
286
F. Montagnini
taurus and Bos indicus respectively). CATIE and many of the other institutions mentioned previously have promoted the use of mixed breeds that can combine greater
rusticity and adaptability to tropical conditions with higher productivity (beef and
dairy products).
In semiarid and sub-humid regions of the Neotropics, other animals are used
successfully for meat and dairy production, and wool in the case of sheep. In many
regions, goats and sheep are also used with success. In many cases these animals
are part of small-scale, self-subsistence systems. However, they are also used at
larger scales for commercial purposes. For example the pelibuey, a type of sheep
commonly raised in the Caribbean and Mexico, successfully replaces sheep and
goats of other breeds due to higher reproduction rates, better production of wool,
and greater adaptability and rusticity (Fig. 13.4).
13.5.2 Agrosilvopastoral Approaches to Sustain Pasture
Productivity and Avoid Degradation
Agrosilvopastoral systems—the combination of timber, fuelwood or fruit trees with
animals, and with or without crops—are practiced at many scales. A large scale
system may include timber plantations with grazing to control weeds and to obtain
a more immediate return from the sale of animal products. Cattle-raising can also
Fig. 13.4 In Atotonilco, in the state of Hidalgo, Mexico, mixed breeds of pelibuey are grown using
electric fencing for pasture rotations, on fields sown with African star grass (Cynodon nlemfluensis
and C. plectostachyus). The owner and other farmers in the region apparently consider this a successful type of operation due to high productivity of the pelibuey and soil conservation. Photo: F.
Montagnini
13 Sustainability and Restoration of Degraded Pastures
287
complement subsistence agriculture, with animals often integrated in home gardens or in systems of fodder production to feed animals in stables (Montagnini
1992). In some regions, the incorporation of trees—especially multiple-purpose
trees (MPTS)—can change cattle raising from an inefficient use of land to a more
ecologically and economically feasible activity. The incorporation of trees can
improve system productivity either by increasing pasture yields through more efficient nutrient recycling, or through the production of fodder from leaves and fruits.
The introduction of cattle in plantations, generally when the trees are big enough
so that cattle cannot damage them, can be a more efficient use of land than plantations alone. Cattle can also help to control weeds and early economic returns
from cattle products help to pay for the costs of reforestation (Montagnini et al.
2003) (Fig. 13.5). In some cases young calves are preferred over adult cattle because
they cannot damage the trees as easily. The smaller animals also decrease soil compaction problems. Cattle prefer some tree species over others for browsing and can
also damage young trees when they scratch against them or strip their bark.
Fig. 13.5 Agrosilvopastoral system established to offset costs of reforestation: beef cattle (bulls)
grazing eucalypt plantations at CATIE in Turrialba, Costa Rica. Photo: F. Montagnini
288
F. Montagnini
Generally, in agrosilvopastoral systems there are short and long-term benefits
from cattle and tree products. Economically, the system is favored through product
diversification. In semiarid ecosystems, fodder trees have the additional advantage
over grasses, of being able to reach deeper layers of soil thus remaining green for
longer periods and providing fodder at critical times.
Depending on the relative proportion of trees to cattle and the emphasis given
to the principal productive objective of the systems, three main types of agrosilvopastoral systems can be found: 1-Trees in pastures, 2-Cattle grazing plantations
or forest, and 3-Integrated agrosilvopastoral systems.
13.5.2.1 Trees in Pastures
In this type of agrosilvopastoral system, the principal objective is cattle production
(e.g., beef, milk) with timber, fruit or fuelwood as secondary products. Often the
trees in the pasture are remnants from the previous forest, or have regenerated in
the pasture and the farmers keep them to provide shade for the cattle, improve the
pasture grasses, and obtain tree products. For example guacimo (Guazuma ulmifolia) trees grow naturally in pastures in Hojancha, in the Pacific region of Costa
Rica. In pasture recovery projects in Hojancha, dense plantations of guacimo are
established to replace the old trees (personal observation). The examples cited in
previous sections (4.2.3) on isolated trees of Psidium guajaba and Inga spp. in
pastures in Ecuador, and on other species in Chiapas, Mexico, and in Monteverde,
Costa Rica (4.2.1), are also agrosilvopastoral systems of this category. In all these
cases the trees serve multiple functions in terms of soil improvement, shelter for
cattle, and also contribute to biodiversity as they are used as perches by birds.
Nitrogen-fixing trees are often planted in degraded pastures for the main purpose of soil restoration, and also for their fodder. In degraded agricultural fields
near Guapiles, Costa Rica, researchers from CATIE use a restoration system they
call “Poró acostado” (“laid poró”). They plant stakes of an N-fixing tree, (“poró”),
Erythrina poeppigiana, laid horizontally in a trench to obtain rapid growth and
nitrogen inputs to the soil (personal observations). This tree can sprout quickly
from a horizontal stake giving multiple stems that will form clumps of young
trees in the pasture, where they can shelter the cattle, improve growth of native
or planted grasses and also their fodder can be harvested and fed to dairy or beef
cattle.
13.5.2.2 Cattle Grazing in Managed Forests and Tree Plantations
In these types of systems, the principal objective is to obtain tree products; cattle
serve to control weeds and to obtain an earlier financial return from the system.
Cattle are often used to complement reforestation projects in the Caribbean lowlands
of Costa Rica, such as calves used to graze plantations of native timber species
(Hyeronima alchorneoides) in Horquetas, Costa Rica; and beef cattle (bulls) grazing
eucalypt plantations at CATIE in Turrialba, Costa Rica (Fig. 13.5).
13 Sustainability and Restoration of Degraded Pastures
289
Examples under forest include beef cattle grazing under natural forest that has
been exploited and later thinned to allow for pasture growth in Santiago de Liniers,
in the subtropical forest of Misiones, Argentina. Axonopus spp. (“pasto Jesuita”) is
a native grass that is shade-tolerant and thus can grow well and stand the grazing
when planted under the forest canopy. Cattle density is high with about 200 calves
(250–400 kg each) per hectare. At these high densities calves are kept grazing in
the forest for just half a day or a day at most, and electric fencing is used to allow
for pasture rotations (personal observations). In another location near the town of
Puerto Rico in the same region of subtropical forest, fully grown (about 20 years
old) Araucaria angustifolia plantations are also underplanted with Axonopus spp.
and grazed by beef calves (120–150 kg each). These calves are also fed a corn
supplement (personal observations). In this region, the araucaria trees, which are
native, take about 30–35 years to reach full size for timber harvest, and income
from the cattle contributes to make the plantations more profitable. In these forest
or plantation grazing systems, the forest or plantations are the principal management objective of the farmer, and the cattle are only introduced after the plantations are already established, or the forest has already been managed for timber
production.
13.5.2.3 Integrated Agrosilvopastoral Systems
These systems are designed with the purpose of having an integrated system from
the start, where cattle production is enhanced by planting pastures in association
with trees that can benefit the pastures and cattle and at the same time provide
important tree products. Tree planting distances, pasture species and cattle management are all carefully selected with these objectives in mind. Examples from
the subtropical region of Misiones, Argentina include plantations of Toona australis
(Australian cedar) planted at 9 m×3 m, a planting distance that gives enough room
between the lines of trees to allow plantations of a shade-intolerant grass for the
cattle. In this case the combination was with Brachiaria spp. grass to feed beef cattle
(personal observation). In this same region, close to the locality of Puerto Rico in
Misiones, Paulownia spp. trees (“kiri”), 9 years old, planted at 9 m ×6 m were
seen in combination with Brachiaria brizantha grass to feed beef cattle. Both the
Australian cedar and kiri trees (exotic) are highly profitable, and although neither of
the two species tends to increase yields of grass, they do provide shelter for the cattle
especially during the hot summer months. The trees, if grown in a pure plantation,
could be planted at higher densities and thus the owner would obtain higher profits
from selling a higher volume of timber at the end of the rotation. However, the
farmers prefer to sacrifice some of the timber profits that they would obtain only
at the end of a rotation in order to include the grasses and cattle and obtain earlier
returns on the investment.
In other examples of this same type of system, trees are planted to improve the
growth of grasses and not necessarily for timber or other products. For example, near
Guapiles in the Caribbean lowlands of Costa Rica, projects by CATIE researchers
for restoration of degraded pastures include plantations of Acacia mangium, an
290
F. Montagnini
N-fixing tree planted for soil erosion control. A. mangium is associated with
native or planted pasture species such as Ischaemum ciliare, Brachiaria brizantha, and the herbaceous legume Arachis pintoi (Amézquita et al. 2005). Arachis
pintoi, a type of peanut, is often planted to control soil erosion because it
forms a dense carpet covering the soil, helps to recover soil nutrients, and also
improves pasture and cattle productivity (Jansen et al. 1997, Bouman et al. 1999)
(Fig. 13.6).
In the same region, Gliricidia sepium, an N-fixing tree that is an important firewood species and a good soil improver, has been found to have excellent qualities
for fodder production and thus is often planted as a living fence. This is another
type of integrated agrosilvopastoral system where the trees are purposefully chosen
to serve their function as a fence and at the same time to improve soils, control
erosion and serve as fodder (often in a cut-and-carry system).
Still another type of integrated agrosilvopastoral system is the “forage bank” or
“fodder bank” generally composed of tree species with highly nutritious foliage for
cut-and-carry systems. The trees are planted very densely to encourage foliage production. They are pruned periodically (often weekly), ground, occasionally mixed
with cut grasses and then fed to cattle. Examples of tree species that can sustain
high fodder productivity in this condition include Trichanthera gigantea, Morus
spp., Erythrina edulis, E. berteroana, E. fusca, Boehmeria nivea, which are used
with success in Colombia and in Costa Rica (CIAT 1976–2000, Amézquita et al.
2005).
Fig. 13.6 Arachis pintoi, a type of peanut, is often planted in the humid regions of Costa Rica to
control soil erosion because it forms a dense carpet covering the soil and contributes to soil nitrogen
and organic matter inputs. In other regions of Costa Rica and Colombia it is also intercropped with
trees or grasses to improve pasture and cattle productivity. Photo: F. Montagnini
13 Sustainability and Restoration of Degraded Pastures
291
Agrosilvopastoral systems can also be conceived at a landscape scale, where the
trees are kept in a managed forest and cattle raising makes use of trees in a complementary manner. For example, models of low-external input agrosilvopastoral
systems that incorporate farmer preferences, and use well-adapted grass-legume
pastures, rotational grazing and the management of natural forest regeneration have
been proposed for regions of the Amazon basin (Loker 1994). The systems provide farmers with the benefits of both annual crops and cattle raising, avoid land
degradation and effectively incorporate trees in the production system.
Agrosilvopastoral systems are important components of the economies of Tropical America’s countries (Amézquita et al. 2005). When improved and well managed,
they can become key land-use systems for the provision of environmental services,
particularly the recovery of degraded areas and carbon sequestration. In addition,
they have the capacity to provide viable economic alternatives to farmers.
13.6 Conclusions
In the Neotropics, pastures represent a high proportion of the total agricultural land,
and they are generally located on poor, acid soils. Conversion of forests to crops
and pastures has been a predominant land–use change in the Neotropics during the
second half of the twentieth century. Pasture development has been ranked one of
the least desirable land uses of tropical forest regions because of land degradation
caused by overgrazing and mismanagement. However, cattle ranching is important
for local and regional economies, therefore it is crucial to understand the factors
limiting pasture productivity, and to design ways of improving productivity and
sustainability.
Abandoned pastures can remain in a state of arrested succession due to a number
of factors that delay or impede their recovery. The factors limiting tree regeneration
in abandoned pastures can include scarcity of nutrients, soil compaction, lack of or
excess soil humidity, high solar radiation, and intra- and inter-specific competition.
To restore abandoned pastures, strategies have to consider these factors limiting
tree regeneration. Restoring abandoned pastures can involve planting trees in lines,
groups or blocks. Tree plantations can improve soil conditions, ameliorate microclimatic conditions and serve as perches for birds, bats and other seed dispersers.
Native tree species are preferred over exotics, except in cases where the performance
of native species is still not well known.
A number of strategies can be used to manage tropical pastures to increase sustainability and avoid degradation. Several local and international research institutions, such as EMBRAPA (Brazil), CIAT (Colombia), CATIE (Costa Rica) and
others, have designed pasture management strategies, mostly focusing on the use
of improved pasture grasses, grass and leguminous herb mixtures, and proper management techniques. The use of the right animal breed and size is also key to avoid
overgrazing and to sustain productivity. Agrosilvopastoral systems, or the combination of animals with trees in the landscape, offer many of the advantages of sound
management and provide environmental services, such as carbon sequestration and
conservation of biodiversity.
292
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Chapter 14
Conclusions, Synthesis, and Future Directions
Randall W. Myster
14.1 Conclusions
In Chapter 1, we saw how individual plants—critical to the functioning of an
ecosystem—respond to various abiotic and biotic environment cues that are often
“bundled together” on gradients and after disturbances (Myster 2001b). These
responses combine to define a plant species’ niche, and these probabilistic niches
express themselves fundamentally by determining plant-plant replacements over
time and space. Replacement of individual plants creates changes in plant abundances, which may also lead to changes in plant species composition. Consequently,
over larger spatial and temporal scales plant-plant replacements produce such plant
patterns as species distributions, coexistence of species, associations among species,
and succession.
Also in Chapter 1, the analysis of data from Neotropical permanent plots after
pasture and after banana and sugarcane cultivation suggested that facilitation is more
common than competition in early old field succession, that grasses are strong competitors after abandonment especially if they persist as the past crop, and that old
field plant communities are best thought of as loosely-structured assemblages of
species rather than as a higher-order “unit” of interacting species.
Many of these themes—plot data and successional patterns, change in plant
resources, plant tolerances and mechanisms—were discussed and developed in the
ensuing chapters. The remainder of the first section of the book underscored how
influential the past crop of a field can be in affecting the field’s environmental
conditions after abandonment. We saw that after pasture above-ground carbon (C)
accumulates quickly, leading to secondary forests with more above-ground C than
primary forests, and that this accumulation may be even faster if exotics are common. Below ground, however, soil C has much more modest gains, which may be
correlated not with age after abandonment, but with years the field was used as
pasture. As another example, after sugarcane cultivation in Puerto Rico an entire
new forest type dominated by Spathodea campanulata has arisen. Such forests have
high basal area, medium tree density, but low species richness. Finally, conversion of
forests to either pastures or agricultural use reduces the soil biota of fungi, microbes,
and earthworms, both in raw numbers and in species diversity. In particular, while
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14 Conclusions, Synthesis, and Future Directions
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earthworms prefer moist fields/pastures, termites prefer drier sites and exotic earthworms are more common after tree plantations than after pastures and crops.
The second section of the book involved reference to the multiple mechanisms
and tolerances that determine the process of plant-plant replacement in old fields.
The authors discussed the importance of dispersal in Neotropic old fields, with
major emphasis on the behavior of dispersing animals and the role of remnant
“nurse” trees. Cattle and horses are an understudied dispersal vector in pastures
and areas recovering from agriculture in the Neotropics, especially in tropical dry
forest. They mainly disperse large, dry indehiscent fruits of the Legume family.
Remnant trees greatly increase density and diversity of dispersed seeds in pastures
by forming vegetation “islands.” Seed predation, competition, and facilitation were
also discussed as first-tier mechanisms that are critical to old fields derived both
from wet and dry Neotropical forests. In Panama pastures, for example, the exotic
grass Saccharum spontaneum invades, competes against, and inhibits the germination and growth of old field plants. Experiments demonstrate that facilitation is
common at the early stage of early old field succession when neighbor vegetation
frequently improves the performance of tree seedlings. For dry forest, in particular,
resprouting may be an important regeneration option because trees in those areas
are specifically adapted for above-ground mortality and drought. Further speciesspecific investigation into the sources of variation in the workings of the discussed
mechanisms is needed.
The last section of the book opened with the analysis of soil microbes (controlled
by soil C which is more easily accessible on coarser soils), their availability after
agriculture (soil microbes stabilize 10 years after abandonment), and their possible
role in restoration. Then authors offered management options for manipulating succession so as to speed up restoration to natural forest. One of the options is to keep
isolated trees in pastures and plant trees in blocks or in plantations in order to reforest abandoned pasture land, to ameliorate microclimatic conditions and to attract
seed dispersal agents. In addition, measures can be taken to either delay pasture
degradation or convert unproductive pasture into more sustainable land uses. For
example, several types of agrosilvopastoral systems can contribute to increased productivity and sustainability while also helping to conserve biodiversity and sequester
carbon in vegetation and soils.
14.2 Synthesis
At least two things are true of Neotropic old fields. First, after abandonment they
have a low level of advanced regeneration from either seedling banks, sapling banks,
or from resprouting, which suggests that the analysis of Neotropic old field succession should focus on the dispersal and fate of seeds and seedlings (Grubb 1977).
Second, the dynamics of the structurally and compositionally dominant woody
plants, especially the functionally dominant trees, needs to be brought to the fore
in old field investigations in areas that were once forest, which is the most common
scenario.
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R. W. Myster
This focus on tree regeneration highlights the limiting mechanisms, or “barriers,” to tree invasion and establishment in old fields (as in the American temperate
zone: Myster 1993). It has been suggested that other plants found during old field
succession may remove some of these barriers and help facilitate tree invasion.
For example, some shrub species (such as Miconia spp.) for the most part recruit
asexually, thereby passing over any dispersal limitations that are common to trees,
especially those tree species that require animals for dispersal. Indeed, any species
that can sprout from rootstocks or from small trees/shrubs never cleared from the
field at the beginning of agriculture would have a similar advantage. As reviewed,
these shrubs, sprouts, remnant “nurse” trees, and downed trees and logs (Peterson
and Haines 2000) have a variety of functions (for instance, acting as bird perches or
seed dispersal foci, shading out grasses and other competitive old field species, and
increasing soil nitrogen levels) that can expedite tree invasion.
Decades of old field research in temperate America allow further synthesis by
providing solid ground for comparison between old field successions in the temperate and tropical zones. Following the outline of the present volume, such a comparison can be made in three ways: plant species patterns, plant mechanisms and
tolerances, and plant restoration. In particular for plant patterns, permanent plots in
temperate and tropical locations (Buell et al. 1971, Uhl 1987, Myster and Pickett
1990a,b, Myster 1993, 2003, 2007) provide the best data.
The temporal patterns in Neotropic old fields of Puerto Rico and Ecuador and
in the temperate fields show similar life-form changes: (1) grass and past crops
dominate, peaking early in succession and then declining; (2) fern and herbaceous
cover peaks early in succession, but at lower levels than grass; (3) woody vines
continue to increase in cover over the first 5 years; and (4) trees and shrubs enter,
peak, and decline in an individualistic pattern.
Both temperate and tropical fields also demonstrate a similar spatial pattern: an
exponential decline of tree density with distance to forest that is followed by a
patchy pattern of tree establishment, especially for bird-dispersed species. Tropical
old fields have a low total number of associations which are mainly positive, with
negative associations involving mostly pasture grasses. There is further correspondence between tropical and temperate fields: the proportion, number, and level of
significant associations decline with time; plants of short life-cycles are involved
in more associations, and exotic and rare species are more likely to be involved in
associations.
Key plant-plant replacement mechanisms and tolerances in the Neotropics
include competition with trees, shrubs, and large herbs for light and with grasses for
water and soil nutrients due in part to the high grass root mass per volume of soil,
drought stress, seed predation, facilitation, and seed dispersal. However, seed and
seedling pathogens may be important as well (Myster 2004). This compares well
with studies in the American temperate zone (see Myster 1993) where dispersal
is also key, along with seed predation, herbivory on seedlings (especially when it
interacts with competition), and competition. Whereas light is critical both in the
temperate zone and in the Neotropics, water may be more limiting in the Neotropics.
In terms of restoration, comparisons between neotropic and temperate old fields
are difficult to make because in the temperate zone restoration is generally not
14 Conclusions, Synthesis, and Future Directions
299
needed due to the deep fertile soils of the area that are not degraded by long-term
use without fertilizer as is common in the Neotropics. In the Neotropics, conversely,
restoration is necessary to return land to high productivity in as short a period of
time as possible and then maintain its sustainability. One way to manage succession
towards an end that better satisfies human needs is to retard succession. Another way
is to quicken the rate of recovery, perhaps by increasing facilitating mechanisms,
such as dispersal, and decreasing inhibitory mechanisms, such as predation and
competition (as discussed above), in order to generate rainforest timber to be used
as a sustainable resource or to create rainforest habitat for rare species. One more
option is to hold succession in a state when agriculture is less costly to maintain.
Yet another approach is to alter the natural successional trajectory itself to match
soil conditions for maximum crop production. Mechanisms of tree replacement, as
discussed in the present book, can be manipulated either as inhibitors or facilitators
towards any of these ends.
Whatever restoration approach is used, studies of post-agricultural succession are
indispensable to the success of restoration attempts because they provide information about how Neotropic areas regenerate and function. Management of old field
succession has to have scientific underpinnings in order to be effective. Hence, the
successional processes and mechanisms that have been discussed in this book should
be the bedrock of any restoration effort.
Studies of old field succession in the Neotropics, when combined with decadeslong similar studies in the temperate zone, become instrumental in our understanding of basic ecological paradigms. Despite the fact that the definition of ecosystem
function is generally accepted among ecologists, the definition of ecosystem structure is still hotly debated (Myster 2001a). Plants are the cornerstone of the ecosystem functions of productivity and decomposition, and an important agent in the
functions of energy/carbon dynamics (e.g., tropic exchanges) and nutrient cycling
in the soil (Wedin and Tilman 1993) which must play a critical role in determining
ecosystem structure as well.
Questions of structure and process naturally fall into two ecological paradigms
(Table 14.1). The subject matter for the first paradigm is variation of individuals
within a species, the species itself, and the species’ population. The paradigm’s
common structures include assimulation rates and population growth rates. Among
its mechanisms, which are individualistic in nature, the ecophysiology of individuals, and the interaction of species populations with the environment are essential.
The paradigm’s cardinal process—the combination of mechanisms that produce
the basic nature of plant communities under this perspective—is individual plantplant replacements that create the terrestrial vegetation patterns of plant distribution,
abundance, and association. Scientists who follow this paradigm combine individual
responses of plants to their environment or focus on dominant species in a plant
community in order to define ecosystem function.
The subject matter of the second paradigm revolves around communities and
ecosystems (Table 14.1). The paradigm’s common structures encompass species
richness, interaction coefficients, and biogeochemical cycles. Its mechanisms are
interactive, and interactions among species groups prevail. The paradigm’s key
processes are changes in plant resources and biogeochemical cycling, both of which
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R. W. Myster
Table 14.1 The two paradigms of ecology and their associated characteristics. Plant examples are
often used for illustrative purposes.
Characteristic
Paradigm I
Paradigm II
unit of study
individual variation within species,
species populations
species tolerances
leaf area, biomass, growth rate,
demographic parameters
groups of species
energy, nutrient dynamics
species interactions
species strategies, richness
interaction coefficients
biogeochemical sources/sinks
higher-order interactions
plant resource availability
biogeochemical cycles
productivity, decomposition
other energy/nutrient aspects
prime mechanism
some structures
key process(es)
plant-plant replacements
functions
productivity, decomposition
are thought to drive vegetation patterns. Followers of the second paradigm look
for relationships between “emergent” structures of a higher order and ecosystem
functions, for example, between diversity and productivity.
The two paradigms have recently been relabeled as a “Schimper” world view
(Schimper 1998) and a “Hubbell” world view (Hubbell 2001), for Schimper laid
emphasis on measurable traits of a species’ individual plants (e.g., leaf mass per
area, leaf lifespan, seed size, height at maturity) and plant evolution while Hubbell
emphasizes diversity, ignoring species differences (Westoby and Wright 2005).
Early plant ecology was dominated by the second paradigm. F. E. Clements
(1911) believed that interactions among species produced community and ecosystem structures with the ultimate result of a “super-organism.” Later, H. A. Gleason
(1926) focused on responses and structures of individual plants and on individual
species’ tolerances to environmental conditions. Gleason, having in mind successional plant communities, maintained that the biology of the species and the variation among individuals of that species were most important in determining how
plants grew and reproduced. R. H. Whittaker (1977) was the first to apply this
“individualistic” view to all plant communities, creating a “continuum” view. Some
ecologists today proceed from the more holistic Clementian view and affirm that
species interactions and resource cycling dominate structure and define function,
thus giving priority to physics and chemistry over biology in ecosystems.
Are ecologists doomed to swing back and forth between the two paradigms forever, making the science of ecology circular rather than directional? To answer
this question, we need to measure delineated plant communities over time, and
the analysis of permanent plot data allows us to do just that. In Chapter 1, permanent plot data in post-agricultural areas in Puerto Rico and Ecuador were used
to examine species associations over time and after different types of crops. The
results supported the first paradigm overwhelmingly. This tentative conclusion, of
course, needs further verification with other plant communities and in other locations. Wherever such testing takes place, however, permanent vegetation plots—
diligently monitored over long periods of time—will be critical in ascertaining
the nature of plant communities and, ultimately, the utility of this or any other
ecological paradigm.
14 Conclusions, Synthesis, and Future Directions
301
In addition to viewing a plant community as an “entity” with numerous interactions among its members, studies that are conducted within the framework of
Paradigm II often contain highly synthetic parameters that are not measured directly
(also a concern of Westoby and Wright 2006), such as “stability,” “resilience,” and
“invasibility.” These parameters can only have validity, however, if they are shown to
be the product of plant-plant replacements. For example, one old field may be more
resilient than another, returning to previous functional levels faster than the other
field after agriculture, but mechanisms behind such dissimilarity must stem from
differences in the plant-plant replacements within the fields. These dissimilarities
are determined either by a difference in the pool of species (i.e., their niches) or
by a difference in the local environmental conditions that activate different slices
of plant niches. Changes in the niche pool and/or in the environment cause plantplant replacements to “line up” in a particular way, creating plant patterns of abundance and composition that translate into such higher-level structures as resiliency.
Another old field example is when one field has faster tree invasion than another
field—greater invasibility—which must be caused also by either niche and/or environmental differences between the fields. These two examples illustrate how old
field studies can act as a template in our general understanding of the ways plant
niches are activated to produce plant-plant replacements and, ultimately, higherlevel structures.
During these investigations it would be wrong to frame the important issue of
whether it is the plants that produce environmental changes or visa versa as nothing
more than a difference in philosophical opinion (a “dialectic”: Bazzaz 1996). Priority should always be given to plants, their responses, and their niches because
an ecosystem is first and foremost biological in nature and plants are the vast
majority of that biology. Terrestrial plant communities are dynamic, and plant-plant
replacements are the main agents—the fundamental process—of that continuous
change.
Field experiments that are carried out at community or ecosystem levels provide
ground for further concern about Paradigm II because plots to which treatments are
applied change constantly and significantly throughout an experiment. As stated in
Chapter 1, plant-plant replacements occur in any plot over time, so by the end of a
field experiment many of the plants in the treatment plots have, in fact, either not
received any treatments or have received a reduced application of the treatments,
which in either case makes it impossible to test the plants statistically for any treatment affects. A review of the literature reveals that few researchers have checked if
individual plants in their study plots are old or new inhabitants. Since replacement
levels are often greater than p-values, many results that have been published from
these kind of experiments are invalid. Moreover, as the length of the experiment
increases, the likelihood of its results being invalid also increases, no matter whether
the experiment deals with natural perturbation or involves applied treatments.
Another field practice for these kinds of experiments is to measure one individual
plant of a study species in each plot at each sampling period. Since plants are not
individually marked, however, odds are great that a different plant is being measured
each time! As before, that plant may have been introduced late in the experiment,
receiving reduced treatments or no treatments at all. The usual statistical remedies
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R. W. Myster
of a control plot, more replication, or a “blocked” design will not solve this problem
because replacements and the plant patterns that result from them are probabilistic, not random.
Finally, Paradigm II allows the questionable approach of grouping species by
function or strategy (e.g., Grime 1977) that treats all plant species within a group
as the same, to the point of looking at parts of the world as green packages where
a patch of moss would be equivalent to a tree (the “big leaf” perspective: Bazzaz 1996). This approach accommodates the risk of ignoring evolution because it
downgrades plants to nothing more than temporary way-stations for energy and
ions and thus passes over species adaptations and differences that are important
in any ecosystem. It needs to be remembered that evolution is a process of creating
individual species that have to be functionally different and important within an
ecosystem in order to survive.
It appears logical then that for the purpose of avoiding significant limitations
in research practices of Paradigm II (e.g., unmeasured parameters, replacements in
treatment plots, species groupings) and of conducting accurate and useful experiments in general, researchers should either concentrate their attention on individual marked plants or direct their efforts to short-term projects. The outlined basic
methodological issues, on the one hand, and the results of the analysis of permanent plot data in Chapter 1, on the other, cast considerable doubt on the validity of
Paradigm II. It seems that this paradigm wants to present ecosystems as a “jumble”
of interactions where everything is connected to everything else and all is equally
important (or unimportant), that is to say there is no over-arching structure, set of
priorities, or (ultimately) meaning.
14.3 Future Directions
The conceptual framework and the studies of post-agricultural succession in the
Neotropics that are presented in this book hold the following recommendations for
future research of Neotropic old fields:
(1) set up permanent vegetation plots in more Neotropic ecosystems and monitor
them for decades with the aim of investigating the pace and nature of land
recovery after common past crops, including other aspects of the agricultural
disturbance regime (e.g., the year of abandonment, the season of abandonment,
and whether the field was plowed or not when abandoned);
(2) explore the ecophysiology of Neotropic old field plants (see Burton and Bazzaz
1991, 1995 for similar studies in temperate old fields), the tolerances of their
niches (see Myster and Everham 1996 for the germination niche of tropical
trees), and other aspects of the biology and ecology of tropical plants (see
Brokaw 1998 for an examination of Cecropia schreberiana); combine these
studies with further mechanistic investigation of old field plants in the Neotropics for a complete definition of plant niches, using data from permanent plots
to identify important plants for further investigation;
14 Conclusions, Synthesis, and Future Directions
303
(3) sample environmental factors in study old fields, with a focus on the abiotic
environment (e.g., energy pathways and biogeochemical cycles that lead to
changes in plant resources) and on the abundance, composition, and behavior
of the biotic environment (e.g., dispersers, predators, herbivores, pathogens,
other plants) in the neighborhood;
(4) monitor those components of the environment (e.g., carbon) that indirectly
affect the immediate abiotic and biotic conditions of a plant’s life;
(5) perform restoration experiments in the field.
In this book, two approaches in investigation of plant communities have been
contrasted: a “static” investigation that engages in snapshot growth and allocation
field experiments, focusing on the largest or most common plants present at that
point in time and space, and a “dynamic” view focusing on how plant niches and
the local environment determine plant-plant replacements over time and space. For
all plant communities, a major future challenge for scientists will be to find out how,
under the influence of the local environment, plant niches generate the plant-plant
replacements which produce the earth’s plant patterns. The mere demonstration of
these patterns will not suffice (Myster 2001a). There can be no shortcuts such as the
use of “surrogate” parameters instead of direct measurements of plant function and
structure.
Results from decades-long post-agricultural studies and the conceptual framework from Chapter 1 suggest how to proceed: key species and replacement patterns
can be ascertained by using permanent plots, plant niches can be determined through
various kinds of field and greenhouse experiments, and plant-plant replacements can
be investigated by sampling local environmental factors and combining them with
plant niches. Followers of Paradigm II may continue to look only to the abiotic
environment for answers, yet the primary emphasis should be on the ways in which
plants respond to and interact with both the living and non-living constituents of the
environment.
As the analysis of data from old field permanent plots and the results of old
field experiments and studies in various places throughout North, South, and Central America have shown, plant communities are much more individualistic (Pickett
1982, Myster and Pickett 1988) and much more dynamic than is often assumed.
References
Bazzaz, F. A. 1996. Plants in Changing Environments: Linking Physiological, Population and
Community Ecology. Cambridge University Press. Cambridge, UK.
Brokaw, N. V. L. 1998. Cecropia schreberiana in the Luquillo Mountains of Puerto Rico. The
Botanical Review 64:91–120.
Buell, M. F., Buell, H. F., Small, J. A., and Siccama, T. G. 1971. Invasion of trees and secondary
succession on the New Jersey Piedmont. Bulletin of the Torrey Botanical Club 98:67–74.
Burton, P. J., and Bazzaz, F. A. 1991. Tree seedling emergence on interactive temperature and
moisture gradients and in patches of old-field vegetation. American Journal of Botany 78:
131–149.
Burton, P. J., and Bazzaz, F. A. 1995. Ecophysiological responses of tree seedlings invading
different patches of old-field vegetation. Journal of Ecology 83:99–112.
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Grime, J. P. 1977. Evidence for the existence of three primary strategies in plants and its relevance
to ecological and evolutionary theory. American Naturalist 111:1169–1194.
Grubb, P. J. 1977. The maintenance of species-richness in plant communities: the importance of the
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Princeton University Press.
Myster, R. W. 1993. Tree invasion and establishment in old fields at Hutcheson Memorial Forest.
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Myster, R. W. 2001a. What is Ecosystem Structure? Carribbean Journal of Science 37:132–134.
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Myster, R. W. 2003. Vegetation dynamics of a permanent pasture plot in Puerto Rico. Biotropica
35(3):422–428.
Myster, R. W. 2004. Post-agricultural invasion, establishment, and growth of Neotropical trees.
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Myster, R. W. 2007. Early successional pattern and process after Banana, Sugarcane and Pasture
in Ecuador. The New Zealand Journal of Botany 45:35–44.
Myster, R. W., and Everham, E. M. III. 1999. Termination cues across the distrubance regime in
the Puerto Rican rain forest. Tropical ecology 110:89–98.
Myster, R. W., and Pickett, S. T. A. 1988. Individualistic patterns of annuals and biennials in early
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Forest Center. Permanent Plotter 3:2–3.
Myster, R. W., and Pickett, S. T. A. 1990b. Initial conditions, history and successional pathways in
ten contrasting old fields. American Midland Naturalist 124(2):213–238.
Myster, R. W., and Pickett, S. T. A. 1992a. Effects of palatability and dispersal mode on spatial
patterns of trees in old fields. Bulletin of the Torrey Botanical Club 119(2):145–151.
Myster, R. W., and Pickett, S. T. A. 1992b. Dynamics of associations between plants in ten old
fields during 31 years of succession. Journal of Ecology 80(2):291–302.
Peterson, C. J., and Haines, B. L. 2000. Early successional patterns and potential facilitation of
woody plant colonization by rotting logs in premontane Costa Rican pastures. Restoration
Ecology 8:361–369.
Pickett, S.T.A. 1982. Population patterns through twenty years of old field succession. Vegetatio
49:45–59.
Schimper, A. F. W. 1998. Pflanzengeographic auf Physiologischer Grundlage. G. Fisher.Uhl, C.
1987. Factors controlling succession following slash-and-burn agriculture in Amazonia. Journal of Ecology 75(2):377–407.
Wedin, D. A., and Tilman, D. 1990. Species effects on nitrogen cycling: a test with perennial
grasses. Oecologia 84:433–441.
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in Ecology and Evolution 21:261–268.
Index
Abundances, 296
Acacia, 168–170, 176, 178, 179–180,
182–183, 289
Advanced regeneration, 297
Agriculture, 6
Agroforestry, 104, 106, 269, 275
Agrosilvopastoral systems, 275, 286, 291
Albizia, 36
Amazonia, 106, 248, 249
Andropogon gayanus, 236
Araucaria angustifolia, 223, 289
Arrested succession, 266, 291
Asexual, 216, 298
Association, 9, 11, 296, 298
Axonopus, 289
Bacterial biomass, 101, 102, 103
Basal
area, 89
respiration, 252
Bats, 276
Big leaf perspective, 302
Biodiversity, 272, 291
Biogeochemical change/cycles, 98, 109, 299
Bird perches, 272
Brachiaria, 289
Brazil, 223, 232, 247
Bucida buceras, 74
Buell-Small successional study, 10
Bulk density, 41, 120
Bulldozing, 268, 283
Burning, 266, 283
13 C/14 C,
42, 121, 255
Calcium, 274
Carbon, 15, 296, 303
above/below ground, 30
organic/inorganic, 26
sequestration, 22, 273, 291
standing stock, 24
Cattle, 166–181, 265, 283, 297
Cecropia, 90, 271
Century model, 258
Cerrado, 232
Chiapas, 167, 176–181
Chronosequences, 9, 34, 38, 239, 284
Clements, F. E., 300
Climate change, 5
Cloud forest, 193
Coexistence, 8, 296
Colletotrichum, 234
Colombia, 107
Competition, 11, 142, 143, 144, 149, 150, 156,
221, 233, 237, 277, 296
Composition, 296
Costa Rica, 193
Crop systems, 105
Cuttings, 239
Decay constant, 24
Decomposition, 3, 93, 95, 98, 109, 299
Degradation, 266
Delonix regia, 86
Dialectic, 301
Dichogaster modigliani, 127
Dichotomies, 7
Dipteryx panamensis, 279
Dispersal, 17
Distribution, 9, 296
Disturbances, 3, 296
regime, 5, 302
Drought, 142, 144
Dry forest, 167, 175, 177–178, 231
Earthworms, 17, 95, 104, 107, 108, 115,
285, 296
endogeic/anecic/epigeic, 115
Ecophysiology, 299, 302
Ecosystems, 3
305
306
function, 3, 299
structure, 3, 299
Ecuador, 10
Edge effects, 9
Equilibrium, 8
Erythrina poeppigiana, 86, 288
Estherella, 123
Eucalyptus, 36, 223
Fabaceae, 167, 175, 176, 177
Facilitation, 11, 219, 221, 227, 280,
296, 298
Fallow, 5, 44, 285
Fence, 290
Ferns, 272
Ficus, 280
Fire, 142, 144, 149, 156, 157, 159, 160, 165,
180–181
Forage, 266, 290
Frugivores, 193–210
Functional groups, 109
Fungal biomass, 17, 94, 100, 101, 102, 103,
109, 296
Fusarium, 234
Germination, 166, 167, 168–174, 176, 179,
180, 181, 233
Gleason, H. A., 300
Gliricidia sepium, 290
Glossodrilus, 124
Gradients, 3, 221, 296
Grasses, 143, 144, 149, 159, 161, 227,
237, 277
Grassland, 104, 105, 109
Grazing, 181
Guazuma, 74, 170, 171, 178,
184, 288
Gullies, 265
Heliophillous species, 278
Hemoglobin, 134
Herbicides, 273
Herbivory, 236
H yeronima alchorneoides, 276
Inga, 224, 280, 288
Insects, 219
ants, 219, 234
termites, 297
Intensification of agriculture, 256
Interaction coefficients, 299
Invasibility, 301
Isotope, 42
Karstic, 26, 85
Index
Land conversion, 252
Land-use, 27, 115, 265
Leucaena leucocephala, 275
Light, 229, 236
Limestone, 26, 233
Litter, 37, 77, 115, 216, 219, 249, 277
Livestock, 166–181
Long-term ecological research (LTER), 9
Lower montane wet forest, 193
Luquillo Experimental Forest (LEF), 9,
87, 117
Macrofauna, 17
Magnesium, 274
Mammals, 216, 219, 234
Management, 17
Manilkara, 85
Maquipucuna Reserve, 10
Mechanical clearing, 6
Mechanism, 4, 298, 302
Megascolecidae, 118
Metabolic quotient, 252
Mexico, 167, 176–181
Miconia spp., 216, 298
Microbial biomass, 99, 100, 101
Microorganisms, 94, 95, 96, 99, 100, 115,
134, 297
Millipede, 95, 104, 108
Model, 24, 222
Monoculture, 221
Monteverde Cloud Forest Reserve, 193–194
Mucus, 128
Multivariate analysis, 145, 146, 149, 150, 205
Mycorrhizal, 99
Myriapods, 95, 104
15 N, 123
Neotropics, 14
Niche, 4, 7, 8, 218, 301, 303
Nitrification, 271
Nitrogen fixation, 95
Nitrogen mineralization, 271
Nurse trees, 9, 219, 297, 298
Nutrients, 229, 282
Old field, 4, 299
Organic matter, 26, 88, 120, 251, 273
Overgrazing, 265
Oxisol, 248
Palms, 34
Panama, 149, 160
Panicum, 86, 283
Paradigms, 299
Parthenogenesis, 131
Past crop, 6, 296, 302
Index
Coffee (Coffea spp.), 90, 217
Banana (Musa spp.), 6, 9, 10, 217, 296
signature, 6, 34
Sugarcane (Saccharum officinarum), 6, 9,
10, 13, 73, 217, 296
Pasture, 5, 99, 104, 106, 107, 108, 192–215,
265, 296
Pasture reformation, 256
Pathogens, 233
Paulownia, 289
Perionex excavatus, 127
Permanent plots, 5, 9, 300, 302, 303
Petri dish, 217
Phoma, 234
Phosphorus, 265
Physical habitat structure, 211–212
Pinus, 221, 224, 274
Plantations, 35, 101, 107, 109
Plant-plant replacements, 7, 297, 298, 299,
301, 303
Pleistocene megafauna, 166, 177–178
Polypheretima elongata, 127
Pontoscolex corethrurus, 118, 125
Post-agricultural, 6, 115, 299
Potassium, 274
Predation of earthworms, 122
Predation of seeds, 17, 142, 143, 156, 216, 233
preferences, 216
seed size, 218
Productivity, 3, 33, 272, 282, 299
Psidium guajaba, 74, 86, 288
Pterocarpus, 74
Puerto Rico, 9, 23, 76, 100, 108
Radiocarbon, See Isotope
Rainforest, 175, 177
Recruitment foci, 202, 207–208
Reforestation, 100, 101, 103, 108
Rehabilitation, 272
Remnants, 143, 149, 150, 153, 156, 160,
197–201, 203–205, 206–208, 272,
288, 297
Remote sensing, 24
Removal treatment, 224
Resilience, 301
Response curves, 3
Resprouting, 239
Restoration, 17, 179–180, 272,
298, 299
Rhizosphere, 121
Rhus, 216
Rhyzobium bacteria, 284
Rodents, See Mammals
Root
biomass, 38
307
stocks, 40, 216
suckers, 143, 157, 238
Roystonea borinquena, 79
Saccharum spontaneum, 149, 150, 151,
160, 297
Savanna, 175, 178
Scarification, 167, 175–176, 180
Secondary forest, 101, 105
Seed
bank, 165, 167, 178
dispersal, 142, 149, 151, 152, 156, 160,
165–181, 192–215, 233
syndrome, 167–175, 177
Setaria sphacelata, 6, 10, 13
Sheep, 286
Shelterbelts, 211
Shifting agriculture, 269
Shrubs, 216, 219, 298
Silvopastoral systems, 179
Slash and burn, 6, 34, 252
Snapshot, 8, 303
Soil
classes, 231
compaction, 236, 266, 287
erosion, 266
fertility, 266, 282
fungi, 94, 101, 104
humidity, 128
macrofauna, 93, 95, 97, 105, 134
microbial community, 93, 99, 104
pH, 120, 270
porosity, 115, 271
texture, 120
Spathodea campanulata, 73, 296
Spearman coefficients, 11
Species
composition, 141, 142, 145, 150, 156, 160,
204–205, 208–210
richness, 141, 145, 146, 150, 156
Sprout/resprout, 216, 297
Stability, 301
Stocking, 285
Strategy, 302
Stress, 11, 221
Succession, 6, 141, 143, 149, 296
trajectory, 205, 208–210, 299
Surrogate parameters, 303
Sustainability, 5
Synthetic parameters, 301
Syzygium, 85
Teak, 36
Termites, 95, 104
T hespesia grandiflora, 78–79
308
Thinning state, 7
Tolerances, 4, 298, 302
T oona australis, 289
Tree regeneration, 231, 298
Tropic of Cancer/Capricorn, 14
Tussocks, 238
Ultraviolet radiation, 131
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Index
Vernonia, 11, 229
V ismia, 271
V ochysia ferruginea, 273
Water, 233, 236
Whittaker, R. H., 300
Windbreaks, 272, 282
Zoochory, 165–181