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Regeneration of commercial tree species following controlled burns in a tropical dry forest in eastern Bolivia

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Title:
Regeneration of commercial tree species following controlled burns in a tropical dry forest in eastern Bolivia
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Kennard, Deborah K., 1969-
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English
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viii, 206 leaves : ill. ; 29 cm.

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Subjects / Keywords:
Ashes ( jstor )
Dry forests ( jstor )
Ecology ( jstor )
Forests ( jstor )
Seedlings ( jstor )
Soils ( jstor )
Species ( jstor )
Sprouts ( jstor )
Trees ( jstor )
Tropical forests ( jstor )
Botany thesis, Ph. D ( lcsh )
Dissertations, Academic -- Botany -- UF ( lcsh )
Forest ecology -- Bolivia ( lcsh )
Forest regeneration -- Bolivia ( lcsh )
Prescribed burning -- Bolivia ( lcsh )
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bibliography ( marcgt )
theses ( marcgt )
non-fiction ( marcgt )

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Thesis:
Thesis (Ph. D.)--University of Florida, 2000.
Bibliography:
Includes bibliographical references (leaves 187-205).
General Note:
Printout.
General Note:
Vita.
Statement of Responsibility:
by Deborah K. Kennard.

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University of Florida
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Copyright [name of dissertation author]. Permission granted to the University of Florida to digitize, archive and distribute this item for non-profit research and educational purposes. Any reuse of this item in excess of fair use or other copyright exemptions requires permission of the copyright holder.
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REGENERATION OF COMMERCIAL TREE SPECIES FOLLOWING CONTROLLED
BURNS IN A TROPICAL DRY FOREST IN EASTERN BOLIVIA













by

DEBORAH K. KENNARD


A DISSERTATION PRESENTED TO THE GRADUATE SCHOOL
OF THE UNIVERSITY OF FLORIDA IN PARTIAL FULFILLMENT
OF THE REQUIREMENTS FOR THE DEGREE OF
DOCTOR OF PHILOSOPHY

UNIVERSITY OF FLORIDA


2000




























Dedicated to my parents,

Margaret Kennard and Robert Kennard,

and to my husband,

Josh McDaniel















ACKNOWLEDGMENTS


I sincerely thank everyone who contributed in some way to this dissertation. First, I

would like to thank the members of my committee. Jack Putz, my advisor throughout my years at

the University of Florida, has been very influential in my interest in forest ecology and

management. I thank him for the opportunities he has provided, his time as a teacher and mentor,

and his diligence as an editor. His enthusiasm for ecology will influence me for years to come.

Henry Gholz was a tremendous help with the project design and generously provided the use of

his lab for resin analyses. Kaoru Kitajima patiently helped me with statistics and was instrumental

in the development of the seedling chapter. Insightful suggestions by Earl Stone aided me

throughout every aspect of the project; I was repeatedly inspired by his immense knowledge and

experience. Finally, I would like to thank George Tanner who provided much-needed

encouragement at critical times.

This dissertation would not have been possible without the generous financial and

logistical support of BOLFOR (Proyecto de Manejo Forestal Sostenible). Several members of

BOLFOR were crucial in its completion. First, I would like to thank Todd Fredericksen, the

Forest Ecologist at BOLFOR, for his help conducting the controlled burns, statistical and technical

advice, ready supply of coca leaves, and humor. T. Fredericksen, J. Nittler, and W. Cordero

provided administrative support in Bolivia. I gratefully acknowledge the following people who

generously volunteered their time to assist with fieldwork: J. McDaniel, L. MacDonald, J.

Chuviru, T. Fredricksen, N. Fredricksen, J. Lincona, J. Justiniano, A. Ademar, K. Gould, F.

Fatima, K. Hueberger, B. Flores, M. Toledo, L. Anderson, B. Mostacedo, and the Aberdeen

students. I wish to acknowledge T. Killeen and the herbarium at the Museo de Noel Kempf









Mercado for use of the data collected in the 1995 inventory of Las Trancas. M. Toledo kindly

assisted with plant identification. My husband and I are very grateful to Todd and Nell

Fredericksen for their hospitality in Santa Cruz.

Numerous Chiquitano community members assisted throughout the 18 months of this

project; I will be forever grateful for and impressed by their hard labors. In particular, I would

like to thank Don Juan "Loco" Pesoa for being instrumental in the installation of the treatments;

Don Juan Faldin for sharing his knowledge of plants and their local uses; and, Don Lucas

Salvatierra for his assistance in locating and measuring abandoned agricultural fields. I would

also like to thank the many members of San Lorenzo who welcomed me into their community

during my time away from El Campamento de Las Trancas.

At the University of Florida, I would like to thank J. Bartos at the Analytical Research

Lab for analysis of soil samples. In the lab of H. Gholz, I would like to thank D. Noletti and K.

Clark for their valuable assistance with the resin extractions. I was supported by a teaching

assistantship offered by the Department of Botany and the Department of Biological Sciences.

Comments of several friends greatly improved the final draft: G. Blate, K. Gould, B. Ostertag, T.

Fredericksen, and J. McDaniel. Most importantly, I would like to thank my traveling companion,

field-assistant, Spanish teacher, motorcycle driver, toilet-digger, wasp-magnet, translator, and

husband, Josh McDaniel.
















TABLE OF CONTENTS

pnage

ACKNOWLEDGMENTS................................................................................. iii

A B S T R A C T ............................................................................................................................. vii

CHAPTERS

1 INTRODUCTION

Introdu action ....................................................................................................................... 1
Conservation and Management of Tropical Dry Forests................................................... 2
Management of Tropical Dry Forests in Eastern Bolivia..................................................... 3
The Role of Historic Disturbance Regimes in Forest Management................................... 4
The Potential of Prescribed Burning in the Management of Bolivian Dry Forests................. 7
Scope of D issertation ....................................................................................................... 8

2 STUDY SITE AND TREATMENT DESCRIPTIONS

In trod auction ..................................................................................................................... 11
S tu d y S ite .............................................................................................................. ......... 13
M methods ......................... ....... ................ ......... ... ............... ..... ... 18
Initial T reatm ent R results ................................................................................................ 24


3 EFFECTS OF CANOPY GAP FORMATION, PLANT REMOVAL, AND CONTROLLED
BURNS OF HIGH AND LOW INTENSITIES ON SOIL CHEMICAL AND
PHYSICAL PROPERTIES

Introdu action ............................................................................................ ........................ 2 9
M eth od s ..................................................................... .................................................... 3 1
R e su lts ....... ....................................................... .............................................................. 3 8
D discussion ................................................................................................ .................... 52
C o n clu sio n s ....................................................... ............................................................. 7 1

4 EFFECTS OF CANOPY GAP FORMATION, PLANT REMOVAL, AND CONTROLLED
BURNS OF HIGH AND LOW INTENSITIES ON EARLY REGENERATION OF
COMMERCIAL TREE SPECIES

Introduction ................................................................................................. ................. 72
M eth od s ....................................................... .................................................................. 7 5









Results ....................................................................................... ..................................... 79
Discussion ........................................................................................................................ 96
Im plications for m anagement................................................................................ .......... 111


5 EFFECTS OF CANOPY GAP FORMATION, PLANT REMOVAL, AND CONTROLLED
BURNS OF HIGH AND LOW INTENSITIES ON A DRY FOREST PLANT
COMMUNITY

Introduction ................................................................................................................... 113
M methods ........................................................................................................................ 115
Results ......................................................................................................................... 119
Discussion ..................................................................................................................... 132
Conclusions ................................................................................................................... 147

6 COMMERCIAL TREE SPECIES REGENERATION FOLLOWING AGRICULTURAL
ABANDONMENT IN BOLIVIAN DRY FORESTS

Introduction ................................................................................................................... 149
M methods ........................................................................................................................ 150
Results ......................................................................................................................... 153
Discussion ................................................................................. .................................... 165
Implications: Management potential of secondary forest in Lomerio............................... 172

7 SUMMARY AND CONCLUSIONS

Sum m ary of Study Results............................................................................................. 174
Im plications for M anagem ent......................................................................................... 175


APPEN DIX ............................................................................................................................ 181

REFEREN CES....................................................................................................................... 187

BIOG RAPHICA L SKETCH ............................................................................ ..... ....... 206
















Abstract of Dissertation Presented to the Graduate School
of the University of Florida in Partial Fulfillment
of the Requirements for the Degree of Doctor of Philosophy


REGENERATION OF COMMERCIAL TREE SPECIES FOLLOWING CONTROLLED
BURNS IN A TROPICAL DRY FOREST IN EASTERN BOLIVIA



by

Deborah K. Kennard

May 2000

Chairman: Francis E. Putz
Major Department: Botany

Low levels of disturbance associated with selective logging may be insufficient for the

establishment of many Bolivian dry forest timber species, the majority of which are shade-

intolerant. To examine the ecological potential of prescribed burning as a silvicultural tool, I

compared the effects of canopy opening, plant removal, and controlled bums of high and low

intensities on 1) soil properties; 2) establishment, growth, and survival of commercial tree species;

3) and, plant community structure and composition. To describe commercial tree regeneration over

longer time scales, I characterized tree population structures in abandoned slash-and-bum fields

ranging in age from 1-50 years, and compared these to a mature forest stand.

Both high- and low-intensity bums caused a dramatic but temporary increase in soil

nutrients. High-intensity bums altered several soil physical properties, whereas low-intensity bums

had little effect. Plant removal and canopy opening had little effect on soil chemical and physical

properties.









Three responses to gap treatments were observed among commercial tree species. 1)

Shade-intolerant species regenerating from seed were most successful following high-intensity

bums. 2) Shade-tolerant species were most abundant in treatments where survival of their

advanced regeneration was most likely (gap control and plant removal). Some of these species had

the ability to survive controlled bums by sprouting. 3) Individuals of root sprouting species were

most abundant following plant removal and low-intensity bums.

Sprouts dominated regeneration following canopy opening, plant removal, and low-

intensity bums. In contrast, seedlings dominated following high-intensity bums. High-intensity

bums shifted species composition relative to the less disturbed treatments.

Regeneration of shade-intolerant timber species was most abundant in young slash-and-

bum fallows. Similar tree population structures in older slash-and-burn fallows and the mature

forest stand suggests that the mature forest likely formed following a large-scale disturbance.

Although prescribed burning enhanced the regeneration of shade-intolerant timber species,

it is not likely to become a forest management tool in Bolivia in the near future due to economic

and political factors. Managing secondary forests in Bolivia would provide an alternative to

current attempts to regenerate these species after selective harvesting of mature forest.
















CHAPTER 1
INTRODUCTION


Introduction

Eastern Bolivia contains some of the largest and most diverse tracts of tropical dry forest

in Latin America. Natural forest management for timber, if profitable, is one means of

discouraging conversion of these forests to competing land uses. However, insufficient

regeneration of many commercial timber species presently poses an ecological barrier to sustained

timber yield, prompting forest managers to explore additional silvicultural methods to enhance

regeneration of these species. The low levels of disturbance associated with highly selective

logging may be insufficient for the establishment of many dry forest timber species, the majority of

which are shade-intolerant and likely require moderately intense disturbances for their

establishment. Fire, of both natural and anthropogenic origins, has likely been a pervasive

influence on tropical dry forests, and therefore, prescribed burning may be an effective silvicultural

tool to enhance regeneration of timber species following selective logging.

In this dissertation, I present the results of studies that examined commercial tree

regeneration following disturbances of various intensities in a dry forest in lowland Bolivia,

including harvesting gap formation, controlled burns of high and low intensity, and slash-and-bumrn

agriculture. My goal in carrying out these studies was to determine the regeneration requirements

of these commercial tree species, as well as to examine the effects of potential silvicultural

treatments on forest soils and community structure and composition.









Conservation and Management of Tropical Dry Forests

Tropical dry forests comprise approximately 42% of tropical forest land, more than either

moist or wet tropical forests (Murphy and Lugo 1986). Tropical dry forests also have supported

higher human population densities than wetter tropical forests for centuries (Murphy and Lugo

1986) and, as a result, have suffered more degradation and deforestation (Mooney et al. 1995,

Murphy and Lugo 1995). Efforts to slow conversion rates of dry tropical forest have been

negligible (Mooney et al. 1995). For example, in 1988, less than 2% of the original dry forest on

the Pacific coast of Central America remained intact and less than 0.1% had conservation status

(Janzen 1988). Consequently, tropical dry forests are considered by some ecologists as the most

threatened of the major tropical forest types (Janzen 1988).

Given the extensive use of tropical dry forests by rural people, their strict preservation

may not be a realistic conservation goal. As Johnson and Carbarle (1993) note, most developing

tropical countries rarely have the luxury of opting for forest preservation over forest exploitation.

Consequently, in most tropical countries, conversion of forested land continues to increase while

the establishment of protected areas remains low (FAO 1999). Consensus is emerging among

ecologists that protected areas, due to their small number and size, cannot effectively conserve the

majority of tropical species (Hansen et al. 1991, Heinrich 1995, Bawa and Seidler 1998).

Mounting concern over global declines of biodiversity has prompted many ecologists to look

outside of parks and nature preserves to semi-natural areas that may help maintain, or at least slow

the loss of, biodiversity (Sayer and Wegge 1992, Chazdon 1998).

Natural forest management, the sustainable production of timber from natural forest areas,

has been proposed as a means of maintaining forest value, thereby deterring land owners from

clearing forested land for other more profitable and destructive land uses (Poore et al 1989,

Johnson and Carbarle 1993, Maser 1994, but see Rice etal. 1997). Although definitions of

natural forest management vary, they usually encompass two ideas: first, a sustained yield of forest









products, and second, achieving this sustained yield through means that maintain other

environmental services, such as biodiversity, soil quality, and hydrology (reviewed by Johnson and

Carbarle 1993). By maintaining forests in a semi-natural state, natural forest management is

viewed by some as a critical means of maintaining biodiversity (Hansen et al. 1991, Sayer and

Wegge 1992, Frumhoff 1995, Heinrich 1995, Dickinson et al. 1996, Putz et al. in press, but see

Bawa and Siedler 1998), particularly in regions where forests are in danger of conversion. And,

despite the fact that few modem examples of economically viable natural forest management

projects exist (e.g., Panayotou and Ashton 1992, Johnson and Carbarle 1993, Rice et al. 1997,

Bawa and Siedler 1998, Bowles et al. 1998, but see Leslie 1987), the promotion of sustainable

forest management has become a mainstay in international strategies for the protection of tropical

forests (Bawa and Siedler 1998, Haworth and Cousell 1999). As Haworth and Cousell (1999: 62)

explain "this approach has often been justified on the grounds that it is the result of a difficult

choice between accepting, on the one hand, the inevitability of continued commercial logging of

natural forests, which will cause some damage to the ecosystem, or, on the other hand, facing the

complete loss of the forest to other causes."

Management of Tropical Dry Forests in Eastern Bolivia

The Chiquitania region in eastern Bolivia contains one of the largest and most diverse

tropical dry forests in the neotropics (Gentry 1993, Killeen et aL 1998). Although there is

currently 150,000 to 200,000 km2 of relatively intact forest in Chiquitania, Dinerstein et al. (1995)

identified this area as one of the most endangered ecosystems in the neotropics. Deforestation in

the alluvial soils near the city of Santa Cruz is in excess of 80,000 ha year' (Killeen et aL. 1998).

This conservation threat comes largely from large-scale industrial agriculture, but other economic

activities, such as cattle ranching, contribute to this rapid conversion of forested land. These

trends mimic past events in Argentina, Paraguay, and Eastern Brazil where similar dry forests have

been deforested and fragmented over the past two decades (Killeen et al. 1998).









Due to recent forest policy changes in Bolivia, natural forest management may now be a

practical means of controlling deforestation in Chiquitania. In 1996, a new forestry law was

passed that requires, among other features, management plans for all Bolivian forests (Nittler and

Nash 1999). Bolivian logging companies now operate with management plans on an estimated 5.7

million hectares of forest and a total of 660,000 hectares of Bolivian forest has been certified as

sustainably managed (Nittler and Nash 1999).

The Lomerio Community Forest, located in the center of Chiquitania, was the first

Bolivian forest to be certified. Its 60,800 hectares are owned and managed by 27 communities of

the Chiquitano indigenous people. The Chiquitanos have been managing their forests for 19 timber

species, 5 of which are classified as highly valuable (Table 1-1). Acquiring and maintaining

adequate regeneration of commercial tree species, a challenge faced by all natural forest managers,

is particularly apparent in Lomerio. For example, seedlings and saplings of 12 of the 19

commercial species are rare in forest understories (Table 1-1). A lack of seed sources due to

previous over-harvesting may account for the scarcity of regeneration among highly valued timber

species. However, poor regeneration plagues most of the tree species that have only recently been

harvested (Fredricksen 1999). Apparently, the current harvesting and silvicultural techniques

employed in Lomerio do not create conditions appropriate for the regeneration of these species. A

better understanding of the regeneration requirements of these commercial tree species is critical,

as continued regeneration failures will undoubtedly compromise the long-term sustainable

management of these forests.

The Role of Historic Disturbance Regimes in Forest Management

It is often assumed that forest management is more compatible with long-term

sustainability if timber harvesting and silvicultural techniques are designed to mimic historic

disturbance regimes (e.g., Pickett and White 1985, Oliver and Larson 1996, Attiwill 1994a,

1994b). Although this assumption has rarely been tested, ecologists argue that replacing harvested






5













Table 1-1. Table modified from Pinard et al. 1999 that reports characteristics of 19
commercial tree species of the dry forests of Lomerio. Species were matched to a
general silvicultural system (even- or uneven-aged) based on their regeneration
requirements. Timber value is based on market value in 1999.
Timber Managment Adult Sapling Shade
value system rarity rarity tolerance
1 Amburana cearensis high even 3 3 1
2 Anadenanthera colubrina low even 1 2 1
3 Aspidosperma cylindrocarpon low uneven 2 3 2
4 Aspidosperma rigidum low uneven 1 2 2
5 Astronium urundueva low even 2 3 1
6 Caesalpinia pluviosa low uneven 1 2 2
7 Cariniana estrellensis low uneven 2 2 3
8 Cedrelafissilis high even 3 3 1
9 Centrolobium microchaete high even 1 3 1
10 Copaifera chodatiana low uneven 2 3 3
11 Cordia alliodora high even 3 2 1
12 Hymenea courbaril low even 3 3 1
13 Machaerium scleroxylon high uneven 1 2 3
14 Phyllostylon rhamnoides low uneven 2 2 3
15 Platymiscium ulei low even 3 3 1
16 Schinopsis brasilensis low even 2 3 1
17 Spondias mombin low even 2 3 1
18 Tabebuia impetiginosa low even 1 3 1
19 Tabebuia serratifolia low even 3 3 1
Shade tolerance: 1 = high light only, large gaps; 2 = partial shade, small gaps;
3 = partial or full shade, understory.
Adult rarity (> 20 cm dbh): 1 common (> 5 ha-'); 2 = intermediate (1-5 ha'); 3 = rare (< 1 ha").
Sapling rarity (5-10 cm dbh): 1 common (>20 ha'); 2 = intermediate (5-20 ha'-); 3 = rare (<5 ha').
Even: even-aged management system with group selection.
Uneven: uneven-aged management system with single-tree selection.









trees without irreversibly damaging the residual forest is more likely to occur under conditions

similar to those that formed the original stand (Uhl et al. 1990). The selective cutting systems used

in many tropical forests are often justified on models of gap-phase regeneration in unharvested

forests (e.g., Whitmore 1989, Hartshorn 1989, Gomez-Pompa and Burley 1991).

Gap-phase regeneration, however, is not the most appropriate model for tropical dry

forests. Evidence suggests that single tree-fall gaps are smaller and less frequent in tropical dry

forests than in moist or wet forests (Dickinson 1998). Rather, very large gaps caused by

catastrophic disturbances more likely govern dry forest dynamics. In Central America for

example, mahogany (Swietenia macrophylla) has been noted to regenerate in even-aged stands

after hurricanes and fires (Lamb 1966, Snook 1996). The low level of disturbance created during

highly selective logging appears to be a poor replicate of this disturbance regime, and possibly for

this reason, natural regeneration of mahogany is scarce in most selectively logged areas (Verissimo

et al. 1995, Gullison et al. 1996, Whitman et al. 1997).

In contrast to Central America, the agents of large scale disturbance have not been a topic

of frequent study in Bolivian dry forests (but see Pinard and Huffminan 1997). As hurricanes are

absent in this landlocked country, it is likely that forest fires (both natural and anthropogenic) have

likely been the most pervasive influence on Bolivian forests. Natural fires have historically

influenced vast areas of Amazonian forest (Clark and Uhl 1987), particularly in dry or deciduous

forests where dry fuels may favor lightning fires (Middleton et al. 1997). In fact, most radiometric

dates of charcoal found throughout the Amazon correspond with the expansion of dry forests

during the dry glacial epochs (Saldariagga et al. 1986, Goldammer 1993, Prado and Gibbs 1993).

As is typical in most areas of the tropics, humans likely have been the most common

agents of forest fires in Bolivia. Although most tropical fires are set intentionally by humans for

the purposes of forest conversion, traditional slash-and-bumrn agriculture, or grazing land

management, many of these intentionally set fires escape (Uhl and Buschbacher 1985, Sarre and









Goldhanmmer 1996, Holdsworth and Uhl 1997, Cochrane et al. 1999). Consequently, human-

caused fires presently contribute more to tropical fire regimes than natural fires (Fearnside 1990,

Goldammer 1993. Cochrane and Schultze 1998, Nepstad et al. 1998, 1999, Goldammer 1999).

And, it is likely this was true historically as well, as human population densities in South America

have recovered only in this century to densities present before Europeans arrived (Denevan 1976).

Recent evidence from Bolivia reveals the susceptibility of seasonally dry forests to escaped

human-ignited fires. Over 1 million hectares of Bolivian dry forests burned during a severe dry

season in 1994 (Pinard et al. 1999), and over 3 million hectares burned in one month in 1999 (T.

Fredericksen, personal communication). Evidence also suggests that dry forests are damaged less

by wildfire than moister forest types (Mostacedo et al. 1999), which may also be indicative of the

pervasive role fire has played in the formation of these dry forests.

The Potential of Prescribed Burning in the Management of Bolivian Dry Forests

Although most guidelines for natural forest management focus on ways of reducing

damage to residual stands (Heinrich 1995, Pinard and Putz 1996, Haworth 1999), low-impact

selective logging may not be a sustainable management strategy in dry forests because of the low

levels of disturbance associated with this harvesting technique. In Lomerio for example, roads and

skid trails covered only 2-4% of logged sites and felling gaps were generally only 40-70 nm2ha'

after harvesting operations (Camacho 1996). Likely, this damage does not create sufficient canopy

opening for the regeneration of commercial tree species, 12 of 19 of which were classified as

having shade intolerant regeneration (Table 1-1).

Due to the pervasive influence fire has likely had on the formation and maintenance of

seasonally dry forests in Bolivia, prescribed bums are a promising silvicultural tool for managed

dry forests. Prescribed bums produce several effects that will likely increase regeneration of

shade-intolerant tree species, including vegetation removal, mineral soil exposure, and nutrient

release (Hungerford et al. 1990, Bond and van Wilgen 1996). The use of prescribed burning in









tropical forest management is not a new idea. Ground fires were used as early as the mid-1 800s to

enhance teak (Tectona grandis) regeneration in deciduous forests of South-east Asia (Dawkins and

Philip 1998). Tropical forest managers have recognized the benefits of prescribed bums for

several shade-intolerant timber species in addition to teak, such as sal (Shorea robusta) and several

pine species (Pinus; Goldammer 1994, Rodriquez 1996).

In South America however, the use of prescribed burning to enhance tree regeneration in

broad-leaf forests is rarely practiced. If fire is addressed in forest management policies, it is

primarily in the context of fire prevention or exclusion from protected areas (e.g., Reis 1996, New

Forestry Law in BOLFOR 1997). Particularly in Bolivia, the techniques of prescribed burning are

not well developed and the effects of prescribed bums on dry forest structure and function are not

well known.

Scope of Dissertation

The overall objective of this dissertation is to examine the ecological potential of

prescribed burning for the management of seasonally dry forests in eastern Bolivia. To be a viable

management strategy for the certified forests of Lomerio, prescribed bums must enhance

regeneration of commercial tree species without causing irreversible damage to the residual forest.

The negative effects of prescribed burns are likely to increase with increasing fire intensity

(reviewed in Chapter 2). Therefore, in this dissertation, I compare the effects of harvesting gap

formation, and controlled bums of high and low intensities on commercial tree regeneration, forest

soils, and plant diversity.

The dissertation contains seven chapters. In the second chapter, I briefly review the effects

of fire intensity on plants and soils, introduce the study forest, and describe four treatments that

form the basis of Chapters 3, 4, and 5. The treatments represent the following four disturbance

intensities: harvesting gap formation, above-ground biomass removal, and, controlled bums of low

and high intensity.









In the third chapter, I examine changes in soil physical and chemical properties following

the four treatments. I address the mechanisms underlying these changes by examining

experimentally the separate effects of heat and ash on soil properties. I also discuss how each of

these treatments, through their effects on soil properties, may influence tree seedling growth.

Commercial tree establishment, growth, and survival in each of the four treatments is

evaluated in Chapter 4. As species' responses to disturbance often vary among regeneration

guilds, I discuss the effects of each treatment by illustrating how they affect each guild differently.

I relate these results to the different management strategies that are appropriate for different

species groups.

The effects of silvicultural treatments are primarily aimed at enhancing regeneration of

commercial tree species. Yet the impacts of these treatments on the remaining plant community are

also of concern, particularly in Lomerio where the local indigenous population depends on the

forest for a variety of other uses. In Chapter 5, I examine the response of the plant community to

the four treatments, focusing on changes in the dominance of species, life forms, and regeneration

modes (seedlings or sprouts) among treatments. I discuss these patterns in relation to their

importance for commercial tree regeneration.

The studies presented in Chapters 2 though 5 represent patterns of regeneration over an 18

month period following the treatments. In the sixth chapter, I examine patterns of regeneration

following bums over longer time scales using a chronosequence of secondary forests. I

characterize tree population structures, stand structure, and species richness in abandoned slash-

and-bum fields of 12 different ages, ranging from 1 to 50 years. Comparing these secondary

forests to a nearby mature forest stand, I discuss the possibility that the dominance of shade-

intolerant trees in this region may be the legacy of slash-and-bumrn agriculture.

In the final chapter, I summarize the chapters and discuss the ecological potential of

prescribed bums for the management of Lomerio forests. I discuss how prescribed burning might






10


fit into the current idea of natural forest management in the tropics. I also raise several questions

of the economical and political constraints to implementing controlled bums on a management

scale in Bolivia.
















CHAPTER 2
STUDY SITE AND TREATMENT DESCRIPTIONS


Introduction

Forest disturbances vary widely in their type, intensity, frequency, and scale (Pickett and

White 1985). Despite this variation, disturbances, by definition, hold their most important

character in common: they reduce the dominance of a site by established individuals and create

openings for colonization and growth by new individuals (Canham and Marks 1985). As such,

disturbances are the primary catalyst of forest stand dynamics (Oliver and Larson 1996).

After most forest disturbances, there is a temporary increase the availability of light,

water, and nutrients. There are at least two mechanisms by which forest disturbance may increase

the availability of these resources. The first is the reduction in rates of uptake or use of resources

due to the loss of plant biomass. This effect is most apparent in the enhancement of light levels in

canopy openings (Chazdon and Fletcher 1984) and increased soil moisture in gaps (Vitousek and

Denslow 1987). Disturbances may also increase resource availability indirectly by altering rates

and pathways of nutrient cycling. For example, increased soil moisture and temperatures following

large-scale windthrow may temporarily increase nutrient availability by increasing the rate of

decomposition of soil organic matter (Bormann and Likens 1979).

Fire is an increasingly common disturbance in tropical forests (e.g., Goldamrnmer 1993,

Bond and van Wilgen 1996, Fearnside 1990, Sarre and Goldamnmer 1996). A feature of fire that

may set it apart from other disturbances is its effect below-ground. Fire acts as a rapid

decomposer, returning some nutrients from above-ground biomass to soil more rapidly than other

disturbances (e.g., Humphreys and Craig 1981, Hungerford et al. 1990, Neary etal. 1999). The









usually vigorous growth of seedlings in burned areas is often attributed to fertilization by deposited

ash and increased mineralization due to soil heating (DeBano et al. 1977). Yet, removal of above-

ground biomass can also be far more complete after fires than after other disturbances, such as

canopy gap formation. As such, it is likely a combination of above- and below-ground effects that

make fire a promising management tool for tree species with shade-intolerant regeneration

(Hungerford et al. 1990, Bond and van Wilgen 1996).

The effects of forest fires on above- and below-ground processes may vary widely

depending on their intensity (Moreno and Oechel 1994, Bond and van Wilgen 1996). For example,

low-intensity fires may have a positive effect on regeneration by increasing available soil nutrients

(DeBano et al. 1977, Wright and Bailey 1982), and stimulating flowering (Whelan 1994,

LeMaitre and Brown 1992), resprouting (Zedler et al. 1983, Moreno and Oechel 1994), and

germination of buried seeds (Bradstock and Auld 1995, Schimmel and Granstrom 1996, Enwright

et al. 1997). In contrast, high-intensity fires may be detrimental to regeneration by volatilizing

nutrients (Wright and Bailey 1982), altering soil properties such as texture, cation-exchange

capacity, and water holding capacity (DeBano et al. 1977, Hungerford et al. 1990), killing buried

seeds (Schinmmel and Granstrom 1996), killing species that would otherwise resprout (Moreno and

Oechel 1994), and damaging or killing potential seed trees. Although there is generally a positive

relationship between the size or intensity of disturbance and the subsequent availability of

resources for plant growth (Canham and Marks 1985) this pattern may not apply for fires of

extreme severity. Regeneration on sites of low-intensity fires may be enhanced while areas of high-

intensity fire may be very slow to recover.

I designed an experiment to address the relative importance of canopy opening, above-

ground biomass removal, and controlled bums of high and low intensity on early patterns of

tropical dry forest regeneration. Using these experiments, I examined soil physical and chemical

properties (Chapter 3), establishment and growth of commercial tree seedlings (Chapter 4), and









changes in the plant community (Chapter 5), comparing each response to conditions in undisturbed

forest understories. In this chapter, I introduce the study site and describe the treatments.

Study Site

The studies described in this dissertation were conducted in the seasonally dry forests of

Chiquitania, a geographic region in the eastern lowlands of Bolivia located in the Province of Nuflo

de Chavez, Department of Santa Cruz (1645'S, 6145'W; Figure 2-1). Chiquitania is situated in

a transition zone between the humid forests on the southern rim of the Amazon basin and the thorn

scrub formations of the Gran Chaco. The natural vegetation is classified as tropical dry forest

(sensu Holdridge et al. 1967).

The regional climate is characterized by pronounced seasonality with a strong dry season

that corresponds to the austral winter (Figure 2-2). Most of the canopy trees are seasonally

deciduous, shedding their leaves from June to September. The mean annual temperature at

Concepcion is 24.3C with temperatures that vary between 3 (July) and 38. 1C (October, Killeen

et al. 1990). The mean annual precipitation is 1129 mm and interannual variability is large, with

lows having reached 500 mm and highs 1717 mm per year (Killeen et al. 1990). The landscape is

dominated by low hills composed of granite, gneiss, and metamorphic rocks of Precambrian origin

(Geobold 1981 in Killeen et al. 1990) punctuated by exposed granitic outcrops (inselbergs). The

soils of the area are classified as Inceptisols (suborder: Tropepts, group: Ustropepts) and Oxisols

(suborder:Ustox, group: Eutrusox; Ippore 1996). Elevation varies between 400 and 600 m a.s.l.

Canopies of mature forest range from 12-18 m tall and are dominated by trees of the

Leguminosae (60% of total basal area of trees >10 cm dbh); trees in the families Bignoniaceae,

Anacardiaceae, and Bombacaceae are also abundant (Killeen et al. 1998). Understory trees are

mostly represented by the families Sapindaceae and Myrtaceae. A spiny ground bromeliad,

Pseudananas sagenarius, is distributed over approximately 80% of the forest and occurs in

clumps up to 2000 m2 (MacDonald et al. 1998).







































S Lomerio














Figure 2-1. Location of the study site in the seasonally dry forests ofChiquitania, a geographic
region in the eastern lowlands of Bolivia located in the Province of Nuflo de Chavez, Department
of Santa Cruz (16"45'S, 6145'W). In the enlarged area, points mark the 28 communities in the
political region of Lomerio.















- temperature
E-s precipitation


500




400


E
300 E
0
.S

200 "I



100



0


'93 '94 '95 '96


Figure 2-2. Mean monthly temperature and monthly precipitation at Concepcion,
Santa Cruz (167' S, 6202' W, 490 m a.s.l.), located approximately 100 km from
Lomerio.









Site background

Chiquitania is so named for the Chiquitano indigenous people, the largest of the lowland

indigenous groups in Bolivia, with a population of around 72,500. Lomerio, where this study was

conducted, is a political region within Chiquitania made up of 27 Chiquitano communities with a

total population of around 5,000. The Chiquitanos of Lomerio have been managing their forests

for timber since 1982 with technical and financial support from several international institutions.

BOLFOR (Proyecto de Manejo Forestal Sostenible), a sustainable forest management project with

USAID funding, began working in Lomerio in 1992. The objective of the current management

plan for the forests of Lomerio is to produce timber sustainably while minimizing negative impacts

on other biological and physical resources in the forest (Pinard et al. 1999). Forestry operations

of the Chiquitano communities were certified as sustainable by the SmartWood Program of

Rainforest Alliance in 1995.

The particular forest in which I worked is owned by the Chiquitano community of Las

Trancas and situated approximately 12 km northeast of this village. The Las Trancas forest

contains 400 ha management blocks, Las Trancas '94 and Las Trancas '95, so named for the year

in which forest inventories were conducted (Figure 2-3). Las Trancas '94 was selectively logged in

July-September of 1995. On average, 3-10 m3 ha' (2-5 trees ha1) of timber were extracted from 6

species. Damage to the residual stand was slight, with 6% of the residual trees damaged and 2-4%

of the area covered by roads and extraction routes (Camacho 1996). Two logging methods were

employed in Las Trancas "95. In 1996, approximately 75% of the area was selectively logged. In

1997, the remaining 25% of the area was selectively harvested in strips (40 m x 200 m, each

separated by an unharvested area 60 m wide). All commercially valuable trees were harvested

from these strips at a harvesting intensity of 4.4 m3 ha1. Log extraction routes (skid trails) entered

each logging strip 100 m from the north and 100 m from the south.






17







Las Trancas '94













Las Trancas '95














/ -scale: 1 km









scale: 1 km


Figure 2-3. The '94 and '95 management blocks of the Las Trancas community owned forest.
The enlarged section of Las Trancas '95 represents the trial area for the strip shelter wood system.
The squares within this enlarged area mark the 16 blocks (20 x 20 m gap areas) depicted in Figure
2-4.









Methods

Location of felling gaps

The studies presented in Chapter 3, 4, and 5 were conducted in the selectively harvested

strips in Las Trancas '95. These strips were the only area of logging activity during the dry season

of 1997 and therefore all newly created felling gaps were located in these strips. In June of 1997, I

located 16 felling gaps for study (Figure 2-3). Gap selection was restricted by the following

criteria: canopy gap area between 200-600 m2, slopes no greater than 15, less than 20% rock

outcrops, no trees > 40 cm DBH within gap area, and not located in the path of skid trails.

I located and marked the center of each gap where the midpoints of two perpendicular

transects intersected, the first running the length of the longest axis. Each gap was divided into four

10 x 10 min plots by cardinal axes from the center point (Figure 2-4). Half-meter wide paths around

the perimeter of the gap and along axes were cut by machete. Existing gap area was enlarged to a

uniform 20 m x 20 m area by cutting all vegetation >2 min tall (sensu Brokaw 1985a) by machete or

chainsaw. Because this forest is a timber management area, commercial tree species > 20 cm

DBH located within the 20 m x 20 m gap area were left uncut (this occurred in only 4 of 16 gaps

and standing trees did not exceed 25 cm DBH).

Gap Treatments

One of four treatments was randomly assigned to each 10 x 10 m plot within each block:

1) high-intensity bum; 2) low-intensity bum; 3) plant and coarse debris removal (hereafter referred

to as plant removal); and, 4) canopy gap with vegetation > 2 min tall cut (the gap control). Other

than cutting all vegetation > 2 min tall, vegetation and woody debris in the gap control was not

manipulated. In the plant removal and low-intensity burn treatments, all vegetation was cut at or

near the soil surface and everything >2.5 cm diameter (>: 100 hour fuels) was removed and

distributed as evenly as possible in the high-intensity bum treatment. Tree trunks and large

diameter branches were sawn into smaller sections so that they could be moved more easily and dry











I I I


forest plot


forest subplots


- 15-20 m


Streatmnent
plot








y subplots


I^ __^


10mon


20 m
Figure 2-4. A single block consisting of a 20 x 20 m felling gap and an adjacent forest plot. Each
felling gap was equally partitioned into four 10 x 10 m treatment plots. Each treatment plot was
randomly assigned one of four treatments: gap control, plant removal, low intensity or high
intensity bum. Within each treatment plot, 2 permanent vegetation sampling subplots (2 x 2 each)
were located near the gap center and 2 additional subplots near the gap edge. All soil sampling
was conducted outside of these permanent sampling plots. Two permanent subplots were located
approximately 15-20 m from the edge of each gap in undisturbed forest. Soil sampling in the
forest was conducted outside but within 5 m of these permanent forest subplots. Hereafter, the
400 m2 felling gap and adjacent forest site are "blocks," the 100 m2 treatment areas are
"plots," and the 4 m2 vegetation sampling areas are "subplots."


20 m









rapidly. Pseudananas sagenarius and cacti were not added to the high-intensity bum treatment,

but instead were removed from the block altogether because of their low flammability. Therefore,

after fuels were manipulated and before prescribed bums, the plant removal and low-intensity bum

treatments had similar amounts of litter and woody debris and no above-ground vegetation. The

high-intensity bum treatment plots had roughly 3 times its original fuel load. Slash was left for 5

rainless weeks to dry before prescribed bums were conducted.

Controlled burns

Fuel loads. Pre-bum fuel loads in the low- and high-intensity bum treatments were

measured in randomly located 0.25 m2 circular subplots, with 2 subplots sampled per plot x

intensity treatment (2 fuel plots x 2 bum treatments x 16 plots). All fuel within each subplot was

removed, divided into fuel diameter size classes (live herbs, <6 mm, 6-25 mm, 25-75 mm, and >75

mm) and weighed in the field. Composite subsamples of each fuel size class were taken from the

field, oven dried to constant weight, and used to calculate the wet-to-dry weight conversion factors.

The diameter and length of trunks and large diameter branches that could not be weighed in the

field were measured in order to estimate volumes. Wood densities available from BOLFOR were

used for volume to mass conversions.

Timing of burns. Although little is known about the historic fire regime in these forests,

seasonal patterns in rainfall and relative humidity make wildfires most likely at the end of the dry

season when fuels are dry and lightning strikes most common. The indigenous Chiquitano

population traditionally bum their agricultural fields and cattle pastures at the end of the dry

season as well, shortly before the onset of rains. Predictably, most escaped fires occur during this

season. Because one of the objectives of my experimental bums was to enhance seedling

establishment of commercial tree species, I planned a burning date in late August or early

September, at the end of the dry season and before peak seed fall of most commercial trees

(Justiniano 1997).









Fire breaks. All fuel was removed from a 1 m wide fire break around low-intensity bum

treatment plots. Fire breaks around high-intensity bum treatment plots were 1 to 2 m wide, wider

where danger of fire escape was perceived to be higher. Standing dead trees near firebreaks were

felled and ladder fuels such as liana tangles were removed. On the day of burns, firebreaks were

raked free of newly fallen leaves.

Prescribed burns. Prescribed bums were conducted from August 29 to September 1,

1997, near the end of the 5 month dry season (Table 2-1). Each day, the earliest bums were

started at 10:00 a.m. and the last bums by 3:00 p.m. Temperature at 10:00 a.m. over the four day

period varied from 34 to 36.4 C and relative humidity varied from 29-38 %. Winds were variable,

but usually calm in the morning with convectional wind gusts of up to 11 km/hr in the afternoon.

A circular ignition technique was used for both bum treatments. A spot fire was lit with a

drip torch in the plot center, then the perimeter was lit starting with the downwind side. The center

fires created convection which drew the ring fire on the borders inward. In the low-intensity bum

treatment plots, ring fires often did not carry to the center, therefore spot fires were ignited where

needed.

A minimum of 5 people conducted the bums over the 4 day period. At least one person

with a backpack water sprayer remained at each fire until fires near the borders were extinguished.

Fires near firebreaks or standing dead trees were extinguished before burning crews returned to

camp. Each fire was checked again after dark and the following morning to extinguish any

potentially dangerous smoldering areas.

Maximum soil temperatures, fire intensities and completeness of burns. Maximum soil

temperature and an index of fire intensity were measured in two locations in each bum plot, near

the two subplots where fuel loads were measured. Maximum soil temperature was measured at 0

and 3 cm depth using temperature indicating paints (Tempilaq Tempil Division. Air Liquide


























Table 2-1. Climatic conditions at 10:00 a.m. the morning of high and low intensity
bums for 16 experimental blocks.
Block Date Ambient Relative Wind speed
burned temperature (C) humidity (%) (km/hr)
1 29-Aug 36 29 0 with gusts
19 29-Aug 36 29 0 with gusts
20 29-Aug 36 29 0 with gusts
4 30-Aug 34 34 11
6 30-Aug 34 34 11
8 30-Aug 34 34 11
21 30-Aug 34 34 11
22 30-Aug 34 34 11
2 31-Aug 34 37 4
7 31-Aug 34 37 4
9 31-Aug 34 37 4
11 01-Sep 35 38 0 with gusts
14 01-Sep 35 38 0 with gusts
17 01-Sep 35 38 0 with gusts
18 01-Sep 35 38 0 with gusts









America Corporation, South Plainfield, New Jersey, USA). Paints of 24 different melting points

ranging from 66 to 1093 C were applied as narrow bands on 2 x 30 cm steel strips. At each

location, one painted steel strip was buried at 3 cm soil depth and another placed flat on the soil

surface directly above it. Soil temperatures were measured to a greater depth in one block. Here,

an additional 3 sets of 4 painted strips were placed at 0, 1, 3, and 7 cm depths. After fires, the

highest indicated melting point was recorded.

Fire intensity was estimated by Beaufait's (1966) technique which calculates total energy

output from the amount of water vaporized from cans during burns as:

total energy output = [(80 cal/g water) x (g water)] + [(540 cal/g water) x (g water)]

Where 80 cal arc needed to raise each gram of water from 20" C to the boiling point and 540 cal

are needed vaporize each gram of water (latent heat of vaporization). Two tin cans per bum were

used, each placed on the soil surface of fuel load subplots. Depth of water was measured

immediately before each bum and within 24 hours after. To account for the amount of water lost

due to evaporation, 2 cans were placed in the center of an unburned gap and the amount of water

evaporated within 24-hrs measured.

Soil moisture, which influences heat movement through soil, was measured several hours

before burns. Soil samples from 0-5 and 5-10 cm depths were collected from each plot, weighed,

oven dried to a constant weight, and moisture content expressed as % of soil dry weight.

The week following bums, completeness-of-bum was estimated visually as the percent

area burned

Establishment of permanent vegetation plots

Three weeks following bums, 4 permanent subplots (2 x 2 m each) were established in

each treatment plot, 2 located near the gap center and 2 located near the gap edge (Figure 2-4).

Two additional subplots were established at random points in undisturbed forest 15-20 m from the

edge of each gap. These permanent subplots were used for sampling commercial tree seedling









establishment (Chapter 4) and vegetative cover (Chapter 5). One plot of each pair was used for a

seeding treatment described in Chapter 4.

Treatment effects on canopy cover and microhabitat

Soil temperature to 3 cm depth was measured at a center and edge subplot of each

treatment as well as the forest subplots with a soil thermometer 3 and 6 months following bums.

Percent canopy cover was measured with a spherical densiometer above each gap center, gap edge,

and forest plot 3 months following bums. Litter depth (cm) and percent cover by debris 2-20 cm

and >20 cm diameter were estimated visually for each of the permanent 4 m2 subplots 6 weeks

following bums. Results were analyzed using an analysis of variance, with treatment as a fixed

effect and block as a random effect, followed by Tukey's post-hoc comparisons.

Initial Treatment Results

Pre-burn fuel loads

Pre-burn fuel loads in high-intensity bum treatment subplots ranged from 10.8 to 82.8

kg/m2 and averaged 48 4.9 kg/m2 (mean 1 standard error; Figure 2-5). Almost half of this mass

was comprised of fuels >7.5 cm diameter. Fuel loads in the low-intensity burn treatment subplots

ranged from 0.8 to 4 kg/m2 and averaged 2.2 2.3 kg/m2. Sixty-six percent of the fuel mass in

low-intensity plots was fine fuel, <6 mm diameter.

Burn characteristics

High-intensity burns. Completeness of high-intensity bums was variable, but the

majority of bums consumed all but the thickest (> 20 cm diameter) branches and trunks. Flame

heights ranged from 1.5 to 5 m. Fire intensities ranged from 152 to 3795 kcal and averaged 1627

241 kcal (n = 15). Temperature at the soil surface during high-intensity burns averaged 704 42

C (n = 16). The highest temperature measured was 927C. Temperature at 3 cm depth averaged

227 27 C (n = 16). Where maximum temperature was measured at additional depths of 1 and 7























4000 Fire intensity (kcal) 1000
400 000--- JA -- io


T




HI

i

High Low


Temperature (C)
soil surface 3 cm depth


High Low

High Low


High Low


Figure 2-5. Pre-bum fuel loads, fire intensities, and maximum temperatures at the soil surface
and 3 cm depth during high and low intensity bums. Box plots show medians (center line),
25th and 75th percentiles (top and bottom lines), 10th and 90th percentiles (top and bottom
whiskers), and points greater than the 90th percentile and less than the 19th percentile (dots).


High Low









cm, temperatures averaged 871 C at the soil surface, 358C at 1 cm depth, 218 C at 3 cm depth,

and 135 C at 7 cm depth (n = 2). Although visible flames were extinguished by nightfall, some

logs continued to smolder for several days: fire intensities and soil temperatures under these logs

were likely greater than measured values.

Low-intensity burns. In general, completeness of low-intensity bums was more variable

than high-intensity bums. Flame heights were low, ranging from 10 to 50 cm. Fire intensity

ranged from 22 to 68 kcal and averaged 41 3 kcal (n = 15). Temperatures at the soil surface

averaged 225 33 C (n = 12); the highest temperature measured was 413 C. Elevated

temperatures at 3 cm were only detected in 2 of 16 plots; these averaged 107 7 C (n = 2). Soil

moisture on the day of bums was low and did not differ between the high- and low-intensity bum

plots (0-5 cm depth: P = 0.94, 5-10 cm depth: P = 0.23). Therefore, differences among the 2 bum

treatments in heat conductivity due to soil moisture were likely negligible and are hereafter ignored.

Treatment effects on microhabitat

Treatments had significant effects on the amount of soil exposed, mid-day ambient soil

temperature, litter depth, and area covered by woody debris (Figure 2-6). Canopy cover above

forest plots was 78%, higher than canopy cover above all 4 gap treatments, which averaged 22%

(P < 0.001). Although canopy cover above gap-center and gap-edge plots was not significantly

different (P = 0.6), soil temperatures in gap centers were higher than near gap edges (P < 0.001).

A maximum temperature of 43 C was recorded in the center of one high-intensity bum treatment 3

months following bums. After 6 months, soil temperatures at gap centers and edges were not

different (P = 0.52) and only soil temperatures in the high-intensity bum treatment were

significantly higher than the other treatments (P < 0.001).

High-intensity bums removed all litter and deposited a layer of ash ranging from 0-14 cm

depth (4.8 0.2 cm, n = 16). Not all woody material was consumed in the high-intensity bums;













Litter depth d
P < 0.001
d


a 7777
ab


Debris 2-20 cm
PK< 0.001 C




~aa


Soil temperature
P< 0.001


High Low Plant
intensity intensity removal


Gap
control


a



Forest


Figure 2-6. Litter depth, percent cover of debris 2-20 cm and > 20 cm diameter,
and mid-day soil temperature in four gap treatment plots and forest plots 6 weeks
after bums. Treatments with the same letter are not significantly different.


0.25


0.00
0.25


0.20 -
0.15 -

0.10 -

0.05 -

0.00 -
45


Debris > 20 cm
P< 0.001


b


a


a a









the remaining woody debris covered approximately 12% of the subplots areas. In the low-intensity

bum treatment, an average of 76% of the subplot areas burned to some degree. Burning was not

complete even within these areas as only an average of 30% of the subplot areas had soil exposed.

Small woody debris remained on approximately 3% of the area of low-intensity bum subplots;

most large woody debris had been removed before burning.

Leaf litter or small woody debris covered all of the plant removal treatment subplots, with

no bare soil exposed. As with the low-intensity bum treatment, most large woody debris was

removed. Gap controls were characterized by deep leaf litter (2.9 0.2 cm, n = 16) and small and

large woody debris covering an average of 25% of the subplot areas. Only 20% of the gap control

subplot areas were devoid of either woody material or surviving plants. Forest understories had

the deepest leaf litter (4.0 0.3 cm, n = 16), but small and large woody debris combined covered

an average of only 7% of the subplot areas.
















CHAPTER 3
EFFECTS OF CANOPY GAP FORMATION, PLANT REMOVAL,
AND CONTROLLED BURNS OF HIGH AND LOW INTENSITIES ON
SOIL CHEMICAL AND PHYSICAL PROPERTIES


Introduction

Fire is a rapid decomposer; it compresses the oxidative processes of organic matter decay

into a short time span (Wright and Bailey 1982). The result is a nutrient pulse much larger than

from the normal decomposition of woody debris and litter, at least for the first few months

following fires (Bond and van Wilgen 1996). As such, controlled burns may benefit tree seedling

growth more than unburned treatments, particularly since the timing of nutrient pulses following

fire coincides with maximum light availability. After intense fires, however, the advantages of

increased nutrient availability may be offset by degraded soil structure. Thus, the benefit of

controlled bums for tree seedling growth may ultimately depend on fire intensity. In this chapter, I

examine both soil nutrient availability and soil physical properties following canopy opening, plant

removal, and controlled burns of high- and low-intensity.

There are three primary mechanisms of increased nutrient availability following fire:

nutrients added to the soil as ash; heating of soil organic matter; and, increased rates of biological

mineralization following fire due to increases in soil pH, temperature, and moisture, as well as due

to a reduction in C:N ratios (Wright and Bailey 1982, Pritchett and Fisher 1987). The degree of

increase in nutrient availability following fires depends largely on fire intensity. Most studies of

low to moderately intense fires report increases in available nutrients (reviews by Dunn et aL. 1977,

DeBano et al. 1977, Wells et aL. 1979, Humphreys and Craig 1981, Wright and Bailey 1982,









Hungerford et al 1990, Neary et al. 1999). In contrast, intense fires may cause a net loss of

nutrients (DeBano et al. 1977, Giovannini et al. 1990).

Due to its low temperature of volatilization (200 C; Weast 1988), nitrogen loss is linked

with the consumption of organic matter (e.g., Dunn et al. 1977). Where fuels are completely

consumed and the surface layer of soil organic matter is destroyed, loss of nitrogen through

volatilization can be substantial (e.g., Nye and Greenland 1964, Ewel et al. 1981). Volatilization

of phosphorus and cations are usually minor due to the high volatilization temperatures of these

minerals (>760 C; Weast 1988), however, their loss from severely burned sites may be caused by

surface erosion, leaching, or transport of ash (Wright and Bailey 1982).

Intense burns may also have detrimental effects on soil physical properties by consuming

soil organic matter. Soil organic matter holds sand, silt, and clay particles into aggregates,

therefore a loss of soil organic matter results in a loss of soil structure. Severe fires may also

permanently alter soil texture by fusing clay particles into sand-sized particles (Dymess and

Youngberg 1957, Ulery and Graham 1993). By altering soil structure and texture, severe fires can

increase soil bulk density (DeByle 1981), and reduce soil porosity, water infiltration rates, and

water holding capacity (e.g., Wells et al. 1979). Intense bums may also induce the formation of a

water repellent soil layer by forcing hydrophobic substances in litter downward through the soil

profile (DeBano 1969), reducing water infiltration rates as a consequence (DeBano 1971).

The changes in chemical and physical soil properties caused by fire potentially have

important consequences on tree seedling growth (Johnson 1919). Increased nutrient availability

after fire may benefit plant growth if nutrients are limiting prior to burning (e.g., Hungerford et al.

1990). On the other hand, seedling growth in intensely burned soils may be slowed due to high pH

and toxic levels of minerals (Giovannini et al. 1990). Altered soil physical properties, such as soil

strength, bulk density, and water infiltration rates, may also impair plant growth. Plant uptake of

nutrients and water is slowed in structurally degraded soils through the combined effects of lower









soil moisture and lower soil porosity (Nye and Tinker 1977). Mechanical impedance of root

growth caused by increased bulk density and soil strength (Gerard et al. 1982) also slows nutrient

and water uptake.

In this chapter, I focus on the below-ground effects of the treatments described in Chapter

2. My objectives were to: 1) compare the effects of canopy gap formation, plant removal, and

controlled bums of high and low intensities on soil nutrient availability, soil physical properties,

and fine root mass; 2) compare the relative importance of soil heating and ash-fertilization on soil

nutrient availability; and, 3) discuss how these treatment-induced changes in soil properties

influence tree seedling growth.

Methods

Study site

The studies presented in this chapter were conducted in the treatment and forest plots

described in Chapter 2 (Figure 2-4). All soil sampling was done within the 100 m2 plots but

outside of the 4 m2 subplots. Forest sampling was done within 5 m of the forest subplots.

Mass and chemical characteristics of ash deposited in high-intensity burn plots

Ash mass deposited in high-intensity bum plots was estimated by collecting and weighing

all ash on the soil surface in a 1 m2 area, replicated in three high-intensity bum plots (n = 3). To

characterize variability in the amount, ash depth was measured in 10 randomly located points in

each high-intensity bum plot (n = 16). Composite ash samples were then collected from each plot

and used to measure pH and nutrient concentrations. Ash pH was determined as for soil pH,

described below. To determine nutrient concentrations, 0.5 g of ash was heated in 10 mL of 1 M

HNO3 and then resolubilized in 10 mL of 1 M HC1. Extracts were then analyzed for phosphorus.

potassium, calcium, and magnesium at the Analytical Research Laboratory at the University of

Florida with an ICAP Spectrometer (Thermo-Jarrell Ash Corporation, Franklin, MA).









Soil sampling

Soil samples from 0-8 cm depth were collected 2, 6, 9, 12 and 18 months after bums and

from 8-20 cm depth after 9, 12, and 18 months. These samples were used to assess moisture

contents, pH, organic matter, and extractable elements. In each treatment and forest plot (n = 16

blocks), 4 samples were taken from randomly selected sites with a 10 cm diameter cylindrical

corer. The 4 samples from each treatment were mixed thoroughly in the field and a -300 g

composite subsample bagged (Anderson and Ingram 1993). In one block, the 3-month samples

were bagged separately, rather than composite, to examine intra-treatment variability. Subsample

soil volume was unknown, therefore bulk density and fine root mass were sampled separately as

described below.

Soil pH and air-dry moisture content

The pH of fresh soil was determined by adding 50 ml of distilled water to 20 g of soil and

stirring for 10 minutes (Anderson and Ingram 1993). The mixture then stood for 30 minutes and

pH of the supernatant was measured with a hand-held meter (Oakton pHTestr 3). Soil samples

were then weighed, air-dried to a constant weight, and reweighed to calculate air-dry moisture

content. Air-dried samples were passed through a 2 mm sieve, bagged, and stored in a cool dry

area until transported back to the Analytical Research Lab at the University of Florida for

chemical analyses.

Soil chemical analyses

Phosphorus, potassium, calcium, and magnesium were extracted with Mehlich-I solution:

0.05 M HCI and 0.0125 M H2SO4 (Hanlon et al. 1994). Extracts were then analyzed by ICAP

spectroscopy. Soil organic matter content was analyzed using the Walkley-Black dichromate

methodology (Hanlon et al. 1994). A subset of soil samples was tested for total nitrogen using an

elemental analyzer (Carloerba NCS 2500). Twenty-four samples from all treatments and sampling

periods from the top 8 cm of soil were selected to represent the full range of organic matter content.









Resin-available nitrogen and phosphorus

Resin-available nitrogen (NI--N and NO3"-N) and phosphorus (PO4 3-P) in each

treatment were estimated by burying anion and cation exchange resin bags at 5 cm depth. Resin

bags were prepared by enclosing 5.0 g (moist weight) of either anion exchange resins (Sigma-

Dowex) or cation exchange resins (Fisher Scientific) in bags of nylon stocking material sewn

closed with nylon thread. Before burial, resin bags were hydrated overnight with dionized water.

Four bags of each resin type were buried per treatment plot (4 bags x 2 resin types x 5 treatments x

16 blocks). Three rotations of resins were buried, each for approximately 3 months. Two

rotations included the first and second rainy seasons following bums (November 1997-January

1998 and December 1998-February 1999, respectively). The middle rotation covered the transition

from the first rainy season to the first dry season following burns (May-July 1998). After removal

from the field, resin bags were placed separately in clean plastic bags and kept cool (refrigerated

when possible) until transported to the University of Florida for analysis. For each resin type, the

4 bags per plot were pooled and 12 g extracted in 120 ml of 2 M KCI for 24 hours. Extractions

were analyzed for amonium-N and nitrate-N using automated spectrophotometry (Flow IV Ion

Analyzer, AlpKem (0-I-Analytical), College Station, TX). Extracts from anion exchange resins

were diluted to 1M KCI and analyzed for PO4 3-P using the atomic emission spectrometric method

(Thermo-Jarrell Ash Corp. Franklin, MA).

Statistical analyses

Soil nutrient concentrations, organic matter, pH, moisture content, and resin-available N

and P were analyzed using an ANOVA with repeated measures. Treatment was a fixed effect and

block a random effect in each model. Soil properties were log transformed for analyses when not

normally distributed, but all values presented in the text are non-transformed. Where a significant

time x treatment interaction was found, variables were analyzed separately by month. Statistically

significant differences (P < 0.05) were further analyzed with Tukey's HSD multiple comparisons.









In order to describe variation within plots, 4 soil samples per treatment in one block (block

4) were analyzed separately for extractable nutrients and organic matter. Also, 4 resin bags per

treatment in one block were extracted and analyzed separately for resin-available nitrogen and

phosphorus. Coefficients of variation (Sokal and Rohlf 1981) were calculated to compare variation

of soil sampled within the same 100 m2 plot and among the 16 different plots.

Fine root mass

Fine root mass (roots < 2 mm diameter) was compared among treatments 12 months

following bums in a reduced sample of 10 blocks (n = 10). Soil cores were extracted with a

cylindrical tube (5 cm inside diameter, 7 cm deep) from 3 randomly located points in each

treatment and forest plot. Fine roots were sorted from samples, dried, and weighed. Fine root

mass (live and dead combined) was compared among treatments using an ANOVA followed by a

Tukey's HSD post-hoc test.

Soil bulk density

Soil bulk density (air-dry) was estimated 6 and 12 months following bums in a reduced

sample of 10 blocks (n = 10). Three samples in each treatment and forest plot were collected using

metal cans (5 cm inside diameter, 7 cm deep, 137 cm3). Samples were air-dried to constant weight

and bulk density calculated as:

bulk density (g/cm3) = g air-dried soil / 137 cm3

Differences among treatments were tested using an ANOVA on square-transformed values of bulk

density with treatment and month as fixed effects and blocks as a random effect, followed by

Tukey's HSD multiple comparisons.

Soil strength

Compressive soil strength was estimated with a pocket penetrometer (Forestry Suppliers)

at 2, 6, 9, and 12 months following bums. Soil strength readings were taken at 4 randomly

selected points in each treatment and forest plot from all 16 blocks (n = 16). Soil strength was









analyzed using ANOVA with repeated measures as described above for soil chemical properties.

Water infiltration

Water infiltration rates were estimated in a reduced sample of 4 blocks 8 months following

bums (n = 4). The technique used here was a modified version of the single ring method (Anderson

and Ingrain 1993). Although double ring methods provide better estimates of infiltration rates

because they compensate for lateral flow, a single ring method was chosen for this study because it

used less water (which had to be transported 27 km). In each gap treatment and forest site, a point

was randomly located and cleared of surface litter. A graduated PCV cylinder (10 cm diameter, 25

cm length) was inserted vertically into the soil 10 cm deep and soil pressed around the base to

minimize water leakage. The cylinder was filled with water to 10 cm and timed until the water

level dropped to 5 cm. This process was repeated three times. Infiltration rates were calculated

separately for each repetition (i.e., the first, second, and third 5 cm increments of water which

correspond to 5, 10, and 15 ml cm-2) as the volume flux of water flowing into the soil profile per

unit surface area (Hillel 1982) and expressed as ml cm-2 sec'. Log transformed infiltration rates

were compared among treatments using an ANOVA with treatments and repetitions (i.e., each 5

cm increment) as fixed effects and blocks as random effects.

Soil wettability

Soil wettability was estimated using a modification of the water drop penetration time

method (WDPT; Letey 1969) in a reduced sample of 7 blocks 8 months following burns (n = 7).

In each gap treatment and forest site, four 20 x 20 cm areas were randomly located and cleared of

surface litter. Five drops of water were placed on the soil surface with a dropper and the time

recorded when all 5 drops were completely absorbed. This was repeated at 1, 2, and 3 cm soil

depth by scraping surface soil away with a machete. Soil wettability (log transformed seconds)

was compared among treatments using an ANOVA with treatments and soil depth as fixed effects

and blocks as random effects.









Comparative effects of soil heating and ash addition on soil chemical properties

To compare the effects of soil heating and ash addition on soil chemical properties, I

carried out a 2 x 3 factorial experiment with two levels of ash addition (no ash and ash added) with

three levels of soil heating (no heat, low-intensity heat, and high-intensity heat). The first trial was

conducted in the field in Las Trancas '95. The second trial was conducted in the BOLFOR

greenhouse in the city of Santa Cruz. Anadenanthera colubrina served as a bioassay in the field

experiment. Because Anadenanthera did not fruit in 1998, Caesalpinia pluviosa was used as a

bioassay in the greenhouse study.

Field study. I utilized the plant removal treatment plots described in Chapter 2 for the field

study, conducted at the end of the dry season in October 1997. The design is a complete

randomized block; each plant removal plot was considered as a block (n = 12). In each block, six

1 m2 plots were located in the area between the gap center and edge permanent subplots (Figure 2-

4). Competing vegetation had been cleared from the larger treatment plots the month before, but

some regrowth had already occurred. Therefore the 1 m2 plots were cleared again of any

vegetation and raked of surface litter to expose the soil surface. Each plot was randomly assigned

a treatment combination of soil heating (no heat, low-intensity heat, or high-intensity-heat) and ash

(no ash or ash added). Treatments were applied to a 50 x 50 cm area in the center of the 1 m2 plots,

creating a 25 cm buffer along the edge. Heat was applied using a propane blow torch.

Temperature of the flame was measured with Tempil heat sensitive paints. In the low-intensity

heat treatment, a flame of 150-250 C was applied to the soil surface of the treatment area for 5

minutes. In the high-intensity heat treatment, a flame of 500-800 C was applied to the soil

surface of the treatment area for 20 minutes. The torch required constant adjustment to maintain a

similar flame, therefore temperatures varied within a treatment. I am confident, however, that

temperatures ranges did not overlap between the high- and low-intensity treatments. After heat









treatments, approximately 500 g of ash collected from high-intensity bum plots was distributed as

evenly as possible to plots assigned the ash treatment. The week following treatments, 10 seeds of

Anadenanthera colubrina were placed in each plot and checked for germination after 4 days.

After 2 weeks, most seeds were found to have been removed or eaten and were therefore not

checked again.

Soil samples 0-8 cm depth were collected 3 weeks following treatments. Soil pH, and

phosphorus, potassium, calcium, magnesium, and organic matter concentrations were analyzed

using the methods described above. Resin-available nitrogen (NH4-N and N03-N) was measured

using anion and cation exchange resin bags. One bag of each type was buried at 5 cm depth in

each plot for 85 days (October 31 January 24). Analysis of resins follows that described for the

larger study.

Greenhouse study. The greenhouse study was conducted at the end of the dry season in

1998. Soil used for this trial was collected from Las Trancas '95 to a depth of 10 cm. Soil was

passed through an 8 mm sieve, mixed well, and divided into three equal batches. Each batch was

assigned a soil heating treatment (no heat, low-intensity heat, and high-intensity heat). In the low-

intensity heat treatment, soil was heated in aluminum pots in a conventional oven at 100-150 C for

a total of 10 minutes (mixing after 5 minutes). Soil in the high-intensity treatment was oven heated

at -200 C for 40 minutes (mixing after 20 minutes) then spread 1 cm deep on a metal tray and

heated with a blow torch for 5 minutes at a temperature of 500-800 C. Oven temperature and

torch temperature were both measured using Tempil heat sensitive paints. One composite soil

sample from each heating treatment (control, low-intensity, and high-intensity) was analyzed for

phosphorus, potassium, calcium, magnesium, and organic matter concentrations. Soil from each

heating treatment was used to fill 24 plastic planting containers (7 x 25 cm). Twelve planting

containers per heating treatment were then selected for the ash addition treatment (15 g of ash









added to the soil surface) and the remaining 12 containers served as controls (n = 12). Two seeds

of Caesalpinia pluviosa were placed in each planting container, watered daily, and seedling height

to the terminal bud was measured after 4 months. Resin-available nitrogen (NH4+-N and NO03-N)

was measured in each treatment combination using additional planting containers. Anion and

cation exchange resin bags were buried at 5 cm depth in 3 containers of each treatment

combination (n = 3) and watered daily for 22 days. Resins were extracted and analyzed using the

methods described above.

Results

Mass and chemical characteristics of ash deposited in high-intensity burn plots

Variability in amount of deposited ash was high and depths ranged from 0-14 cm (4.8

0.3 cm, x S.E., n = 16). Ash mass deposited by high-intensity bums averaged 1.5 0.6 kg/m2

(n = 3). Using this value and measured concentrations of individual elements in ash (Figure 3-1)

indicates an average nutrient deposition of 524 g/m2 of Ca, 26 g/m2 of Mg, 83 g/m2 of K, and 7.7

g/m2 of P. Ash samples had an average pH of 10.7 0.1 (n = 16).

Treatment effects on soil nutrients

High-intensity burns significantly increased P, Mg, K, and Ca in the top 8 cm of soil, but

the magnitude and its change over time varied by nutrient (Figure 3-2). These increases were also

detected at 8-20 cm for all elements except Mg (Figure 3-3). Low-intensity burns also

significantly increased P, Mg, K, and Ca in the top 8 cm of soil, although increases were smaller

than in high-intensity bum plots, did not persist as long, and were not detected at 8-20 cm. Plant

removal and gap control treatments had no detectable effect on P, Mg, K, and Ca at either soil

depth. Results of statistical analyses are summarized in Tables 3-1 and 3-2.





















Mg
(mg/g ash) 100


*


Figure 3-1. Box plot diagrams of concentrations of Ca, Mg, K, and P in ash sampled
from high intensity burn plots (n = 9). Box plots show medians (center line), 25th and
75th percentiles (top and bottom lines), 10th and 90th percentiles (top and bottom
whiskers), and observations lying outside of the 10th to 90th percentiles (dots).


500 (mg/gash)


400 -


300 -


200 -


100 -


0 1











100 -

80

60

40 S ---

20 -- -- -
0--- --------------^~ i ~~--- -i --
I I t I I-- -= ~ :~
0
500

400

300

200 -]----------..-..--- --.....


100 -i ,
800 0-- High intensity burn
--- Low intensity burn
600 --- Plant removal
-,- Gap control
400 Forest

200 -

0II I--
600O


4000 -



2000 -


9 12
Months after treatment


Figure 3-2. Extractable soil concentrations of P, Mg, Ca, and K in soil samples
(0- 8 cm depth) in four gap treatments and forest sites at 5 sampling times
over an 18 month period following bums (bars = S.E.).














































.4


~ ~~... .....-........-......... .-4

9 12 18
Months after treatment


Figure 3-3. Extractable P, Mg, K, and Ca in soil sampled from 8-20 cm in four gap
treatments and forest sites at 9. 12, and 18 months following bums. Y-axis scales are
identical to those in Figure 3-2 for soils sampled from 0-8 cm depth (bars = S.E.).


IY'I










~-.--new.-.


100

80


0

500

400

300

200

100


800


600 -

400 -

200 -


0

6000


-*- High intensity bum
-0- Low intensity bum
--v- Plant removal
-V- Gap control
U Forest


4000 -


2000 -


- -- -


100
80














Table 3-1. Results of ANOVAs of soil nutrients, organic matter, water content, and
soil pH of soil sampled 0-8 cm in four gap treatments and forest plots at 5 times
following bums. All variables were log transformed. Where a significant time *
treatment interaction was found, variables were analyzed separately by month.
Treatments with different letters are significantly different at P < 0.05.


No interaction time treatment
Variable Factor F


Post-hoc test results
P Month high low remove control forest


Magnesium Treatment 23.3 <0.001
Time 8.7 < 0.001


Treatment 70.1 < 0.001
Time 3.0 0.026


Water content Treatment 3.9 0.008
Time 390 <0001


Significant time treatment interaction
Variable Month F P


Potassium





Phosphorus





Organic matter




Soil strength


77.4 < 0.001
64.1 <0.001
11.5 <0.001
7.8 < 0.001
3.7 0.009
167.5 < 0.001
60.8 <0.001
45.5 <0.001
58.7 < 0.001
37.2 < 0.001
7.3 < 0.001
15.9 <0.001
4.6 0.003
4.7 0.002
3.9 0.007
12.6 < 0.001
24.3 <0.001
28.8 < 0.001
16.9 < 0.001


3 a b c c c
6 a b bc c c
9 a b bc c c
12 a b c c c
18 a b b b b
3 a b c c c
6 a b bc c bc
9 a b b b b
12 a b b b b
18 a a a a a
3 ab a a a b
6 ab ab a ab b
9 a a a a b
12 a a a a a
18 a a a a a
Post-hoc test results
high low remove control forest
a b c c c
a bc c bc b
a bc c b bc
a b b ab a
ab b b ab a
a b c c c
a a b b b
a b c c c
a b c c c
a b c c c
a b b b b
a b b b b
a b b ab ab
a b ab ab b
ab ab b ab a


a b
a a
a b
a bc


b b b
a b c
b c c
b c c


Calcium














Table 3-2. Results of ANOVAs of soil nutrients, organic matter, water content, and
soil pH in the 8-20 cm depth of soil of four gap treatments and forest plots at 3 times
following bums. All variables were log transformed. Where a significant time *
treatment interaction was found, variables were analyzed separately by month.
Treatments with different letters are significantly different at P < 0.05.


No interaction time treatment


Factor F


Post-hoc test results

P Month high low remove control forest


Phosphorus Treatment 105.4 < 0.001


Time


59.4 < 0.001


Magnesium Treatment 2.1 0.09


Time


7.3 0.001


Treatment 10.5 < 0.001


Time


4.7 0.013


Water content Treatment 5.6 0.001


Time


512 <0.001


Treatment 70.9 <0.001
Time 58.9 < 0.001


Significant time treatment interaction


Month F


9 34.6 < 0.001
12 17.5 < 0.001
18 9.5 <0.001


9 a b bc bc bc
12 a b bc bc bc
18 a b b b b
9
12
18
9 a b b b b
12 a ab b b b
18 a b ab b b
9 a a a a b
12
18 ab ab a ab b
9 a b bc c bc
12 a b c c c
18 a b c c c


Post-hoc test results

high low removal control forest
a b c bc bc
a b b b b
a b b b ab


Organic matter 9


1.7 0.16
2.7 < 0.001
11.3 <0.001


a ab ab b ab
b b a c c


Variable


Calcium


Variable


Potassium











Total soil N was strongly related to soil carbon (R2= 0.93; Figure 3-4) thus, patterns of

total N differences among treatments are expected to follow those for soil organic matter.

Both high- and low-intensity bum treatments significantly increased resin-available NH-

N, N03'-N, and P04-3-P during the first rainy season following bums (Table 3-3, Figure 3-5). This

pulse decreased after the first rainy season. Other than an increase in N03-N in plant removal

treatments during the first rainy season, the remaining treatments had little effect on NI--N, N03-

N, and PO3-P availability.

Coefficients of variation (CV) calculated for soil nutrients and organic matter within one

plot and among the 16 plots for each treatment are displayed in Table 3-4. A pattern emerged that

in the burned treatments, soil nutrients and organic matter were more variable within the one plot

than among all 16 plots. The opposite pattern was true for the plant removal, gap control, and

forest plots. Variation was in general greater among the 16 different plots than within the one plot.

Treatment effects on soil pH, soil organic matter content, and soil water content

Soil pH after high-intensity bums at 0-8 and 8-20 cm was higher than in all other

treatments throughout the 18 month sampling period (Figure 3-6 and 3-7). Soil pH in high-

intensity burn plots was 2 pH units higher than forest soils 2 months following bums. In low-

intensity bum treatments, pH was higher than in the remaining treatments at both depths at all

sampling periods. The plant removal and gap control treatments had little effect on soil pH.

High-intensity burn treatments significantly lowered soil organic matter; 2 months

following high-intensity burns soil organic matter in the top 8 cm of soil was approximately 72%

that of forest soils. By 18 months soil organic matter recovered to levels comparable to the

remaining treatments. Differences among the remaining treatments were small and varied

throughout the sampling periods.
























0.8 -

0.7 R2 = 0.93

,-- 0.6 -

S0.5-
O

" 0.4 -0
O0
0 0
0.3 -

0
H 0.2 -


0.1

0.0 ---------
0 1 2 3 4 5 6 7

Total soil carbon (%)


Figure 3-4. Total soil nitrogen and soil carbon in soil sampled 0-8 cm depth.
Soil samples used for this analyses were chosen from among all treatments
and all times since bums (2, 6, 9, 12, and 18 months) to obtain as wide a range
as possible for carbon content.


























Table 3-3. Results of ANOVAs of resin exchangable NH4, NO3, and P04 in 4 gap treatment
and forest plots measured at 3 time periods following bums. All variables had a significant
time*treatment interaction and therefore all were analyzed separately by month. All variables


were log transformed prior to analyses.
at P< 0.05.


Treatments with different letters are significantly


Analyzed by time period


Variable


F P


NH4-N first wet season
first dry season
second wet season

N03-N first wet season
first dry season
second wet season

P04-P first wet season
first dry season
second wet season


58.7
11.2
2

6.3
4.5
9.9

12.2
12.2
3.2


0
0.11


0
0.003
0

0
0
0.02


Post-hoc test results

high low removal control forest
a b c c c
a a ab be c


a ab ab b b


a a b b b
a bc c b ab
a bc ab ab ab









0.014 -

> 0.012 -
ir 0.010 -
-0 0.008 -
b 0.006 -
0 .0 0 4 .. .. ..... ....... . .
0 0.002 --- -

0.000
0.7 -
1 0.6 -
0.5
I-\
'SP 0.4 -
0.3 -

Z 0.2 -" -


0.0

0.5

0. o-0- High intensity burn
c--G Low intensity bum
0o.3 -- --- Plant removal
o -V-- Gap control
S0.2 *.. Forest
0.1

S0.0


first rainy season first dry season second rainy season

Sampling period

Figure 3-5. Resin-available ammonium, nitrate, and phosphate determined from exchange resins
buried in soil at 5 cm depth during 3 periods following treatments. Resins were buried for
approximately 3 months during each period. Time since bum of sampling periods were:
1st rainy season (2-5 mo), 1'tdry season (8-11 mo), and 2nd rainy season (15-18 mo; bars = S.E.).
















Table 3-4. Coefficients of variations (C V) of soil cations, organic matter, and resin-available nitrogen sampled within a
100 m2 plot ("within") and among the 16 different plots ("among") of each treatment. CV within plots was calculated
from 4 samples taken from each treatment in one block. CV among plots was calculated from the composite samples
taken from each treatment of 16 blocks. All calculations were performed with data taken 3 months following bums.
High intensity burn Low intensity burn Plant removal Gap control Forest
within among within among within among within among within among
Ca 39 18 42 35 28 37 15 53 36 46
Mg 40 36 53 32 19 22 7 31 17 35
K 47 21 53 30 34 31 21 49 17 39
P 84 39 66 38 8 27 23 29 14 54
OM 33 24 37 20 24 25 8 35 24 19
N03-N 27 36 25 71 49 58 75 51 48 51
NH4-N 49 160 54 39 41 114 53 18 41 20
ash depth 46 22
a CV = (standard deviation / mean) 100 (Sokal and Rohlf 1981)
* = coefficient of variation greater within a single treatment plot than among all treatment plots.









2.0
4 5 1 -. High intensity bum
--- -0- Low intensity burn
_,_ -- Plant removal
1.0 .-v- Gap control
7M a ^--* Forest

0.5 .
I '5' ^==^^ ^ ----------

0.0 -------
9 -

8


7 7-""" .
6-~ ~ -a---- _.....
6 -

5 -\i

S16 -



0 8 ...... ........... ..

S 4-

0
5
4 -
0 I-I I--





3

1 2

( 1 -

2 6 9 12 18
Months after treatment

Figure 3-6. Soil pH, air-dry water content, and organic matter content measured in soil (0-8 cm depth)
in the four gap treatments and forest sites at 5 sampling times over an 18 month period following bums.
Soil strength was measured at the soil surface with a soil penetrometer at 4 sampling periods over a 12
month period following bums. Water content is expressed as a percentage of air-dry weight (bars = S.E.).























1 ^.-_ ~ -_--


Months after treatment


Figure 3-7. Soil pH, air-dry water content, and organic matter content in soil sampled
from 8-20 cm depth in four gap treatments and forest sites at 9, 12, and 18 months
following bums. Y-axis scales are identical to those in Figure 3-6 for soil sampled from
0-8 cm depth (bars = S.E.).


-- High intensity burn
--0- Low intensity burn
-V Plant removal
--v- Gap control
-. ..a Forest


W









Although significant differences in soil water content were detected among treatments,

differences were not large and patterns were not consistent over the sampling period. The forest

plots had the lowest soil water content during the first 9 months, but this difference diminished

after 12 months. Larger differences were due to seasonal changes in soil water content, predictably

with the highest water contents in the rainy season of 1999, and the lowest water contents in the

drn season of September 1998.

Treatment effects on fine root mass

Fine root mass 18 months following bums was significantly higher in forest plots than in

high- and low-intensity bum treatments (F = 4.4, P = 0.005, n = 10). Fine root mass (live and

dead combined) in the top 7 cm of soil in was: forest 1.8 0.3 kg/m2, gap control 1.3 0.2 kg/m2,

plant removal 1.3 0.2 kg/m2, low-intensity bum 0.9 0.1 kg/m2, and high-intensity bum 0.7

0.1 kg/m2.

Treatment effects on soil physical properties

Soil bulk density and soil strength. Bulk density in high-intensity bum treatments was

significantly higher than in forest plots after 6 and 12 months (F = 3.1, P = 0.02, n = 10). There

were no significant differences among the remaining treatments. Bulk density averaged 1.3 0.05

g cm-2 in high-intensity bum treatments and 1.2 0.02 g cm"2 in forest plots after 12 months.

Soil strength in high-intensity bum treatments increased during the first 12 months

following burns (Figure 3-6, Table 3-1). Although at 3 months, soil strength was lowest in high-

intensity bum treatments, it was the highest in this treatment by 9 months. Soil strength in the other

treatments also increased during the first 12 months, but to a less extent than in high-intensity bum

plots.

Water infiltration and soil wettability. Water infiltration rates were significantly lower

in the high-intensity bums than in the remaining treatments (F = 31, P < 0.001, Figure 3-8). Soil









wettability significantly differed among treatments (F = 4.6, P = 0.002) and a slight, but non-

significant, difference was found in the wettability of different soil depths (F = 2.4, P = 0.07).

Surface soils in all treatments except the high-intensity bum tended to repel water (Figure 3-9). In

the high-intensity bum treatment, a slightly water-repellent layer was detected at 2-3 cm depths.

Effects of soil heating and ash addition on soil chemical properties

Field study. Due to high seed predation of A. colubrina, I only report results of soil

analyses. Ash addition significantly increased soil concentrations of P, K, Ca, and Mg as well as

soil pH (Table 3-5; Figure 3-10). Ash addition lowered soil organic matter, but did not affect

available resin-available NH4-N and N03-N. Heated soil had lower concentrations of Mg, but did

not have significantly different concentrations of P, K, Ca, organic matter, or resin-available

NH4-N and N03-N.

Greenhouse study. Both soil heating and ash addition decreased growth of C pluviosa

seedlings (Table 3-6; Figure 3-11). Intensely heated soil had higher levels of resin-available NW4-

N than moderately heated or control soil. Resin-available NO3'-N was not detectable in intensely

heated soil, but it was significantly higher in moderately heated soil than control soil. Ash addition

did not affect resin-available NH4-N and N03-N.

Discussion

Effects of high- and low-intensity burns on soil chemical properties

Controlled bums significantly affected on all soil chemical properties examined (soil pH,

soil organic matter, resin-available N and P, and Mehlich-extractable P, Ca, K, and Mg). These

changes, attributable to soil heating and/or ash deposited during bums, were greater after high-

intensity bums than low-intensity bums. High fuel loads combined with relatively complete bums

in the high-intensity bum treatment resulted in an average ash depth of 4.8 cm. Maximum

temperatures reached during high-intensity bums averaged 704 C at the soil surface and 227 C at

3 cm depth (Chapter 2). Little ash was deposited after low-intensity bums, due mostly to the lower



















0.20 -

--0- High intensity burn
-0- Low intensity bum
0.16 Plant removal
-* ,-- Gap control
N' i**" Forest


S0.12 N




S 0.08 ....... ......




0.04 -




0.00 -,,
5 10 15

Cummulative volume of water (ml cm-2)


Figure 3-8. Water infiltration rates of soil in four gap treatments and forest plots.
Infiltration was measured as the time required for the first 5 ml of a 10 ml column
of water to infiltrate soil.






54














0-


.g Ts/ i ---






" --0-^ ^ high intensity bum
2-- low intensity bum
0
S/ plant removal
S-/ 7- gap control
/ U- forest

3



0 5 10 15 20 25

Time for total absorption (s)



Figure 3-9. Soil wettability at the soil surface soil and at 1, 2, and 3 cm depths in four gap
treatments and forest plots. X-axis refers to the time to total abortion of 5 drops of
water applied to the soil surface with a dropper.
























Table 3-5. Results of ANOVAs of a field experiment examining the separate effects of
ash addition and soil heating on soil properties. Two levels of ash (no ash, 500 g ash) and
3 levels of soil heating (no heat, low intensity heat, and high intensity heat) were applied
to 0.25 m2 subplots in the field and soil sampled after 3 weeks. In the low intensity heat
treatment, a flame of 150-250C was applied to the soil surface of the treatment area for
5 minutes with a propane torch. In the high intensity heat treatment, a flame of 500-800C
was applied to the soil surface of the treatment area for 20 minutes. Ammonium and nitrate
were measured using anion and cation exchange resins buried at a depth of 5 cm for 3
months (n = 12).


Variable


Calcium
Potassium*
Magnesium*
Phosphorus*
NH4-N*
N03-N*
OM
pH


df F


615.8
65.5
75.6
23.5
2.7
1.2
11.7
210.3


* log transformed prior to analyses


P
< 0.001
< 0.001
< 0.001
< 0.001
0.11
0.28
0.001
< 0.001


Factors
Heat
F P
3.1 0.05
1.2 0.32
4.7 0.01
0.2 0.81
1.3 0.29
1.8 0.17
3.8 0.03
2.7 0.08


Ash heat
F P
3.3 0.05
0.5 0.59
1.0 0.38
2.8 0.07
0.4 0.96
2.3 0.11
2.3 0.11
0.4 0.96











Calcium


8000
7000
6000
5000
4000
3000
2000
1000
0

700
600
500
400
300
200
100
0

6

5
4

3

2

1
0

0.10

0.08

0.06

0.04

0.02


control ash added

Organic matter









control ash added

-+
N'H4


1400 7
1200 -
CA 1000-
800-
600 -
400-
200 -
0

50 -


0-I

10

8-

6

4

2

0-



0.4 -
o

.00.3

S0.2

Z0.1
0
z 0.0 -


Potassium

-control
E low heat
- high heat


control ash added


control ash added


pH


control ash added


NO3"


control ash added


control ash added


Figure 3-10. Results of field study of soil that had received combinations of heating (no heat, low
intensity heat, and high intensity heat) and ash addition (no ash, ash added). All soil was sampled
from 0-8 cm depth, 3 weeks after treatments. Resin available NH+ and N03 were measured
using anion and cation exchange resins buried to a depth of 5 cm for 3 months (n=12; bars=S.E.).


control ash added


Magnesium


TI


0.00 1






















Table 3-6. Results of ANOVAs of a greenhouse experiment examining the separate effects o
ash addition and soil heating on available nitrogen and seedling growth of Caesalpinia
pluviosa. Two levels of ash (no ash, 15 g ash added) and 3 levels of soil heating (no heat, lov
intensity heat, and high intensity heat) were applied to soil used to fill planting bags. In the lo,
intensity heat treatment, soil was heated in a conventional oven at 150-200 C for 10 minutes.
In the high intensity\ heat treatment, soil was heated in a conventional oven at -250 C for 40
minutes and fired with a propane torch for 5 minutes at a flame temperature of 500-800 C.
Ammonium and nitrate weremeasured using anion and cation exchange resins buried at a depth
of 5 cm for 3 months. Seedling height was measured 4 months after planting (n = 12).
Factors
Ash Heat Ash heat
Variable df F P F P F P
NH4-N 5 6.2 0..023 291.0 < 0.001 1.8 0.20
N03-N 5 1.7 0.20 23.3 < 0.001 0.6 0.58
Seedling height 4.8 0.012**


*Significant interaction for seedling height therefore analyzed separately by heat treatment.
Heat treatment F P tallest
Seedling height no heat 11.4 0.003 w/o ash
low heat 11.7 0.002 w/oash
high heat 0.15 0.7


















o'3












2
4-








.o
3
z


0








aM)


control ash added


Figure 3-11. Results of greenhouse study of Caesalpinia pluviosa seedlings planted in
soil that been heated (no heat, low intensity heat, and high intensity heat) and had ash
added (no ash, 15 g ash added). Seedling height was measured 4 months after planting
(n = 12). Ammonium and nitrate were measured using anion and cation exchange resins
buried in individual bags and watered for 22 days (n = 3). Bars are S.E.


2.00


1.501


0.08/
0.06 -
0.04 -
0.02 -
0.00

0.6

0.5 -

0.4 -

0.3 -

0.2 -

0.1 -


SIno heat
ilow heat
high heat


I


I


0.0 1









fuel loads but also to incomplete combustion of these fuels. Maximum temperatures

reached during low-intensity bums averaged 2250 C at the soil surface; elevated temperatures at 3

cm depth were mostly undetectable.

Soil pH. Increased soil pH is a general effect forest fires (e.g., DeBano et al. 1977,

Wright and Bailey 1982, Kutiel et al. 1990, Hungerford et al. 1990, Stromgaard 1992, Neary et al.

1999). High concentrations of basic cations in ash deposited following fires (e.g., Ca, K, Mg, Na)

is the major mechanism of increased soil pH (Kutiel et al. 1990, Stromgaard 1992). Although soil

heating may also increase pH by releasing basic cations from soil organic matter (Giovannini et al.

1990), results of the soil heating and ash addition field study (Figure 3-11) revealed that ash

addition significantly increased soil pH while soil heating had only a slight but non-significant

effect.

Soil organic matter. High-intensity bums caused a net loss of organic matter in surface

soils, a predictable consequence of intense fire (e.g., DeBano et. al. 1977, Hungerford et al. 1990,

Neary et al. 1999). Several studies conducted in the tropics have found decreased soil organic

matter following slash burning (Amazonia: Uhl and Jordan 1984, Mackensen et al. 1996,

Australia: Rab 1996). Experimental studies have shown that soil organic matter loss is a direct

effect of soil heating (e.g., Hosking 1938, in Humphreys and Craig 1981, Giovannini et al. 1990),

with distillation of volatile organic compounds occurring between soil temperatures of 100-300 C

and near complete loss of soil organic matter at temperatures >450 C. Soil organic matter

contents following high-intensity bums (Figure 3-6), averaged over 0-8 cm, likely do not reflect

larger losses in soil organic matter that occurred in the first several centimeters. Consumption of

soil organic matter was probably complete at the soil surface during the high-intensity bums,

where maximum soil temperatures averaged 683C. Due the sharp decrease in soil temperature

with depth (Chapter 2), organic matter consumption was probably negligible below the top several

cm of soil and not detected at 8-20 cm depth.









Average surface temperatures during low-intensity bums (160 C) were not hot enough for

the consumption of soil organic matter, hence average soil organic matter contents of low-intensity

bum plots were not significantly lower than those of forest soils. In fact, average soil organic

matter contents after low-intensity bums were higher than those of adjacent forest soils. Increases

in soil organic matter have been shown to occur during light to moderate bums (e.g., Hungerford et

al. 1990) due to the incorporation ofunbumrned or partially burned slash fragments into soil. For

example, Stromgaard (1992) attributed increased soil carbon following slash burning in miombo

woodlands to charcoal accumulation or small organic particles washed in from ash.

Soil organic matter in high-intensity bum plots recovered to levels higher than those in

forest plots within 18 months following bums. High daytime soil temperature in burned plots and

high soil moisture within gap treatments may have contributed to this rapid recovery by increasing

decomposition rates. Though generally rapid, recovery of soil organic matter following slash and

bum vary among and within tropical forests. For example, at the same site in the Venezuelan

Amazon, Montagnini and Buschbacher (1989) reported recovery of soil organic matter within 6

months, while Uhl and Jordan (1984) reported that recovery required 5 years. Other than

differences in climate and site productivity, fire intensity and land use following burns also affect

recovery of soil organic matter, thus making comparisons among studies difficult.

Total soil nitrogen. Total soil N was linearly related to soil organic carbon, hence the

greatest declines in soil N occurred in high-intensity bum plots where organic matter in the top 8

cm of soil decreased an average of 28% from adjacent forest soils. Similarly, total soil N was

reported to decrease following slash and bum of tropical forest in Costa Rica (Ewel et al. 1981)

and in the Venezuelan Amazon (Uhl and Jordan 1984). As with soil organic matter, this average

(28%) underestimates N losses from the top several centimeters of soil or from more intensely

burned patches; N losses reached 84% in the top 8 cm of soil and likely approached 100% in

scorched surface soils. Comparably high N losses were reported in chaparral soil heated to 500 C









(80% N loss; Dunn and DeBano 1977) and Mediterranean soils heated to 600 C (86% N loss;

Kutiel et al. 1990).

Losses of total soil N during low-intensity burns were negligible as indicated by low losses

of soil organic matter. In fact, slightly higher soil organic matter contents in low-intensity bum

plots relative to adjacent forest soil suggest that total soil N increased after this treatment. This is

likely due to the mixing of slash fragments into surface soils. Increases in total N in surface soils

has also been found following slash burning in the tropical forests (Montagnini and Buschbacher

1989, Stromgaard 1992) and temperate forests (Gholz et al. 1985).

Resin-available nitrogen. In contrast to decreases in total soil N, amounts of resin-

available nitrogen (NH4-N and NO03-N) increased after bums of high-intensity. These findings

agree with those of Matson et al.(1987) and Montagnini and Buschbacher (1989), who also

reported increases of ammonium and nitrate following slash burning of tropical forest in Costa

Rica and Venezuela, respectively. In the present study, low-intensity bums increased resin-

available N levels as well, although to a lesser degree than high-intensity bums. Temperate zone

studies have also noted that increases in inorganic N are dependent on fire intensity (Dunn and

DeBano 1977, Giovannini etal. 1990, Kutiel et al. 1990, Rice 1993, Weston and Attiwill 1996,

McMurtrie and Dewar 1997). Dunn and DeBano (1977) demonstrated that the greatest increases

in ammonium and nitrate for chaparral soils occurred at soil temperatures up to 300 C, due to the

mineralization of organic N. At soil temperatures of>500 C, inorganic N decreases due to

volatilization (Dunn and DeBano 1977). Similar results were reported in soil heating studies

conducted by Giovannini et al. (1990).

Increased ammonium availability following bums may be enhanced by soil microbial

death, which occurs at temperatures as low as 50-121 C (Neary et al. 1999). Soil microbial

death was likely substantial in high-intensity bum plots and may have occurred within small well-

burned patches in low-intensity bum plots as well. In their study of nitrogen transformations









following slash and bum on volcanic soils in Costa Rica, Matson et al. (1987) found that the

amount of nitrogen that disappeared from microbial biomass after bums was similar to the

concurrent increase in apparent net nitrogen mineralization.

Matson et al. (1987) also attributed increased nitrate concentrations to enhanced

nitrification rates. Burning generally creates favorable conditions for nitrification, such as raised

pH values and base saturation (Pritchett and Fisher 1987). Increased nitrification rates were also

reported following slash burning of Venezuelan forests (Montagnini and Buschbacher 1989). In

contrast, other studies have shown that nitrification rates are reduced by fire, due to a decreased

biomass ofnitrifiers (e.g., Dunn and DeBano 1977, Stromgaard 1992). Reduced nitrification rates

would cause an accumulation of ammonium, which is less subject to leaching than nitrate. This

effect may explain why nitrate was undetectable in the intensely heated soil of the greenhouse study

while levels of ammonium were very high (Figure 3-12).

Elevated concentrations of ammonium and nitrate following high and low-intensity bums

were short-lived in my study. Within 8 months of bums (after the first rainy season), inorganic N

in burned plots declined to levels found in adjacent forest. This result is similar to the rate of

decline of inorganic N following slash burning of a wet forest in Costa Rica (Matson et al. 1987).

Phosphorus. High- and low-intensity bums increased both extractable P and resin-

available PO43-P. Inorganic P additions in ash likely contributed to these increases. In the soil

heating and ash addition field study (Figure 3-11), ash addition significantly increased extractable

P, while soil heating had little effect. Similarly, Stromgaard (1992) attributed increases in

extractable P after slash burning of miombo woodlands to ash deposition, and Rice (1993) found

that soil P043-P concentrations in Californian chaparral following fire were correlated with ash

depth but not fire intensity.

Though possibly less important than ash deposition, soil heating can increase extractable P

by mineralizing organic P, as may have occurred with resin-available N. Giovannini et al. (1990)









found an increase in inorganic P accompanied by an equivalent decrease in organic P in soil

samples heated to 4600 C; at temperatures >460C, all organic P was destroyed and only inorganic

P remained. Soil heating may have been comparatively more important in low-intensity bum

treatments, where relatively little ash was deposited. Other studies have reported increases in

inorganic P following fires of low-intensity. For example, in a study of soil nutrient levels

associated with shifting agriculture in the Asian tropics, Andriessse and Koopmans (1984) found

available P increased almost 300% after heating to 200 C, which they attributed to mineralization

of organic P.

Cations. High-intensity bums significantly increased soil concentrations of extractable

cations (Ca, K, and Mg). Low-intensity burns also increased cation concentrations, although not

as dramatically. Similar to P, results of the soil heating and ash addition field study suggest that

increases in extractable cation concentrations by burning is mostly due to ash deposition.

Significant increases in extractable Ca, Mg, and K after fire in miombo woodland (Stromgaard

1992) and Brazilian cerrado (Coutinho 1990) were also attributed to ash deposition.

As with P, soil heating may also increase extractable Ca, K, and Mg through

mineralization of organic forms (Giovannini et al. 1990). Extractable Ca and Mg peaked in soil

heated to 200 C and declined at higher temperatures; extractable K peaked in soil heated to 700 C

(Giovannini et al. 1990). Results of the soil heating and ash addition study conform with this

pattern; soil heating significantly decreased extractable Ca and Mg concentrations and had no

effect on extractable K concentration.

Decreases in cation concentrations over the 18 month post-bum period are similar to those

following bums in temperate forests (e.g., DeBano et al. 1977, DeRonde 1990, Kutiel and Shaviv

1992, Hemrnandez el al. 1997) and tropical forests (e.g., Uhl and Jordan 1984, Coutinho 1990,

Stromgaard 1992, Mackensen 1996). The order of decrease (K> Mg > Ca) corresponds with

cations' mobility and susceptibility to leaching. In high-intensity bum plots, plant uptake was









probably not important in cation decreases during the first year, as plant cover remained less than

25% (Chapter 5). Plant uptake may have been more important during the second year, as

vegetative cover reached 60% after 18 months.

Variation of soil nutrients within and among plots. Based on the comparison of one

plot. soil nutrients and organic matter in burned treatments appeared to be more variable within

plots than among plots of the same treatment (Table 3-4). This pattern suggests that natural

variation in soil nutrients was increased due to heterogeneity of bums. The opposite pattern was

true for the plant removal, gap control, and forest plots. Variation in soil properties was in general

greater among the 16 different plots than within the same 100 nm2 plot, suggesting variations in soil

fertility in the absence of fire are expressed at larger scales.

Actual variation in soil properties after high-intensity bums may have been greater than

reflected by random sampling. For example, nutrient and organic matter contents of severely

scorched soil differed greatly from averages of high-intensity bum plots. Mineral concentrations of

such scorched soil sampled to 5 cm (with percentages of high-intensity bum plot averages) were:

6415 mg/kg Ca (133%); 197 mg/kg Mg (56%); 71 mg/kg K (11%); 0.25 mg/kg P (0.3%); and

0.1% organic matter (4%). These extremely scorched soils were not common (< 1% high-intensity

bum plots), but potentially affect plant colonization by providing microsites different from less

intensely burned areas.

Increased soil heterogeneity after bums has been observed by other authors. Christensen

(1985) noted that soil nutrient concentration in chaparral is considerably more variable after fire

than before, due to local variation in fire intensity and the uneven distribution of ash.

Heterogeneity in soil nutrients potentially has important consequences for colonizing plants. For

example, Rice (1993) observed that even small scale patterns in fire intensity and ash distribution

were reflected in later establishment of chaparral shrubs.









Changes in fine root mass following burns

Lower fine root biomass after high-intensity bums was likely due to a combination of rapid

decomposition of dead roots as well as direct oxidation by fire. Experimental studies have shown

that fine roots are desiccated or killed at soil temperatures of 48-54 C (Neary et al. 1999).

Temperatures in high-intensity bum plots (61-399 C at 3 cm depth) were not only well above this

range, but were likely high enough in places to completely oxidize fine roots. Even during low-

intensity burns, temperatures in surface soils (160 C average at soil surface) were sufficiently high

to kill fine roots. Fine root mortality during bums was likely greater than death of larger roots, not

only because of their small size, but also due to their concentration near the soil surface.

Fine root mortality potentially has important effects on soil fertility, as their decomposition

may increase soil nutrient concentrations. Nutrient input from roots has been hypothesized to be

an important pathway for nutrient cycling, particularly in tropical dry forests, due to their larger

store of biomass below-ground (Martinez-Yrizar 1996, Jaramnillo and Sanford 1995). Decreased

fine live-root mass is also expected to contribute to higher water and nutrient availability due to

reduced uptake. However, average soil moisture contents in high-intensity bum plots were not

significantly different than those in other gap treatments. Possibly, lower water uptake was offset

by decreased water holding capacity of intensely bum soil caused by a loss of soil organic matter.

Effects of high- and low-intensity burns on soil physical properties

Changes in soil strength, bulk density, and water infiltration rates in high-intensity bum

plots were substantial. The decrease in soil organic matter in high-intensity bum plots likely

influenced these observed changes in soil physical properties. Organic matter influences soil

structure through aggregate formation; a decrease in organic matter decreases total porosity,

particularly macro-pore spaces (> 0.6 mm). The increase in surface soil strength during the first

year following high-intensity bums was likely due to the settling of soil minerals and ash into









spaces left void by organic matter and fine roots. This settling of soil particles would also

contribute to higher soil bulk densities.

Decreased macro-pore space would also contribute to the lower infiltration rates observed

in high-intensity bum plots. However, these lowered infiltration rates caused by high-intensity

bums did not result in any observable surface runoff. The lowest infiltration rate recorded in a

high-intensity burn plot was 5 times faster than the rate needed to absorb a 5 cmhr" rainfall (0.002

cm3cm2sec"). Further, most plots were located on level ground, therefore if a rain event exceeded

the soil's rate of infiltration, the accrued water would not run-off.

Increased wettability of surface soils after high-intensity bums conforms to studies that

report soil temperatures >288 C destroy water-repellent layers (Neary et al. 1999). Soil

temperatures of 176-288 C reportedly form water-repellent layers (Neary et al. 1999), explaining

the presence of a slightly water-repellent layer at 2-3 cm depth. However, the decreased wettability

of this soil layer does explain the lower infiltration rates in high-intensity burn treatments, as

surface soils in the remaining treatments had similar wettability properties. Possibly, a more

water-repellent layer was formed deeper than 3 cm in high-intensity bum plots, but was undetected

due to the sampling strategy.

Soil strength, bulk density, water infiltration, and water repellency of low-intensity bum

plots were not different from those in the unburned treatments. Again, this pattern may reflect the

influence of organic matter on soil physical characteristics; the lower temperatures during low-

intensity burns (mean 120 C), did not decrease soil organic matter.

Potential effects of high- and low-intensity burns on tree seedling growth

Soil heating and ash addition significantly affected Caesalpinia seedling growth, although

in a manner opposite than expected; ash addition decreased seedling growth. This result suggests

that the quantity of ash added to soil may have been at toxic levels for this species. Also, tree

seedlings were shorter in soil heated at both low and high intensities. This result only partially









corresponds with a similar study by Giovannini et al. (1990), which examined the effects of soil

heating on wheat (Triticum aestivum) seedling growth. They found that while soil heated to 170

C had no effect on plant growth, soil heated to 220 C and 460 C increased seedling height and

biomass, whereas soil heated to 700 C and 900 C had detrimental effects on seedling growth.

The authors attributed increased growth in moderately heated soil to greater ammonium and

available phosphorus concentrations. Lowered growth in intensely heated soil was attributed to the

sharp increase in soil pH and release of Ca and K to toxic levels, as seedlings in this treatment of

their study showed symptoms of nutritional disorder.

In my greenhouse study, decreased seedling growth in intensely heated soil may be the

result of degraded soil structure or toxic levels of cations. However, it is unclear why seedling

growth was impaired in the lightly heated soil. It is important to consider that although seedling

heights significantly differed among treatments, maximum height differences were only 3 cm. This

slight difference after 4 months of growth may not be biologically significant. Perhaps the effects

of soil heating and ash addition on seedling growth would have been more apparent if a shade-

intolerant species had been used as a bioassay. Caesalpinia is partially shade-tolerant and

exhibited slow growth rates in the field as well (Chapter 4).

Despite the potentially negative effects of increased bulk density and soil strength, lowered

infiltration rates, and possibly toxic effects of cations on plant growth, seedling heights of shade-

intolerant species were greatest in high-intensity bum plots (Chapter 4). This increased growth in

intensely burned soils may be due to several factors. Initially, soil strength in high-intensity burn

plots was the lowest of all treatments, therefore early colonizing seedlings should not have

experienced mechanical impedance of root growth. Secondly, nutrient concentrations were highest

in high-intensity burn plots which may have offset decreased movement of nutrients through the

soil. Also, toxic levels of cations may only have been a factor in small areas of high ash deposition

or severely scorched soils; seedlings may not have been able to establish in these small areas and









therefore the effects on growth were not observed. Most importantly, the density of plants

colonizing high-intensity bum plots was low (Chapters 5), so that established tree seedlings likely

benefited from reduced competition for soil water and nutrients.

Effects of plant removal and canopy gap formation on soil chemical and physical properties

Soil moisture content was higher in all of the gap treatments than forest plots for the first 9

months following bums. Higher soil moisture within tree fall gaps than under adjacent forest is a

pattern repeatedly found in tropical forest studies (e.g., Vitousek and Denslow 1986) and has been

attributed to decreased transpiration within gaps due to less vegetation. The difference in soil

moisture content between forest and gap plots diminished over the first year as the amount of

vegetation in gaps increased.

Plant removal and gap control treatments did not significantly change soil chemical or

physical properties from those in adjacent intact forest. Although it is hypothesized that the

increased soil temperatures, moisture, and litter depth in tree fall gaps will increase nutrient

availability (i.e. Bazzaz 1980), conclusive evidence to suggest this is true has not been reported.

For example, in a study of natural tree fall gaps in lowland moist forest in Costa Rica, Vitousek

and Denslow (1986) found that nitrogen mineralization did not increase in tree fall gaps and slight

phosphorus increases were not significant. The only difference they detected was within gap

microhabitats; the root throw zone had significantly less N and P than the crown zone. Luizao et

al. (1998) found similar results in a study of artificial gaps ranging in size from 40-2500 m2 in

Brazilian rain forest. No differences in microbial biomass, soil respiration, and nitrogen

mineralization or nitrification were found between gap and forest sites.

Most of the variation observed within the plant removal, gap control, and forest plots over

time was due to seasonal changes. Soil moisture content varied predictably with changes in rainfall

and N03 availability declined slightly during the dry season. This observation agrees with the few

studies of nutrient cycling conducted in tropical dry forests which have shown that nitrification









rates are highest during the rainy season and lowest at the end of the dry season (Singh et al. 1989,

Garcia-Mendez et al. 1991, see also Smith et al. 1998).

Longer term effects of controlled burns on soil properties

The duration of this study limits its conclusions to only short-term treatment effects. A

similar study conducted in Las Trancas in 1995 (Stanley 1995) however, reveals slightly longer

term effects of burning on these forest soils. Although not identical, the treatments applied in this

earlier study were comparable to those used in mine: a gap control; gap vegetation slashed and

removed; gap vegetation slashed and burned; and gap enlarged by 30%, vegetation slashed, and

burned (Stanley 1995). Fires in the enlarged gaps were likely more intense than fires in gaps that

were not enlarged, due to a greater amount of fuel. I measured soil nutrient concentrations and soil

pH at 0-8 and 8-20 cm depths in April 1998, 3 years following experimental bums and found no

significant differences among gap treatments at either depth (Figure 3-12). The soil property with

the most distinct trend was P concentration (P = 0.11), followed by Ca concentration (P = 0.19),

soil pH (P= 0.22), K concentration (P = 0.27), Mg concentration (P = 0.32), and organic matter

content (P = 0.35).

These results suggest that soil chemical changes following bums are relatively short-lived.

However, the lack of significant results may have been more indicative of the large variation found

within treatments rather than the lack of variation found among treatments. Felling gaps included a

diverse array of habitats, such as rock outcrops and stream side areas. Importantly, there was no

indication of declining soil nutrient concentrations in the cleared and burned areas of this pilot

study after 3 years. A widely held notion about the recovery of tropical ecosystems following

disturbance is that severe nutrient losses following deforestation limit forest regeneration (e.g.,

Allen 1985, Buchbacher et al. 1988). Clearly, longer-term sampling of the burned plots is needed

before conclusions can be drawn about the long-term sustainability of severely burned soils at this

forest site.













o+ Cb
CD W





a






0 -


o -





~CD











0 00








ot
* 0 0























(7 0
Cr~i

















( o00


5 Ui H C A




iT,
u% -s


Organic matter


0 w1J -. /Ih O


Mg (mg kg")

s
k/i 0 k/i 0 kI 0


0 0 0


u


000









pH

! !00


S. 4 00 oo
0 00 00 0~ 0~ 0% 0


Ca (mg kg1)
--, M ,
0 0 0 0


P (mg kg1") K (mg kg"')











Conclusions

The experimental bums had mixed effects on physical, chemical, and biological soil

properties. High-intensity bums increased levels of available nitrogen and phosphorus, but these

pulses quickly decreased. Ash deposited after high-intensity bums increased cation concentrations,

but these pulses declined over time as well. Loss of organic matter during high-intensity bums

likely altered surface soil structure. This effect was apparent in increased bulk density and soil

strength, and decreased infiltration.

Low-intensity bums increased cation concentrations and available forms of N and P. Soil

structural changes were not as marked as after high-intensity bums, because little soil organic

matter was consumed. Therefore, soil chemical properties can be altered during low-intensity fires,

but changes in soil physical properties may only occur during intense fires. Plant removal, gap

control, and forest plots had no significant effects on soil chemical or physical properties.

Greenhouse studies suggested that loss of soil structure caused by soil heating and toxic

levels of cations created by large quantities of ash may hinder tree seedling growth. However,

these potentially negative effects were not apparent in the field. Shade-intolerant seedlings

established in high-intensity bum plots grew faster than those in all other treatments. Evidently, the

increase in soil nutrients caused by high-intensity burns was not offset by altered soil structure.

Importantly, the choice of tree species used as bioassays in these studies likely affected the results.

An important feature of the increase in resource availability produced by bums is the

transient nature of the increases. Therefore, the first plants to become established after bums

should benefit most from greater resource availability. Bums also greatly increased soil resource

heterogeneity. Therefore, successful establishment and vigorous growth of plants after bums may

be greatly influenced by chance, i.e., seeds dispersed into patches of well burned soil will exhibit

higher growth than those dispersed into unburned soil or severely scorched soil.
















CHAPTER 4
EFFECTS OF CANOPY GAP FORMATION, PLANT REMOVAL, AND
CONTROLLED BURNS OF HIGH AND LOW INTENSITIES ON
EARLY REGENERATION OF COMMERCIAL TREE SPECIES


Introduction

The most basic principle of sustainable forest management is that rates of timber

harvesting should not exceed the rate at which timber volume accumulates (e.g., Johnson and

Carbarle 1993, Dawkins and Philip 1998). This criterion requires sufficient regeneration of

harvested species and, in many neotropical forests, poses the greatest barrier to sustainable forest

management (e.g., Wyatt-Smith 1987, Verrisimo 1995, Gullison et al. 1996). There are several

silvicultural means of improving poor regeneration, ranging from intensive techniques such as

prescribed burning, to less intensive techniques such as selective harvesting without further

treatment. Choice of technique must be knowledgeably based on the natural regeneration

requirements of the target species. Presently, lack of information of the autoecology of many

harvested tree species is one of the largest deterrents to sustainable timber production (Bazzaz

1990, Bazzaz and Pickett 1980, Fox 1976, Gomez-Pompa and Burley 1991, Hall 1996, Whitmore

1989).

For many decades, seedling ecology was a minor part of tropical forestry (Hall 1996);

more recently, tropical seedling ecology has become the focus of much ecological research (e.g.,

Garwood 1983, Augspurger 1984a, 1984b, Swaine 1996, Kitajima 1996, Fenner and Kitajima

1998). The seedling stage is critical to regeneration as most mortality occurs early in the life of a

tree (Lieberman 1996, Li et al. 1996). Understanding this life stage is also important for forest

managers as species vary widely in their seedling ecology (Hall 1996), arid the division of tree

72









species into regeneration guilds is based on seed germination and seedling establishment

requirements. The most well defined group are species that require high light conditions for seed

germination and seedling establishment and often colonize following disturbances (i.e., early-

successional, Bazazz 1979; pioneer or secondary, Budowski 1965; or shade-intolerant, Swaine and

Whitmore 1988). At the other end of the spectrum are species that can germinate and persist in the

low light of forest understories (i.e., late-successional, Bazazz 1979; climax or primary, Budowski

1965; or shade-tolerant, Swaine and Whitmore 1988).

Tree species with different regeneration strategies require different silvicultural treatments

to enhance their regeneration. For example, many shade-tolerant species have advanced

regeneration in forest understories (Brokaw 1985b, Hartshorn 1989), therefore management

techniques would mostly entail ensuring the survival of this regeneration during harvesting and

enhancing its growth to mature stages. Enhancing regeneration of shade-intolerant species that do

not have seedling banks in forest understories involves creating sites suitable for seed germination

and seedling establishment and promoting safe arrival of seeds to these sites (Dickinson 1998).

In the seasonally dry forests of Bolivia, commercial tree species are represented in both

shade-tolerant and shade-intolerant groups (Guzman 1997). Consequently, a mixed-management

system was proposed for the community-owned forests of Lomerio (Pinard et al. 1999). Single-

tree selection was recommended to enhance regeneration of shade-tolerant trees. Group selection,

harvesting groups of trees to foster the development of even-aged patches, was suggested to

improve regeneration of shade-intolerant species.

Prescribed burning of logging gaps has also been suggested as an additional treatment for

the management of shade-intolerant species (Stanley 1995). Prescribed bums may enhance

seedling establishment and growth of shade-intolerant species in a number of ways, including

removing litter, reducing logging slash, and slowing vine proliferation (Stanley 1999). Abundant

regeneration of shade-intolerant species after wildfire in Lomerio (Mostacedo et al 1999, Gould et









al. 1999) lends further support to the promise of prescribed bums as a silvicultural tool for the

management of dry forests in Bolivia.

Although the pioneer-climax dichotomy provides an often useful paradigm for ecologists,

regeneration strategies of many tree species fall between the two extremes of completely shade-

intolerant or shade-tolerant (Augspurger 1984b, Condit et al. 1996). In fact, most rain forest tree

species are both shade-tolerant and gap-dependent, meaning they have the ability to persist in a

seedling bank in forest understories but require canopy opening to reach maturity (Hartshorn

1989). And, as there is a continuum of species regeneration strategies, there is also a gradient of

disturbance intensities among potential silvicultural treatments. The interaction of individual

species' biology and silvicultural treatments of varying intensities is inherently complex. Before

management techniques can be prescribed on a large scale, the effects of these techniques on all of

the species in question need to be examined. For example, what are the effects of more intense

disturbances, such as high-intensity fires, on the advance regeneration shade-tolerant trees? Or,

what is the minimum disturbance level required for the regeneration of shade-intolerant trees?

In this chapter, I examine the range of responses of commercial timber tree species to

experimental canopy opening, above-ground biomass removal, and controlled bums of high and

low intensities and compare these responses to those in forest understories. My objective was to

determine which gap treatment provided the best conditions for the establishment and growth of

each species. I also address the question of whether regeneration of each of these species is limited

by seed dispersal or by sites suitable for their establishment and growth. I approached these

questions by studying the effects of gap treatments on seed germination, seed predation, vegetative

regeneration, and seedling and sprout growth and mortality. Finally, I discuss results in relation to

silvicultural management options for each species.









Methods

I conducted this study in the four gap treatments (high-intensity burn, low-intensity burn,

plant removal, and gap control) applied to the 400 m2 blocks described in Chapter 2. Trees were

sampled in the paired 4 m2 subplots positioned near the gap center and gap edge of each 100 m2

treatment plot as well as the paired forest subplots (see Figure 2-4, Chapter 2).

Seed addition treatment

One randomly selected subplot (4 m2) of each pair was assigned a seed addition treatment.

Seeds of 5 commercial timber species were used in this treatment: Anadenanthera colubrna,

Astronium urundueva, Centrolobium microchate, Copaifera chodatiana, and Schinopsis

brasiliensis (hereafter referred to by genera; Table 4-1). I chose these species based on their

commercial importance and seed availability in 1997. Seeds of all species were collected in the '95

coupe of Las Trancas from the forest floor and stored in cloth bags under a shelter until sown in

plots. I removed all seeds that appeared to be damaged by predators or fungi. Twenty seeds each

(= 5 seeds/m2) of Anadenanthera, Astronium, and Copaifera, and 10 seeds each (= 2.5 seeds/m2)

of Centrolobium and Schinopsis were placed in each plot on the litter or soil surface, depending on

the soil surface conditions in each treatment (Chapter 2).

Viability of collected seeds

Viability of seeds collected for the seed addition treatment was determined in germination

trials conducted over a 60 day period (November-January) in Santa Cruz. One-hundred seeds of

each of the 5 species were planted in 5 trays (20 seeds per tray) of a 50:50 mix of sand and soil.

Trays were placed in a location receiving morning shade and afternoon sun and were watered each

morning. Newly germinated seeds were counted and removed daily.



















Table 4-1. Characteristics of tree species used in seed addition treatment or those species with sufficient regeneration for statistical
analysis. All but Acosmium are commercial timber species. Ecological classifiaction of shade tolerance of regeneration taken from
Pinardet al. 1998.

Shade Dispersal
Species name Family tol. Dispersal unit Fruit characteristics Seed size
Acosmium cardenasii Papilinoideae 3 wind seed legume, 10 x 15 mm, 1 mm thick
Anadenanthera colubrina Mimosoideae 1 gravity seed legume, 10-25 cm long 12 x 10 mm, 1 mm thick
Aspidosperma rigidum Apocynaceae 3 wind seed pod, 6 x 5 cm 25 x 20 mm, 1 mm thick
Astronium urundeuva Anacardiaceae 1 wind fruit small dried drupe, calyx to 1 cm 3 mm diam.
Caesalpinia pluviosa Caesalpinaceae 2 gravity seed legume, 10-15 cm long 10 x 10 mm 2-3 mm thick
Centrolobium microcheate Papilinoideae 1 wind fruit samara, 8-10 cm long x 3-4 cm wide 12 x 2 mm
Copaifera chodatiana Caesalpinaceae 3 animal seed dry pod, 2 x 3 cm seeds with oily aril 10 x 5 mm, 4 mm thick
Schinopsis brasiliensis Anacardiaceae 1 wind fruit samara, 15-20 mm long x 5 mm wide 10 x 3 mm
Shade tolerance: 1 = shade intolerant; 2 = partially shade-tolerant; 3 = shade-tolerant









Density, height, and relative height growth rate measurements

Seedlings have been variously defined as individuals still dependent on seed reserves (e.g.,

Garwood 1996), to individuals up to 2.7 m tall (Whitmore 1996). In this study, I did not use size

or physiology as a defining character, rather I define seedlings as individuals originating from seed

as opposed to those regenerating as sprouts. I measured seedling and/or sprout density and height

in each 4 m2 subplot (both seeded and unseeded) at 1.5, 3, 6, 9, 12, and 18 months after bums. At

each sampling period, all commercial species within subplots (of both seeded and unseeded

species) were identified as sprouts or seedlings, tagged, and height to the apical meristem

measured. Sprouts originating from the stem or root collar were easily identified because scars

were visible. Root sprouts (root suckers) were more difficult to identify but were recognizable

because the first leaves generally differed from the first true leaves of seedlings. Relative height

growth rates (hereafter referred to as RGR) was calculated as:

RGR = [In (height t2) In (heightti)] / (t2 t)

where tj and t2 are two measurement periods. Seedlings of Anadenanthera were extremely

abundant in 1997, therefore a maximum of 3 randomly selected individuals per subplot were

tagged for height measurements and the remaining individuals counted. Additionally, a maximum

of 3 randomly selected individuals of Acosmium cardenasii were tagged in each plot and the

remaining individuals counted. Although Acosmium is not commercially valuable due to its

susceptibility to heart rot, I included it because it has the most abundant tree regeneration of any

canopy tree species in Las Trancas '95. Many of the Anadenanthera seedlings had been browsed

between the 6 and 9 month assessments, therefore I noted presence or absence of browsing of

tagged individuals.









Statistics analyses

Seedling densities of most tree species could not be normalized, therefore seedling densities

were analyzed using Kruskall-Wallis non-parametric tests (SPSS 1997). Separate tests were run

for each species by month, testing for the effects of the seeding treatment and gap treatments on

densities. For species that regenerated from both seeds and sprouts, these regeneration modes were

analyzed separately. Square-root transformed densities of Acosmium and Anadenanthera were

normally distributed, therefore densities of these species were compared using repeated measures

ANOVAs. For Anadenanthera, the seeding treatment and gap treatments were factors in the

ANOVA model. For Acosmium, gap treatments and regeneration mode were used as factors, as

this species was not used in the seeding treatment but regenerated from both seeds and sprouts.

Seedling height and RGR were analyzed separately for each species using repeated

measures ANOVAs with gap treatments and, for the species that sprouted, regeneration mode as

factors. The effects of location within gap on seedling height and RGR was tested for

Anadenanthera; other species were not sufficiently abundant for this test. Also, the effect of

treatment on proportion of Anadenanthera seedlings browsed was tested with an ANOVA. Blocks

were random effects in each of the above models.

Effects of treatments on seed predation

A seed predation study was conducted to compare rates of seed predation among

treatments and at varying distances from gap centers. Two species were chosen for this study:

Centrolobium and Copaifera. In each treatment, 2 seeds of each species were placed at each of

five stations 1, 3, 5, 7, and 9 m from the gap center. Seeds were inspected after 2 and 9 weeks for

removal or signs of predation. In this study, I assumed that removal indicated predation and

therefore use "predation" as the sum of removed seeds and damaged seeds. The effects of

treatment and distance from gap center on seed predation were tested using Kruskal-Wallis tests on

proportions of seeds remaining and undamaged after 9 weeks.









Effects of the low-intensity burn treatment on seed germination

To assess the effect of low-intensity bums on seed germination of 3 commercial tree

species, 5 seeds each of Copaifera chodatiana, Centrolobium microchaete, and Schinopsis

brasiliensis were placed on leaf litter in low-bum plots just prior to controlled bums. After bums,

remaining seeds were retrieved and transported to the nearby community of San Lorenzo for

germination trials. Seeds from burned plots were placed in plastic trays with sand and watered

daily for 2 months. To detect if germination was related to degree of bum damage, seeds were

inspected before germination trials and assigned a damage score. Seeds of Centrolobium and

Schinopsis, which are protected by a dry husk, had 6 damage categories relating to the degree of

damage to the fruit: 0 (no damage) to 5 (fruit completely burned). Copaifera, which is dispersed

with only a fleshy aril, had 3 damage categories relating to the degree of visible damage to the seed.

The percentage of seeds germinating within each damage category were compared using a Chi-

squared test for independence.

Results

Seed viability of species used in seed addition treatment

Seeds of Copaifera demonstrated the highest viability in greenhouse germination trials,

followed by Anadenanthera, Astronium, Schinopsis, and Centrolobium (Figure 4-1). Seeds of

Anadenanthera and Astronium germinated most rapidly; of their viable seed, 80% and 83%,

respectively, germinated within 4 days.

Effects of treatments on seed predation

Seed predation was uncommon in gaps and there were no differences among treatments

(Copaifera P = 1.00; Centrolobium P = 0.30) or distance from gap edge (Copaifera P = 0.91;

Centrolobium P = 0.73). Overall, 17% of Copaifera and 5% of Centrolobium seeds were

removed or had evidence of predation over the 9 week observation period.












100 100

80 ---80------------------- 80
""^ ^^^^_^^ ~,t-~~ -'"

--o-- Anadenc
S...v... Astroniu
S60 60 -- Centrol
q---_.- 'Schinop.
.,vv A v vi i .............A Acosmiu
40 -40
A
Uv
20 20


0 0
0 10 20 30 40 50 60
Days after planting


Figure 4-1. Greenhouse seed germination trial of the 5 commercial tree species used in the seeding treatment. Results
represent the cumulative percentages of 100 seeds germinated each day over a 60 day germination trial.


nthera
m
ibium
is
'm









Effects of low-intensity burns on seed germination

Low-intensity bums either decreased or did not affect seed germination of Copaifera,

Centrolobium, and Schinopsis. Copaifera seeds placed in low-intensity bums had an overall

germination rate of 35%, less than half the germination rate of seeds used in the seed viability

germination trial (88%). Percent germination of Copaifera seeds was dependent on the degree of

bum damage (X2 = 9.5, P < 0.05). Germination rates of Schinopsis seeds placed in low-intensity

burns were also lower than unburned seeds (4% compared with 26%), however numbers were too

low for analysis. Centrolobium seeds placed in low-intensity bum plots had germination rates of

4%, only slightly lower than the germination of unburned seeds (5%).

Treatment effects on seedling densities of commercial tree species

Effects of seed addition treatment. Seed addition significantly increased seedling

densities of Centrolobium and Copaifera but not those of Anadenanthera or Astronium (Tables 4-

2, 4-3). Only 9 individuals of Schinopsis were recorded in all subplots; of these individuals, 6

were in seeded plots.

Effects of gap treatments. Patterns of commercial tree density among treatments and

over the 18 month sampling period varied among species. For simplicity, I have displayed

commercial species that had similar density patterns over the 18 months together as groups in

Figure 4-2. Significant differences among treatments are reported separately for each species in

Tables 4-3 and 4-4 and treated in more detail in following sections.

Density of the first group (Anadenanthera and Astronium), all true seedlings, peaked

within 3 months after burns and declined thereafter. Density of the second group (Copaijera.

Aspidosperma, and Caesalpinia) differed according to regeneration mode. While seedling density

of this group gradually increased throughout the 18 month observation period, sprout density

remained fairly constant after the first 3 months. The third group consists of one species

(Centrolobium) that regenerated predominately as root sprouts. Seedlings of Centrolobium, which










Table 4-2. Summary of seed fall and seedling densities of 7 commerical tree species and a non-commerical tree species (Acosmium cardenasii).
Seed fall refers to seedfall before or after the prescribed bum treatments (August 30-September 2, 1997). Significance values for the seeding
treatment are only given for seeded species. Due to the delayed germination of several species, the effects of the seedling treatment was not
detected until 12 months following treatments. For these species, densities and significance values are reported for 12 months. For species
not used in the seeding treatment, total density (seedlings and sprouts combined) at the end of the study (18 months) is reported with the
exception of the 1st year cohort of Caesalpinia, and Acosmium, whose densities are reported for 12 months. Pre-treatment density is based
on a pre-logging inventory conducted in 1995 in the same logging coupe before logging activities (Killeen et al. 1998).
Seed fall Seed addition Seedling density ______
density of unseeded seeded Months Sig. of Pre-treatment
before or after added seeds density density following seeding density
Species burns (mi2) (m"2) (m-2) treatment treatment (m-2)


Seeded species
Anadenanthera
Astronium
Centrolobium
Copaifera
Schinopsis
unseeded species
Caesalpinnia (1s1 year cohort'
Caesalpinnia (2nd year cohort
Aspidosperma
Acosmium


both
after
before
before
both


1.9
0.05
0
0.04
0.003


before
after
after
after


2.7
0.09
0.04
0.16
0.009


NS
NS


N/Aa


0.05 b
0.002
0
0.15
0


0.03
0.06
0.05
0.6


0.025

0.08
0.71


*** P< 0.001, ** 0.001 < P < 0.01, 0.01 < P < 0.05, NS = not significant
'Numbers too low for statistical analysis, justification described in text.
b Likely seedlings germinated in 1995, the year of the census, as no seedlings of these species > 1-yr old were found in forest plots in this study.























Table 4-3. Statistical analyses testing for the effect of the seed addition treatment.
Seedling densities ofAstronium. Centrolobium, and Copaifera could not be
normalized, therefore the effects of the seeding treatment was tested with a Kruskall
Wallis non-parametric test for each month. Square root transformed seedling densities
of Anadenanthera were normally distributed, therefore a repeated measures ANOVA
was used to test the effect of the seeding treatment and gap treatments for this species.
Because a significant time*treatment interaction was found for Anadenanthera an
ANOVA was used to analyze each month separately.
Astronium Copaifera Centrolobium
Month X2 P X2 P X2 P
3 0.9 0.34 1.1 0.29 1.0 0.32
6 1.3 0.25 13.4 0.00 6.2 0.01
9 0.7 0.40 20.0 0.00 15.2 0.00
12 1.0 0.31 26.0 0.00 15.2 0.00
18 0.7 0.40 23.7 0.00 3.0 0.08


Anadenanthera
seeding treatment gap treatments
Month F P F P
3 3.4 0.07 6.1 0.000
6 3.5 0.06 2.9 0.03
9 3.3 0.07 2.8 0.03
12 2.7 0.10 3.8 0.01
18 1.8 0.17 5.1 0.001









5

4

3

2

1

0
0.5

C-4 0.4
E
3 0.3
0.2

.> 0.1
.o
" 0.0
.4 0.2


0.1



0.0
0.10


0.05

0.00
0.050

0.025

0.000


Group 1. seedlings 0 high intensity burn
--0- low intensity burn
plant removal
--v- gap control
." ".... f forest
.;-~-- -

"//7 "" .-..i ........... z. --.
ff '-- ----- -- --- :- ..s ...........i



Group 2. seedlings


.. . . . . . .;





Group 2. sprouts




4-. i .......... ............ i .. .. ..........................


1.5mo 3 mo 6mo 9mo 12 mo 18 mo


Figure 4-2. Densities of commercial species over the 18 month sampling period following
experimental burns. Species are grouped according to similar regeneration strategies.
Group 1 (Anadenanthera and Astronium) were found predominately as seedlings, Group 2
(Aspidosperma, Copaifera, and Caesalpinia) were found as both seedlings and sprouts,
and Group 3 (Centrolobium) was found predominately as root sprouts (bars = S.E.; n = 16).


S I









were rare, established slowly over the first year and many died by 18 months. In contrast, density

of Centrolobium root sprouts remained relatively constant throughout the sampling period.

The prolonged seedling establishment and sprouting of species in these last 2 groups

restricted the calculations of RGR to the later part of the 18 month sampling period. For these

species, I calculated RGRs from no earlier than 6 or 9 months due to the low seedling and sprout

densities at 3 months. This method of calculating RGR may have limited detection of treatment

differences. If the most rapid growth occurs during the first several weeks following germination

or sprouting, then differences in RGRs measured after this initial growth spurt may not be

detectable. For this reason, seemingly contradictory results were obtained for some species in the

following results sections (i.e., Aspidosperma and Copaifera), where seedling heights were

significantly different among treatments whereas RGRs were not.

Treatment effects on seedling densities, heights, RGRs, and survival of seeded species

Anadenanthera. Seedlings of Anadenanthera were the most abundant of all commercial

tree species with an average density throughout the treatments of 1.1 0.4 seedlings/m2 (mean 1

S.E.) 18 months following treatments. All Anadenanthera seedlings in these analyses are from the

1997 cohort. Seed production of Anadenanthera in 1997 was larger than most years according to

locals. I did not encounter any seedlings > 1 yr-old in subplots, and due to extremely low seed

production in 1998, 1 also did not encounter Anadenanthera seedlings germinating in 1998.

Although Anadenanthera has the ability to coppice (pers. obs.), I did not encounter any sprouts in

the 4 m2 subplots.

Three months after bums, Anadenanthera seedling density was highest in forest plots (F =

6.1, P < 0.001; Figure 4-3). Density declined in forest plots due to high mortality and by 12

months, seedling densities were highest in the plant removal treatment (F = 3.8, P = 0.02).

Anadenanthera seedling density was lowest in gap controls throughout the study.










4 A.

rq 3
E

f 2 -

12- -j-- -



3 mo 6-m- 9 MS 12 in 18- Mn


0 1. ..____ "* ** *. .... ..."^
--V






3 mo 6mo 9 mo 12 mo 18 mo







120
C.
100 -4-- high intensity
--80 0- low intensity
u- plant removal
60 ---v- gap control-
S-forest

S40
20 "-
S. ... ... .... .. .
U. U
6 mo 9 mo 12 mo 18 mo

0.012 --
TD.




0.010 D.

0.008 -
0. 006-

0.004 -
0002 ........

0.000 ":- "
0-6 mo 6-12 mo 12-18 mo


Figure 4-3. A. Seedling density, B. percent seedling survival, C. seedling height, and
D. seedling relative height growth rates for Anadenanthera colubrina in the four gap
treatments and forest plot. All graphs follow the legend shown in graph C (bars= SE..
n= 16).









Height of Anadenanthera seedlings averaged 49 5 cm after 18 months. The tallest

seedling measured was 3 m and was found in a high-intensity bum plot. Seedlings in high and low-

intensity bum treatments were significantly taller than seedlings in the gap control or forest

understory; seedling height in the plant removal treatment was intermediate (F = 15.4, P < 0.001).

Correspondingly, differences in RGRs among treatments were significant and patterns followed

those for height (F= 18.4, P < 0.001). Anadenanthera seedlings were taller and had higher RGRs

in gap centers than near gap edges (F = 21.8, P = < 0.001; F = 16.0, P = < 0.001). After 18

months, seedlings in gap centers averaged 25 cm taller than seedlings near gap edges.

A mean of 21% (at 9 months) and 12% (at 12 months) of Anadenanthera seedlings were

browsed (tracks suggest by brocket deer, Mazama sp.). Seedlings in high-intensity bum treatments

suffered the highest rates of browsing while seedlings in the forest experienced the lowest rates (F

= 11.16, P< 0.001).

Astronium. Pattern of Astronium seedling density was strongly related to disturbance

intensity (Table 4-4). Throughout the 18 month sampling period, the highest seedling densities

were found in the high-intensity bum treatment plots followed by the low-intensity bum and plant

removal treatments (Figure 4-4). Mortality rates in these treatments was moderately high. Only 2

Astronium individuals were found in all 16 gap control plots (but these died during the 18 month

sampling period) and no Astronium seedlings were found in forest plots. I encountered only one

sprout of Astronium in the permanent subplots (a low-intensity bum treatment) and did not include

it min these analyses.

Height of Astronium seedlings averaged 110 14 cm after 18 months, the tallest mean

seedling height among species. Although heights and RGRs of Astronium seedlings were not

significantly different among treatments (F = 4.0, P = 0.11; F = 1.0, P = 0.40, respectively), there

was a distinct trend of taller Astronium seedlings with increasing disturbance intensity. Mean










Density at 18 months
0.4 -
0. Astronium urundueva*

0.2

0I jj*
0.0
04
., Centrolobium microchaete*


03 Caesalpinnia pluviosa
0 (pre-'98 cohorts)
0.1
Oi ~ w l j^;; ifW "''"'-Aa


1"
03
0.2
0- 1


.3] Copaifera chod&


Height at 18 months
300
400
300
200
2100
0


C.)

23


0 --

400
300
200


500
Soo
400

'00
100
0


500


1 2 3 4 5


1 2 3 4 5


Figure 4-4. Densities and heights of seedlings (black bars) and sprouts (grey bars) of 6 species
18 months following treatments (* = seeded species). Treatment codes along the x-axis are:
1= high intensity bum, 2 = low intensity bum, 3 = plant removal, 4 = gap control, 5 = forest
(bars = S.E.; n = 16).


Caesalpinnia pluviosa
'98 cohort


m i -_ i


Aspidosperma rigida


iinii


B ,i^ -n "f*














Table 4-4. Statistical analyses testing for differences among gap treatments and forest understory plots. Seedling densities of
Astronium, Aspidosperma, Caesalpinia, Centrolobium, and Copaifera could not be normalized, therefore the effect of gap
treatments was tested using a Kruskall Wallis non-parametric test for each month. Seedlings and sprouts were also analyzed
separately for these species (if applicable). Square-root transformed densities of Acosmium were normally distributed, therefore
a repeated measures ANOVA was used to test of the effect of the gap treatments and regeneration mode (seedling or sprout) for
this species. A significant regeneration mode*treatment was found, therefore seedlings and sprouts were then analyzed separately.
Astronium Aspidosperma Caesalpinia Centrolobium Copaifera
seedlings onl> seedlings sprouts seedlings sprouts seedlings root suckers seedlings sprouts
Mo X2 P X2 P X2 P X2 P X2 P X2 P X2 P X2 P X2 P
3 25.1 <0.001 9.9 0.04 7.7 0.10 21.2 0 8.6 0.07 4.0 0.41 11.1 0.03 17.5 0.002 6.2 0.18
6 34.7 <0.00 1 9.9 0.04 6.7 0.16 14.6 0.006 8.6 0.07 16.2 0.003 11.2 0.02 12.8 0.01 6.2 0.18
9 37.6 <0.001 4.0 0.40 12.1 0.02 11.6 0.021 6.8 0.15 11.8 0.02 23.2 <0.001 12.3 0.02 5.2 0.26
12 37.0 <0.001 5.5 0.24 12.0 0.02 11.6 0.021 5.6 0.23 5.6 0.24 18.5 0.001 10.4 0.03 6.1 0.19
18 37.5 <0.001 7.5 0.11 12.1 0.02 4.3 0.37 5.6 0.23 2.0 0.73 21.6 <0.00 11.5 0.02 6.1 0.20

Acosmium Acosmium seedlings sprouts
Source of variation df F P Source of variation df F P F P
Block 15 2.2 0.01 Block 1 1.1 0.40 1.2 0.29
Regeneration mode 1 0.0 0.93 Treatment 15 12.1 <.001 11.3 <.001
Treatment 4 3.6 0.01 Error 4
Mode Treatment 4 4.0 0.01
Error 65









height of Astronium seedlings in high-intensity bum treatments was 150 cm, more than twice the

height of seedlings in plant removal treatments (65 cm). Also, the tallest Astronium seedling (4 m)

was found in a high-intensity bum plot.

Copaifera. Copaifera regenerated from both seeds and sprouts, although the overall

density of seedlings was more than 10 times higher than sprouts (Table 4-5). Copaifera seedlings

were most abundant in gap control plots; sprout density did not differ among treatments (Table 4-

4, Figure 4-4). Mortality was low throughout the study period, particularly for sprouts (Table 4-6).

Sprouts of Copaifera were taller than seedlings (F = 8.4, P = 0.01), although their RGRs

did not differ (F = 1.8, P = 0.20). Height and RGRs did not differ among treatments (F = 0.3, P =

0.47; F = 0.3, P = 0.84, respectively).

Centrolobium. Centrolobium regenerated both from seeds and root sprouts. Density of

root sprouts was higher than seedling density at 6 months (Table 4-5). Apparently none of the

seedlings arose from naturally dispersed seeds; natural regeneration of this species was composed

entirely of root sprouts. At 3 months, Centrolobium root sprouts were most abundant in the plant

removal treatment; from 6 to 18 months they were most abundant in the low-intensity burn

treatment (Table 4-4, Figure 4-2). At 6 months, Centrolobium seedlings were most abundant in

the high-intensity burn treatment; at 9 months they were most abundant in the burned and plant

removal treatments. Seedling mortality was greater than root sprout mortality (Table 4-6).

Centrolobium sprouts averaged 267 cm tall after 18 months, more than 7 times the mean

height of Centrolobium seedlings (37 cm; F = 39.2, P = 0.003; Figure 4-4). Similarly, RGRs of

Centrolobium sprouts were also higher than those of seedlings (F = 13.6, P = 0.01). No

differences in heights at 18 months or RGRs were detected among treatments (F = 1.0, P = 0.49; F

= 0.1, P = 0.97. respectively).

Schinopsis brasilensis. A total of 9 Schinopsis seedlings were recorded in all subplots,

although the maximum number at any one sampling period was 7 (Table 4-7). After 18 months, 6




























Table 4-5. Statistical analyses comparing seedling and sprout densities ofAspidosperma,
Caesalpinia, Centrolobium, and Copaifera. Seedling and sprout densities of these
species could not be normalized, therefore a Kruskall-Wallis non-parametric test was
used to analyze each month.

Copaifera Centrolobium Aspidosperma Caesalpinnia
Month X2 P X2 P X2 P X2 P
3 5.6 0.02 3.7 0.06 7.7 0.006 0.13 0.72
6 22.2 0.00 5.9 0.02 11.2 0.001 0.32 0.57
9 27.6 0.00 2.6 0.11 10.2 0.001 0.04 0.84
12 31.5 0.00 2.2 0.13 15.2 0.000 0.001 0.98
18 35.9 0.00 3.1 0.08 10.2 0.001 15.6 0.00


















Table 4-6. Killing power, a parameter similar to mortality rate, for 5 commercial tree species in the four gap treatments and forest
understory plots. Killing power (k) is calculated separately for seedlings and sprouts of each species, except for Astronium, for which
no sprouts were found. Killing power is calculated as (logloax loglo0ak+), where a, represents the number of individuals in the
first cohort following treatments, and a,+I is the number of these individuals surviving into the next year (Begon and Mortimer 1981).
The numbers of individuals that died during this period are listed in columns "#."
high intensity low intensity plant gap control forest All individuals
Species bum bum removal ____________
k # k # k # k # k # k #
Aspidosperma seedlings 0.20 3 0.30 1 0.30 1 0.60 3 0.15 2 0.25 10
resprouts 0.00 0 0.22 9 0.05 1 0.48 4 1 0.17 15
Caesalpinia seedlings 0.00 0 0.00 0 0.07 1 0.22 2 0.00 0 0.09 3
resprouts 0.00 0 0.11 2 0.00 0 0.06 1 ______ 0.05 3
Centrolobium seedlings 0.11 2 0.12 4 0.06 1 1 2 0.14 10
resprouts 0.18 2 0.00 0 0.11 2 0.12 1 _______ 0.07 5
Copaifera seedlings 0.09 3 0.10 2 0.09 6 0.06 4 0.13 5 0.09 20
resprouts 0.00 0 0.00 0 0.08 1 0.00 0 0.00 0 0.02 1
Astronium seedlings 0.11 10 0.13 9 0.18 9 1 0.14 29
* = where a,+I was equal to 0 (all individuals died), k equals infinity




Full Text

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REGENERATION OF COMMERCIAL TREE SPECIES FOLLOWING CONTROLLED BURNS IN A TROPICAL DRY FOREST IN EASTERN BOLMA by DEBORAH K. KENNARD A DISSERTATION PRESENTED TO THE GRADUATE SCHOOL OF THE UNIVERSITY OF FLORIDA IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF DOCTOR OF PfllLOSOPHY UNIVERSITY OF FLORIDA 2000

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Dedicated to m y parents, Margaret Kennard and Robert Kennard and to my husband, Josh McDaniel

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ACKNOWLEDGMENTS I sincerely thank everyone who contributed in some way to this dissertation. First I would like to thank the members of my committee. Jack Putz, my advisor throughout my years at the University of Florida, has been very influential in my interest in forest ecology and management. I thank him for the opportunities he has provided, his time as a teacher and mentor and his diligence as an editor. His enthusiasm for ecology will influence me for years to come Henry Gholz was a tremendous help with the project design and generously provided the use of his lab for resin analyses Kaoru Kitajima patiently helped me with statistics and was instrwnental in the development of the seedling chapter. Insightful suggestions by Earl Stone aided me throughout every aspect of the project; I was repeatedly inspired by his immense knowledge and experience. Finally I would like to thank George Tanner who provided much-needed encouragement at critical times This dissertation would not have been possible without the generous financial and logistical support of BOLFOR (Proyecto de Manejo Foresta! Sostenible). Several members of BOLFOR were crucial in its completion First I would like to thank Todd Fredericksen, the Forest Ecologist at BOLFOR, for his help conducting the controlled bums statistical and technical advice ready supply of coca leaves, and humor. T. Fredericksen, J. Nittler and W. Cordero provided administrative support in Bolivia I gratefully acknowledge the following people who generously volunteered their time to assist with fieldwork : J. McDaniel L. MacDonald J. Chuviru, T. Fredricksen, N. Fredricksen, J Lincona, J. Justiniano, A. Ademar K. Gould, F Fatima, K. Hueberger B. Flores M Toledo L. Anderson, B Mostacedo and the Aberdeen students. I wish to acknowledge T Killeen and the herbarium at the Museo de Noel Kempf llJ

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Mercado for use of the data collected in the 1995 inventory of Las Trancas M Toledo kindl y assisted with plant identification My husband and I are very grateful to Todd and Nell Fredericksen for their hospitality in Santa Cruz Numerous Chiquitano community members assisted throughout the 18 months of thi s project ; I will be forever grateful for and impressed by their hard labors In particular I would like to thank Don Juan Loco Pesoa for being instrumental in the installation of the treatments ; Don Juan Faldin for sharing his knowledge of plants and their local uses ; and Don Lucas Salvatierra for his assistance in locating and measuring abandoned agricultural fields. I would also like to thank the many members of San Lorenzo who welcomed me into their community during my time awa y from El Campamento de Las Trancas At the University of Florida, I would like to thank J Bartos at the Analytical Research Lab for analysis of soil samples In the lab of H Gholz I would like to thank D. N oletti and K Clark for their valuable assistance with the resin extractions I was supported by a teaching assistantship offered by the Department of Botany and the Department of Biological Science s. Comments of several friends greatly improved the final draft : G Blate K. Gould, B. Ostertag T. Fredericksen and J. McDaniel. Most importantly I would like to thank my traveling companion, field-assistant Spanish teacher motorcycle driver toilet-digger wasp-magnet, translator and husband, Josh McDaniel. I V

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TABLE OF CONTENTS ACKNOWLEDGMENTS .. .. .. . . . .. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . iii ABSTRACT ........... ...... ............................ ... ... . .................. . . .. .... ... .. .. .. .. ..... ................... ..... vii CHAPTERS 1 INTRODUCTION Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1 Conservation and Management of Tropical Dry Forests ...... .. ............. .. ..... .. .... . .. .. . ........ 2 Management of Tropical Dry Forests in Eastern Bolivia .. ............. .. ..... .... .. .. . ... .... ... . .. ... 3 The Role of Historic Disturbance Regimes in Forest Management ..... . .... . .. .. ... . .......... ... .. 4 The Potential of Prescribed Burning in the Management of Bolivian Dry Forests . .. ... ........ . 7 Scope of Dissertation ... .. ....... .. . ....... . ... .... .. .. ...... ... .. .. .......... .. ............. . ... .. . ............ 8 2 STUDY SITE AND TREATMENT DESCRJPTIONS Introduction ......... .... .... .. ... ... .... . ....... .. ... ..... . ........ ..... ............. .................................. . 11 Study Site ........ ....... ........ ... .. .... . ..... . .. ...... .. .. ............ ... ...... . .. . ... ..... ... .... ... .......... .. 13 Methods ......... .................. .... .. ....... ... .. ... ....... .. ..... .. .... ." ... ........ .. .. ........ .... . .... ... .. . ... 18 Initial Treatment Results ........... ..... ... .... .. .... .. ....... .... .. ...... .. ... ......... ..... ................... 24 3 EFFECTS OF CANOPY GAP FORMATION PLANT REMOVAL AND CONTROLLED BURNS OF HIGH AND LOW INTENSITIES ON SOIL CHEMICAL AND PHYSICAL PROPERTIES Introduction ...................... ...... . ................. . ... .. ..... ......... ...... ... ......... .... ............ .. ...... .. .. 29 Methods ................. ... ... .. .... ....... .. ..... ......... .... .... .. ....... ... .. .. . ... ...... .. .... . .......... . .. ... .... 31 Results . ...... . ...... .. ........ .. .... ....... ......... ...... ... . ..... ...... .. .... .. ......... .. ... . ... .. . . .............. 38 Discussion ............ .. ......... ........ .......... .. ... ...... . . ........ ..... . ..... .......... .. ... .. ....... . .... ..... 52 Conclusions . . ...... ... ....... .. .. .... ......... ... .. ....... . . ... ............ ... .. .. .......... ... .. ................. .. 71 4 EFFECTS OF CANOPY GAP FORMA TlON, PLANT REMOVAL AND CONTROLLED BUR.il\JS OF HIGH AND LOW INTENSITIES ON EARLY REGENERATION OF COMMERCIAL TREE SPECIES Introduction ......... .. .......... ...... .. .. .. .. ....... ... . . ... .. . .. .. .... .... . ..... .. .... ............. .............. 72 Methods ........... ................... .. .. .. ... .... .... ... . ............................. .......... .......... ...... .... ....... 75 V

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Results ... .. ...... . .. .. . .... .. ................ ..... ... ......... .................. .. ..... .. ... ...... .................... . ... 79 Discussion ................ ........... .. ............................................. . ... .................. .............. ... . 96 Implications for management. ........ .. ...... ... .. ... .. .. .... . ................ ......................... . ..... 111 5 EFFECTS OF CANOPY GAP FORMATION PLANT REMOVAL AND CONTROLLED BURNS OF HIGH AND LOW INTENSITIES ON A DRY FOREST PLANT COMMUNITY Introduction ........................ ................ ....................... .. .. ........ .. ..... .. . ................. ..... .. 113 Methods . . . .. . . . . . . . . . . . . . . . . . . . . . . . . .. . . . . .. . . . . . .. .. . . . . . . . . . .. . . .. .. .. .. 115 Results . .. .. . . . . .. . . . . .. .. .. .. .. .. . . . . . . .. .. .. .. .. . . . . . . . . . . .. . . . . . .. .. .. .. . . . . . . . . .. 119 Discussion ........ ......... ....... .... ... ........................ ... . ......... ............ .............. ... ...... .. ... .. 132 Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14 7 6 COMMERCIAL TREE SPECIES REGENERATION FOLLOWING AGRICULTURAL ABANDONMENT IN BOLMAN DRY FORESTS Introduction ................ .. ...... . .. ................. . ........ ....... .. .............. ... ............... .. .......... ... 149 Methods ...... .... ....... ... ...... ................. ........... .. ............................ ............................ .. .... 150 Results . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15 3 Discussion .... ................... .... . ..... .. .. ........... ... .... ... .. ... ........ . ...... .................. ............... 165 Implications : Management potential of secondary forest in Lomerio ... ..... ..................... .. 172 7 SUMMARY AND CONCLUSIONS Summary of Study Results ....... ............................................... .... ...... ... . .. .................... 17 4 In1plications for Management ........... ............................ .. .. ... ... .. .. .. ................... .. .. .. .. .. .... 175 APPENDIX ........................ .. .... ......... ... .. ..................................................... .......... ............. 181 REFERENCES .......... .... ... .................................. ... ......................... ....... ....................... .... ...... 187 BIOGRAPHICAL SKETCH ...... .. . ....................... ......... ............... ........... .................... .... ..... 206 vi

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Abstract of Dissertation Presented to the Graduate School of the University of Florida in Partial Fulfillment of the Requirements for the Degree of Doctor of Philosophy REGENERATION OF COMMERCIAL TREE SPECIES FOLLOWING CONTROLLED BUR.l~S IN A TROPICAL DRY FOREST IN EASTERN BOLIVIA Chairman: Francis E Putz Major Department: Botany by Deborah K. Kennard May 2000 Low levels of disturbance associated with selective logging may be insufficient for the establishment of many Bolivian dry forest timber species the majority of which are shade intolerant. To examine the ecological potential of prescribed burning as a silvicultural tool, I compared the effects of canopy opening, plant removal and controlled burns of high and low intensities on 1) soil properties ; 2) establishment, growth, and survival of commercial tree species ; 3) and plant community structure and composition. To describe commercial tree regeneration over longer time scales, I characterized tree population structures in abandoned slash-and-bum fields ranging in age from 1-50 years, and compared these to a mature forest stand Both highand low-intensity bums caused a dramatic but temporary increase in soil nutrients. High-intensity bums altered several soil physical properties whereas low-intensity bums had little effect. Plant removal and canopy opening had little effect on soil chemical and physical properties vii

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Three responses to gap treatments were observed among commercial tree species. 1) Shade-intolerant species regenerating from seed were most successful following high-intensity bums 2) Shade-tolerant species were most abundant in treatments where survival of their advanced regeneration was most likely (gap control and plant removal) Some of these species had the ability to survive controlled bums by sprouting 3) Individuals of root sprouting species were most abundant following plant removal and low-intensity bums Sprouts dominated regeneration following canopy opening plant removal and low intensity bums In contrast seedlings dominated following high-intensity bums High-intensity bums shifted species composition relative to the less disturbed treatments Regeneration of shade-intolerant timber species was most abundant in y oung slash-and bum fallows Similar tree population structures in older slash-and-bum fallows and the mature forest stand suggests that the mature forest likely fom1ed following a large-scale disturbance Although prescribed burning enhanced the regeneration of shade-intolerant timber species it is not likely to become a forest management tool in Bolivia in the near future due to economic and political factors Managing secondary forests in Bolivia would provide an alternative to current attempts to regenerate these species after selective harvesting of mature forest. viii

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CHAPTER 1 INTRODUCTION Introduction Eastern Bolivia contains some of the largest and most diverse tracts of tropical dry forest in Latin America. Natural forest management for timber, if profitable, is one means of discouraging conversion of these forests to competing land uses. However insufficient regeneration of many commercial timber species presently poses an ecological barrier to sustained timber yield, prompting forest managers to explore additional silvicultural methods to enhance regeneration of these species. The low levels of disturbance associated wirh highl y selective logging may be insufficient for the establishment of many dry forest timber species the majority of which are shade-intolerant and likely require moderately intense disturbances for their establishment. Fire of both natural and anthropogenic origins has likely been a pervasive influence on tropical dry forests and therefore, prescribed burning may be an effective silvicultural tool to enhance regeneration of timber species following selective logging In this dissertation, I present the results of studies that examined commercial tree regeneration following disturbances of various intensities in a dry forest in lowland Bolivia including harvesting gap formation, controlled bums of high and low intensity and slash-and-bum agriculture My goal in carrying out these studies was to determine the regeneration requirements of these commercial tree species as well as to examine the effects of potential silvicultural treatments on forest soils and community structure and composition

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Conservation and Management of Tropical Dry Forests Tropical dry forests comprise approximately 42% of tropical forest land more than either moist or wet tropical forests (Murphy and Lugo 1986) Tropical dry forests also have supported higher human population densities than wetter tropical forests for centuries (Murphy and Lugo 1986) and, as a result have suffered more degradation and deforestation (Mooney et al. 1995 Murphy and Lugo 1995). Efforts to slow conversion rates of dry tropical forest have been negligible (Mooney et al 1995). For example, in 1988 less than 2% of the original dry forest on the Pacific coast of Central America remained intact and less than O 1 % had conservation status (Janzen 1988). Consequently, tropical dry forests are considered by some ecologists as the most threatened of the major tropical forest types (Janzen 1988) 2 Given the extensive use of tropical dry forests by rural people, their strict preservation may not be a realistic conservation goal. As Johnson and Carbarle ( 1993) note, most developing tropical countries rarely have the luxury of opting for forest preservation over forest exploitation Consequently, in most tropical countries, conversion of forested land continues to increase while the establishment of protected areas remains low (F AO 1999) Consensus is emerging among ecologists that protected areas due to their small number and size, cannot effectively conserve the majority of tropical species (Hansen et al. 1991, Heinrich 1995, Bawa and Seidler 1998). Mounting concern over global declines of biodiversity has prompted many ecologists to look outside of parks and nature preserves to semi-natural areas that may help maintain or at least slow the loss of, biodiversity (Sayer and Wegge 1992, Chazdon 1998). Natural forest management, the sustainable production of timber from natural forest areas, has been proposed as a means of maintaining forest value thereby deterring land owners from clearing forested land for other more profitable and destructive land uses (Poore et al 1989 Johnson and Carbarle 1993, Maser 1994 but see Rice et al. 1997). Although definitions of natural forest management vary, they usually encompass two ideas: first a sustained yield of forest

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3 products and second achieving this sustained yield through means that maintain other environmental services such as biodiversity soil quality and hydrology (reviewed by Johnson and Carbarle 1993) By maintaining forests in a semi-natural state, natural forest management is viewed by some as a critical means of maintaining biodiversity (Hansen et al 1991 Sayer and Wegge 1992 Frumhoff 1995 Heinrich 1995 Dickinson et al 1996, Putz et al. in press but see Bawa and Siedler 1998), particularly in regions where forests are in danger of conversion And despite the fact that few modem examples of economically viable natural forest management projects exist (e g Panayotou and Ashton 1992, Johnson and Carbarle 1993, Rice et al. 1997 Bawa and Siedler 1998, Bowles et al. 1998 but see Leslie 1987) the promotion of sustainable forest management has become a mainstay in international strategies for the protection of tropical forests (Bawa and Siedler 1998 Haworth and Cousell 1999). As Haworth and Cousell (1999: 62) explain "this approach has often been justified on the grounds that it is the result of a difficult choice between accepting on the one hand the inevitability of continued commercial logging of natural forests which will cause some damage to the ecosystem or on the other hand facing the complete loss of the forest to other causes Management of Tropical Dry Forests in Eastern Bolivia The Chiquitania region in eastern Bolivia contains one of the largest and most diverse tropical dry forests in the neotropics (Gentry 1993, Killeen et al. 1998). Although there is currently 150 000 to 200 000 km 2 ofrelatively intact forest in Chiquitania Dinerstein et al. (1995) identified this area as one of the most endangered ecosystems in the neotropics Deforestation in the alluvial soils near the city of Santa Cruz is in excess of 80 000 ha year" 1 (Killeen et al 1998). This conservation threat comes largely from large-scale industrial agriculture but other economic activities such as cattle ranching contribute to this rapid conversion of forested land These trends mimic past events in Argentina Paraguay and Eastern Brazil where similar dry forests have been deforested and fragmented over the past two decades (Killeen et al 1998)

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4 Due to recent forest policy changes in Bolivia natural forest management may now be a practical means of controlling deforestation in Chiquitania In 1996 a new forestry law was passed that requires among other features management plans for all Bolivian forests (Nittler and ash 1999) Bolivian logging companies now operate with management plans on an estimated 5 7 million hectares of forest and a total of 660 000 hectares of Bolivian forest has been certified as sustainably managed (Nittler and Nash 1999) The Lomerio Community Forest, located in the center of Chiquitania, was the first Bolivian forest to be certified Its 60,800 hectares are owned and managed by 27 communities of the Chiquitano indigenous people. The Chiquitanos have been managing their forests for 19 timber species, 5 of which are classified as highly valuable (Table 1-1). Acquiring and maintaining adequate regeneration of commercial tree species a challenge faced by all natural forest managers is particularly apparent in Lomerio For example, seedlings and saplings of 12 of the 19 commercial species are rare in forest understories (Table 1-1 ) A lack of seed sources due to previous over-harvesting may account for the scarcity of regeneration an1ong highly valued timber species However poor regeneration plagues most of the tree species that have only recently been harvested (Fredricksen 1999) Apparently the current harvesting and silvicultural techniques employed in Lomerio do not create conditions appropriate for the regeneration of these species A better understanding of the regeneration requirements of these commercial tree species is critical as continued regeneration failures will undoubtedly compromise the long-term sustainable management of these forests The Role of Historic Disturbance Regimes in Forest Management It is often assumed that forest management is more compatible with long-term sustainability if timber harvesting and silvicultural techniques are designed to mimic historic disturbance regimes (e g ., Pickett and White 1985 Oliver and Larson 1996 Attiwill 1994a 1994b) Although this assumption has rarely been tested ecologists argue that replacing harvested

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Table l-1 Table modified from Pinard et al. 1999 that reports characteristics of 19 commercial tree species of the dry forests of Lomerio Species were matched to a general silvicultural system (evenor uneven-aged) based on their regeneration requirements. Timber value is based on market value in 1999. Timber Managment Adult Sapling value system rarity rarity Shade tolerance l Amburana cearensis high even 3 3 2 Anadenanthera colubrina low even 1 2 3 Aspidosperma cylindrocarpon low uneven 2 3 4 Aspidosperma rigidum low uneven 1 2 5 Astronium urundueva low even 2 3 6 Caesalpinia pluviosa low uneven 1 2 7 Cariniana estrellensis low uneven 2 2 8 Cedrela jissilis high even 3 3 9 Centrolobium microchaete high even 1 3 10 Copaifera chodatiana low uneven 2 3 11 Cordia alliodora high even 3 2 12 Hymenea courbaril low even 3 3 13 Machaerium scleroxylon high uneven 1 2 14 Phyllostylon rhamnoides low uneven 2 2 15 Platymiscium ulei low even 3 3 16 Schinopsis brasilensis low even 2 3 17 Spondias mombin low even 2 3 18 Tabebuia impetiginosa low even 1 3 19 Tabebuia serratifolia low even 3 3 Shade tolerance : l = high light only large gaps ; 2 = partial shade, small gaps ; 3 = partial or full shade understory 1 1 2 2 1 2 3 1 1 3 1 1 3 3 1 1 1 1 1 5 Adult rarity(> 20 cm dbh) : 1 common(> 5 ha1 ); 2 = intermediate (1-5 ha \ 3 =rare(< 1 ha -\ Sapling rarity (5-10 cm dbh) : 1 common (>20 ha-\ 2 = intermediate (5-20 ha1 ) ; 3 = rare (<5 ha-\ Even : even-aged managment system with group selection Uneven : uneven-aged managment system with single-tree selection

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6 trees without irreversibly damaging the residual forest is more likely to occur under conditions similar to those that formed the original stand (Uhl et al. 1990). The selective cutting s y stems used in many tropical forests are often justified on models of gap-phase regeneration in unharvested forests (e g Whitmore 1989 Hartshorn 1989 Gomez-Pompa and Burle y 1991) Gap-phase regeneration however is not the most appropriate model for tropical dry forests Evidence suggests that single tree-fall gaps are smaller and less frequent in tropical dry forests than in moist or wet forests (Dickinson 1998). Rather very large gaps caused b y catastrophic disturbances more likely govern dry forest dynamics In Central America for example, mahogany (Swi etenia macrophyila) has been noted to regenerate in even-aged stands after hurricanes and fires (Lamb 1966 Snook 1996) The low level of disturbaJice created during highly selective logging appears to be a poor replicate of this disturbance regime and possibly for this reason natural regeneration of mahogany is scarce in most selectively logged areas (Verissimo et al. 1995 Gullison et al. 1996 Whitman et al. 1997) In contrast to Central America the agents of large scale disturbance have not been a topic of frequent stud y in Bolivian dry forests (but see Pinard and Huffman 1997) As hurricanes are absent in this landlocked country it is likely that forest fires (both natural and anthropogenic) have likely been the most pervasive influence on Bolivian forests Natural fires have historicall y influenced va st areas of Amazonian forest (Clark and Uhl 1987) particularl y in dry or deciduous forests where d ry fuels ma y fa v or lightning fires (Middleton et al. 1997) In fact most radiometric dates of charcoal found throughout the Amazon correspond with the expansion of dry forests during the dry glacial epochs (Saldariagga et al. 1986 Goldammer 1993 Prado and Gibbs 1993) As is typical in most areas of the tropics humans likely have been the most common agents of forest fires in Bolivia Although most tropical fires are set intentionally b y humans for the purposes of forest conversion traditional slash-and-bum agriculture or grazing land management man y of these intentionally set fires escape (Uhl and Buschbacher 1985 Sarre and

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Goldhammer 1996 Holdsworth and Uhl 1997 Cochrane et al. 1999) Consequently, human caused fires presently contribute more to tropical fire regimes than natural fires (Feams i de 1990 Goldamm er 19 93. Cochrane and Schultze 1998 Nepstad e t al. 1998 I 999 Goldammer 1999) And it is likely this was true historically as well as human population densities in South America have recovered onl y in this century to densities present before Europeans arrived (Denevan 1976). 7 Recent evidence from Bolivia reveals the susceptibility of seasonally dry forests to escaped human-ignited fires Over 1 million hectares of Bolivian dry forests burned during a severe dry season in 1994 (Pinard et al. 1999) and over 3 million hectares burned in one month in 1999 (T. Fredericksen, personal communication) Evidence also suggests that dry forests are damaged less b y wildfire than moister forest types (Mostacedo et al. 1999) which may also be indicative of the pervasive role fire has played in the formation of these dry forests The Potential of Prescribed Burning in the Management of Bolivian Dry Forests Although most guidelines for natural forest management focus en ways of reducing damag e to residual stands (He i nrich 1995 Pinard and Putz 1996 Haworth 1999) low-impact selective logging may not be a sustainable management strategy in dry forests because of the low levels of disturbance associated with this harvesting technique In Lomerio for example roads and skid trails covered only 2-4% oflogged sites and felling gaps were generally only 40-70 m 2 ha1 after harvesting operations (Camacho 1996) Likely, this damage does not create sufficient canop y opening for the regeneration of commercial tree species 12 of 19 of which were classified as having shade intolerant regeneration (Table 1-1 ) Due to the pervasive influence fire has likely had on the formation and maintenance of seasonally dry forests in Bolivia, prescribed bums are a promising silvicultural tool for managed dry forests Prescribed bums produce several effects that will likely increase regeneration of shade-intolerant tree species including vegetation removal mineral soil exposure and nutrient release (Hungerford et al. 1990 Bond and van Wilgen 1996) The use of prescribed burmng in

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8 tropical forest management is not a new idea Ground fires were used as early as the midl 800s to enhance teak (Tectona grandis) regeneration in deciduous forests of South-east Asia (Dawkins and Philip 1998) Tropical forest managers have recognized the benefits of prescribed bums for several shade-intolerant timber species in addition to teak such as sal (Shorea robusta) and several pine species (Pinus ; Goldammer 1994 Rodriquez 1996) In South America however, the use of prescribed burning to enhance tree regeneration in broad-leaf forests is rarely practiced. If fire is addressed in forest management policies, it is primarily in the context of fire prevention or exclusion from protected areas (e g Reis 1996, New Forestry Law in BOLFOR 1997). Particularly in Bolivia the techniques of prescribed burning are not well developed and the effects of prescribed bums on dry forest structure and function are not well known Scope of Dissertati o n The overall objective of this dissertation is to examine the ecological potential of prescribed burning for the management of seasonally dry forests in eastern Bolivia. To be a viable management strategy for the certified forests of Lomerio, prescribed bums must enhance regeneration of commercial tree species without causing irreversible damage to the residual forest. The negative effects of prescribed burns are likely to increase with increasing fire intensity (reviewed in Chapter 2). Therefore, in this dissertation, I compare the effects of harvesting gap fonnation, and controlled bums of high and low intensities on commercial tree regeneration forest soils, and plant diversity The dissertation contains seven chapters. In the second chapter, I briefly review the effects of fire intensity on plants and soils, introduce the study forest, and describe four treatments that form the basis of Chapters 3, 4 and 5 The treatments represent the following four disturbance intensities : harvesting gap formation, above-ground biomass removal and controlled bums of low and high intensity

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In the third chapter I examine changes in soil physical and chemical properties following the four treatments I address the mechanisms underlying these changes by examining experimentally the separate effects of heat and ash on soil properties I also discuss how each of these treatments through their effects on soil properties may influence tree seedling growth 9 Commercial tree establishment, growth and survival in each of the four treatments is evaluated in Chapter 4. As species' responses to disturbance often vary among regeneration guilds I discuss the effects of each treatment by illustrating how they affect each guild differently. I relate these results to the different management strategies that are appropriate for different species groups. The effects of silvicultural treatments are primarily aimed at enhancing regeneration of commercial tree species Yet the impacts of these treatments on the remaining plant community are also of concern particularly in Lomerio where the local indigeno~s population depends on the forest for a variety of other uses In Chapter 5 I examine the response of the plant community to the four treatments focusing on changes in the dominance of species life forms and regeneration modes (seedlings or sprouts) an1ong treatments. I discuss these patterns in relation to their importance for commercial tree regeneration. The studies presented in Chapters 2 though 5 represent patterns of regeneration over an 18 month period following the treatments. In the sixth chapter I examine patterns of regeneration following bums over longer time scales using a chronosequence of secondary forests I characterize tree population structures, stand structure and species richness in abandoned slash and-bum fields of 12 different ages, ranging from 1 to 50 years Comparing these secondary forests to a nearby mature forest stand I discuss the possibility that the dominance of shade intolerant trees in this region may be the legacy of slash-and-bum agriculture In the final chapter I summarize the chapters and discuss the ecological potential of prescribed bums for the management of Lomerio forests I discuss how prescribed burning might

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fit into the current idea of natural forest management in the tropics I also raise several questions of the economical and political constraints to implementing controlled bums on a management scale in Bolivia 10

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CHAPTER2 STUDY SITE AND TREATMENT DESCRIPTIONS Introduction Forest disturbances vary widely in their type, intensity, frequency and scale (Pickett and White 1985) Despite this variation disturbances by definition, hold their most important character in common : they reduce the dominance of a site by established individuals and create openings for colonization and growth by new individuals (Canham and Marks 1985) As such disturbances are the primary catalyst of forest stand dynamics (Oliver and Larson 1996) After most forest disturbances there is a temporary increase the availability of light water and nutrients There are at least two mechanisms by which forest disturbance may increase the availability of these resources The first is the reduction in rates of uptake or use of resources due to the loss of plant biomass This effect is most apparent in the enhancement of light levels in canopy openings (Chazdon and Fletcher 1984) and increased soil moisture in gaps (Vitousek and Denslow 1987) Disturbances may also increase resource availability indirectly by altering rates and pathways of nutrient cycling For example increased soil moisture and temperatures following large-scale windthrow may temporarily increase nutrient availability by increasing the rate of decomposition of soil organic matter (Bormann and Likens 1979) Fire is an increasingly common disturbance in tropical forests (e.g. Goldammer 1993 Bond and van Wilgen 1996 Fearnside 1990 Sarre and Goldammer 1996) A feature of fire that may set it apart from other disturbances is its effect below-ground Fire acts as a rapid decomposer returning some nutrients from above-ground biomass to soil more rapidly than other disturbances (e.g Humphre y s and Craig 1981 Hungerford et al. 1990 Neary e t al. 199 9 ). The 11

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12 usually vigorous growth of seedlings in burned areas is often attributed to fertilization b y deposited ash and increased mineralization due to soil heating (DeBano et al. 1977) Yet removal of above ground biomass can also be far more complete after fires than after other disturbances such as canopy gap formation As such it is likely a combination of aboveand below-ground effects that make fire a promising management tool for tree species with shade-intolerant regeneration (Hungerford et a l. 1990 Bond and van Wilgen 1996). The effects of forest fires on aboveand below-ground processes may vary w idel y depending on their intensity (Moreno and Oechel 1994 Bond and van Wilgen 1996) For example low-intensity fires ma y have a positive effect on regeneration by increasing available soil nutrients (DeBano et al 1977 Wright and Baile y 1982) and stimulating flowering (Whelan 1994 LeMaitre and Brown 1992) resprouting (Zedler et al. 1 983 Moreno and Oechel 1994) and germination of buried seeds (Bradstock and Auld 1995 Schimmel and Granstrom 1996 Enwright et al 1997) In contrast high-intensity fires may be detrimental to regeneration by volatilizing nutrients (Wright and Bailey 1982) altering soil properties such as te::>..'tllre cation-e x chan g e capaci ty, and water holding capacity (DeBano et al 1977 Hungerford e t al. 1990 ) killin g buried seeds (Schimmel and Granstrom 1996) killing species that would otherwise resprout (Moreno and Oechel 1994 ) and damaging or killing potential seed trees. Although there is generall y a positive relationship between the size or intensity of disturbance and the subsequent availability of resources for plant growth (Canham and Marks 1985) this pattern may not appl y for fires of extreme severity Regeneration on sites of low-intensity fires ma y be enhanced w hile areas of high intensity fire ma y be very slow to recover I designed an experiment to address the relative importance of canop y opening above ground biomass removal and controlled bums of high and low intensity on earl y patterns of tropical dry forest regeneration Using these experiments I examined soil ph ysi cal and chemical properties (Chapter 3) establishment and growth of commercial tree seedlings (Chapter 4), and

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13 changes in the plant community (Chapter 5), comparing each response to conditions in undisturbed forest understories In this chapter I introduce the study site and describe the treatments Study Site The studies described in this dissertation were conducted in the seasonally dry forests of Chiquitania a geographic region in the eastern lowlands of Bolivia located in the Provinc e ofNuflo de Chavez Department of Santa Cruz (l 6 45 S 61 45 W ; Figure 2-1 ) Chiquitania is situated in a transition zone between the humid forests on the southern rim of the Amazon basin and the thorn scrub formations of the Gran Chaco The natural vegetation is classified as tropical dry forest (sensu Holdridge et al. 1967) The regional climate is characterized by pronounced seasonality with a strong dry season that corresponds to the austral winter (Figure 2-2) Most of the canopy trees are seasonally deciduous, shedding their leaves from June to September. The mean annual temperature at Concepcion is 24.3 C with temperatures that vary between 3 (July) and 38 1 C (October Killeen et al. 1990) The mean annual precipitation is 1129 mm and interannual variability is large with lows having reached 500 mm and highs 1717 mm per year (Killeen et al. 1990) The landscape is dominated by low hills composed of granite gneiss and metamorphic rocks of Precambrian origin (Geobold 1981 in Killeen et al 1990) punctuated by exposed granitic outcrops (inselbergs) The soils of the area are classified as Inceptisols (suborder : Tropepts group : Ustropepts) and Oxisols (suborder : Usto x, group : Eutrusox ; lppore 1996). Elevation varies between 400 and 600 m a s l. Canopies of mature forest range from 12-18 m tall and are dominated by trees of the Leguminosae (60% of total basal area of trees> 10 cm dbh) ; trees in the families Bignoniaceae Anacardiaceae and Bombacaceae are also abundant (Killeen et al 1998) Understory trees are mostly represented by the families Sapindaceae and Myrtaceae A spiny ground bromeliad Pseudananas sagenarius is distributed over approximately 80% of the forest and occurs in clumps up to 2000 m 2 (MacDonald et al. 1998)

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14 Las T rancas San Lorenzo Lomerio Figure 2-1. Location of the study site in the seasonally dry forests of Chiquitania, a geographic region in the eastern lowlands of Bolivia located in the Province of Nuflo de Chavez, Department of Santa Cruz (l6 45'S 61 W) In the enlarged area, points mark the 28 communities in the political region of Lomerio

PAGE 23

6' u "-' Q) I-, ::, ..... cd I-, Q) 0.. E Q) ..... ;;,-., :2 ..... C 0 E C cd Q) ::E _._ temperature j:::: ,:,:::::::,:,, ,: j precipitation 50 -,-------~------,-------,------------.------. 500 40 30 20 10 400 300 200 100 '93 '94 '95 '96 Figure 2-2 Mean monthly temperature and monthly precipitation at Concepcion Santa Cruz (16' S, 62' W, 490 m a s l.), located approximately 100 km from Lomerio 15 E E "-' C 0 .:: cd u Q) Ip.,

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16 Site background Chiquitania is so named for the Chiquitano indigenous people the largest of the lowland indigenous groups in Bolivia with a population of around 72 500 Lomerio where this study was conducted is a political region within Chiquitania made up of 27 Chiquitano communities with a total population of around 5 000 The Chiquitanos ofLomerio have been managing their forests for timber since 1982 with technical and financial support from severai international institutions BOLFOR (Proyecto de Manejo Foresta! Sostenible) a sustainable forest management project with USAID funding began working in Lomerio in 1992 The objective of the current management plan for the forests of Lomerio is to produce timber sustainably while minimizing negative impacts on other biological and physical resources in the forest (Pinard et al. 1999) Forestry operations of the Chiquitano communities were certified as sustainable by the SrnartWood Program of Rainforest Alliance in 1995 The particular forest in which I worked is owned by the Chiquitano community of Las Trancas and situated approximately 12 km northeast of this village The Las Trancas forest contains 400 ha management blocks Las Trancas 94 and Las Trancas 95 so named for the year in which forest inventories were conducted (Figure 2-3) Las Trancas 94 was selectively logged in July-September of 1995 On average 3-10 m 3 ha 1 (2-5 trees ha. 1 ) of timber were extracted from 6 species. Damage to the residual stand was slight with 6% of the residual trees damaged and 2-4% of the area covered by roads and extraction routes (Camacho 1996) Two logging methods were emplo y ed in Las Trancas 95 In 1996 approximately 75% of the area was selectivel y logged In 1997 the remaining 25% of the area was selectively harvested in strips (40 m x 200 m each separated by an unharvested area 60 m wide) All commercially valuable trees were harvested from these strips at a harvesting intensity of 4.4 m 3 ha 1 Log extraction routes (skid trails) entered each logging strip 100 m from the north and 100 m from the south

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. : : . /. .. : : . : : . . : 17 Las Trancas '94 scale: 1 km scale: 1 km Figure 2-3. The '94 and '95 management blocks of the Las Trancas community owned forest. The enlarged section of Las Trancas '95 represents the trial area for the strip shelter wood system The squares within this enlarged area mark the 16 blocks (20 x 20 m gap areas) depicted in Figure 2-4

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18 Methods Location of felling gaps The studies presented in Chapter 3, 4 and 5 were conducted in the selectively harvested strips in Las Trancas '95 These strips were the only area oflogging activity during the dry season of 1997 and therefore all newly created felling gaps were located in these strips. In June of 1997 I located 16 felling gaps for study (Figure 2-3). Gap selection was restricted by the following criteria : canopy gap area between 200-600 m 2 slopes no greater than 15 less than 20% rock outcrops no trees> 40 cm DBH within gap area and not located in the path of skid trails I located and marked the center of each gap where the midpoints of two perpendicular transects intersected, th e first running the length of the longest axis Each gap was divided into four 10 x 10 m plots by cardinal axes from the center point (Figure 2-4). Half-meter wide paths around the perimeter of the gap and along axes were cut by machete Existing gap area was enlarged to a uniform 20 m x 20 m area by cutting all vegetation >2 m tall (sensu Brokaw 1985a) by machete or chainsaw Because this forest is a timber management area commercial tree species> 20 cm DBH located within the 20 m x 20 m gap area were left uncut (this occurred in only 4 of 16 gaps and standing trees did not exceed 25 cm DBH) Gap Treatments One of four treatments was randomly assigned to each 10 x l O m plot within each block: 1) high-intensity bum ; 2) low-intensity burn ; 3) plant and coarse debris removal (hereafter referred to as plant removal) ; and 4) canopy gap with vegetation> 2 m ta11 cut (the gap control) Other than cutting all vegetation > 2 m tall vegetation and woody debris in the gap control was not manipulated In the plant removal and low-intensity bum treatments all vegetation was cut at or near the soil surface and everything ~2 5 cm diameter(~ 100 hour fuels) was removed and distributed as evenly as possible in the high-intensity bum treatment. Tree trunks and large diameter branches were sawn into smaller sections so that they could be moved more easily and dry

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19 .... .... __ : I I : forest sub~----__ J ---__ __ /, forest plot ~ 15-20 m l 10m 20m 20m Figure 2-4 A single block consisting of a 20 x 20 m felling gap and an adjacent forest plot. Each felling gap was equally partitioned into four 10 x 10 m treatment plots Each treatment plot was randomly assigned one of four treatments : gap control, plant removal low intens i ty or high intensity bum Within each treatment plot, 2 permanent vegetation sampling subplots ( 2 x 2 each ) were located near the g ap center and 2 additional subplots near the gap edge All soil sampl in g was conducted outside of these permanent sampling plots Two permanent subplots were located approximately 15-20 m from the edge of each gap in undisturbed forest. Soil samplin g in the forest was conducted outside but within 5 m of these permanent forest subplots Hereafter the 400 m 2 felling gap and adjacent forest site are blocks ," the 100 m 2 treatment areas are plots ," and the 4 m 2 vegetation sampling areas are subplots

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20 rapidly Pseudananas sagenarius and cacti were not added to the high-intensity bum treatment but instead were removed from the block altogether because of their low flammability Therefore after fuels were manipulated and before prescribed bums the plant removal and low-intensity bum treatments had similar amounts of litter and woody debris and no above-ground vegetation The high-intensity bum treatment plots had roughly 3 times its original fuel load Slash was left for 5 rainless weeks to dry before prescribed bums were conducted Controlled burns Fuel loads. Pre-bum fuel loads in the lowand high-intensity bum treatments were measured in randoml y located 0 25 m 2 circular subplots, with 2 subplots sampled per plot x intensity treatment (2 fuel plots x 2 bum treatments x 16 plots). All fuel within each subplot was removed divided into fuel diameter size classes (live herbs <6 mm, 6-25 mm 25-75 mm, and > 75 mm) and weighed in the field. Composite subsamples of each fuel size class were taken from the field oven dried to constant weight and used to calculate the wet-to-dry weight conversion factors The diameter and length of trunks and large diameter branches that could not be weighed in the field were measured in order to estimate volumes Wood densities available from BOLFOR were used for volume to mass conversions Timing of burns. Although little is known about the historic fire regime in these forests seasonal patterns in rainfall and relative humidity make wildfires most likely at the end of the dry season when fuels are dry and lightning strikes most common The indigenous Cbiquitano population traditionall y bum their agricultural fields and cattle pastures at the end of the dry season as well, shortly before the onset of rains. Predictably, most escaped fires occur during this season Because one of the objectives of my experimental burns was to enhance seedling establishment of commercial tree species I planned a burning date in late August or early September at the end of the dry season and before peak seed fall of most commercial trees (Justiniano 1997)

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Fire breaks. All fuel was removed from a I m wide fire break around low-intensi ty bum treatment plots Fire breaks around high-intensity bum treatment plots were 1 to 2 m wide wider where danger of fire escape was perceived to be higher Standing dead trees near firebreaks were felled and ladder fuels such as liana tangles were removed. On the day of bums firebreaks were raked free of newl y fallen leaves 21 Prescribed burns Prescribed bums were conducted from August 29 to September 1 1997 near the end of the 5 month dry season (Table 2-1). Each day the earliest bums were started at 10 : 00 a m and the last bums by 3 : 00 p m Temperature at 10 : 00 a m over the four da y period varied from 34 to 36.4 C and relative humidity varied from 29-38 %. Winds were v ariable but usually calm in the morning with convectional wind gusts of up to 11 km/hr in the afternoon A circular ignition technique was used for both bum treatments A spot fire was lit with a drip torch in the plot center then the perimeter was lit starting with the downwind side. The center fires created convection which drew the ring fire on the borders inward In the low-intensity bum treatment plots ring fires often did not carry to the center therefore spot fires were ignited where needed A minimum of 5 people conducted the bums over the 4 day period. At least one person with a backpack water sprayer remained at each fire until fires near the borders were extinguished Fires near firebreaks or standing dead trees were extinguished before burning crews returned to camp Each fire was checked again after dark and the following morning to extinguish an y potentially dangerous smoldering areas Maximum soil temperatures, fire intensities and completeness of burns. Maximum soil temperature and an index of fire intensity were measured in two locations in each burn plot near the two subplots w her e fuel loads were measured Maximum soil temperature was measured at 0 and 3 cm depth using temperature indicating paints (Tempilaq Tempi! Divis i on Air Liquide

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22 Table 2-l. Climatic conditions at 10 : 00 a m. the morning of high and low intensity bums for 16 experimental blocks. Block Date Ambient Relative Wind speed burned temQerature ( C) humidity(%) (km/hr) 1 29-Aug 36 29 0 with gusts 19 29-Aug 36 29 0 with gusts 20 29-Aug 36 29 0 with gusts 4 30-Aug 34 34 11 6 30-Aug 34 34 11 8 30-Aug 34 34 11 21 30-Aug 34 34 11 22 30-Aug 34 34 11 2 31-Aug 34 37 4 7 31-Aug 34 37 4 9 31-Aug 34 37 4 11 01-Sep 35 38 0 with gusts 14 01-Sep 35 38 0 with gusts 17 01-Sep 35 38 0 with gusts 18 01-Sep 35 38 0 with gusts

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23 America Corporation South Plainfield New Jersey USA) Paints of24 different melting points ranging from 66 to 1093 C were applied as narrow bands on 2 x 30 cm steel strips At each location one painted steel strip was buried at 3 cm soil depth and another placed flat on the soil surface directly above it. Soil temperatures were measured to a greater depth in one block Here an additional 3 sets of 4 painted strips were placed at 0 1 3 and 7 cm depths After fires the highest indicated melting point was recorded Fire intensity was estimated by Beaufait' s ( 1966) technique which calculates total energy output from the amount of water vaporized from cans during burns as : total energy output= [(80 cal/g water) x (g water)] + [(540 cal/g water) x (g water)] Wh e r e 80 ca l a r c n ee d e d t o r aise e ach gram of water from 20 C to the boiling point and 540 cal are needed vaporize each gram of water (latent heat of vaporization) Two tin cans per bum were used, each placed on the soil surface of fuel load subplots. Depth of water was measured immediately before each bum and within 24 hours after To account for the amount of water lost due to evaporation 2 cans were placed in the center of an unburned gap and the amount of water evaporated within 24-hrs measured Soil moisture which influences heat movement through soil was measured several hours befo e burns Soil samples from 0-5 and 5-10 cm depths were collected from each plot weighed, oven dried to a constant weight and moisture content expressed as % of soil dry weight. The week following bums completeness-of-bum was estimated visually as the percent area burned Establishment of permanent vegetation plots Three weeks following bums 4 permanent subplots (2 x 2 m each) were established in each treatment plot 2 located near the gap center and 2 located near the gap edge (Figure 2-4) Two additional subplots were established at random points in undisturbed forest 15-20 m from the edge of each gap. These permanent subplots were used for sampling commercial tree seedling

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establishment (Chapter 4) and vegetative cover (Chapter 5) One plot of each pair was used for a seeding treatment described in Chapter 4 Treatment effects on canopy cover and microhabitat 24 Soil temperature to 3 cm depth was measured at a center and edge subplot of each treatment as well as the forest subplots with a soil thermometer 3 and 6 months following bums Percent canopy cover was measured with a spherical densiometer above each gap center gap edge and forest plot 3 months following bums Litter depth (cm) and percent cover by debris 2-20 cm and >20 cm diameter were estimated visually for each of the permanent 4 m 2 subplots 6 weeks following bums Results were analyzed using an analysis of variance with treatment as a fixed effect and block as a random effect followed by Tukey s post-hoc comparisons Initial Treatment Results Pre-burn fuel loads Pre-bum fuel loads in high-intensity bum treatment subplots ranged from 10.8 to 82.8 kg/m 2 and averaged 48 4.9 kg/m 2 (mean 1 standard error ; Figure 2-5) Almost half of this mass was comprised of fuels >7 5 cm diameter. Fuel loads in the low-intensity bum treatment subplots ranged from 0.8 to 4 kg/m 2 and averaged 2.2 2 3 kg/m 2 Sixty-six percent of the fuel mass in low-intensity plots was fine fuel <6 mm diameter. Burn characteristics High-intensity burns. Completeness of high-intensity bums was variable but the majority of bums consumed all but the thickest (> 20 cm diameter) branches and trunks. Flame heights ranged from 1.5 to 5 m Fire intensities ranged from 152 to 3795 kcal and averaged 1627 241 kcal (n = 15) Temperature at the soil surface during high-intensity burns averaged 704 42 C (n = 16) The highest temperature measured was 927C. Temperature at 3 cm depth averaged 227 27 C (n = 16) Where maximum temperature was measured at additional depths of 1 and 7

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25 Temperature (C) 2 40 0 0 Fire intensity (kcal soil surface Fuel load (kg/m ) 3 cm depth 100 10 00 8 0 3000 75 0 60 T I 2000 5 0 0 40 i 1000 250 20 ......... --+0 0 0 High Low High Low High Low High Low Figure 2-5. Pre-bum fuel loads fire intensities and maximum temperatures at the soil surface and 3 cm depth during high and low intensit y bums Bo x plots show medians (c e nter line) 25th and 75th percentiles (top and bottom lines) 10th and 90th percentiles (top and bottom w hiskers) and points g reater than the 90th percentile and less than the 19th percentile (dots)

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26 cm temperatures averaged 871 C at the soil surface 358 C at 1 cm depth 218 C at 3 cm depth and 135 Cat 7 cm depth (n = 2) Although visible flames were extinguished b y nightfall some logs continued to smolder for several da y s : fire intensities and soil temperatures under these logs were likely greater than measured values Low-intensity b u rns. In general completeness of low-intensity burns was more va riable than high-intensity bums. Flame heights were low ranging from 10 to 50 cm Fire intensity ranged from 22 to 68 kcal and averaged 41 3 kcal (n = 15) Temperatures at the soil surface averaged 225 33 C (n = 12) ; the highest temperature measured was 413C. Elevated temperatures at 3 cm were only detected in 2 of 16 plots ; these averaged 107 7 C (n = 2) Soil moisture on the day of bums was low and did not differ between the highand low-intensity bum plots (0-5 cm depth : P = 0.94 5-10 cm depth : P = 0 23) Therefore differences among the 2 bum treatments in heat conductivity due to soil moisture were likely negligible and are hereafter ignored. Treatment effects on microhabitat Treatments had significant effects on the amount of soil exposed mid-day ambient soil temperature litter depth and area covered by woody debris (Figure 2-6) Canopy cover above forest plots was 78% higher than canopy cover above all 4 gap treatments which averaged 22% (P < 0 001) Although canopy cover above gap-center and gap-edge plots was not significantly different (P = 0 6) soil temperatures in gap centers were higher than near gap edges (P < 0 001) A maximum temperature of 43 C was recorded in the center of one high-intensity bum treatment 3 months following burns After 6 months soil temperatures at gap centers and edges were not different (P = 0.52) and only soil temperatures in the high-intensity bum treatment were significantly higher than the other treatments (P < 0 00 l) High-intensi ty bums removed all litter and deposited a la y er of ash ranging from 0-14 cm depth (4 8 0 2 cm n = 16) Not all woody material was consumed in the high-intensity bums ;

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6 5 ,--... E 4 (.) ,_ ..c: 3 ...... 0.. 8 0 15 .... c:: 0 0 15 (.) .... c:: 20 cm P < 0.001 b a a d C a I b b a 0 40 e Soil temperature P < 0 001 '--' 20 cm diameter and mid-day soil temperature in four gap treatment plots and forest plots 6 weeks after bums. Treatments with the same letter are not significantly different. 27

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28 the remaining woody debris covered approximately 12% of the subplots areas In the low-intensity burn treatment an average of 76% of the subplot areas burned to some degree Burning w as not complete even within these areas as only an average of30% of the subplot areas had soil exposed Small woody debris remained on approximately 3 % of the area of low-intensity burn subplots ; most large w ood y debris had been removed before burning Leaf litter or small woody debris covered all of the plant removal treatment subplots with no bar e soil ex posed As with the low-intensit y bum treatment most large wood y debris w as removed Gap controls were characterized b y deep leaf litter (2.9 0 2 cm n = 16) and small and large woody debris covering an average of 25% of the subplot areas Only 20 % of the gap con t rol subplot areas were d e void of either wood y material or surviving plants Forest understories had the deepest leaflitter (4 0 0 3 cm n = 16) but small and large wood y debris combined co v ered an average of onl y 7% of the subplot areas

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CHAPTER3 EFFECTS OF CANOPY GAP FORMATION PLANT REMOVAL AND CONTROLLED BURNS OF I-IlGH AND LOW INTENSffiES ON SOIL CHEMICAL AND PHYSICAL PROPERTIES Introduction Fire is a rapid decomposer ; it compresses the oxidative processes of organic matter deca y into a short time span (Wright and Ba iley 1982) The result is a nutrient pulse much larger than from the normal decomposition of woody debris and litter at least for the first few months following fires (Bond and van Wilgen 1996) As such controlled bums may benefit tree seedling growth more than unburned treatments particularly since the timing of nutrient pulses following fire coincides with maximum light availability After intense fires however the advantages of increased nutrient availability may be offset by degraded soil structure Thus the benefit of controlled bums for tree seedling growth may ultimately depend on fire intensity In this chapter I examine both soil nutrient availability and soil physical properties following canop y opening plant removal and controlled burns of highand low-intensity There are three primary mechanisms of increased nutrient availability following fire : nutrients added to the soil as ash ; heating of soil organic matter ; and increased rates of biological mineralization following fire due to increases in soil pH temperature and moisture as well as due to a reduction in C : N ratios (Wright and Bailey 1982 Pritchett and Fisher 1987) The degree of increase in nutrient availability following fires depends largel y on fire intensity Most studies of low to moderately intense fires report increases in available nutrients (reviews b y Dunn et al. 1977 DeBano et al. 1977 Wells et al. 1979 Humphreys and Craig 1981 Wright and Baile y 1982 29

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Hungerford et al 1990 Neary et al. 1999) In contrast intense fires may cause a net loss of nutrients (DeBano et al 1977 Giovannini et al. 1990) 30 Due to its low temperature of volatilization (200 C ; Weast 1 9 88) nitrogen loss is linked with the consumption of organic matter (e g. Dunn et al. 1977) Where fuels are completel y consumed and the surface la y er of soil organic matter is destroved loss of nitrogen through volatilization can be substantial (e g ., Nye and Greenland 1964 Ewe! et al 1981) Volatilization of phosphorus and cations are usually minor due to the high volatilization temperatures of these minerals (>760 C ; Weast 1988) however their loss from severely burned sites ma y be caused b y surface erosion leaching or transport of ash (Wright and Baile y 1982) Intense burns may also have detrimental effects on soil physical properties b y consuming soil organic matter. Soil organic matter holds sand, silt and clay particles into aggregates therefore a loss of soil organic matter results in a loss of soil structure Severe fires may also permanently alter soil texture by fusing clay particles into sand-sized particles (Dyrness and Youngberg 1957 Ulery and Graham 1993). By altering soil structure and texture severe fires can increase soil bulk density (DeByle 1981 ) and reduce soil porosity water infiltration rates and water holding capacity (e g. Wells et al. 1979). Intense burns may also induce the formation of a water repellent soil layer by forcing hydrophobic substances in litter downward through the soil profile (DeBano 1969) reducing water infiltration rates as a consequence (DeBano 1 971 ). The changes in chemical and physical soil properties caused b y fire potentiall y ha v e important consequences on tree seedling growth (Johnson 1919). Increased nutrient availabilit y after fire may benefit plant growth if nutrients are limiting prior to burning (e g. Hunge1ford et al 1990) On the other hand seedling growth in intensely burned soils ma y be slowed due to high pH and toxic levels of minerals (Giovannini et al. 1990) Altered soil physical properties such as soil strength bulk density and water infiltration rates ma y also impair plant growth Plant uptake of nutrients and water is slowed in structurally degraded soils through the combined effects of lower

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31 soil moisture and lower soil porosity (Nye and Tinker 1977) M;;:chanical impedance of root growth caused by increased bulk density and soil strength (Gerard et al. 1982) also slows nutnent and water uptake In this chapter I focus on the below-ground effects of the treatments described in Chapter 2 My objectives were to : l) compare the effects of canopy gap formation plant removal and controlled bums of high and low intensities on soil nutrient availability soil physical properties and fine root mass ; 2) compare the relative importance of soil heating and ash-fertilizat i on on soil nutrient availability ; and 3) discuss how these treatment-induced changes in soil properties influence tree seedling growth Methods Study site The studies presented in this chapter were conducted in the treatment and forest plots described in Chapter 2 (Figure 2-4) All soil sampling was done within the 100 nt2 plots but outside of the 4 m 2 subplots Forest sampling was done within 5 m of the forest subplots Mass and chemical characteristics of ash deposited in high-intensity burn plots Ash mass deposited in high-intensity bum plots was estimated by collecting and weighing all ash on the soil surface in a 1 m 2 area replicated in three high-intensity burn plots (n = 3) To characterize variability in the amount, ash depth was measured in l 0 randoml located points in each high-intensity bum plot (n = 16) Composite ash samples were then collected from each plot and used to measure pH and nutrient concentrations. Ash pH was determined as for soil pH described below To determine nutrient concentrations, 0.5 g of ash was heated in 10 rnL of 1 M HN03 and then resolubilized in 10 rnL of 1 M HCl. Extracts were then analyzed for phosphorus potassium calcium and magnesium at the Analytical Research Laboratory at the University of Florida with an ICAP Spectrometer (Thermo-Jarrell Ash Corporation Franklin MA)

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32 Soil sampling Soil samples from 0-8 cm depth were collected 2 6 9 12 and 18 months after burns and from 8-20 cm depth after 9 12 and 18 months These samples were used to assess moisture contents pH organic matter and e>..1ractable elements In each treatment and forest plot (n = 16 blocks), 4 samples were taken from randomly selected sites with a 10 cm diameter cylindrical corer The 4 samples from each treatment were mixed thoroughly in the field and a ~300 g composite subsample bagged (Anderson and Ingram 1993) In one block the 3-month samples were bagged separately, rather than composited to examine intra-treatment variability Subsample soil volume was unknown, therefore bulk density and fine root mass were sampled separately as described below. Soil pH and air-dry moisture content The pH of fresh soil was determined by adding 50 ml of distilled water to 20 g of soil and stirring for 10 minutes (Anderson and Ingram 1993) The mixture then stood for 30 minutes and pH of the supernatant was measured with a hand-held meter (Oakton pHTestr 3) Soil samples were then weighed air-dried to a constant weight and reweighed to calculate air-dry moisture content. Air-dried samples were passed through a 2 mm sieve, bagged and stored in a cool dry area until transported back to the Analytical Research Lab at the University of Florida for chemical analyses Soil chemical analyses Phosphorus potassium, calcium and magnesium were extracted with Mehlich-1 solution : 0.05 M HCl and 0.0125 M H 2 SO4 (Hanlon et al. 1994). Extracts were then analyzed b y ICAP spectroscopy Soil organic matter content was analyzed using the Walkley-Black dichromate methodology (Hanlon et al. 1994) A subset of soil samples was tested for total nitrogen using an elemental analyzer (Carloerba NCS 2500) Twenty-four samples from all treatments and sampling periods from the top 8 cm of soil were selected to represent the full range of organic matter content.

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33 Resin-available nitrogen and phosphorus Resin-available nitrogen (NI-4 +N and NO 3 --N) and phosphorus (PO/-P) in each treatment were estimated by burying anion and cation exchange resin bags at 5 cm depth Resin bags were prepared by enclosing 5 0 g (moist weight) of either anion exchange resins (Sigma Dowex) or cation exchange resins (Fisher Scientific) in bags of nylon stocking material sewn closed with nylon thread Before bmial resin bags were hydrated overnight with dionized water. Four bags of each resin type were buried per treatment plot ( 4 bags x 2 resin types x 5 treatments x 16 blocks) Three rotations of resins were buried each for approximately 3 months Two rotations included the first and second rainy seasons following bums (November 1997-January 1998 and December 1998-February 1999, respectively) The middle rotation covered the transition from the first rainy season to the first dry season following bums (May-Jul y 1998) After removal from the field resin bags were placed separately in clean plastic bags and kept cool (refrigerated when possible) until transported to the University of Florida for analysis For each resin type the 4 bags per plot were pooled and 12 g ex1racted in 120 ml of 2 M KCl for 24 hours Extractions were analyzed for amonium-N and nitrate-N using automated spectrophotometry (Flow IV Ion Analyzer AlpKem (O-1-Analytical) College Station TX) Extracts from anion exchange resins were diluted to l M KCl and analyzed for PO 4 3 -P using the atomic emission spectrometric method (Thermo-Jarrell Ash Corp Franklin MA) Statistical analyses Soil nutrient concentrations organic matter, pH, moisture content and resin-available N and P were analyzed using an ANOVA with repeated measures Treatment was a fixed effect and block a random effect in each model. Soil properties were log transformed for anal yse s when not normally distributed, but all values presented in the text are non-transformed. Where a significant time x treatment interaction was found variables were analyzed separately b y month. Statisticall y significant differences (P < 0 05) were further analyzed with Tuke y' s HSD multiple comparisons

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34 In order to describe variation within plots 4 soil samples per treatment in one block (block 4) were analyzed separately for extractable nutrients and organic matter Also 4 resin bags per treatment in one block were extracted and analyzed separately for resin-available nitrogen and phosphorus Coefficients of variation (Sokal and Rohlf 1981) were calculated to compare variation of soil sampled within the same 100 m 2 plot and among the 16 different plots Fine root mass Fine root mass (roots < 2 mm diameter) was compared among treatments 12 months following burns in a reduced sample of 10 blocks (n = 10). Soil cores were extracted with a cylindrical tube (5 cm inside diameter 7 cm deep) from 3 randomly located points in each treatment and forest plot. Fine roots were sorted from samples dried and weighed Fine root mass (live and dead combined) was compared among treatments using an AN OVA followed b y a Tukey s HSD post-hoc test. Soil bulk density Soil bulk density (air-dry) was estimated 6 and 12 months following burns in a reduced sample of 10 blocks (n = 10) Three samples in each treatment and forest plot were collected using metal cans (5 cm inside diameter 7 cm deep 137 cm 3 ). Samples were air-dried to constant weight and bulk densit y calculated as : bulk density (g/cm 3 ) = g air-dried soil/ 137 cm 3 Differences among treatments were tested using an ANOVA on square-transformed values of bulk density with treatment and month as fixed effects and blocks as a random eftect followed by Tukey s HSD multiple comparisons Soil strength Compressive soil strength was estimated with a pocket penetrometer (Forestry Suppliers) at 2 6 9 and 12 months following burns Soil strength readings were taken at 4 randomly selected points in each treatment and forest plot from all 16 blocks (n = 16) Soil strength was

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analyzed using AN OVA with repeated measures as described above for soil chemical properties Water infiltration 35 Water infiltration rates were estimated in a reduced sample of 4 blocks 8 months following burns (n = 4) The technique used here was a modified version of the single ring method (Anderson and Ingram 1993) Although double ring methods provide better estimates of infiltration rates because they compensate for lateral flow a single ring method was chosen for this study because it used less water (which had to be transported 27 km) In each gap treatment and forest site a point was randomly located and cleared of surface litter. A graduated PCV cylinder (10 cm diameter 25 cm length) was inserted vertically into the soil IO cm deep and soil pressed around the base to minimize water leakage The cylinder was filled with water to IO cm and timed until the water level dropped to 5 cm This process was repeated three times Infiltration rates were calculated separately for each repetition (i e., the first second and third 5 cm increments of water which correspond to 5 10, and 15 ml cm 2 ) as the volume flux of water flowing into the soil profile per unit surface area (Hillel 1982) and expressed as ml cm 2 sec 1 Log transformed infiltration rates were compared among treatments using an ANOVA with treatments and repetitions (i e each 5 cm increment) as fixed effects and blocks as random effects Soil wettability Soil wettability was estimated using a modification of the water drop penetration time method (WDPT ; Letey 1969) in a reduced sample of7 blocks 8 months following burns (n = 7) In each gap treatment and forest site four 20 x 20 cm areas were randomly located and cleared of surface litter. Five drops of water were placed on the soil surface with a dropper and the time recorded when all 5 drops were completely absorbed This was repeated at 1 2 and 3 cm soil depth by scraping surface soil away with a machete Soil wettability (log transformed seconds) was compared among treatments using an ANOVA with treatments and soil depth as fixed effects and blocks as random effects

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36 Comparative effects of soil heating and ash addition on soil chemical properties To compare the effects of soil heating and ash addition on soil chemical properties I carried out a 2 x 3 factorial experiment with two levels of ash addition (no ash and ash added) with three levels of soil heating (no heat, low-intensity heat and high-intensity heat) The first trial was conducted in the field in Las Trancas 95 The second trial was conducted in the BOLFOR greenhouse in the city of Santa Cruz Anadenanthera colubrina served as a bioassay in the field experiment. Because Anadenanthera did not fruit in 1998 Caesalpinia pluviosa was used as a bioassay in the greenhouse study. Field study. I utilized the plant removal treatment plots described in Chapter 2 for the field study, conducted at the end of the dry season in October 1997 The design is a complete randomized block ; each plant removal plot was considered as a block (n = 12) In each block six 1 m 2 plots were located in the area between the gap center and edge permanent subplots (Figure 24). Competing vegetation had been cleared from the larger treatment plots the month before but some regrowth had already occurred Therefore the 1 m 2 plots were cleared again of any vegetation and raked of surface litter to expose the soil surface Each plot was randomly assigned a trcatm~nt combination of soil heating (no heat low-intensity heat or high-intensity-heat) and ash (no ash or ash added) Treatments were applied to a 50 x 50 cm area in the center of the 1 m 2 plots, creating a 25 cm buffer along the edge. Heat was applied using a propane blow torch Temperature of the flame was measured with Tempil heat sensitive paints. hi the low-intensity heat treatment, a flame of 150-250C was applied to the soil surface of the treatment area for 5 minutes In the high-intensity heat treatment, a flame of 500-800 C was applied to the soil surface of the treatment area for 20 minutes The torch required constant adjustment to maintain a similar flame, therefore temperatures varied within a treatment. I am confident however that temperatures ranges did not overlap between the highand low-intensity treatments After heat

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37 treatments approximately 500 g of ash collected from high-intensity bum plots was distributed as evenly as possible to plots assigned the ash treatment. The week following treatments l O seeds of Anadenanthera colubrina were placed in each plot and checked for germination after 4 da y s After 2 weeks most seeds were found to have been removed or eaten and were therefore not checked again Soil samples 0-8 cm depth were collected 3 weeks following treatments Soil pH, and phosphorus potassium calcium magnesium and organic matter concentrations were anal y zed using the method s d e scribed abo v e Resin-available nitrogen (NH / -N and N0 3-N) w as measured using anion and cation exchange resin bags One bag of each type was buried at 5 cm depth in each plot for 85 da y s (October 31 January 24) Anal y sis ofresins follows that described for the larger stud y. Greenhouse study. The greenhouse stud y was conducted at the end of the dry season in 1998. Soil used for this trial was collected from La s Trancas 95 to a depth of 10 cm So il was passed through an 8 mm sieve mixed well and divided into three equal batches Each batch was assigned a soil heating treatment (no heat low-intensity heat and high-intensit y heat) In the lo w intensity heat treatment soil was heated in aluminum pots in a conventional oven at 10 0 1 5 0 C for a total of 1 0 m i nut e s (mixing after 5 minutes) Soil in the high-intensi ty treatment was o ve n heat e d at ~ 200 C for 4 0 minutes (mi x ing after 20 minutes) then spread I cm deep on a metal tra y and heated with a blow torch for 5 minutes at a temperature of 500-800 C Oven temperature and torch temperature were both measured using Tempi! heat sensitive paints One compos i te soil sample from each heating treatment (control low-intensity and high-intensity) was analyzed for phosphorus potass i um calcium magnesium and organic matter concentrations Soil from each heating treatment was used to fill 24 plastic planting containers (7 x 25 cm) Twelve planting containers per heating treatment were then selected for the ash addition treatment ( 15 g o f ash

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38 added to the soil surface) and the remaining 12 containers served as controls (n = 12) Two seeds of Caesalpinia pluviosa were placed in each planting container watered daily and seedling height to the terminal bud was measured after 4 months Resin-available nitrogen (NH / -N and N0 3 --N) was measured in each treatment combination using additional planting containers Anion and cation exchange resin bags were buried at 5 cm depth in 3 containers of each treatment combination (n = 3) and watered dail y for 22 days Resins were e>..1:racted and anal y zed using the methods described above Results Mass and chemical characteristics of ash deposited in high-intensity burn plots Variability in amount of deposited ash was high and depths ranged from 0-14 cm ( 4 8 0.3 cm x S E ., n = 16) Ash mass deposited by high-intensity bums averaged 1.5 0 6 kg/m 2 (n = 3) Using this value and measured concentrations of individual elements in ash (Figure 3-1) indicates an average nutrient deposition of 524 g/m 2 of Ca 26 g/m 2 of Mg 83 g/m 2 ofK and 7 7 g/m 2 of P Ash samples had an average pH of 10 7 0 1 (n = 16) Treatment effects on soil nutrients High-intensity bums significantly increased P Mg K and Ca in the top 8 cm of soil but the magnitude and its change over time varied by nutrient (Figure 3-2) These increases were also detected at 8-20 cm for all elements except Mg (Figure 3-3) Low-intensity bums also significantly increased P Mg K, and Ca in the top 8 cm of soil although increases were smaller than in high-intensity bum plots did not persist as long and were not detected at 8-20 cm Plant removal and gap control treatments had no detectable effect on P, Mg K and Ca at either soil depth Results of statistical analyses are summarized in Tables 3-1 and 3-2

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39 Ca Mg p 500 m ash 50 m I ash 100 K (mg/gash) 10 (m g / g as h ) 400 300 200 100 i . o~--4 0 30 20 10 0 --'-------8 75 T 6 50 4 2 5 2 o~----~ 0 _.__ __ _ Figure 3-1 Box plot diagrams of concentrations of Ca Mg K and P in ash sampled from high intensit y bum plots (n = 9) Bo x plots sho w medians (center line) 2 5th and 75th percentiles (top and bottom lines) 10th and 90th percentiles (top and bottom whiskers) and observations l y ing outside of the 10th to 90th percentiles (dots)

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,,-.., bl) s .._, A-4 ,,-.., bl) s .._, bl) ,,-.., bl) s .._, '@ 5 ro u 100 80 60 40 20 0 500 400 300 200 100 800 600 400 200 0 6000 4000 2000 Q--~--,___ __ ----0----0 -------........:::-~-=--I --=w=-=-=-=-=-----=--:....:1 2 _._ High intensity burn -0Low intensity bum -YPlant removal ---'V"Gap control Forest 6 9 12 18 Months after treatment Figure 3-2 Extractable soil concentrations of P Mg Ca and Kin soil samples (08 cm depth) in four gap treatments and forest sites at 5 sampling times over an 18 month period following bums (bars= S.E ) 40

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41 100 80 ,--... 0.0 60 0.0 6 4 0 ..._; 20 I 0 500 ,--... 400 bl} 300 s ..._.., bl} 200 ::E 100 800 ---High intensi ty bum -(>Lo w intensi ty bum ,--... 600 --vPlant remo va l 0.0 400 s -SQGap control Fores t ..._; 200 ~--==9=-= -< 111 < 11 I _,. ~ 0 I 6000 4000 8 2000 ~.;::;::;;:;;. .. .. . 777". .... . .. . 9 12 18 Months after treatment Figure 3-3 Extractable P Mg K and Ca in soil sampled from 8-20 cm in four gap treatm e nts and forest sites at 9 12 and 18 months following bums Y-axis scales ar e identical to those in Figure 3 2 for soils san1pled from 0-8 cm depth (bars = S.E.)

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42 Table 3-1 Results of ANOV As of soil nutrients organic matter water content and soil pH of soil sampled 0-8 cm in four gap treatments and forest plots at 5 times following bums. All variables were log transformed Where a significant time treatment interaction was found, variables were analyzed separately by month Treatments with different letters are significantly different at P < 0 05 No interaction time* treatment Post-hoc test results Variable Factor F p Month high low rem ova control forest Magnesium Treatment 23 3 < 0.001 3 a b e e e Time 8 7 < 0.001 6 a b be e e 9 a b be C C 12 a b C e C 18 a b b b b Calcium Treatment 70.1 < 0 001 3 a b C e e Time 3 0 0 026 6 a b be e be 9 a b b b b 12 a b b b b 18 a a a a a Water content Treatment 3 9 0 008 3 ab a a a b Time 39 0 < 0 001 6 ab ab a ab b 9 a a a a b 12 a a a a a 18 a a a a a Significant time treatment interaction Post-hoc test results Variable Month F p high low r emova control forest Potassium 3 77.4 < 0.001 a b C e C 6 64 1 < 0.001 a be e be b 9 11.5 < 0 001 a be e b be 12 7 8 < 0 001 a b b ab a 18 3 7 0 009 ab b b ab a Phosphorus 3 167 5 < 0 001 a b C e e 6 60.8 < 0 001 a a b b b 9 45 5 < 0 001 a b C C C 12 58 7 < 0.001 a b e C e 18 37 2 < 0 001 a b C C C Organic matter 3 7.3 < 0 001 a b b b b 6 15 9 < 0 001 a b b b b 9 4 6 0 003 a b b ab a b 12 4 7 0.002 a b ab ab b 18 3 9 0 007 ab ab b ab a Soil strength 3 12 6 < 0 001 a b b b b 6 24 3 < 0 001 a a a b C 9 28 8 < 0.001 a b b C C 12 16.9 < 0 001 a be b e C

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4 3 Table 3-2 Results of ANOVAs of soil nutrients organic matter water content and soil pH in the 8-20 cm depth of soil of four gap treatments and forest plots at 3 times following bums All variabl e s were log transformed Where a significant time* treatm e nt int e ract i on w as found v ariables w ere anal y zed separate! b y month Treatments w ith different letters are significantl y different at P < 0 05 No interaction time treatment Post-hoc test results Variable Factor F p Month high low remova control forest Phosphorus Treatment 105.4 < 0 001 9 a b be be be Time 59.4 < 0 001 12 a b be be be 18 a b b b b Magnesium Treatment 2.1 0 09 9 Time 7 3 0 001 12 18 Calcium Treatment 10 5 < 0.001 9 a b b b b Time 4 7 0 013 12 a ab b b b 18 a b ab b b Water content Treatment 5 6 0 001 9 a a a a b Time 51.2 < 0 001 12 18 ab ab a ab b pH Treatment 70 9 < 0 001 9 a b be C be Time 58 9 < 0 001 12 a b C C C 18 a b C C C Significant time treatment interaction Post-hoc test results Variable Month F p high low removal control forest Potassium 9 34 6 < 0 001 a b C be be 12 17.5 < 0 001 a b b b b 18 9 5 < 0 001 a b b b ab Organic matter 9 1.7 0 16 12 2 7 < 0 001 a ab ab b a b 18 11.3 < 0 001 b b a C C

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Total soil N was strongly related to soil carbon (R 2 = 0 93 ; Figure 3-4) thus patterns of total N differences among treatments are expected to follow those for soil organic matter 44 Both highand low-intensity burn treatments significantly increased resin-available NH / N NO 3 -N and PO 4 3 -P during the first rainy season following bums (Table 3-3 Figure 3-5) This pulse decreased after the first rainy season Other than an increase in NO 3 -N in plant removal treatments during the first ra-iny season the remaining treatments had little effect on ~ + -N NO 3N and PO / -P availability. Coefficients of variation (CV) calculated for soil nutrients and organic matter within on e plot and among the 16 plots for each treatment are displa y ed in Table 3-4 A pattern emerged that in the burned treatments soil nutrients and organic matter were more variable within the one plot than among all 16 plots The opposite pattern was true for the plant removal gap control and forest plots Variation was in general greater among the 16 different plots than within the one p l ot. Treatment effects on soil pH, soil organic matter content, and soil water content Soil pH after high-intensity bums at 0-8 and 8-20 cm was higher than in all other treatments throughout the 18 month sampling period (Figure 3-6 and 3-7) Soil pH in high intensity bum plots was 2 pH units higher than forest soils 2 months following burns In low intensity burn treatments pH was higher than in the remaining treatments at both depths at all sampling periods The plant removal and gap control treatments had little effect on soil pH. High-intensity bum treatments significantly lowered soil organic matter ; 2 months following high-intensity bums soil organic matter in the top 8 cm of soil was approximatel y 72% that of forest soils By 18 months soil organic matter recovered to levels comparable to the remaining treatments Diff erences among the remaining treatments were small and varied throughout the sampling periods.

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0 8 0.7 R 2 = 0.93 ,-.. 0 .6 ,.,_,, :::: 0 5 V OJ) 0 .b a 0.4 0 en 0 .3 (ij ... 0 0.2 0.1 0 .0 0 2 3 4 5 6 Total soil carbon (%) Figure 3-4. Total soil nitrogen and soil carbon in soil sampled 0-8 cm depth Soil samples used for this analyses were chosen from among all treatments 7 and all times since bums (2 6 9, 12 and 18 months) to obtain as wide a range as possible for carbon content. 45

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46 Table 3-3 Results of ANOVAs of resin exchangable NH 4 NO 3 and PO 4 in 4 gap treatment and forest plots measured at 3 time periods following bums. All variables had a significant time*treatment interaction and therefore all were analyzed separately by month All variable~ were log transformed prior to analyses Treatments with different letters are significantly at P < 0 05 Analyzed by time period Post-hoc test results Variable F p high low removal control forest NH 4 -N first wet season 58.7 0 a b C C C first dry season 11.2 0 a a ab be C second wet season 2 0.11 NOrN first wet season 6 3 0 a ab ab b b first dry season 4 5 0 003 ab a a ab b second wet season 9 9 0 a ab be be C PO 4 -P first wet season 12 2 0 a a b b b first drv season 12 2 0 a be C b ab second wet season 3.2 0 02 a be ab ab ab

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0 014 -0.012 "'O I:: 0 010 cii Q) 1-, 00 0.008 -0.006 a -......, P-c 0.004 I '
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--N 8 bl) c t 0 .b C/) 0.. ---._, .... 0 .... i:: 0 Q 1-, ---._, 1-, 8 <.) -~ o.O 1-, 0 2 0 1.5 1.0 0 5 0 0 9 8 7 6 5 16 12 8 4 0 5 4 3 2 1 _._ High intensity bum -0Low intensity bum --TPlant removal ~ Gap control Forest 0----0-. -t -0--0.---. ......._ -o ~ .. y~ -~':i ." . -~ + --y -ii..._, '--'-"-' .:.: . :. : ... . . .. II ----v ----4---==-=-. ~ ,~ -2 6 9 12 18 Months after treatment 49 Figure 3-6. Soil pH, air-dry water content and organic matter content measured in soil (0-8 cm depth) in the four gap treatments and forest sites at 5 sampling times over an 18 month period following bums Soil strength was measured at the soil surface with a soil penetrometer at 4 sampling periods over a 12 month period following bums Water content is expressed as a percentage of air-dry weight (bars= S E )

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9 8 7 0.. 6 5 20 ,_ '-._/ 15 .... i::: Cl) 1:: 10 0 u I-, Cl) 5 0 5 ,_ '-._/ I-, Cl) 4 8 3 u }j 2 oI) I-, 0 I ---High intensity burn -0Low intensity burn --vPlant removal -'vGap control Forest Q----Q-------0 I=:~ :-:--:-:-: -:-:-=. t:--:::::.. :::-:.,:-= ~--=.: =-::.:..~ =m I L_ ..___ I I ---~ . ~ . ......... 9 12 18 Months after treatment 50 Figure 3-7 Soil pH, air-dry water content, and organic matter content in soil sampled from 8-20 cm depth in four gap treatments and forest sites at 9 12, and 18 months following bums. Y-axis scales are identical to those in Figure 3-6 for soil sampled from 0-8 cm depth (bars= S.E ).

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51 Although significant differences in soil water content were detected among treatments differences were not large and patterns were not consistent over the sampling period. The forest plots had the lowest soil water content during the first 9 months but this difference diminished after 12 months Larger differences were due to seasonal changes in soil water content predictabl y with the highest water contents in the rainy season of 1999 and the lowest water contents in the di) season of S eptembe r 1 998 Treatment effects on fine root mass Fine root mass 18 months following bums was significantly higher in forest plots than in highand low-intensi ty bum treatments (F = 4.4 P = 0 005 n = 10) Fine root mass (l ive and dead combined) in the top 7 cm of soil in was : forest 1.8 0 3 kg/m 2 gap control 1.3 0 .2 kg/m 2 plant removal 1.3 0 2 kg/m 2, low-intensity bum 0 9 0.1 kg/m 2, and high-intensity bum 0.7 0 1 kg/m 2 Treatment effects on soil physical properties Soil bulk density and soil strength. Bulk density in high-intensity bum treatments was significantl y higher than in forest plots after 6 and 12 months (F = 3 1 P = 0 02 n = 10) There were no significant differences among the remaining treatments. Bulk density averaged 1 .3 0 05 g cm 2 in high-intensit y bum treatments and 1 2 0 02 g cm 2 in forest plots after 12 months Soil strength in high-intensity bum treatments increased during the first 12 months following bums (Figure 3-6 Table 3-1) Although at 3 months soil strength was lowest in high intensity bum treatments it was the highest in this treatment by 9 months Soil strength in the other treatments also increased during the first 12 months but to a less extent than in high-intensity bum plots Water infiltration and soil wettability. Water infiltration rates were significantl y lower in the high-intensity bums than in the remaining treatments (F = 31 P < 0 001 Figur e 3-8). So i l

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52 wettability significantl y differed among treatments (F = 4 6 P = 0 002) and a slight but non significant differ ence was found in the wettability of different soil depths (F = 2.4 P = 0 .07 ) Surface soils in all treatments except the high-intensity bum tended to repel water (Figure 3-9). In the high-intensity bum treatment a slightly water-repellent layer was detected at 2-3 cm depths Effects of soil heating and ash addition on soil chemical properties Field study. Due to high seed predation of A colu brina I onl y report results of soi l analyses. Ash addition significantly increased soil concentrations of P, K, Ca and Mg as well as soil pH (Table 3-5 ; Figure 3-10) Ash addition lowered soil organic matter but did not affect available resin-available NH / -N and NO 3 -N Heated soil had lower concentrations of Mg but did not have significantly different concentrations of P K, Ca organic matter or resin-available Nl-L / -N and NO 3-N Greenhouse study. Both soil heating and ash addition decreased growth of C. pluviosa seedlings (Table 3-6 ; Figure 3-11) Intensely heated soil had higher levels ofresin-available NH N than moderatel y heated or control soil. Resin-available NO 3 -N was not detectable in intensely heated soil but it was significantly higher in moderately heated soil than control soil. Ash addition did not affect resin-available NH / -N and NO 3 -N Discussion Effects of highand low-intensity burns on soil chemical properties Controlled bums significantly affected on all soil chemical properties examined (soil pH, soil organic matter resin-available N and P and Mehlich-extractable P Ca K and Mg) These changes, attributable to soil heating and/or ash deposited during burns were greater after high intens ity bums than low-intensity bums High fuel loads combined with relativel y complete bums in the high-intensity bum treatment resulted in an average ash depth of 4 8 cm Maximum temperatures reached during high-intensity bums averaged 704 C at the soil surface and 227 C at 3 cm depth (Chapter 2) Little ash was deposited after low-intensity burns due mostl y to the lower

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53 0 20 -----------------------------, ,-...., ,......, I u (!) f/l N I 8 u g 0 16 0 12 i:::: 0.08 0 ..... .b j 0 04 '''-'-'----High intensity burn -}Low intensity burn -. Plant removal -WGap control Forest '-W----. . ---------. .. --w .. ... ~ -......__,_ .... .. ..... ... ~--y "" ------------4 .___ ___ ---. 0 00 -"------.-----------,---------.------' 5 IO 15 Cummulative volume of water (ml cm2 ) Figure 3-8 Water infiltration rates of soil in four gap treatments and forest plots. Infiltration was measured as the time required for the first 5 ml of a 10 ml column of water to infiltrate soil.

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,--.,. s u ...__, 0.. Q) "'O 0 r/) 54 0 .---------------Gl----------'1.,._------t ~'-----~ 1 2 3 0 5 10 15 --high intensity burn -low intensity bum -V plant removal ....szygap control forest 20 Time for total absorption (s) 25 Figure 3-9 Soil wettability at the soil surface soil and at l 2 and 3 cm depths in four gap treatments and forest plots X-axis refers to the time to total aborption of 5 drops of water applied to the soil surface with a dropper.

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Table 3-5 Results of ANOVAs of a field experiment examining the separate effects of ash addition and soil heating on soil properties Two levels of ash (no ash 500 gash) and 3 levels of soil heating (no heat, low intensity heat a.Tld high intensity heat) were applied to 0 25 m 2 subplots in the field and soil sampled after 3 weeks In the low intensity heat treatment, a flame of l50 250C was applied to the soil surface of the treatment area for 55 5 minutes with a propane torch In the high intensity heat treatment a flame of 500-800C was applied to the soil surface of the treatment area for 20 minutes Ammonium and nitrate were measured using anion and cation exchange resins buried at a depth of 5 cm for 3 months (n = 12) Factors Ash Heat Ash* heat Variable df F p F p F p Calcium 15 615 8 < 0.001 3 1 0 05 3.3 0 05 Potassium* 15 65 5 < 0 001 l.2 0 32 0 5 0.59 Magnesium* 15 75 6 < 0 001 4 7 0 01 1.0 0 38 Phosphorus* 15 23 5 < 0 001 0 2 0 81 2 8 0 07 NH 4 -N* 15 2 7 0 11 1.3 0 29 0.4 0 96 NOrN* 15 1.2 0.28 1.8 0 17 2 3 0 11 OM 15 11.7 0 001 3 8 0 03 2 3 0 11 pH 15 210 3 < 0.001 2 7 0.()8 0.4 0 96 log transformed prior to analyses D

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8000 -,-------------, ;:::-7000 6000 Calcium (/J !f 5000 W 4000 ';' 3000 U 2000 1000 0---'---control ash added 700 --------------, ::::-600 Magnesium 500 i 400 i::; 300 '-' 200 100 0 ---'---control ash added 6-----------, 5 4 3 2 Organic matter 0 --'---control ash added 0 10 -,--------------, "Cl C: 'ci, 0 08 V ... i 0 .0 6 .__,, 0 04 NH+ 4 0 .0 2 0 00 --'----control ash added 56 1400 -----------, ;:::--1200 1000 bl) 800 E 600 '-' 400 200 Potassium control EI:E] low heat 11111111111111 high heat 0 -'---control ash added so-------------, ;:::-40 5 Phosphorus (/J 30 8 20 '-' 10 0 -'---control ash added 10 --------------, 8 6 4 2 pH 0 -'--0 5 "Cl ] o.4 NO V 3 ... 0 3 ::, 0 2 '-' z '.., 0.1 0 z 0 0 ....L..... __ control ash added control ash added Figure 3-10 Results of field study of soil that had received combinations of heating (no heat, low intensity heat and high intensity heat) and ash addition (no ash ash added) AU soil was sampled from 0-8 cm depth 3 weeks after treatments Resin available NH / and NQ 3were measured using anion and cation exchange resins buried to a depth of 5 cm for 3 months (n=l2 ; bars=S E ).

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57 Table 3-6 Results of ANOVAs of a greenhouse experiment examining the separate effects o ash addition and soil heating on available nitrogen and seedling growth of Caesalpinia pluviosa Two levels of ash (no ash 15 g ash added) and 3 levels of soil heating (no heat lov intensity heat and high intensity heat) were applied to soil used to fill planting bags In the lo, intensity heat treatment soil was heated in a coventional oven at 150-200 C for 10 minutes In th e high intensity heat treatment soil was heated in a coventional oven at ~2 50 C for 40 minutes and fired with a propane torch for 5 minutes at a flame temperature of 500-800 C Ammonium and nitrate weremeasured using anion and cation exchange resins buried at a dept of 5 cm for 3 months Seedling height was measured 4 months after planting (n = 12) Variable NH 4 -N NOrN Seedling height df 5 5 F 6.2 1.7 Ash p 0 .. 023 0 20 Factors Heat F 291.0 23 3 p < 0.001 < 0 001 Ash* heat F p 1.8 0 6 4 8 0.20 0.58 0.012** **Significant interaction for seedling height therefore analyzed separately by heat treatment. Heat treatment F P tallest Seedling height no heat 11.4 0 003 w/o ash low heat 11.7 0 002 w/o ash high heat 0 15 0.7

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2 00 .,....._ :>.. C'd -0 c 1.50 Vl I Q) '0() -..... 0.08 ::s -0 06 ss::t" 0 04 0 02 0 00 0 6 .,....._ -0 0 5 c oo no heat Q) 0 4 I0() CJJZD low heat -..... 0 3 high heat ::s '-' z 0 2 I c-1") 0 1 0 z 0 0 18 15 G ...__, ...... 12 (1.) 9 ..c:= 00 .s 6 "O (1.) (1.) 3 C/) 0 control ash added Figure 3-11 Results of greenhouse study of Caesalpinia pluviosa seedlings planted in soil that been heated (no heat, low intensity heat and high intensity heat) and had ash added (no ash, 15 gash added). Seedling height was measured 4 months after planting (n = 12) Ammonium and nitrate were measured using anion and cation exchange resins buried in individual bags and watered for 22 days (n = 3) Bars are S E 58

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59 fuel loads but also to incomplete combustion of these fuels. Maximum temperatures reached during low-intensity bums averaged 225 C at the soil surface ; elevated temperatures at 3 cm depth were mostl y undetectable Soil pH. Increased soil pH is a general effect forest fires (e g. DeBano et al 1977 Wright and Bailey 1982 Kutiel et al 1990 Hungerford et a/.1990 Stromgaard 1992 Neary et al 1999) High concentrations of basic cations in ash deposited following fires (e g Ca K Mg Na) is the major mechanism of increased soil pH (Kutiel et al 1990, Stromgaard 1992) Although soil heating may also increase pH by releasing basic cations from soil organic matter (Giovannini et al. 1990) results of the soil heating and ash addition field study (Figure 3-11) revealed that ash addition significantly increased soil pH while soil heating had only a slight but non-significant effect. Soil organic matter High-intensity bums caused a net loss of organic matter in surface soils, a predictable consequence of intense fire (e.g. DeBano et. al. 1977 Hungerford et al 1990 Neary et al. 1999) Several studies conducted in the tropics have found decreased soil organic matter following slash burning (Amazonia : Uhl and Jordan 1984, Mackensen et al 1996 Australia : Rab 1996) Experimental studies have shown that soil organic matter loss is a direct effect of soil heating (e.g Hosking 1938 in Humphreys and Craig 1981 Giovannini et al. 1990) with distillation of volatile organic compounds occurring between soil temperatures of 100-3 00 C and near complete loss of soil organic matter at temperatures >450 C. Soil organic matter contents following high-intensity bums (Figure 3-6) averaged over 0-8 cm likely do not reflect larger losses in soil organic matter that occurred in the first several centimeters Consumption of soil organic matter was probably complete at the soil surface during the high-intensity bums where maximum soil temperatures averaged 683 C. Due the sharp decrease in soil temperature with depth (Chapter 2) organic matter consumption was probably negligible below the top several cm of soil and not detected at 8-20 cm depth

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60 Average surface temperatures during low-intensity burns (160 C) were not hot enough for the consumption of soil organic matter hence average soil organic matter contents of low-intensi ty bum plots were not significantl y lower than those of forest soils In fact averag e soil organic matter contents after low-intensity bums were higher than those of adjacent forest soils Increases in soil organic matter have been shown to occur during light to moderate burns (e .g., Hungerford e t al. 1990) due to the incorporation of unburned or partiall y burned slash fragments into soi l. For example Stromgaard ( 1992) attributed increased soil carbon following slash burning in miombo w oodlands to charcoal accumulation or small organic particles washed in from ash Soil organic matter in high-intensity bum plots recovered to levels higher than those in forest plots within 18 months following bums High daytime soil temperature in burned plots and high soil moisture w ithin gap treatments ma y have contributed to this rapid recove ry b y increas in g decomposition rates Though generally rapid recovery of soil organic matter following slash and bum vary among and within tropical forests For example at the same sit e in the Venezuelan Amazon Montagnini and Buschbacher (1989) reported recovery of soil organic matter within 6 months while Uhl and Jordan (1984) reported that recovery required 5 years Other than differences in climate and site productivity fire in t ensity and land use following bums also affect recovery of soil organic matter thus making comparisons among studies difficult. T o tal soil nitroge n. Total soil N was linearly related to soil organic carbon hence the greatest declines in soil N occurred in high-intensity bum plots where organic matter in the top 8 cm of soil decreased an average of 28% from adjacent forest soils Similarly total soil N was reported to decrease following slash and bum of tropical forest in Costa Rica (Ewe! et a l 1981 ) and in the Venezuelan Amazon (Uhl and Jordan 1984) As with soil organic matter this average (28%) underestimates N losses from the top several centimeters of soil or from more i ntensel y burned patches ; N losses reached 84 % in the top 8 cm of soil and likel y approached 100% in scorched surface soils Comparably high N losses were reported in chaparral soil heated to 500 C

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(80% N loss; Dunn and DeBano 1977) and Mediterranean soils heated to 600 C (86% loss ; Kutiel et al.1990) 61 Losses of total soil N during low-intensity bums were negligible as indicated by low losses of soil organic matter. In fact slightly higher soil organic matter contents in low-intensity bum plots relative to adjacent forest soil suggest that total soil N increased after this treatment. This is likely due to the mixing of slash fragments into surface soils Increases in total N in surface soils has also been found following slash burning in the tropical forests (Montagnini and Buschbacher 1989, Stromgaard 1992) and temperate forests (Gholz et al 1985) Resin-available nitrogen. In contrast to decreases in total soil N amounts of resin available nitrogen (NH / -N and N0 3 --N) increased after bums of high-intensity These findings agree with those of Matson et al.(1987) and Montagnini and Buschbacher (1989) who also reported increases of ammonium and nitrate following slash burning of tropical forest in Costa Rica and Venezuela respectivel y In the present study low-intensity bums increased resin available N levels as well although to a lesser degree than high-intensity bums Temperate zone studies have also noted that increases in inorganic N are dependent on fire intensity (Dunn and DeBano 1977 Giovannini et al 1990 Kutiel et al. 1990 Rice 1993 Weston and Attiwill 1996 McMurtrie and Dewar 1997) Dunn and DeBano (1977) demonstrated that the greatest increases in anunoniurn and nitrat e for chaparra l soils occurred at soil temperatures up to 3 00 C du e to th e mineralization of organic N At soil temperatures of>500 C inorganic N decreases due to volatilization (Dunn and DeBano 1977) Similar results were reported in soil heating studies conducted by Giovannini et al (1990) Increased ammonium availability following bums ma y be enhanced by soil microbia l d e ath w hich occurs at temperatures as low as 50 -121 C (Neary et al. 1999) Soil microbial death was likel y substantial in high-intensity bum plots and ma y have occurred within small well burned patches in low intensity bum plots as well. In their study of nitrogen transfom1ations

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following slash and bum on volcanic soils in Costa Rica Matson et al ( 1987) found that the amount of nitrogen that disappeared from microbial biomass after bums was similar to the concurrent increase in apparent net nitrogen mineralization 62 Matson et al ( 1987) also attributed increased nitrate concentrations to enhanced nitrification rates Burning generally creates favorable conditions for nitrification such as raised pH values and base saturation (Pritchett and Fisher 1987) Increased nitrification rates were also reported following slash burning of Venezuelan forests (Montagnini and Buschbacher 1989) In contrast other studies ha ve shown that nitrification rates are reduced b y fire due to a decreased biomass ofnitrifiers (e.g ., Dunn and DeBano 1977 Stromgaard 1992) Reduced nitrification rates would cause an accumulation of ammonium which is less subject to leaching than nitrate This effect may explain why nitrate was undetectable in the intensely heated soil of the greenhouse study while levels of ammonium were very high (Figure 3-12) Elevated concentrations of ammonium and nitrate following high and low-intensity bums were short-lived in my study Within 8 months of burns (after the first rainy season) inorganic N in burned plots declined to levels found in adjacent forest. This result is similar to the rate of decline of inorganic N following slash burning of a wet forest in Costa Rica (Matson et al. 1987). Phosphorus. Highand low-intensity burns increased both extractable P and resin available PO / -P. Inorganic P additions in ash likely contributed to these increases. In the soil heating and ash addition field stud y (Figure 3-11 ) ash addition significantl y increased extractable P while soil heating had little effect. Similarly Stromgaard ( 1992) attributed increases in extractable P after slash burning of miombo woodlands to ash deposition and Rice (1993) found that soil P0 3 -P concentrations in Californian chaparral following fire were correlated with ash depth but not fire intensity Though possibly less important than ash deposition soil heating can increase extractable P by mineralizing organic P as may have occurred with resin-available N Giovannini et al. (1990)

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63 found an increase in inorganic P accompanied by an equivalent decrease in organic P in soil samples heated to 460 C ; at temperatures >460C, all organic P was destroyed a.rid only inorganic P remained Soil heating may have been comparatively more important in low-intensity burn treatments where relatively little ash was deposited Other studies have reported increases in inorganic P following fires of low-intensity. For example in a study of soil nutrient levels associated with shifting agriculture in the Asian tropics Andriessse and Koopmans (1984) found available P increased almost 300% after heating to 200 C which they attributed to mineralization of organic P Cations. High-intensity burns significantly increased soil concentrations of extractable cations (Ca K and Mg) Low-intensity burns also increased cation concentrations although not as dramatically. Similar to P, results of the soil heating and ash addition field study suggest that increases in extractable cation concentrations by burning is mostly due to ash deposition. Significant increases in extractable Ca Mg and K after fire in miombo woodland (Stromgaard 1992) and Brazilian cerrado (Coutinho 1990) were also attributed to ash deposition. As with P soil heating may also increase extractable Ca K and Mg through mineralization of organic forms (Giovannini et al.1990) Extractable Ca and Mg peaked in soil heated to 200 C and declined at higher temperatures ; extractable K peaked in soil heated to 700 C (Giovannini et al 1990). Results of the soil heating and ash addition study conform with this pattern ; soil heating significantly decreased e"-'tractable Ca and Mg concentrations and had no effect on extractable K concentration Decreases in cation concentrations over the 18 month post-bum period are similar to those following burns in temperate forests (e g ., DeBano et al 1977 DeRonde 1990 Kutiel and Shaviv 1992 Hernandez et al 1997) and tropical forests (e.g ., Uhl and Jordan 1984 Coutinho 1990 Stromgaard 1992 Mackensen 1996) The order of decrease (K> Mg> Ca) corresponds with cations mobility and susceptibility to leaching In high-intensity burn plots plant uptake was

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64 probabl y not important in cation decreases during the first y ear as plant cover remained l e ss than 25% (Chapter 5) Plant uptake may have been more important during the second y ear as vegetative cover reached 60% after 18 months Variation of soil nutrients within and among plots. Based on the comparison of one plot. soil nutrients and organic matter in burned treatments appeared to be more variable w ithin plots than among plots of the same treatment (Table 3-4) This pattern suggests that natural variation in soil nutrients was increased due to heterogeneity of burns. The opposite pattern w as true for the plant removal gap control and forest plots Variation in soil properties was in general greater among the 16 different plots than within the same 100 m 2 plot suggesting variations in soil fertility in the absence of fire are expressed at larger scales Actual v ariation in soil properties after high-intensity bums may have been greater than reflected by random sampling For example nutrient and organic matter contents of severel y scorched soil differed greatly from averages of high-intensity bum plots Mineral concentrations of such scorched soil sampled to 5 cm (with percentages of high-intensity bum plot averages) were : 6415 mg/kg Ca (133%) ; 197 mg/kg Mg (56%) ; 71 mg/kg K (11%) ; 0 25 mg/kg P (0 3%) ; and 0 1 % organic matter (4%) These extremel y scorched soils were not common( < 1 % high-intensi ty bum plots) but potentially affect plant colonization b y providing rnicrosites different from less intensely burned areas Increased soil heterogeneity after bums has been observed b y other authors Christensen ( 1985) noted that soil nutrient concentration in chaparral is considerably more variable after fire than b e fore. du e to local variation in fire intensity and the uneven distribution of ash Heterogeneity in soil nutrients potentiall y has important consequences for colonizing plants For example Rice ( 1993) observed that even small scale patterns in fire intensity and ash distribution were reflected in later establishment of chaparral shrubs

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65 Changes in fine root mass following burns Lower fine root biomass after high-intensity bums was likely due to a combination of rapid decomposition of dead roots as well as direct oxidation by fire. Experimental studies have shown that fine roots are desiccated or killed at soil temperatures of 48 -54 C (Neary et al. 1999) Temperatures in high-intensity bum plots (61 -399 Cat 3 cm depth) were not only well above this range but were likel y high enough in places to completel y oxidize fine roots Even during low intensity bums temperatures in surface soils ( 160 C average at soil surface) were sufficiently high to kill fine roots Fine root mortality during bums was likely greater than death of larger roots not only because of their small size but also due to their concentration near the soil surface Fine root mortality potentially has important effects on soil fertility as their decomposition may increase soil nutrient concentrations Nutrient input from roots has been hypothesized to be an important pathwa y for nutrient c y cling particularl y in tropical dry forests due to their larger store of biomass below-ground (Martinez-Yrizar 1996 Jaramillo and Sanford 1995) De c reased fine live-root mass is also expected to contribute to higher water and nutrient availability due to reduced uptake. However average soil moisture contents in high-intensity bum plots were not significantly different than those in other gap treatments Possibl y, lower water uptake was offset by decreased water holding capacity of intensely bum soil caused by a loss of soil organic matter. Effects of highand low-intensity burns on soil physical properties Changes in soil strength bulk density and water infiltration rates in high-intensity burn plots w ere substantial. The decrease in soil organic matter in high-intensit y burn plots lik e l y influenced these observed changes in soil physical properties Organic matter influen c es so i l structure through ag g r e gate formation ; a decrease in organ i c matter decreases total porosi ty particularly macro-pore spaces ( > 0 6 mm) The increase in surface soil strength during th e first ye ar following high-int e nsit y burns was likel y due to the settling of soil minerals and ash in t o

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spaces left void by organic matter and fine roots This settling of soil particles would also contribute to higher soil bulk densities 66 Decreased macro-pore space would also contribute to the lower infiltration rates observed in high-intensity bum plots. However these lowered infiltration rates caused by high-intensity bums did not result in any observable surface runoff. The lowest infiltration rate recorded in a high-intensity bum plot was 5 times faster than the rate needed to absorb a 5 cm hr" 1 rainfall (0.002 cm 3 cm 2 sec 1 ) Further most plots were located on level ground therefore if a rain event exceeded the soil s rate of infiltration the accrued water would not run-off. Increased wettability of surface soils after high-intensity bums conforms to studies that report soil temperatures >288 C destroy water-repellent layers (Neary et al. 1999). Soil temperatures of 176-288 C reportedly form water-repellent layers (Neary et al. 1999) explaining the presence of a slightly water-repellent layer at 2-3 cm depth. However the decreased w ettability of this soil la y er does explain the lower infiltration rates in high-intensity bum treatments as surface soils in the remaining treatments had similar wettability properties Possibl y, a more water-repellent layer was formed deeper than 3 cm in high-intensity bum plots but was undetected due to the sampling strategy Soil str e ngth. bulk density water infiltration and water repellenc y oflow-intensity bum plots were not different from those in the unburned treatments Again this pattern ma y reflect the influence of organic matter on soil physical characteristics ; the lower temperatures during low intensity bums (mean 120 C), did not decrease soil organic matter Potential effects of highand low-intensity burns on tree seedling grow1h Soil heating and ash addition significantly affected Caesalpinia seedling growth although m a manner opposite than expected ; ash addition decreased seedling growth This result suggests that the quantity of ash added to soil may have been at toxic levels for this species. Also tree seedlings were shorter in soil heated at both low and high intensities This result only partially

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67 corresponds with a similar study by Giovannini et al ( 1990) which examined the effects of soil heating on wheat (Triticum aestivum) seedling growth They found that while soil heated to 170 C had no effect on plant growth, soil heated to 220 C and 460 C increased seedling height and biomass, whereas soil heated to 700 C and 900 C had detrimental effects on seedling growth The authors attributed increased growth in moderately heated soil to greater ammonium and available phosphorus concentrations. Lowered growth in intensely heated soil was attributed to the sharp increase in soil pH and release of Ca and K to toxic levels, as seedlings in this treatment of their study showed symptoms of nutritional disorder. In my greenhouse study decreased seedling growth in intensely heated soil ma y be the result of degraded soil structure or toxic levels of cations However it is unclear wh y seedling growth was impaired in the lightly heated soil. It is important to consider that although seedling heights significantl y differed among treatments maximum height differences were only 3 cm This slight difference after 4 months of growth may not be biologically significant. Perhaps the effects of soil heating and ash addition on seedling growth would have been more apparent if a shade intolerant species had been used as a bioassay. Caesalpinia is partially shade-tolerant and exhibited slow growth rates in the field as well (Chapter 4) Despite the potentially negative effects of increased bulk density and soil strength lowered infiltration rates and possibly toxic effects of cations on plant growth seedling heights of shade intolerant species were greatest in high-intensity burn plots (Chapter 4) This increased growth in intensely burned soils may be due to several factors Initially, soil strength in high-intensity burn plots was the lowest of all treatments therefore early colonizing seedlings should not have experienced mechanical impedance of root growth Secondly nutrient concentrations were highest in high-intensity burn plots which may have offset decreased movement of nutrients through the soil. Also, toxic levels of cations may only have been a factor in small areas of high ash deposition or severely scorched soils ~ seedlings may not have been able to establish in these small areas and

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6 8 therefore the effects on growth were not observed Most importantl y, the density of plants colonizing high-intensity burn plots was low (Chapters 5) so that established tree seedlings likel y benefited from reduced competition for soil water and nutrients Effects of plant removal and canopy gap formation on soil chemical and physical properties Soil moisture content was higher in all of the gap treatments than forest plots for the first 9 months following bums. Higher soil moisture within tree fall gaps than under adjacent forest is a pattern repeatedly found in tropical forest studies (e.g. Vitousek and Denslow 1986) and has been attributed to decreased transpiration within gaps due to less vegetation The difference in soil moisture content between forest and gap plots diminished over the first year as the amount of vegetation in gaps increased Plant r emova l and gap control treatments did not significantl y chang e soil chemical or physical properties from those in adjacent intact forest. Although it is hypothesized that the increased soil temperatures moisture and litter depth in tree fall gaps will increase nutrient availability (i e Bazzaz 1980) conclusive evidence to suggest this is true has not been reported For example in a study of natural tree fall gaps in lowland moist forest in Costa Rica Vitousek and Denslow (1986) found that nitrogen mineralization did not increase in tree fall gaps and slight phosphorus increases were not significant. The only difference they detected was within gap rnicrohabitats ; the root throw zone had significantly less N and P than the crown zone Luizao et al. (1998) found similar results in a study of artificial gaps ranging in size from 40-2500 m 2 in Brazilian rain forest. No differences in microbial biomass soil respiration and nitrogen mineralization or nitrification were found between gap and forest sites Most of the variation observed within the plant removal gap control and forest plots over time was due to seasonal changes Soil moisture content varied predictably with changes in rainfall and NO 3availability declined slightly during the dry season This observation agrees with the few studies of nutrient cycling conducted in tropical dry forests which have shown that nitrification

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69 rates are highest during the rainy season and lowest at the end of the dry season (Singh et a l. 1989 Garcia-Mendez et al 1991 see also Smith et al 1998) Longer term effects of controlled burns on soil properties The duration of this study limits its conclusions to only short-term treatment effects A similar study conducted in Las Trancas in 1995 (Stanley 1995) however reveals slightl y longer term effects of burning on these forest soils Although not identical the treatments applied in this earlier study were comparable to those used in mine: a gap control ; gap vegetation slashed and removed ; gap vegetation slashed and burned ; and gap enlarged by 30% vegetation slashed and burned (Stanley 1995). Fires in the enlarged gaps were likely more intense than fires in gaps that were not enlarged, due to a greater amount of fuel I measured soil nutrient concentrations and soil pH at 0-8 and 8-20 cm depths in April 1998 3 years following experimental bums and found no significant differences among gap treatments at either depth (Figure 3-12) The soil property with the most distinct trend was P concentration (P = 0 11 ) followed by Ca concentration (P = 0 19) soil pH (P= 0.22) K concentration (P = 0 27) Mg concentration (P = 0.32) and organic matter content (P = 0.35) These results suggest that soil chemical changes following bums are relatively short-lived However the lack of significant results may have been more indicative of the large variation found within treatments rather than the lack of variation found among treatments Felling gaps included a diverse array of habitats such as rock outcrops and stream side areas Important! there was no indication of declining soil nutrient concentrations in the cleared and burned areas of this pilot study after 3 years A widely held notion about the recovery of tropical ecosystems following disturbance is that severe nutrient losses following deforestation limit forest regeneration (e.g., Allen 1985 Buchbacher et al. 1988) Clearly longer-term sampling of the burned plots is needed before conclusions can be drawn about the long-term sustainability of severely burned soils at this forest site

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0 r-300 250 ,-.._ 200 OJ) ..!iG OJ) 150 8 '-" OJ) 100 50 0 t[ I I t li 2 3 4 5 6 5 [i 4 ...., E (.) 3 a c,j OJ) 2 I0 W i t: o iJJJ ft : I 1 ii:i 2 3 4 5 180 160 140 --120 OJ) 100 ..!iG OJ) E 80 '-" 60 40 20 0 50 40 ,-.._ '-;' on 30 ..!iG a '-" 20 p.. 1 ,1 :;::: LI /il!i j :::::: 1 I I 2 if 3 4 5 0-8 c m t: :,:,:,,,,,,,,, ;::j 8 2 0 cm lO J 1111I 2 3 4 5 6000 ~---------, 5000 4000 ... a 3000 '-" c,j U 2000 -1 1000 -1 u. ti 0. J r I r r~ 2 3 4 5 8 ~--------, 7 :r:: 0.. !Ill 6 ,I ~IJIJI J l 2 3 4 5 Figur e 3-12 Conc e ntrations of extractable Mg K Ca and P and percent organic matter and soil pH of soil sampled from logging gaps with th e following tr e atm e nts : 1) enlarged and burned ; 2) burne ; 3) slash removed ; 4) control ; and 5) forest understory These treatments w e r e applied ( 3 y rs prior to sampling) to 6 gaps each in Las Trancas '94 Soil was sampled 0-8 cm (black bars) and 8-20 cm (gra y bars). Dott e d lines r e present th e range of means of soil sampled from all treatments in m y study 18 months following treatments (n = 6 ; bars = S E )

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71 Conclusions The experimental bums had mixed effects on physical, chemical and biological soil properties. High-intensity bums increased levels of available nitrogen and phosphorus but these pulses quickly decreased Ash deposited after high-intensity bums increased cation concentrations but these pulses declined over time as well Loss of organic matter during high-intensity bums likely altered surface soil structure This effect was apparent in increased bulk density and soil strength, and decreased infiltration Low-intensity bums increased cation concentrations and available forms ofN and P Soil structural changes were not as marked as after high-intensity bums because little soil organic matter was consumed. Therefore, soil chemical properties can be altered during low-intensity fires but changes in soil physical properties may only occur during intense fires Plant removal gap control and forest plots had no significant effects on soil chemical or physical properties Greenhouse studies suggested that loss of soil structure caused by soil heating and toxic le v els of cations created by large quantities of ash may hinder tree seedling grov.rth However these potentially negative effects were not apparent in the field. Shade-intolerant seedlings established in high-intensity bum plots grew faster than those in all other treatments Evidently the increase in soil nutrients caused by high-intensity burns was not offset by altered soil structure Importantly the choice of tree species used as bioassays in these studies likely affected the results An important feature of the increase in resource availability produced by bums is the transient nature of the increases Therefore, the first plants to become established after bums should benefit most from greater resource availability Bums also greatly increased soil resource heterogeneity Therefore successful establishment and vigorous growth of plants after bums may be greatly influenc e d b y chance i.e ., seeds dispersed into patches of well burned soil will exhibit higher growth than those dispersed into unburned soil or severely scorched soil.

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CHAPTER4 EFFECTS OF CANOPY GAP FORMATION PLANT REMOVAL AND CONTROLLED BURNS OF HIGH AND LOW INTENSITIES ON EARLY REGENERATION OF COMMERCIAL TREE SPECIES Introduction The most basic principle of sustainable forest management is that rates of timber harvesting should not exceed the rate at which timber volume accumulates (e g Johnson and Carbarle 1993 Dawkins and Philip l 998) This criterion requires sufficient regeneration of harvested species and in many neotropical forests poses the greatest barrier to sustainable forest management (e.g Wyatt-Smith 1987 Verrisimo 1995 Gullison et al. 1996) There are several silvicultural means of improving poor regeneration ranging from intensive techniques such as prescribed burning to less intensive techniques such as selective harvesting without further treatment. Choice of technique must be knowledgeably based on the natural regeneration requirements of the target species Presently lack of information of the autoecology of man y harvested tree species is one of the largest deterrents to sustainable timber production (Bazzaz 1990 Bazzaz and Pickett 1980 Fox 1976 Gomez-Pompa and Burley 1991 Hall 1996 Whitmore 1989) For many decades seedling ecology was a minor part of tropical forestry (Hall 1996) ; more recently tropical seedling ecology has become the focus of much ecological research (e.g ., Garwood 1983, Augspurger 1984a 1984b Swaine 1996 Kitajin1a 1996 Fenner and Kitajima 1998) The seedling stage is critical to regeneration as most mortality occurs early in the life of a tree (Lieberman 1996 Li et al 1996) Understanding this life stage is also important for forest managers as species vary widely in their seedling ecology (Hall 1996) and the division of tree 72

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73 species into regeneration guilds is based on seed germination and seedling establishment requirements The most well defined group are species that require high light conditions for seed germination and seedling establishment and often colonize following disturbances (i e. early successional Bazazz I 979 ; pioneer or secondary Budowski I 965 ; or shade-intolerant Swaine and Whitmore 1988) At the other end of the spectrum are species that can germinate and persist in the low light of forest understories (i.e ., late-successional Bazazz 1979 ; climax or primary Budowski 1965 ; or shade-tolerant Swaine and Whitmore 1988) Tree species with different regeneration strategies require different silvicultural treatments to enhance their regeneration. For example many shade-tolerant species have advanced regeneration in forest understories (Brokaw 1985b Hartshorn 1989), therefore management techniques would mostly entail ensuring the survival of this regeneration during harvesting and enhancing its growth to mature stages Enhancing regeneration of shade-intolerant species that do not have seedling banks in forest understories involves creating sites suitable for seed germination and seedling establishment and promoting safe arrival of seeds to these sites (Dickinson 1998) In the seasonally dry forests of Bolivia, commercial tree species are represented in both shade-tolerant and shade-intolerant groups (Guzman 1997) Consequently a mixed-management system was proposed for the community-owned forests of Lomerio (Pinard et al. 1999) Single tree selection was recommended to enhance regeneration of shade-tolerant trees Group selection harvesting groups of trees to foster the development of even-aged patches was suggested to improve regeneration of shade-intolerant species. Prescribed burning of logging gaps has also been suggested as an additional treatment for the management of shade-intolerant species (Stanley 1995) Prescribed bums may enhance seedling establishment and growth of shade-intolerant species in a number of ways including removing litter reducing logging slash and slowing vine proliferation (Stanley 1999) Abundant regeneration of shade-intolerant species after wildfire in Lomerio (Mostacedo et al 1999 Gould et

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al 1999) lends further support to the promise of prescribed bums as a silvicultural tool for the management of dry forests in Bolivia 74 Although the pioneer-climax dichotomy provides an often useful paradigm for ecologists, regeneration strategies of many tree species fall between the two extremes of completely shade intolerant or shade-tolerant (Augspurger 1984b Condit et al 1996) In fact most rain forest tree species are both shade-tolerant and gap-dependent, meaning they have the ability to persist in a seedling bank in forest understories but require canopy opening to reach maturity (Hartshorn 1989) And as there is a continuum of species regeneration strategies there is also a gradient of disturbance intensities among potential silvicultural treatments The interaction of individual species biology and silvicultural treatments of varying intensities is inherently complex Before management techniques can be prescribed on a large scale the effects of these techniques on all of the species in question need to be examined For example what are the effects of more intense disturbances such as high-intensity fires on the advance regeneration shade-tolerant trees ? Or what is the minimum disturbance level required for the regeneration of shade-intolerant trees? In this chapter I examine the range of responses of commercial timber tree species to experimental canop y opening above-ground biomass removal and controlled bums of high and low intensities and compare these responses to those in forest understories My objective was to determine which gap treatment provided the best conditions for the estabiishment and growth of each species I also address the question of whether regeneration of each of these species is limited b y seed dispersal or b y sites suitable for their establishment and growth I approached these questions by studying the effects of gap treatments on seed germination seed predation vegetative regeneration and seedling and sprout growth and mortality Finally I discuss results in relation to silvicultural management options for each species

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75 Methods I conducted this study in the four gap treatments (high-intensity bum low-intensi ty bum plant removal and gap control) applied to the 400 m 2 blocks described in Chapter 2 Trees were sampled in the paired 4 m 2 subplots positioned near the gap center and gap edge of each 100 m 2 treatment plot as well as th e paired forest subplots (see Figure 2-4 Chapter 2) Seed addition treatment One randomly selected subplot ( 4 m 2 ) of each pair was assigned a seed addition treatment. Seeds of 5 commercial timber species were used in this treatment: Anadenanthera colubrina Astronium urundueva Centrolobium microchate Copaifera chodatiana and Schinopsis brasiliensis (hereafter referred to by genera ; Table 4-1) I chose these species based on their commercial importance and seed availability in 1997. Seeds of all species were collected in the 95 coupe of Las Trancas from the forest floor and stored in cloth bags under a shelter until sown in plots I removed all seeds that appeared to be damaged by predators or fungi. Twenty seeds each (= 5 seeds/m 2 ) of Anadenanthera Astronium and Copaifera and 10 seeds each(= 2.5 seeds/m 2 ) of Centrolobium and Schinopsis were placed in each plot on the litter or soil surface depending on the soil surface conditions in each treatment (Chapter 2) Viability of collected seeds Viability of seeds collected for the seed addition treatment was determined in germination trials conducted over a 60 day period (November-January) in Santa Cruz One hundred seeds of each of the 5 species were planted in 5 trays (20 seeds per tray) of a 50:50 mix of sand and soil. Trays were placed in a location receiving morning shade and afternoon sun and were watered each morning Newly germinated seeds were counted and removed dail y.

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Table 4-1. Characteristics of tree species used in seed addition treatment or those species with sufficient regeneration for statistical analysis All but Acosmium are commercial timber species Ecological classifiaction of shade tolerance of regeneration taken from Pinardetal. 1998 Shade Dispersal Seecies name Famil:t tot. Diseersal unit Fruit characteristics Seed size Acosmium cardenasii Papilinoideae 3 wind seed legume 10 x 15 mm 1 mm thick Anadenanthera colubrina Mimosoideae 1 gravity seed legume 10-25 cm long 12 x 10 mm 1 mm thick Aspidosperma rigidum Apocynaceae 3 wind seed pod 6 x 5 cm 25 x 20 mm 1 mm thick Astronium urund e uva Anacardiaceae 1 wind fruit small dried drupe calyx to 1 cm 3 mm diam Ca e salpinia pluvio s a Caesalpinaceae 2 gravity seed legume 10-15 cm long 10 x 10 mm 2-3 mm thick Centrolobium microcheate Papilinoideae 1 wind fruit samara 8-10 cm long x 3-4 cm wide 12 x 2 mm Copaifera chodatiana Caesalpinaceae 3 animal seed dry pod 2 x 3 cm seeds with oily aril 10 x 5 nun 4 mm thick Schinoe.sis brasiliensis Anacardiaceae 1 wind fruit samara 15-20 nun long x 5 mm wide 10x3nun Shade tolerance : 1 = shade intolerant ; 2 = partially shade-tolerant ; 3 = shade-tolerant

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77 Density, height, and relative height growth rate measurements Seedlings have been variously defined as individuals still dependent on seed reserves (e g Garwood 1996) to individuals up to 2. 7 m tall (Whitmore 1996) In this stud y, I did not use size or physiology as a defining character rather I define seedlings as individuals originating from seed as opposed to those regenerating as sprouts I measured seedling and/or sprout density and height in each 4 m 2 subplot (both seeded and unseeded) at 1.5 3 6 9 12 and 18 months after burns At eac h sampling p e riod all commercial species within subplots (of both seeded and unseeded species) were identified as sprouts or seedlings tagged and height to the apical meristem measured Sprouts originating from the stem or root collar were easily identified because scars were visible Root sprouts (root suckers) were more difficult to identify but were recognizable because the first leaves generally differed from the first true leaves of seedlings. Relative height growth rates (hereafter referred to as RGR) was calculated as : RGR = [In (height 1 2 ) 1n (heightt1)] / (t 2 t1) where t1 and t 2 are two measurement periods Seedlings of Anadenanthera were extremel y abundant in 1997 therefore a maximum of 3 randomly selected individuals per subplot were tagged for height measurements and the remaining individuals counted Additionall y, a maximum of 3 randomly selected individuals of Acosmium cardenasii were tagged in each plot and the remaining individuals counted Although Acosmium is not commercially valuable due to its susceptibility to heart rot I included it because it has the most abundant tree regeneration of any canopy tree species in Las Trancas 95 Many of the Anadenanthera seedlings had been bro w sed between the 6 and 9 month assessments therefore I noted presence or absence of browsing of tagged individuals.

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78 Statistics analyses Seedling densities of most tree species could not be normalized therefore seedlin g densities were analyzed using Kruskall-Wallis non-parametric tests (SPSS 1997) Separate tests were run for each species by month testing for the effects of the seeding treatment and gap treatments on densities For species that regenerated from both seeds and sprouts these regeneration modes w ere analyzed separately Square-root transformed densities of Acosmium and Anadenanthera were normally distributed therefore densities of these species were compared using repeated measures ANOVAs. For Anad ena nthera the seeding treatment and gap treatments were factors in the ANOVA model. For Acosmium gap treatments and regeneration mode were used as factors as this species was not used in the seeding treatment but regenerated from both seeds and sprouts Seedling height and RGR were analyzed separately for each species using repeated measures ANOVAs with gap treatments and for the species that sprouted regeneration mode as factors The effects of location within gap on seedling height and RGR was tested for Anadenanthera ; other species were not sufficiently abundant for this test. Also the effect of treatment on proportion of Anadenanthera seedlings browsed was tested with an ANOV A. Blocks were random effects in each of the above models Effects of treatments on seed predation A seed predation study was conducted to compare rates of seed predation among treatments and at varying distances from gap centers Two species were chosen for this study : Centrolobium and Copaifera In each treatment 2 seeds of each species were placed at each of five stations 1 3 5 7, and 9 m from the gap center. Seeds were inspected after 2 and 9 weeks for removal or signs of predation In this study, I assumed that removal indicated predation and therefore use predation as the sum of removed seeds and damaged seeds The effects of treatment and distance from gap center on seed predation were tested using Kruskal-Wallis tests on proportions of seeds remaining and undamaged after 9 weeks

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79 Effects of the low-intensity burn treatment on seed germination To assess the effect of low-intensity bums on seed germination of 3 commercial tree species 5 seeds each of Copaifera chodatiana Centrolobium microchaete and Schinop s is brasiliensis were placed on leaf litter in low-bum plots just prior to controlled burns After bums remaining seeds were retrieved and transported to the nearb y community of San Lorenzo for germination trials Seeds from burned plots were placed in plastic trays with sand and watered daily for 2 months To detect if germination was related to degree of bum damage seeds were inspected before gennination trials and assigned a damage score Seeds of Centrolobium and Schinopsis which are protected by a dry husk, had 6 damage categories relating to the degree of damage to the fruit: 0 (no damage) to 5 (fruit completely burned) Copaifera which is dispersed with only a fleshy aril had 3 damage categories relating to the degree of visible damage to the seed The percentage of seeds germinating within each damage category were compared using a Chi squared test for independence. Results Seed viability of species used in seed addition treatment Seeds of Copaifera demonstrated the highest viability in greenhouse germination trials followed by Anadenanthera Astronium Schinopsis and Centrolobium (Figure 4-1) Seeds of Anadenanthera and Astronium germinated most rapidl y; of their viable seed 80 % and 83 % respectivel y, germinated within 4 days. Effects of treatments on seed predation Seed predation was uncommon in gaps and there were no differences among treatments ( Copaifera P = 1 00 ; Centrolobium P = 0 .30) or distance from gap edge ( Copaifera P = 0 91 ; Centrolobium P = 0 73). Overall 17% of Copaifera and 5% of Centrolobium seeds were removed or had evidence of predation over the 9 week observation period.

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,.-... '-" C 0 -~ C Cl) 00 Cl) ;> ~ "3 E ::s u 100 100 80 80 --Anadenanthera
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81 Effects of low-intensity burns on seed germination Low-intensity bums either decreased or did not affect seed germination of Copaifera Centrolobium and Schinopsis Copaifera seeds placed in low-intensity bums had an ov e rall gennination rate of 35% less than half the gennination rate of seeds used in the seed viability gennination trial (88%). Percent germination of Copaifera seeds was dependent on the degree of bum damage (X 2 = 9 5, P < 0 05) Germination rates of Schinopsis seeds placed in low-intensity bums were also lower than unburned seeds (4% compared with 26%) however numbers were too low for anal y sis Cent rolobium seeds placed in low-intensity bum plots had germination rates of 4% only slightly lower than the germination of unburned seeds (5%) Treatment effects on seedling densities of commercial tree species Effects of seed addition treatment. Seed addition significantly increased seedling densities of Centrolobium and Copaifera but not those of Anadenanthera or Astronium (Tables 42 4-3). Only 9 individuals of Schinopsis were recorded in all subplots ; of these individuals 6 were in seeded plots Effects of gap treatments. Patterns of commercial tree density among treatments and over the 18 month sampling period varied among species For simplicity I have displayed commercial species that had similar density patterns over the 18 months together as groups in Figure 4-2 Significant differences among treatments are reported separately for each species in Tables 4-3 and 4-4 and treated in more detail in following sections Density of the first group (Anadenanthera and Astronium) all true seedlings peaked within 3 months after bums and declined thereafter. Density of the second group (Copaifera Aspidosperma, and Caesa/pinia) differed according to regeneration mode. While seedling density of this group graduall y increased throughout the 18 month observation period sprout density remained fairly constant after the first 3 months The third group consists of one species (Centrolobium) that regenerated predominately as root sprouts Seedlings of Centrolobium, which

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Table 4-2 Summary of seed fall and seedling densities of 7 commerical tree species and a non-commerical tre e species (Acosmium cardena s ii ). Seed fall refers to seedfall before or after the prescribed bum treatments (August 30-September 2 1997). Significance values for the seeding treatment are only given for seeded species. Due to the delayed germination of several species the effects of the seedling treatment was not detected until 12 months following treatments For these species densities and significance values are reported for 12 months For species not used in the seeding treatment total density (seedlings and sprouts combined) at the end of the study (18 months) is reported with the exception of the 1st year cohort of Caesalpinia and Acosmium whose densities are reported for 12 months Pre-treatment density is based on a pre-logging inventory conducted in 1995 in the same logging coupe before logging activities (Killeen e t al. 1998) Species Seeded species Anadenanthera Astronium Centro/obi um C opaifera Schinopsis unseeded species Caesalpinnia ( l s t year cohort : Caesalpinnia (2 n d year cohort Aspidosperma Acosmium Seed fall Seed addition Seedling density before or after bums both after before before both before after after after density of unseeded seeded Months added seeds density density following (m. 2 ) (m. 2 ) (m. 2 ) treatment 5 l.9 2 7 3 5 0.05 0 09 3 2 5 0 0 04 12 5 0 04 0 16 12 2 5 0 003 0 009 12 0.03 12 0.06 18 0 05 18 0.6 12 *** P < 0 001 ** 0 001 < P < 0.01 0 01 < P < 0 05 NS = not significant a Numbers too low for statistical analysis justification described in text. Sig of seeding treatment NS NS *** *** N/Aa Pre-treatment density (m.2) 0 05b 0 002 0 0 15 0 0 025 0 08 0 71 b Likely seedlings germinated in 1995 the year of the census as no seedlings of these species > Iy r old were found in forest plots in this stud y 00 N

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Table 4-3 Statistical analyses testing for the effect of the seed addition treatment. Seedling densities of Astronium Centrolobium and Copaifera could not be normalized therefore the effects of the seeding treatment was tested with a Kruskall Wallis non-parametric test for each month. Square root transformed seedling densities of Anadenanthera were normally distributed therefore a repeated measures AN OVA was used to test the effect of the seeding treatment and gap treatments for this species Because a significant tirne*treatment interaction was found for Anadenanthera an ANOVA was used to analyze each month separately Month 3 6 9 12 18 Astronium Copaifera x2 P x2 P 0 9 0.34 1.1 0 29 1.3 0.25 13.4 0 00 0 7 0.40 20.0 0 00 1.0 0.31 26 0 0 00 0 7 0.40 23 7 0 00 Anadenanthera seeding treatment gap treatments Month F p F 3 3.4 0 07 6 1 6 3.5 0 06 2 9 9 3.3 0 07 2.8 12 2 7 0 10 3 8 18 1.8 0.17 5.1 Centrolobium x2 P 1 0 0.32 6.2 0 01 15 2 0.00 15 2 0 00 3 0 0 08 p 0 000 0 03 0 03 0 01 0 001 83

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84 5~--------------------------, Group 1 seedlings -----high intensi ty burn 4 I -0---lo w intensity burn -vpl a nt removal 3 -vgap control II forest 2 -------w 0 0 5 ,_ 0.4 N E I-. 0 3 Q.) 0.. 0 1 :.a c:: :.=, 0 0 >-, 0 2 ..... v.i c:: Q.) 0 Group 2 seedlings j l . . . ..... l .. .......... I . . : : : L~-_j=-~----===i---.-?I--~::-p : : : = : -: : --Group 2 sprouts -0 1 -~_.;..----4----==----=--~ ... i .. . ........ -ii . . .. .. .. ii ... . . . . j .. .. .. . . .. . . . . .. iii 0 0 ....l.---9"'-'-----r----~-------,-----r----------,--------' 0 10 -r--------------------------------, Group 3 root sprouts 0 05 L--/J= . ---0 0 050 -,---------------------------, o.o2s J aroup3.seedlings 0.000 ] :--~_ ~ ~ 1.5 mo 3 mo 6mo 9mo 12 mo 18 mo Figure 4-2 Densities of commercial species over the 18 month sampling period follo w ing experimental bums Species are grouped according to similar regeneration strategies Group 1 (Anadenanthera and Astronium) were found predominately as seedlings Group 2 (Aspidosperma Copaifera, and Caesalpinia) were found as both seedlings and sprouts and Group 3 (Centrolobium) was found predominatel y as root sprouts ( bars = S E. ; n = 16)

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85 were rare established slowl y over the first year and many died by 18 months In contrast density of Centrolobium root sprouts remained relatively constant throughout the sampling period The prolonged seedling establishment and sprouting of species in these last 2 groups restricted the calculations of RGR to the later part of the 18 month san1pling period For these species I calculated RGRs from no earlier than 6 or 9 months due to the low seedling and sprout densities at 3 months This method of calculating RGR may have limited detection of treatment differences If the most rapid growth occurs during the first several weeks following germination or sprouting, then differences in RGRs measured after this initial growth spurt may not be detectable For this reason seemingly contradictory results were obtained for some species in the following results sections (i.e ., Aspidosperma and Copaifera), where seedling heights were significantly different among treatments whereas RGRs were not. Treatment effects on seedling densities, heights, RGRs and survival of seeded species Anadenanthera. Seedlings of Anadenanthera were the most abundant of all commercial tree species with an average density throughout the treatments of 1 1 0.4 seedlings / m 2 (mean 1 S.E.) 18 months following treatments. All Anadenanthera seedlings in these analyses are from the 1997 cohort. Seed production of Anadenanthera in 1997 was larger than most years according to locals I did not encounter any seedlings> 1 yr-old in subplots and due to extremely low seed production in 1998 I also did not encounter Anadenanthera seedlings germinating in 1998 Although Anadenanthera has the ability to coppice (pers obs ) I did not encounter an y sprouts in the 4 m 2 subplots Three months after bums Anadenanthera seedling density was highest in forest plots (F = 6 1 P < 0.001 ; Figure 4-3) Density declined in forest plots due to high mortality and b y 12 months seedling densities were highest in the plant removal treatment (F = 3 8 P = 0 02) Anadenanthera seedling density was lowest in gap controls throughout the stud y

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86 4 A. "' 3 E --Cll 0() 2 =i rt: .s :a (I) (I) Cll 1 0 ~ -i:---... ::. . .. ~ ... ..... ~~:.i 3 mo 6mo ~ 9 mo 12 mo :w.._ . . ----vi----~ --v 18 mo . ... --v . . 8. 0 1 ...1..____.,____J'---~---'------'---'----'--~ ---'----'-----'----'-----'----'------' 3 mo 6mo 9mo 12 mo 1 8 mo 120 -,---------------------------, 100 e &o --high intensity --0low intensity -Tplant removal C. u __... .E 60 -v-gap control 0() a:3 40 ..c: forest -if .. -20 ~-~-=-~::::::--II 0 ...L_ ______________________ _ __ __J 6mo 9mo 12 mo 18 mo 0 01 2 ~ ----------------------, 0 0 1 0 0 00 8 0 006 p::: 0 00 2 D. 0 004 1 0 0 0 0 -+----~ ----~----1----~ 0-6 mo 6-12 mo 12-18 mo Figure 4-3 A Seedling density B. percent seedling survival C seedling height and D. seedling relative height growth rates for Anadenanthera co lubrin a in the fou r gap treatments and forest plot. All graphs follow the legend shovm in graph C (bar s= SE .; n = 16)

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87 Height of Anadenanthera seedlings averaged 49 5 cm after 18 months The tallest seedling measured was 3 m and was found in a high-intensity bum plot Seedlings in high and low intensity bum treatments were significantly taller than seedlings in the gap control or forest understory ; seedling height in the plant removal treatment was intermediate (F = 15 .4 P < 0 00 l) Correspondingly differences in RGRs among treatments were significant and patterns followed those for height (F= 18.4 P < 0 001) Anadenanthera seedlings were taller and had higher RGRs in gap centers than near gap edges (F = 21.8 P = < 0 001 ; F = 16 0 P = < 0 001). After 18 months seedlings in gap centers averaged 25 cm taller than seedlings near gap edges A mean of 21 % (at 9 months) and 12% (at 12 months) of Anadenanthera seedlings we re browsed (tracks suggest by bracket deer Mazama sp ) Seedlings in high-intensity bum treatments suffered the highest rates of browsing while seedlings in the forest experienced the lowest rat es (F = 11.16 P < 0 001) Astronium. Pattern of Astronium seedling density was strongly related to disturbance intensity (Table 4-4). Throughout the 18 month sampling period the highest seedling densities were found in the high-intensity bum treatment plots followed by the low-intensity burn and plant removal treatments (Figure 4-4). Mortality rates in these treatments was moderately high. Only 2 Astronium individuals were found in all 16 gap control plots (but these died during the 18 month sampling period) and no Astronium seedlings were found in forest plots I encountered only one sprout of Astronium in the permanent subplots (a low-intensity bum treatment) and did not include it in these analyses Height of Astronium seedlings averaged 110 14 cm after 18 months the tallest mean seedling height among species Although heights and RGRs of Astronium seedlings were not significantly different among treatments (F = 4 0 P = 0.11 ; F = 1 0 P = 0.40 respectively) there was a distinct trend of taller Astronium seedlings with increasing disturbance intensity Mean

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Density at 18 months 0.4 ~----------0 J Astronium urundueva* 0 2 0.1 0 0 0 4 .--------------, O l Centrolobium microchaete 0 2 0 1 II O __ 11;.;;IIS ::,Z _ ., r-,,, ::\ '-_ _, .,.._"";.;. ; ,:_ .~ ---1 0. 4 ------------, 0 J 0 2 0 1 0 3 0 2 0 1 C aesalpinnia pluviosa (pre1 98 cohorts) C aesalpinnia pluviosa 1 98 cohort Aspidosperma rigida 0 0 L_ -~-~--= --l':lli -llJ,'5 -----' 0 4 ~---------, 0 3 0 1 a a .L._ -~ -=-1 o --------------, 0 8 Acosmium cardenasii 2 3 4 5 88 Height at 18 months 500 -----'-----------, 400 JOO 50 -E (.) '-' ..... 25 ..c:: 00 a:3 ..c:: 500 400 JOO 200 100 0 500 400 300 i I :~ I 500 400 JOO 200 1 00 1 2 3 4 5 Figure 4-4. Densities and heights of seedlings (black bars) and sprouts (gre y bars) of 6 sp e cies 18 months following treatments(*= seeded species). Treatment codes along the x-axis are : l = high intensity bum 2 = low intensity bum 3 = plant remo v al 4 = gap control 5 = forest (bars= S E .; n = 16)

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Table 4-4 Statistical analyses testing for differences among gap treatments and forest understory plots Seedling densities of A s tronium Aspidosperma C a es alpinia Ce ntrolob i um and C opaifera could not be normalized therefore the effect of gap treatments was tested using a Kruskall Wallis non-parametric test for each month Seedlings and sprouts were also analyzed separately for these species (if applicable) Square-root transformed densities of Acosmium were normally distributed therefore a repeated measures ANOVA was used to test of the effect of the gap treatments and regeneration mode (seedling or sprout) for h . A . fi d d th edl. d th l d 1 t 1s species s1gm cant regeneration mo e treatment was oun ere ore se mgs an sprouts were en ana yze separate y Astronium Aspidosperma Caesa/pinia Centro/obium Copaifera seedlings onl) seedlings sprouts seedlings sprouts seedlings root suckers seedlings sprouts Mo x2 p x 2 p x2 p x2 p x2 p x2 p x 2 p x2 p x 2 p 3 25 1 < 0 001 9.9 0 04 7 7 0 10 21.2 0 8 6 0 07 4 0 0.41 11.1 0 03 17 5 0 002 6 2 0.18 6 34 7 < 0 00] 9 9 0 04 6 7 0 16 14 6 0.006 8.6 0 07 16 2 0 003 11.2 0 02 12.8 0.01 6 2 0 18 9 37 6 < 0 001 4.0 0.40 12 1 0 02 11.6 0 021 6 8 0 15 11.8 0 02 23 2 < 0.001 12 3 0 02 5 2 0 26 12 37.0 < 0 001 5 5 0 24 12 0 0 02 11.6 0 021 5 6 0 23 5 6 0.24 18 5 0 001 10.4 0 03 6 1 0 19 18 37 5 < 0.001 7 5 0 11 12.1 0 02 4 3 0 37 5 6 0.23 2 0 0 73 21.6 < 0 001 11.5 0 02 6 1 0 20 Acosmium Acosmium seedlings sprouts Source of variation df F p Source of variation df F p F p Block 15 2.2 0 01 Block 1 1.1 0.40 1.2 0 29 Regeneration mode 1 0.0 0 93 Treatment 15 12.1 <. 001 11.3 <. 001 Treatment 4 3 6 0 01 Error 4 Mode* Treatment 4 4 0 0 01 Error 65 00

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90 height of Astronium seedlings in high-intensity bum treatments was 150 cm more than twice the height of seedlings in plant removal treatments (65 cm) Also the tallest A st ronium seedling (4 m) was found in a high-intensity bum plot. Copaifera. Copaifera regenerated from both seeds and sprouts although the overall density of seedlings was more than 10 times higher than sprouts (Table 4-5) Copaifera seedlings were most abundant in gap control plots ; sprout density did not differ an10ng treatments (Table 44 Figure 4-4) Mortality was low throughout the study period particularly for sprouts (Table 4-6). Sprouts of Co paifera were taller than seedlings (F = 8.4 P = 0 01) although their RGRs did not differ (F = 1 8 P = 0 20) Height and RGRs did not differ among treatments (F = 0 3 P = 0.47 ; F = 0 3 P = 0 84 respectively) Centrolobium. Ce ntrolobium regenerated both from seeds and root sprouts Densi ty of root sprouts was higher than seedling density at 6 months (Table 4-5). Apparently none of the seedlings arose from naturall y dispersed seeds ; natural regeneration of this species was composed entirely of root sprouts At 3 months Centrolobium root sprouts were most abundant in the plant removal treatment ; from 6 to 18 months they were most abundant in the low-intensity bum treatment (Table 4-4 Figure 4-2) At 6 months Centrolobium seedlings were most abundant in the high-intensity bum treatment ; at 9 months they were most abundant in the burned and plant removal treatments Seedling mortality was greater than root sprout mortality (Table 4-6) Centrolobium sprouts averaged 267 cm tall after 18 months more than 7 times the mean height of Centrolobium seedlings (37 cm ; F = 39.2, P = 0 003 ; Figure 4-4) Similar! _, RGRs of Centrolobium sprouts were also higher than those of seedlings (F = 13 6 P = 0.01 ) No differences in heights at 18 months or RGRs were detected among treatments (F = 1.0 P = 0.4 9 ; F = 0 1 P = 0 97. respectivel y ) Schinopsis brasilensis A total of 9 Schinopsis seedlings were recorded in all subplots although the maximum number at any one sampling period was 7 (Table 4-7) After 18 months 6

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T a ble 4 5 Statistical analyses comparing see dl ing and s p rout densities of Aspidosperma C aesalpinia Ce ntrolobium and C opaifera. See dl ing an d s p ro u t d e n sities of these s p ecies could not be normalized, therefore a Kruskall-Wallis non parametric test was used to analyze each month 91 C opaifera C entrolobium Aspido s perma C ae s alp i nnia Month x 2 p x 2 p x 2 p x 2 p ,, 5 6 0 02 3 7 0 06 7 7 0 006 0 13 0.72 .J 6 22.2 0 00 5 9 0.02 11.2 0.001 0.32 0 57 9 27.6 0.00 2.6 0.11 10.2 0 001 0 04 0 84 12 31.5 0 00 2 2 0.13 15 2 0.000 0 001 0 98 18 35 9 0 00 3.1 0.08 10 2 0 001 15 6 0.00

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Table 4-6 Killing power a parameter similar to mortality rate for 5 commercial tree species in the four gap treatments and forest understory plots Killing power (k) is calculated separately for seedlings and sprouts of each species except for A s tronium for which no sprouts were found Killing power is calculated as (log 10 <1.x log 10 <1.x + i) where
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93 seedlings remained for an overall density of 0 005 seedlings/m 2 With such a small sample it is difficult to determine if seed addition significantly increased seedling density Although the majority of seedlings were in seeded plots ( 6 of the total 9) distribution among gaps was distinctly clumped with all 9 seedlings found in only 4 gaps and 4 seedlings in a gap with a mature Schinopsis nearby These observations suggest that most of the seedlings established from naturally dispersed seeds Although Schinopsis was not statistically analyzed due to its small sample size it appears to favor establishment in burned areas Of the 9 seedlings 55% were in high-intensity burn plots 25% in low-intensity burn plots and one seedling each in a plant removal treatment and forest plot (although both of these seedlings died). Average height of Schinopsis seedlings surviving at 18 months was 43 cm Treatment effects on seedling densities, heights, RGRs, and survival of non-seeded species Acosmium cardenasii. Acosmium was the second most abundant species with an average density of 0 6 individuals/m 2 18 months following treatments Roughly half (57%) of Acosmium individuals were sprouts (Table 4-4 ) Patterns of seedling and sprout densities among treatments differed Seedling densities showed a negative response to disturbance intensity (Figure 4-4). The lowest seedling densities were found in high-intensity burn treatments while the highest densi ties were in gap control and forest plots In contrast sprout density was highest in the plant remo v al and low-intensity burn treatments There were no differences among seedlings and sprouts in height or RGR (F = 2 7 P = 0 11 ; F = 0 7 P = 0.42 respectively) although heights and RGRs of both seedlings and sprouts differed slightly among treatments Both seedlings and sprouts were tallest in gap controls (F = 4 8 P = 0 01 ) a pattern that likely resulted because individuals already established in these plots prior to gap formation were cut to a height of 2 m in the gap control whereas in the other treatments the y were cut to soil level or burned. Actual RGR was higher in high and low-intensity burn treatments than gap controls (F = 6 0 P = 0 002)

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94 Caesalpinia pluviosa Caesalpinia had abundant seed rain the second y ear of the study (1998) Approximately 0 06 new seedlings/m 2 were added to the existing seedling bank of 0 03 seedlings/m 2 C aesalpinia regenerated from both seeds and sprouts and before the abundant seed fall in 98 seedlings and sprouts occurred with roughl y equal frequency (0 015 seedlings v s 0 016 sprouts/m 2 ) After the seed rain in 1998 seedlings were 3 times more common than sprouts ( Table 4-5) Excluding the 98 cohort Caesalpinia seedlings were most abundant in forest plots ; there was no difference in sprout density among treatments (Table 4-4 Figure 4-4) Among seedlings of the 98 cohort there were no differences in density among treatments Mortality was low throughout the stud y p e riod for both seedlings and sprouts (Table 4-6) C aesalpin i a sprouts were taller than seedlings (F = 9 3 P = 0 01) although RGRs did not differ (F = 0 2 P = 0 66) There were no treatment differences in heights or RGRs among individuals in the pre' 98 cohort (F = 0 1 P = 0 95 ; F = 0 6 P = 0 63 respectivel y ) or among seedlings in the 98 cohort (F = 1.0 P = 0.44 ; F = 1.0 P = 0.44 respectivel y ) Aspidosperma rigidum. Aspidosperma sprouts were more abundant than seedlings (Table 4-5) Differences in seedling density among treatments were only found at 3 and 6 months ; at these sampling periods seedlings were most abundant in the forest understory (Table 4-4 Figure 4-4) Sprout density was most abundant in the low-intensity bum treatment after 9, 12 and 18 months Mortality of both seedlings and sprouts was moderate (Table 4-6) Aspido s perma sprouts were taller than seedlings (F = 14.8 P = 0 003) RGRs among sprouts and se e dlings did not differ (F = 0.01 P = 0 84). There was no pattern in height or RGRs among treatments (F = 3.4 P = 0 07 ; F = 1.3 P = 0 32) Additional commercial tree species found at low densities. Four additional commercial species were found at very low densities in treatment subplots : Cordia alliodora Phyllos ty lon rh a mnoid es, Pith e c e llobium sp ., and Platymiscium ul e ii (Table 4-7). Combined densit y of th e se species and Schinopsis a seeded species was highest in high-intensity bum subplots followed b y

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Table 4-7 Densities and heights of commercial tree species found in the four gap treatments and forest understory plots at low frequencies Total number is the number of individuals in all treatments at 18 months following treatments. Percent sprouting is based on the total number of individuals Densities and heights within treatments represents values 18 months following treatments. Treatments High intensity Low intensity Plant removal Gap control Forest Total Percent density height density height density height density height density height Seecies number serouting (#/m2) {cm} (#/m2) {cm} (#/m 2 ) {cm} (#/m 2 ) {cm} (#/m 2 ) {cm) seeded species Schinop s is brasili e nsi s 6 0% 0.020 45 0 004 35 0 0 0 unseeded species Cordia alliodora 1 100% 0 004 190 0 0 0 0 Phylostylon rhamniodes 3 0% 0 008 18 0.004 5 0 0 0 Pithecellobium sp 4 50% 0 004 25 0 008 65 0 004 135 0 0 Platymiscium ulei 7 100% 0 004 110 0 012 163 0 012 210 0 0 Averages per treatment 0 005 86 0 006 67 0.004 135 0 0

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96 low-intensity burn and plant removal subplots. None of these species were found in gap control or forest subplots Discussion Effects of treatments on seed predation The rates of seed predation of Copaifera ( 17%) and Centrolobium (5%) observed in this study are much lower than those reported in the literature Crawley ( 1992) found that post dispersal seed predation reviewed over a variety of ecotypes averaged 45-50% and often approached l 00%. Among studies conducted in tropical forests, reported rates of seed predation are even higher (36-98% DeSteven and Putz 1984 ; 50-90% Schupp 1988 ; 27-98% Alvarez Buyella and Martinez-Ramos 1990; 25-94% Hammond 1995 ; 20-100% Holl and Ludlow 1997). The low seed predation found for these 2 species in this study may have been due to several factors Seeds removed from areas of high seed density near parent trees and placed beyond the typical range of seed dispersal may escape predation by specialists (Janzen 1970 Connell 1971 but see Augspurger and Kitajima 1992). Both Copaifera and Centrolobium appeared to suffer high rates of seed predation on or near parent trees Fredericksen et al. (2000) reported that 34 % of Copaifera and 41 % of Centrolobium seeds beneath parent trees were damaged. Curculionid beetles and rodents were observed eating Copaifera seeds and Cebus monkeys parrots, and squirrels were observed eating Centrolobium seeds (Fredericksen et al. 2000) Centrolobium also suffered high rates of pre-dispersal predation: of a sample of 110 recently dispersed fruits, 28% had either exit holes or larvae inside (pers. obs ) The low rates of seed predation observed in this study may also be due to the fact that only the gap control treatment offered any protection for potential seed predators. Most woody debris was removed or burned in the other three treatments, and colonizing vegetation was not thick until a few months after seeds were placed out. Schupp et al. (l 989) found that seed predation was

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higher in disturbed areas than forest understories due to the number of small rodents that take refuge in brush 97 I found no effect of distance from forest edge on seed predation a result which agre e s with those of Holl and Ludlow (1997) Overall these low predation rates suggest that seed predation had little effect on seed densities of Copaifera and Centrolobium after their dispersal or placement in gap treatments However seed predation may have been important for other species or during other seasons For example, in a different experiment (Chapter 3) more than half of 720 Anadenanthera se eds we re damaged or removed primaril y b y leaf cutter ants (Atta sp ) This occurred in November after the start of the rainy season when seeds has imbibed water and w ere starting to germinate In contrast I did not observe any predation of Anadenanthera seeds that were placed out in October before rains began Effects of controlled burns on seed germination of selected tree species Decreased germination rates of seeds subjected to low-intensity bums indicates this treatment may have contributed to lower seedling densities for species with peak seed fall before bums (e.g ., Copaifera Centrolobium) Low seedling densities from seed banks sampled after low intensity bums (Chapter 5) suggests most species in this forest have decreased germination after exposure to low-intensity fire. Extremely low seedling densities from seed banks sampled after high-intensity bums (Chapter 5) indicate that most burned seeds are killed during intense fires Similar results were reported by Brinkmann and Vieira (1971) in their study of the effects of fire on regeneration from buried seeds in the Amazon rain forest. They found that seeds of 31 tropical tree species were killed during slash burning These findings indicate that the timing of controlled bums relative to peak seed fall can alter amounts of viable seed and therefore is an important factor in commercial seedling density

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98 Effects of seeding treatment on tree seedling densities Seed addition increased seedling densities of 2 species (Centrolobium and Co paifera ) but did not affect seedling densities in the remaining 3 (Anadenanthera Astronium and Schinopsis) Increased seedling densities following seed addition indicates that seeds are limiting either because seed production is low or because seeds fail to reach all sites suitable for their germination and establishment (i e ., dispersal limitation ; Dalling 1998). Both Copaifera and Centrolobium produced abundant seed in 1997 (Fredericksen et al 2000) so it is unlikely that seed production was limiting during the year of this study. Both species show characteristics of dispersal limitation due to either their dispersal mode or the distribution of mature trees For example Copaifera seeds mostly fall beneath the parent tree or may be carried short distances b y ants or small mammals (Fredericksen and Justiniano 1998). Only 4 of 16 gaps had mature Copaifera within 15 m of the gap edge so it is likely that naturally dispersed seeds were scarce in the remaining 12 gaps. Centrolobium produces large samaras fruits that can easily reach gap centers aided by the strong winds of the dry season (Chapter 5) However only 9 of the 16 gaps had mature Centrolobium trees within 15 m of the gap edge so it is likely seeds of this species were scarce in the remaining 7 gaps The timing of seed dispersal relative to bums also may have contributed to the significant effect of the seed addition treatment for Copaifera and Centrolobium. Seed fall of both species occurred before controlled bums Naturally dispersed seeds present in plots before bums were likely damaged or killed which would contribute to higher seedling densities in seeded plots of the burned treatments Astronium and Schinopsis seedling densities were apparently not increased b y seed addition Seedling densities of Astronium and Schinopsis were highest in the intensel y burned treatments and absent from forest and control plots This pattern suggests that for these 2 species seedling recruitment is limited more by the a v ailabilit y of sites suitable for germination and

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99 establishment (establishment limitation ; Dailing 1998) than b y seed production or dispersal. Although it is unclear from this study it can be argued that higher Astronium and Schinopsi s seedling densities in the intensely disturbed treatments were due more to higher seedling survival than to enhanced seed germination. Although intense disturbances produce several effects which have been shown to promote seed germination of some species including litter removal (Vazquez Yanes et al. 1990) soil exposure (Putz 1983) increased light (Vazquez-Yanes and Smith 1982) and increased soil temperature (Probert 1992, Cervantes 1996) regulation of germination by these cues are more characteristic of a specialized group of short-lived pioneer species such as Cecropia spp Seeds of most tropical tree species lack a period of dormancy (Vazquez-Yanes and Orozco Segovia 1993 Garwood 1989) ; those that exhibit dormancy usually germinate in response to increased moisture (Vazquez-Yanes and Orozco-Segovia 1993) particularly in seasonal forests (Garwood 1983). It is likely that seeds of Astronium and Schinopsis germinated but quickl y died and were missed in the first census Seedling mortality is reported to be higher in forest understorys than gaps (Veenendall et al. 1995) particularly for shade-intolerant species (Kitajima 1994 Li et al. 1996) and the greatest mortality due to unsuitable habitat conditions occurs very early in the life of seedlings (Li et al. 1996). In contrast to the other 4 seeded species Anadenanthera recruitment does not appear to be limited by seed production seed dispersal, or suitable germination and establishment sites As previously discussed seed production of Anadenanthera was abundant in 1997 And although Anadenanthera seeds are primarily gravity dispersed (Justiniano and Fredericksen 1998) all 16 gaps had at least one mature Anadenanthera within 15 m and seeds were dispersed into most gaps (Chapter 5) Also seedling establishment does not appear to be controlled by microsite availability as Anadenanthera seedlings were present in all treatments Although Anadenanthera seedling recruitment was abundant during the year ofthis study it may be limited in other y ears Taking into account the scarcity of seeds produced in 1998 Anadenanthera s reproductive phenology fits

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100 characteristics of a masting species (Kelly 1994) In non-masting years seed production and subsequent seedling r e cruitment of masting species is low (Whitmore 1984 ) Therefore for th i s species synchrony of harvesting activities and silvicultural treatments with y ears of seed production is necessary to ensure sufficient regeneration ; silvicultural treatments applied during years of low seed production will not result in abundant seedling recruitment. Timber species with irregular reproductiv e phenologies may complicate long-term management planning due to the difficulty of predicting masting years. Gap treatment effects on tree regeneration Regeneration strategies. Gap treatments had significant effects on the regeneration of all of the tree species studied but each species responded differently to the treatments While seedling density and height of some species varied among treatments the mode of regeneration (from seeds or sprouts) of other species was more affected The response of these tree species to the treatments highlights their unique regeneration strategies ; no single pattern neatl y describes the w hole However I will attempt to generalize the responses according to the predominate regeneration mode of the species (from seeds sprouts or both) Based on re g eneration mode three patterns of species response to treatments can be identified : species that regenerated primarily from seed (Anadenanthera and Astronium) species that regenerated from both seeds and sprouts (Caesalpinia Copaifera Acosmium Aspidosp e rma) and species that regenerated b y root sprouting or suckering (Centrolobium) These groups correspond to groups 1 2 and 3 in Figure 4-2 (Acosmium was not included in this graph because it is not commerciall y valuable) A summary of treatment effects on the density height and RGRs of these groups is displa y ed in Table 4-8 The response of these groups to treatments reflects their pre-gap seedling density which is more-or-less related to a species shade tolerance The species that regenerated b y both seeds and sprouts are for the most shade-tolerant or partially shade-tolerant (Pinard et al 1999)

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Table 4-8. Summary of treatment and regeneration mode effects on densities heights and RGRs of 7 commercial tree species and one non-commercial tree species (Acosmium cardenasii ) Species are grouped into 3 classes : I) species that regenerated from seeds ; 2) species that regenerated from both from seeds and sprouts ; and 3) one species that regenerated from root sprouts Significant differences among treatments or mode of regeneration arc noted as "S Non-significant differences are noted as "NS Where results are significant the treatment or mode of regeneration with the highest values (i e most abundant tallest, or highest RGR) are noted Treatments codes are : 1 = high intensit y bum ; 2=low intensity bum ; 3 = plant removal ; 4 = gap control: and 5=forest understory Regeneration modes are symbolized by seedlings (sd) sprouts (sp) and root sprouts (rs) Result s of statistical tests summarized in th.is table are displayed in Tables 4-3 4-4 and 4-5 or in the te>..'1 Species and regeneration groups 1) Seedlings on(v Anadenanthera colubrina Astronium urundueva 2) Both seedlings and sprouts C aesalpinnia pluviosa J st co h o rt seedlings sprouts C aesalpinnia pluviosa 2 " c ohort Copaiferachodatiana seedlings sprouts Aspidosperma rigidum seedlings sprouts Density3 trt with s1g most s 5 . 3 c s I s 3 NS NS s 3 4 5 NS s 5 s 2 Treatment effects Heightb trt with s1g tallest s 1 NS NS" NS NS NS RGR s1g s NS NS NS NS NS trt with highest Mode of regeneration Densit s1g N I A NIA NS NIA s s abundant mode sd sd sd sd Sp Height tallest s1g mode s sp s sp s sp RGR h.ighest s1g mode NS NS NS 0 .._

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Table 4-8 (cont.) Treatment effects Seedlings vs. resprouts Density Height RGR Density abundant Height RGR trt with trt with trt with tallest highest Species and regeneration groups s1g most sig tallest s1g highest s1g mode s1g mode s1g mode Aco smium cardenasii seedlings sprouts 3) Root sprouts s s 5 4 3,2 s 4 s I 2 NS NS Centrolobium microchaete NS NS S rs S rs seedlings (sown) S 1 root sprouts S 2 3 a = Kruskall-Wallis tests were used to compare densities of all species except Anadenanthera and Acosmium b = ANO VA models with treatment and mode of regeneration as factors were used to test heights and RGR's If the treatment*mode of regeneration interaction was not significant seedlings and saplings were not analyzed separatel y c = most abundant in the forest after 3 months most abundant in plant removal treatment after 18 months "NIA" Statistics not performed all seedlings NS s rs 0 N

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103 Characteristic of shade-tolerant species they had seedling banks in forest understories before treatments (Table 4-2) and therefore a pool of individuals from which to sprout. Anadenan t h e ra and Astronium show shade-intolerant characteristics and their pre-gap seedling densities were low (Table 4-2) offering no structures from which to sprout. These species also had abundant seed fall in 1997 and moderate to high germination rates creating a large seedling cohort the first y ear of this study Centro/obium which regenerated primarilyas root suckers provides an exception to these 2 groups Sprouts of Centrolobium did not originate from the seedling bank (pre-gap seedling densities of Centro/obium were very lownone were encountered in the 1995 inventory ; Killeen et al. 1998), but from the root systems of mature trees. Although seedlings of Centro/obium did establish in all treatments, apparently none of these seedlings were from naturall y dispersed seeds ; natural regeneration of Centrolobium consisted only of root suckers Similarly regeneration of Platymiscium u/ei was only encountered as root suckers and may have the same regeneration strategy as Centrolobium Gap treatment effects on density and survival Species regenerating as seedlings only. Because the conditions that favor seedling establishment may not be the same as those that favor sprouting density of regeneration among species varied according to their predominant mode of regeneration The species that regenerated as seedlings Anadenanthera and Astronium were most abundant in treatments that favored their seedling establishment and survival although initially the optimal treatment differed among these two species. Astronium seedlings were most abundant after high-intensity bums and absent from forest and gap control plots a pattern that suggests Astronium seedlings are unable to establish or survive under shade or need exposed soils to germinate and establish. Although canopy cover > 2 m tall in gap control plots was similar to the other gap treatments (Chapter 2), colonizing plant

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104 cover< 2 m tall was almost 100% within less than 6 months following burns (Chapter 5) Cover by low vegetation potentially has a more negative impact on newly emerging seedlings than cover higher in the canopy (Marquis et al. 1986). Similarly Kennedy and Swaine (1992) showed that experimental removal of advance regeneration resulted in increased survival of seedlings that colonize gaps from seed in lowland dipterocarp forests Initial plant cover was zero in treatment plots with Asrronium recruitment (plant removal and burn treatments) In high-intensity burn treatments, where Astronium recruitment was greatest total plant cover remained low( < 25%) throughout the first year (Chapter 5). In contrast Anadenanthera seedlings were initially most abundant in forest plots. This density pattern is likely a combination of higher seed fall into forests than gaps (Chapter 5) and, unlike Astronium, the ability of Anadenanthera seedlings to persist in shade However due to high mortality of Anadenanthera seedlings in forest understories, patterns of seedling distribution among treatments may resemble those of Astronium after several years During the first year of this study, forest seedling densities declined to levels similar to gap treatments. Also although a density of 0 05 seedlings/m 2 was found in 1995 in this same forest (Killeen et al. 1998) I did not encounter any seedlings older than the 1997 cohort in forest plots supporting the idea that seedlings of Anadenanthera are short-lived in the shade. Seedling survival in the second year of the study was highest in high-intensity burn treatments possibly due to the larger size of these seedlings (Lieberman 1996). Therefore seedling density after several years may be highest in this treatment. Enhanced seedling establishment or survival of Astronium and Anadenanthera in the burned and plant removal treatments can be attributed to a variety of factors including reduced competition, removal of slash and litter and/or soil disturbance In other neotropical forests the abundance of species colonizing from seed have been shown to increase in canopy gaps that have been further disturbed by either logging or fire In dry forests of Mexico for example Dickinson

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105 ( 1998) found that colonizing seedling density in logging gaps traversed by skidders was more than double that in logging gaps not disturbed by skidders. Similarly high densities of pioneer seedlings were reported to establish in canopy opening after fire in Amazonian Brazil (Holdsworth and Uhl 1997) However in these studies the effects of soil disturbance and litter removal by fire or logging were not independent of reduced competition due to removal of established vegetation (Dickinson 1998) Removal of advanced regeneration from gaps in my study (plant removal treatment) resulted in higher seedling densities of Anadenanthera than burned treatments but lower densities of Astronium. Therefore the effects of rnicrosite on seedling establishment vary even among species with similar regeneration strategies Although early seedling density is a function of seed germination and establishment long tenn seedling densit y is more affected by survival. As previously discussed greater survival of Anadenanthera seedlings after high-intensity burns will likely result in higher seedling densities in this treatment. Hence rnicrosite conditions that favor germination and establishment do not necessarily benefit subsequent growth and survival. Similar conclusions were reached by Stanley ( 1999) who found that although seed germination of Swietenia macrophylla was lower in burned than unburned gaps subsequent seedling growth rates were higher. Species regenerating as root sprouts. Natural regeneration of Centrolobium was most abundant in the plant removal and burn treatments suggesting these treatments favor root sprouting Similarly root sprouts of Platymiscium uleii were found onl y in the plant removal and bum treatments Root sprouting in Fagus grandifolia (American beech) developed after roots wer e i1~ured or exposed to air or elevated temperatures (Jones and Rayna! 1988) conditions likely to occur in the plant removal and burn treatments Similarly Dickinson (1998) found that traversal by skidders during logging operations promoted subsequent root sprouting by several dry forest tree species in Mexico Root damage can lead to the formation of callus tissue from which adventitious buds and sprouts arise (Jones and Rayna! 1988). Root-sprouting ma y also be

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106 promoted by the loss of apical dominance (e g for Populus tremuloides Schier 1975) This mechanism would account for high densities of root sprouts in gap control plots where stumps of harvested Ce ntr olobium we re located Loss of apical dominance would also explain intermediate frequencies of root sprouts in logging gaps that were not traversed by skidders in Dickinson s study (1998). Species regenerating as both seedlings and sprouts. Density patterns of species regenerating as both seedlings and sprouts (Acosmium, Copaifera Caesalpinia, Aspidosperma) differed according the regeneration mode Of this group Acosmium revealed the most distinct response made more clear perhaps by its high abundance Acosmium had the highest pre treatment density (0 7 seedlings/m 2 in forest plots in a 1995 inventory ; Killeen et al. 1998) The density of seedlings and sprouts in my study largely reflect what happened to this seedling bank during treatments Seedlings were mostly undamaged in forest understorys and gap controls therefore the highest seedling densities were found in these areas All seedlings were damaged in the low-intensity bum and plant removal treatment rendering sprouts more common in these treatments. Seeds and stems were likely killed in the high-intensity burn treatment resulting in the lowest densities for both seedlings and sprouts in this treatment. Seedling densities of Copaifera, Aspidosperma and Caesalpinia were also highest in the less disturbed treatments Patterns of regeneration among these species stresses the important role advance regeneration pla y s in determining the success of shade-tolerant species following disturbances Where advance regeneration is abundant, less intense disturbances such as canopy gap formation will favor these species However more intense disturbances such as severe fire may kill advance regeneration thus favoring species that colonize from seed Dickinson ( 1998) reported similar results in his study in Mexico Shade tolerant species were most abundant in forest understories and natural canop gaps but in logging gaps, skidder damag e reduced the abundance of these species

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107 Interestingly sprout density of Aspidosperma was highest in the burned treatments This pattern may be due to its ability to survive fire Based on its bark thermal properties and thickness Aspidosperma was classified as the most fire resistant among the species focusc::d on in my study (Pinard and Huffman 1997) Survival by sprouting among the other shade-tolerant species in high intensity bum plots was comparatively low again possibly due to their bark characteristics Acosmium Copaifera and Caesalpinia were classified among the least fire-resistant among canopy trees in Lomerio (Pinard and Huffman 1997) Gap treatment effects on heights and RGRs Treatment effects on the height and RGR of individuals of a particular species also are more clearly interpreted in terms of the species regeneration strategy by seeds, sprouts or a combination of both In general, species that regenerated primarily by seed (Anad e nanthera Astronium) revealed patterns of treatment effects on height, while species that regenerated by sprouting (Centrolobium) or a combination (Caesalpinia Copaifera Aspidosperma) did not show significant treatment effects (Table 4-8) Species regenerating as seedlings only. Seedlings of both Anadenanthera and Astronium were tallest in high-intensity burn plots and average height decreased with decreasing treatment intensity Similarly Stanley (1999) found that relative height growth of Swietenia seedlings was greater in burned than unburned gaps in Guatemala High growth rates of seedlings establishing after high-intensity burns is not unexpected because this treatment had the slowest recovery of vegetation among all gap treatments (Chapter 5) and therefore less aboveand below ground competition. In addition to less below-ground competition, soil sampled from the high intensity bum treatment had higher availability of ammonium nitrate and phosphorous as well as higher cation concentrations (Ca Mg and K ; Chapter 3) It is not clear from this study which factor had more of an effect on seedling growth increased light availability or increased soil nutrients Seedlings of Anadenanthera were taller in

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108 gap centers than gap edges, which suggests that even slight increases in light availability can enhance seedling growth On the other hand by comparing Anadenanthera seedling height in the low-intensity bum and plant removal treatments, an argument can also be made for a positive effect of increased soil nutrients on seedling height. Recovery of vegetation after low-intensity bum and plant removal treatments was not significantly different (Chapter 5) but soils sampled after low-intensity burns had significantly higher concentrations of soil nutrients (Chapter 3) The greater height of Anadenanthera seedlings in low-intensity bum treatments suggests that below ground effects may also significantly increase seedling growth. Many studies have addressed the importance of abovevs below-ground effects on seedling growth In greenhouse studies, Kitajima ( 1992) showed that light availability was the primary factor limiting seedling growth in a cross-factorial experiment controlling nutrient or water supply with light availability Conclusions drawn from field studies reveal that the relative importance of below-ground effects vary according to the site and species in question In dry forests, reduced below-ground competition appears to benefit seedling growth. For example in a study conducted in a dry forest in Costa Rica, Gerhardt (1995, 1996) reported that root trenching increased seedling growth of 4 tree species during the rainy season and increased their survival during the dry season In a similar study conducted in a dry forest in Mexico Dickinson ( 1998) found that seedling height of Swietenia macrophylla increased in response to reduced root competition. Studies conducted in infertile sites also conclude that below-ground competition may limit seedling growth (e.g., Putz and Canham 1992 Coomes and Grubb 1998). In contrast to the results from dry or infertile forest sites studies conducted in La Selva a "ct forest \\ ith high soil fortility report that soil resources do not significantly limit seediing growth (Ostertag 1998 Denslow et al. 1990). Correspondingly, Pooter ( 1998) working in a moist Amazonian forest in Bolivia found that watering tree seedlings during the dry season did not enhance growth

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109 Species regenerating as both seedlings and sprouts. Among most of the species regenerating as both seedlings and sprouts there was no effect of treatments on height or RGR (Table 4-8). This pattern may be because any possible treatment effects were obscured by the tremendous height differences between seedlings and sprouts. Sprouts were taller than conspec1fic seedlings therefore the height of an individual was more dependent on if it was a seedling or sprout rather than the treatment where it was located. This lack of significance may also have reflected a real lack of difference in seedling or sprout growth among treatments Growth of sprouts may have been less influenced by treatments because they depend on stored carbohydrate reserves in roots and therefore are less affected by aboveand below-ground competition during early developmental stages Seedling growth of these particular species may have been less affected by treatments than seedlings of Anadenanthera and Astronium because they are more shade-tolerant. Shade-tolerant species which have slower acclimation rates than shade-intolerant species (Kitajima 1996) grow more slowly even in high light (e.g ., Denslow 1987, Denslow et al. 1990) Perhaps due to their slower growth treatment effects on seedling height and RGR were not apparent after 18 months for these shade-tolerant species Also differences in below-ground effects among treatments ma y have been less important for shade-tolerant species due to their slower growth rates The relative importance of below-ground effects differs among shade-tolerant and intolerant species In Dickinson's study ( 1998) reduced root competition did not increase seedling height of a shade tolerant species as it did for Swietenia Also, Denslow et al. (1998) reported that whereas light demanding species responded better to increased soil nutrients in high light shade tolerant species showed no response Species regenerating as root sprouts. There were a greater differences in height and RGR between seedlings and root sprouts of Centrolobium than observed in the other species that regenerated from both seeds and sprouts Centrolobium root sprouts were more than 7 times taller

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110 than conspecific seedlings Centrolobium sprouts also were taller than sprouts of other commercial species and there were more Centrolobium sprouts > 2 5 m tall in the treatment plots than of any other tree species commercial or non-commercial (Chapter 5). Centrolobium sprouts may grow more rapidly than sprouts of other species because they sprout from the root systems of mature trees and therefore have larger carbohydrate reserves to utilize than sprouts of other species which mostl y originated from juveniles. Advantages of early colonization for seedling growth and survival Seedling density height and RGR of the 98 cohort of Caesalpinia did not differ among treatments possibly reflecting tha t after l year colonizing vegetation had obscured treatment effects Within 18 months plant cover < 2 m was greater than 60% in all gap treatments (Chapter 5) A rapid regrowth of vegetation is often observed after canopy gap formation or other disturbances in many tropical forests For example, Thompson et al. (1998) reported that gap area (as defined by vegetation> 2 m tall) was reduced by 85% within onl y 4 y ears after gap formation in an Amazonian forest. Due to this rapidly colonizing vegetation light temperature and relative humidity often to return to pre-gap levels within the first or second year after gap formation (e g ., Denslow et al. 1987) Where areas opened b y disturbance are quickly recolonized earl y establishment is vital to the success of regenerating individuals Individuals establishing early often enjo y greater resource availability than those establishing later (Canham and Marks 1985) and as a result growth and survival rates are reported to be higher for plants establishing soon after canopy gap formation in the tropics For example in treefall gaps in Panama earl y recruits of both pioneers and prima ry forest species grew faster than later recruits (Brokaw 1985a) Often it is the individuals that were present before disturbance (i e ., advance regeneration) that are the most successful. For example Uhl et al. (1988b) found that advance regeneration accounted for 97% of all trees > 1 m tall in single treefall gaps and 83% in multiple treefall gaps 4

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111 years after gap formation in Amazonia. Similarly, Brown and Whitmore (1992) reported that the most important determinant of seedling survival and growth in treefall gaps in Borneo was seedling size at the time of gap creation regardless of species The patterns of seedling and sprout density among the shade-tolerant species in this study suggest that these species rely mostly on advance regeneration following disturbances of low to medium intensity The general success of early recruits has important implications for studies such as this one, that examine regeneration only for short time periods Early survivorship has been shown to be a good indicator of subsequent survivorship for periods up to 6 years of age (Li et al. 1996) To the extent that individuals or species that are successful during the first year or two will continue to dominate over later recruits the future success of species in this stud y can be predicted based on their survival rates and height 18 months following gap treatments For example seedlings of Anadenanthera and Astronium will likely be most successful in the high-intensity burn treatment. Root sprouts of Centrolobium will likely dominate individuals of other commercial tree species particularly in gaps formed by a harvested mature Centrolobium. Shade-tolerant species (Acosmium Caesalpinia Aspidosperma, and Copaifera) will be most successful in gap controls and sprouts of these species will likely be more dominant (i.e. larger) than conspecific seedlings. Implications for management Variation among species responses to gap treatments in this study reinforces the conclusion of Pinard et al. ( 1999) that one management system cannot be applied to enhance the regeneration of all timber species within this forest site They suggested a mixed management scheme is most appropriate with this mixture of species, proposing an even-aged managem e nt system for the shade-intolerant species (Anadenathera Astronium Centrolobium and Schinopsis) involving the creation of large multiple-treefall gaps and an uneven-aged management system for the shade-tolerant species (Caesalpinia Aspidosperma and Copaifera) involving the creation of smaller single-treefall gaps

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112 However the results ofthis study suggest that even large multiple-treefall gaps may not provide suitable habitat for shade-intolerant species Poor recruitment of Anadenathera and A s rronium seedlings in gap control plots suggest that advance regeneration logging debris and deep litter in treefall gaps limit suitable sites for establishment and survival of these speci e s Therefore, selective logging without additional silvicultural treatments such as slash removal or prescribed burning may not be sufficient to improve regeneration of these shade-intolerant species Growth and survival of Anadenathera and Astronium improved with increasing treatment intensity However high-intensity burns may not be the most appropriate silvicultural treatment for regenerating these species High-intensity burns will negatively affect species that depend on seedling banks for regeneration Growth and survival of shade-intolerant species after low intensity burns while lower than after high intensity burns may be sufficient for management purposes. If burns are prescribed they must be timed before peak seedfall as burns of high and low-intensity decrease seed viability of seeds in the seed bank This study also supports conclusions drawn by other researchers that advanced regeneration is the important source of post-disturbance regeneration among shade-tolerant species In this study most advance regeneration was killed by high-intensity burns The importance of advance regeneration creates a conflict between promoting regeneration of shade-intolerant species with intense silvicultural treatments while trying to maintain regeneration of shade-tolerant species If advanced regeneration of shade-tolerant species is abundant within a particular area then intensive silvicultural treatments such as prescribed burns would not be appropriate Other less intense treatments that wou l d release seedlings/saplings of competition such as weeding and/or thinning would be more suitable

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CHAPTERS EFFECTS OF CANOPY GAP FORMATION PLANT REMOVAL AND CO TROLLED BURNS OF HIGH AND LOW INTE SITIES ON A DRY FOREST PLANT COMi\1UNITY Introduction Patterns of forest regeneration following natural or anthropogenic disturbances are determined by interactions between the disturbance regime (i e ., intensity frequenc _, scale) and the biologies of species (i.e ., life history physiology behavior ; Pickett and White 1985) Disturbances that differ from the historic disturbance regime either in type frequenc intensity or scale ma y not only decrease species diversity but may also shift the community to one dominated by a few tolerant species or life forms (e.g. Connell 1978 Denslow 1995 Roberts and Gilliam 1995). For example vines often dominate tropical forests after uncontrolled logging (Dawkins 1961) and grasses may dominate for decades after pasture abandonment (Uhl et al. 1988a) Shifts in species composition following disturbances arise in part from differences in species-specific modes of persistence through these disturbances Modes of regeneration i e. from sprouts seedling banks seed banks or dispersed seeds ma y influence the success or dominance of a species in successional communities Although regeneration from seed has received more attention from researchers in tropical forests sprouting is also a common means of persistence through disturbance (Stocker 1981 Uhl and Jordan 1984 Putz and Brokaw 1989 Kaufmann and Uhl 1990 Kaufmann 1991). In tropical dry forests frequencies of sprouting are thought to be even higher than in wet tropical forests (Ewel 1980, Sampiao et al. 1993 Murph y and Lugo 1986 Hardesty 1988 Miller and Kaufmann 1998a b) 113

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114 The importance of sprouting as a means of survival varies within a site as well among forest types depending on the intensity of disturbance Several studies show that frequency of sprouting declines with increasing severity of disturbance. For example in both wet and dry tropical forests sprouts were less common after slash burning than after less intense disturbances (Sampiao et al 1993 Miller and Kaufmann 1998a b Kaufmann 1991 Uhl et al 1981 Uhl and Jordan 1984) Regeneration from seed may also be affected by disturbance intensity While less intense disturbances such as canopy opening may stimulate many seeds to germinate (e g ., Vazquez-Yanes and Orozco-Segovia 1993) more severe disturbances such as intense fires kill seeds buried in surface soils (Brinkmann and Vieira 1971 Uhl et al. 1981 ) Theri::fore it is often colonizing species that regenerate via widely dispersed seeds that dominate following disturbances of high intensity (Uhl et al 1981 Schimmel and Granstrom 1996) Modes of regeneration not only determine survival through a disturbance but also influence growth and survival following disturbance Due to their larger carbohydrate reserves sprouts may form larger taller crowns more rapidly than seedlings (Miller and Kauffman 1998b) And, although few studies have directly compared survival of seedlings and sprouts their survival rates are also likely to differ. While mortality of both seedlings and sprouts may be affected by herbivores (Moreno and Oechel 1994) or fungal pathogens (Wenger 1951 Augsperger 1983) mortality of seedlings is more dependent on external factors such as light availability (Veenendall et al. 1995, Thompson et al. 1998 Kitajima 1994) and soil moi!:ture (Gerhardt 1996) Sprout mortality at least initially is correlated with internal factors such as the size or age of the original stem, height and extent of stem damage and number of sprouts sharing the same root system (Wenger 1951 Blake 1983) In this chapter I characterize early successional patterns following canopy opening, plant removal, and controlled bums of high and low intensity I focus on the dominance of species, life forms and alternate regeneration modes (i e ., from seed or by sprouting) In these studies I define

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115 dominance b y va rious measures of size (stem height and diameter) or cover (percent cover and crown area). Specifically m y objectives are: 1) to examine the effects of gap formation above ground plant removal and high and low-intensity bums on soil seed banks and to quantify new seed fall into gaps following these treatments ; 2) to compare the effects of these treatments on the relative cover of different life forms ; 3) to document the species lifeform and mode of regeneration of dominant individuals in these treatments ; 4) to compare the relative dominance of seedlings and sprouts among these treatments ; and 5) to compare species richness among the treatments. Methods The studies described in this chapter were carried out in the high-intensi ty bum low intensity bum plant removal gap control and forest plots described in Chapter 2 Seed bank and seed rain studies Effects of gap treatments on soil seed banks. To examine treatment effects on seeds stored in the soi~ seed bank samples were collected from each treatment the week following controlled burns Four subsamples (20 x 20 cm 3 cm depth) were collected from each gap treatment plot and transported to the nearby community of San Lorenzo for germination Soil samples were spread on 10 x 50 cm trays watered twice daily, and examined every 7 to 10 da y s for germinated seeds Trays received direct sunlight during the morning and evening hours. After 11 weeks, all seedlings were collected and pressed Most seedlings other than commercial tree seedlings we r e difficult to identify and were grouped b y morphospecies. The total number of germinated seeds and the number of germinated seeds of commercial tree species were compared among treatments using AN OVA with treatments as fixed effects and blocks as random effects followed by Tukey s HSD post-hoc comparisons

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116 Seed rain into gap centers, gap edges, and forest understories. To compare seed rain densities into forest understories gap edges, and gap centers seed traps were placed at each site in each block and monitored for 9 weeks following bums. Gap center and gap edge traps were located 2 5 and 7.5 m from gap centers respectivel y, and forest traps were located within 2-5 m of the forest subplots Traps were constructed of 50 x 50 cm PVC frames with a net of woven plastic sewn to the sides. Traps were elevated 50 cm above the soil surface Seeds were collected every 1 to 2 weeks counted and identified to species or morphospecies The total numbers of seeds commercial tree seeds liana seeds and species were compared among trap locations (gap center gap edge and forest understory) using ANOVA with trap location as a fixed effect and blocks as random effects followed b y Tuke y' s HSD post-hoc comparisons Effects of gap treatments on cover of different life forms To compare relative cover of life forms among the four gap treatments and fores t plots I estimated percent cover 1.5 3 6 9 12 and 18 months after bums In each 4 m 2 subplot total plant cover and percent cover of bromeliads vines, grasses herbs shrubs and trees was estimated visually. In forest plots only cover < 2 m tall was estimated Vine cover included both herbaceous and woody vines (lianas) Maximum total cover was limited to l 00% but where leaf co ve r of different life forms overlapped the sum of their percent covers could total > 100 %. Total cover and percent cover b y life form were compared among treatments using repeated measures anal y ses of variance with treatments as fixed effects and blocks as random effects followed by Tukey's HSD post-hoc comparisons Regeneration mode, life form and species of tallest individuals To identify which species life forms and regeneration modes were dominant in each gap treatment I selected the tallest individual in each 4 m 2 subplot at 1.5, 3, 6 9, 12 and 18 months following identified it to species determined its mode of regeneration (seedling or sprout) and

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117 measured its height. Additionall y, after 9 and 18 months I identified all individuals within each I 00 m treatment plot > 2 5 m tall determined their modes of regeneration (seedling or sprout) and measured their heights Average heights of the tallest individual per 4 m 2 subplot were compared among treatments using repeated measures ANOVA with treatments as fixed effects and blocks as random effects followed by Tuke y' s HSD post-hoc comparisons A Kruskall-Wallis test was used to compare life form and regeneration mode of the tallest individual per 4 m 2 subplot among treatments For this test proportions were calculated for each treatment plot using the four 4 m 2 subplots The total number of individuals > 2 5 m tall were compared among treatments using a Kruskall-Wallis test at 9 months and an ANOVA at 18 months The proportions of tree seedlings tree sprouts liana seedlings and liana sprouts among of the total number of individuals > 2 5 m tall were compared among treatments using a Kruskall-Wallis test. Comparison of dominant seedlings and sprouts Seedlings were not commonly the tallest individual per 4 m 2 subplot and few seedlings reached heights> 2 5 m. Therefore I conducted an additional survey 9 months following bums to compare sizes of dominant seedlings and dominant sprouts In this survey sampling was restricted to the two center subplots of each gap treatment (two 4 m 2 subplots = 8 m2). Although one plot of this pair received a seed addition treatment (Chapter 4) in this survey only size rather than density of seedlings and sprouts are compared and therefore the seed addition treatment has minimal effect on the results. In each 8 m 2 area the 5 tallest individuals of each of the following groups were identified : tree sprouts tree seedlings liana sprouts and liana seedlings For each individual I measured the height basal diameter crown length (L) crown width (W) and for sprouting individuals, the number of stems diameter of the largest sprout and sprout origin (root root collar, or stem) Crown areas were estimated as :

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118 Area = (L W 3 1416) / 4 Crown widths and lengths were difficult to measure for lianas that had one-to-man y long climbing or scrambling stems For these lianas the measurement of crown length (L) was estimated b y multiplying the number of stems by the estimated average length of these stems The a vera ge width ofleaves on stems was used as crown width (W) The sizes of seedlings and sprouts sampled in the 8 m 2 plots were compared using ANOVAs with treatment regeneration mode and life form as factors For these tests mean heights crown areas stem diameters and numbers of stems were log transformed prior to analyses Blocks were treated as a random effect in each model. To compare the origin of sprouting (root root collar stem) among treatments frequencies of each were compared among gap treatments using Kruskal-Wallis tests Effects of gap treatments on plant species richness and similarity Species richness was assessed in each gap treatment 9 months following bums (June 1998) All species within each 100 m 2 gap treatment plot were identified for all blocks Unknown species were collected pressed and identified at the herbarium in the Museo de Historia Natural de Noel Kempff Mercado in Santa Cruz. The mean number of species and the average number of grass shrub herb, tree and woody and herbaceous vine species per 100 m 2 plot were compared among treatments using ANOVAs with treatments as fixed effects and blocks as random effects. Sorenson's similarity index (Sc) was used to compare percent similarity of species among treatments : Sc= 2c/(a+b) where c is the number of species in common be tween 2 treatments and a and b are the total number of species found in the 2 treatments

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119 Results Seed bank and seed rain studies Effects of gap treatments on soil seed banks. Total seedling density and commercial tree seedling density were highest in seed banks sampled from gap controls and lowest in high-intensity burn plots (Figure 5-1) An average of32 (5-164) seeds m 2 germinated in seed banks sampled from gap controls including 6 seeds m 2 of commercial tree species In seed banks sampled from high-intensity burn plots, only 5 germinants of an unidentified species and no commercial tree species were found in all 64 seed bank samples (representing a total area of 2 56 m 2 ) Ru elli a spp (Acanthaceae) perennial herbs were the most common species germinating in seed banks sampled from gap control plant removal and low-intensity burn treatments A total of 25 morphospecies including four commercial tree species germinated in seed banks sampled from all treatments Seed rain into gap centers, gap edges, and forest understories. An average of 94 (24212) seeds m2 were caught b y forest seed traps during the 9 week collection period from September 11 to November 13 1997 Seed rain peaked the first week of October but by the third week had fallen to 2 seeds m2 week 1 Seed rain into gaps was only 25% of that collected in forest traps (Figure 5-1) Roughly equal numbers of seeds fell into gap center and gap edge traps. Anadenanthera colubrina a commercial species had the most abundant seed fall Of the total 181 Anadenanthera seeds collected 77% fell into forest traps 14 % into gap edge traps and 9% into gap center traps Distribution of Astronium urundueva seed rain also a commercial species was more even ; of the total 94 Astronium seeds collected 29% fell in gap center traps 35% into gap edge traps and 36% into forest traps Thiloa paraguariensis (Combretaceae ) had the most abundant seed rain of vines ; of the total 154 seeds collected 66% fell into forest traps 12 % into gap edge traps and 21 % into gap center traps.

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120 A. Seed fall 120 ..,....--------------------------"7 VJ ... 0 N I E "O
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121 Effects of gap treatments on cover of different life forms Total plant cover. Total cover increased greatl y in all gap treatments o v er the 18 month post-bum sampling period and significantl y differed among treatments (Table 5-1 ) Total cover in the high-intensit y burn treatment remained the lowest at all sampling periods increasing fr om l % at 6 w eeks to 6 5% at 18 months (Figure 5-2) Total cover in gap controls remained the highest increasing from 47 % at 6 weeks to 100% at 18 months Relative increases in total cover in the low-intensity bum and plant removal treatments were the greatest ; the y increased from 7 % to 90 % and 14% to 95 %, respectively Plant cover< 2 min forest understory plots did not incr e ase o v er the duration of the stud y but varied seasonall y, ranging from 46 % during the dry season to 62 % during the wet season. Percent cover of life forms There were significant differences among treatments in percent cover of trees herbaceous and wood y vines shrubs herbs and ground bromeliads (Table 5-1 and Figure 5-2) Percent cover of ground bromeliads was highest in forest and gap control plots Also for e st plots tended to have higher shrub cover and less herb vine and tree co v er than gap treatments Differ e nces in percent cover of herbs vines and trees among gap treatments closely reflected differences in total cover. For example if total cover was highest in gap control plots so was percent cover of herbs vines and trees Grass cover was low in all treatments( < 5%). Two herbaceous species showed positive responses to burning One species Co mmelina sp a perennial monocot that sprouts vegetativel y from subterranean bulbs covered an a v erage of 14% (75% of total plant cover) oflow-intensity bum plots 6 months after burns significantl y more than in an y other treatment (P < 0 001) Cover of this species declined in successi ve sampling periods Another species an unidentified annual herb in the Euphorbiaceae established onl y on the edges of the high-intensity bum plots and in the more intensel y burned patches of the low-intensity burn plots sugg e stin g its seeds ar e stimulated b y

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122 Table 5-l R es ults of A OVA and Tuke y' s post-hoc tests of total cover percent cover b y lifeform and height of the tallest individual in 4 m 2 subplots at 5 measurement periods after burns All v ariables showed a significant treatment*time interaction w hen tested with a repeated measures model therefore each was tested separately b y month Distribution of bromeliad percent cover was not normal therefore this variable was tested with a Kruskall Wallis test. Cover at 6 weeks wa s not tested due to the high number of cells containing zeros Treatments with different letters are significantly different at P < 0 05 high low Total cover intensity intensity plant gap Months F p burn burn removal control forest 3 98 7 0 a b b C C 6 70 6 0 a b be C C 9 77.4 0 a b e d b 12 58 7 0 a b C d b 18 36 7 0 a b C b a Tree cover Months F p high low removal control forest 3 3 0 025 a ab b ab ab 6 7 9 0 000 a b b ab a 9 8.4 0 000 ab C C be a 12 5.4 0 001 ab b b b a 18 10 0 000 ab b b b a Herbaceous and woody vine cover Months F p high low removal control forest 3 16 7 0 000 a ab b C ab 6 14 6 0.000 a b b b a 9 7 5 0 000 a b b b a 1 2 10 5 0 000 a b b b a 18 12.4 0 000 b C be be a Herb cover Months F p high low removal control forest 3 5.8 0 000 a ab ab C be 6 21.3 0 000 a be cd d b 9 39 3 0 000 a b C C b 12 35 9 0 000 a a b b a 18 10.4 0 000 ab be C C a

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Table 5-1. continued Shrub cover Months 3 6 9 12 18 F p 31.4 0.000 26 3 0 000 16 2 0.000 16 9 0 000 15 8 0 000 Bromeliad cover 2 Months X 3 56.2 6 52.9 9 56 12 55 9 18 53.2 high a a a a a p 0.000 0 000 0 000 0 000 0 000 Tallest individual per 4 m 2 subplot Months F p high 1 .5 55 9 0.000 a 3 48 7 0.000 a 6 22 1 0 000 a 9 21.7 0.000 a 12 16.1 0.000 a 18 1.3 0 .2 8 123 low removal control forest a a a b ab b b C ab ab b C ab b b C ab ab b C plant low removal control b C d b C d b b be b C be b b b

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124 1.00 -,-----------------~ 0 75 0. 5 0 0 2S o oo _L _ _:::x::_ --r::e~~::1.l -1e~~ UL 1.5 months 1.00 ~----------------~ 0 7 5 0 50 0 25 1. 00 0 75 0.75 0 50 0.25 to tal co v er c::J trees lianas n t:'tt J shrubs ml herbs E22] bromeliads 0 00 ...J.._ ___J l-" ____ __._._ 3 months 6 months 1.00 -,------------------~ 0 75 0.50 0 25 1. 00 0. 7 5 0. 50 0 25 hi g h intensity lo w intensity burn burn plant remo v al gap c ontrol 12 months forest Figure 5 2 Percent co v er of bromeliads herbs shrubs lianas trees and total cover ( e mp ty bo x ) in 4 m 2 subplots at 1.5 6 9 12 and 18 months following treatments Bars on t ot al cover are standard errors (n = 16) Where cover of different lifeforms o v erlapped th e s um their percent cover could e x ceed total plant co v er.

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fires of intermediate intensity (treatment difference : P < 0.001) Individuals of this sp ecie s died after the y set seed during 6-9 months following burns Regeneration mode, life form, and species of tallest individuals 125 Tallest individuals in 4 m 2 subplots. At 6 weeks average height of the tallest individuals per 4 m 2 subplot was greatest in gap controls (Figure 5-3 Table 5-1); after 18 months there were no differences among treatments After 9 months sprouts were most frequently the tallest individual per subplot in the gap control plant removal and low-intensity burn plots ; in high intensity burn plots seedlings were most frequentl y the tallest (Figure 5-3, Table 5-2) After 6 months commercial tree species were more frequently the tallest individuals in high-intensity burn plots than in other treatments (Figure 5-4 Table 5-2) After 9 months non-commercial tree species were more frequentl y the tallest individuals in gap control plant removal and lo w-in tensit y burn plots than in high-intensity burn plots Sprouts of Mimosa sp. a scrambling shrub, we re most frequentl y the tallest individuals in the 4 m 2 plots (Table 5-3) The second most frequent species was Centrolobium microchaete a commercial tree species that suckered from roots. Individuals> 2.5 m tall in 100 m 2 treatment plots. All of individuals > 2 5 m tall 9 months following treatments were sprouts (Figure 5-5) Of these sprouts 90% were trees the remaining I 0% we re lianas At 9 months the plant removal treatment had the most individuals > 2 5 m with 3 p er 10 0 m 2 (X 2 = 8 8 P = 0 03) After 18 months the number of individuals> 2.5 m averaged 6 per 100 m 2 and did not differ among treatments (F = 0 8 P = 0 50) Seedlings comprised 40% of individuals > 2.5 min high-intensit y burn plots a proportion slightl y but not significantly higher than other treatments (X 2 = 7 7 P = 0 051) The most common individuals > 2 5 m tall were Centrolo bium microchaete (Table 5-4) The second most common species> 2 5 m tall was Mimosa sp . Anadenanthera colubrina the third most abundant species was found predominatel y as seedlings Other common species whose

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rn = 0 I-< 0. rn 4-; 0 Q 0 ..... t:: 0 0. 0 I-< 3 2 126 A. Average height of tallest plants in each subplot t _l-::::-~:: ~ --=-~ ::::::: .. / .. ......-:: J.;. .. Y ::;..-. .. .. .. 'y' . ffe,. . / y~/ .. e / . y tr' --high intensity burn @ low intensity burn -Vplant removal -W gap control 0 -+----~---~--~------,--------.-----,---------' 0 3 6 9 12 15 18 Months following treatments B. Proportion of sprouts among tallest plants in each subplot 1.00 ~------------------------0.75 k-~~-4t // . -t .. .. .. .. ~: .-:-:-: -:-:-: --:-:--;_-:-:,..~ --------. . :: .;. / .. . 0.50 0.25 0 00 +----~---~--~------r------,------r-------' 0 3 6 9 12 15 18 Months following treatments Figure 5-3 A Average height and (B) proportion of sprouts among tallest individuals per 4 m 2 subplot. Bars represent standard errors (n = 16).

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Table 5-2 Kruskall-Wallis analyses testing whether the likelihood of a patticular lifefonn being the tallest individual per 4 m 2 subplot differs atnong treatments ( df == 3 ) Sprouts similarly analyzed Month 1.5 3 6 9 12 18 commercial trees x2 P 1.4 0 69 5 5 0.14 19 6 0 000 13.9 0.003 10 7 0 01 14 0 003 non-comm. trees x2 P 5.6 0.13 4.4 0.23 6 1 0.11 8 3 0.04 9.3 0.03 15.5 0.001 lianas x2 P 5.4 0.14 1.4 0 72 5 1 0 16 2.8 0.42 1.8 0.61 2.9 0.41 herbs shrubs sprouts x2 P x2 P x2 P 19.6 0.0 3 7 0.30 3 7 0.29 9 5 0 02 8 5 0.04 12.5 0 01 11 0 01 8 2 0 04 7.5 0 06 10.4 0 02 1.5 0.69 10 5 0 02 3.1 0 38 3 6 0.31 17 5 0 001 7 7 0 05 6 1 0 11 18.4 0 000 N -.J

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.... 0 0.. .D ::l VJ .... Q) 0.. ca ::l -0 s: :.a .5 .... VJ Q) 3 VJ t'd VJ E .... c.8 <+-< 0 >, u C Q) ::l c::r Q) .... i., 128 1.00 0.75 l.5mo 0 50 0 25 0 00 1.00 0 75 3mo 0.50 0 25 0.00 1.00 0 75 6mo 0.50 0 25 0 00 1.00 0 75 9mo 0.50 0 25 0 00 1.00 0 75 12mo :,;i,'l. ~. _,:_, 'i'. 0 50 .. "!:' 0 25 0 00 1.00 0 75 18mo 0 50 0.25 0 00 high intensity low intensity b plant removal gap control shrubs ,-=~ lianas herbs ""L ...................... trees ~I --commercial Treatments Figure 5-4. Frequency of commercial trees, non-commercial trees, herbs, lianas, and shrubs as the tallest individual per 4 m 2 subplot at 1.5, 6, 9, 12, and 18 months following treatments

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129 Table 5-3. Species recorded as the tallest individual per 4 m 2 subplot 18 months following treatments Total number percentage of sprouts, average height, and the frequency of each species in gap treatments are listed Frequency (%) LifeTotal Sprouts Ht. gap plant low high Seecies form # % !m) cont. rem. int. int. Mimosa sp. LN 22 100 2 9 36 18 45 0 Ce ntrolobium microchaete CT 17 100 4.1 6 38 31 25 unknown herbs HB 16 0 1.8 6 6 31 56 Ruellia sp. HB 16 0 1.5 54 23 0 23 Acosmium cardenasii TR 15 100 3 5 47 47 0 7 Acacia /orentensis TR 12 100 5.4 17 42 25 17 Thiloa paraquariensis LN 12 8 2 0 0 8 17 75 Anadenanthera colubrina CT 10 20 2 2 0 0 20 80 Casearia arborea TR 9 100 2 8 11 44 33 11 Mendoncia sp. LN 8 63 1.9 25 38 38 0 Casearia gossyiosper111a TR 8 100 2.4 29 43 29 0 Neea hermaphrodita TR 8 88 2 .0 63 25 13 0 Spondias mombin TR 6 100 3 0 33 0 50 17 Pogonopus tubu/osus SH 6 100 3.8 17 17 50 17 unknown lianas LN 5 67 1.7 17 0 67 17 Caesalpinnia pluviosa CT 6 83 2 5 50 0 33 17 Urera baccifera TR 6 100 2 2 17 50 17 17 Trema micrantha TR 6 0 2 8 0 67 17 17 Luehea paniculata TR 5 100 3 2 0 40 40 20 unknown pioneer tree TR 5 0 2 9 0 40 20 40 Bignoniaceae sp LN 4 50 1.7 50 25 0 25 Astronium urundueva CT 4 0 35 1 0 0 0 100 Bauhinia sp SH 3 100 4.3 0 67 0 33 lvf anihot sp SH 3 100 3 0 100 0 0 0 Cordia alliodora CT 3 100 2.0 0 0 0 100 Casearia aculata TR 3 100 3.2 0 0 100 0 unknown vines VN 3 100 2.0 0 0 100 0 Ga/ipea sp TR 2 100 2 3 0 100 0 0 Prokia crucis SH 2 1 00 2 3 0 0 100 0 A rrabidaea fagoides LN 2 50 1.5 50 50 0 0 Zanthoxylum hasslerianum TR 2 100 2 5 50 50 0 0 rnmbretum lepr osu m TR 100 4.0 100 0 0 0 Ce /tis pubescens TR l 100 4 .0 100 0 0 0 Pseudananas sagenarius BR 1 0 1.0 100 0 0 0 Serjania marginat a LN 0 2 0 100 0 0 0 A spidosperma rosado CT 1 100 3 0 0 0 0 100 Platypodium elegans TR 1 100 7 0 100 0 0 0 Passi.flora sp. VN 1 0 1.5 0 0 100 0 Allophyllus edulis TR 1 100 2 0 100 0 0 0 Copaifera chodatiana CT 1 100 1.8 0 0 100 0 Platymiscium u/ei TR l 100 1.4 0 0 100 0 Chorisia speciosa TR 1 100 2 5 100 0 0 0 Trigonia boliviana LN 1 100 1.5 0 100 0 0 unknown 14 CT = commercial tree HB = herb LN = liana SH = shrub TR = non-commercial tree

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13 0 NS 4~-----------------------, 0 0 ...... --3 s If) N 2 I\ (I) ca ;:j "'O ;; g liana seedlings liana sprouts "'O o_L_-----~==----.S 9 months 8~--------------------------7 N 9 6 0 0 ...... --5 1 high intensity burn low intensity burn plant removal Gap treatments 18 months gap control Figure 5-5 Number of individuals > 2 5 m tall in each 100 m 2 gap treatment plo t after 9 and 18 months Individuals are divided into classes based on regeneration mod e (seedling or sprout) and life form (tree or liana)

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131 Table 5-4 Species recorded> 2.5 tall in 100 m 2 treatment plots 18 months after bums Total number, percentage of sprouts average height, and frequency in gap treatments are listed for each species An additional 13 species had 1 individual> 2.5 m but are not listed Frequency (%) LifeTotal Sprout Ht. gap plant low high Seecies form # % {m} cont. rem. int. int. Centrolobium microchaete CT 67 100 4.2 1.8 24 15 42 Mimosa sp. LN 35 100 3 3 29 37 34 0 Anadenanthera colubrina CT 31 16 3 5 6 3 16 74 Acosmium cardenasii TR 26 100 4 1 38 46 8 8 Casearia gossypiosperma TR 26 96 3 2 32 36 0 36 Acacia loretensis TR 25 100 5 5 16 40 40 4 Casaeria arborea TR 14 100 3 1 14 43 36 7 Spondias mombin TR 9 100 3 0 11 0 56 33 Zanthoxylum hasslerianum TR 9 100 3 3 33 67 0 0 Neea hermaphrodita TR 15 100 3 1 26 40 33 0 Trema micrantha TR 8 0 4 8 0 50 25 25 Pogonopus tubulosus TR 8 100 3.8 0 40 40 20 Caesalpinnia p/uviosa CT 7 100 3.0 71 0 0 29 Platymiscium ulei CT 7 100 3 0 0 43 57 0 Urera baccifera TR 7 71 3 3 43 43 14 0 He/iocarpus sp TR 7 0 4 0 0 29 29 43 Mendoncia sp. LN 4 100 2.5 25 25 50 0 Gallesia integrefolia TR 4 100 3 9 25 50 0 25 Cordia alliodora CT 4 100 3.4 50 0 25 25 Bauhinia rufa TR 4 50 4 9 0 25 0 75 Astronium urundueva CT 4 0 3 3 0 0 0 100 Bignoniaceae sp LN 3 100 3.4 67 0 33 0 Galipea trifoliata TR 3 100 3.2 67 33 0 0 Luehea panicu/ata TR 3 100 4.7 0 33 33 33 Coursetia hass/eri TR 3 100 3.4 0 33 33 33 Platypodium elegans TR 3 100 2 8 0 66 33 0 Prokia crucis SH 3 100 2 6 0 33 67 0 Al/ophyllus edulis TR 3 100 2.5 0 100 0 0 Capparis prisca TR 2 100 3 0 0 50 50 0 Manihot sp. TR 2 100 4.0 50 50 0 0 Combretum leprosum TR 2 100 8.0 0 0 0 100 Bougainvillea modesta SH 2 100 4 0 33 33 33 0 Casearia arborea TR 2 100 3 3 0 50 50 0 Tabebuia impetiginosa CT 2 100 2 8 50 0 50 0 Chorisia speciosa TR 2 100 3 3 100 0 0 0 Trigonia boliviana LN 2 100 2 8 50 0 50 0 Ce/tis pubescens TR 2 50 3 0 50 0 50 0 Thiloa paraguariensis LN 2 0 2 5 33 0 0 67 CT commercial tree HB = herb LN = liana SH = shrub, TR = non-commercial tree

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seedlings reached heights > 2.5 m tall were: the trees Trema micrantha Heliocarpus sp .. and Astronium unmdueva and the vine Thiloa paraguariensis. Comparison of dominant seedlings and sprouts 132 Tree and Iiana sprouts were taller (F = 8 9 P = 0.003) had more stems per individual (F = 32 1 P < 0 001) larger crown areas (F = 42 8 P < 0 001) and larger basal diameters (F = 13.4 P < 0 001 ; Figure 5-6) than tree and liana seedlings Among sprouts the origin of sprouting differed among treatments (Figure 5-7) Sprouting from roots was more common in burned plots (P < 0.001 X 2 = 25.2) sprouting from the root collar was more common in the low-intensity bum and plant removal treatments (P < 0 001 X 2 = 23.8), and sprouting from stems was the more common in the plant removal and gap control plots (P < 0 001 X 2 = 23 8) Effects of gap treatments on plant species richness and similarity Two-hundred and sixty-nine species and morphospecies were collected from all treatments including : 100 vines 64 trees 61 herbs 25 shrubs 8 grasses 2 palms 3 cacti 1 fem and 1 bromeliad Total species richness was lowest in gap controls (161 species Figure 5-8) However per 100 m 2 plot fewer species were collected in high-intensity bum treatments than other treatments (F = 16 2 ; P < 0 001 ; Figure 5-9) Low-intensity burn and plant removal treatments had the highest species richness (192 and 191 species respectively) Sorenson s similarity index revealed that the plant removal and low-intensity burn treatments shared the most species in common (161 species) while species in gap control and high-intensity bum plots were the least similar (124 species shared : Table 5-5) Frequencies of each species b) gap treatment are provided i.n the Appendix Discussion Seed bank density and seed rain following treatments High soil temperatures created by high-intensity bums reduced densities of viable seed by an average of 94%. Average temperatures during high-intensity bums at the soil surface

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-N E ,.__,, ro V c:: 0 'u V oO ro 'V > ro -E u ,.__,, ..... ..c:: bl) -~ ..c:: -E E ,.__,, 'V ..... V :.a ca en ro .D ca ::1 "O :~ "O c:: :.::::: en E V ..... en :it: 8 6 4 2 0 200 150 100 5 0 0 20 15 10 5 0 10 8 6 4 2 0 tree sprout [El vine sprout 11111111111 tree seedling f:: ::::::::::::::: : : : : j vine seedling high intensity low intensity plant removal Treatments gap control 1 -, -, .).) Figure 5-6 Average crown areas heights basal diameters of the largest ne w stem and number of stems per individual of tree and vine sprouts or seedlings in each of the 4 gap treatments Variables were measured on the 5 most dominant individuals of each group in the paired 4 m 2 center plots in each treatment of 11 blocks 9 months following burns

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Cl) ..;.., ;:j 0 I-< 0. C/) Cl) i:: .,... > '"O Cl) Cl) .ti ro 4-t 0 0 t: 0 0. 0 I-< P-4 1.0 0.9 0 8 0 7 0.6 0.5 0.4 0 3 0 2 0.1 0 0 high intensity low intensity plant removal gap control Gap treatments Figure 5-7 Frequency of sprouting from roots root collars or stems in the 4 gap treatments Proportion of each sprout origin was taken from the entire sample of tree and vine sprouts measured in the paired 4 m 2 center plots of 11 blocks 9 months following burns 134

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r:,i Cl.I c Cl.I Q., r:,i 0 250 ----------------------, 200 ,,,:;v-, .... _..,,-v1 .. @. 150 Cl.I ..c e = = Cl.I ;;.. -,e 100 z
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N s 0 0 r;/) Q) C) Q) 0.. r;/) (.;... 0 I--, Q) 1 70 GRASS EE] HERB 1111111111 LIANA 60 llliiillil SHRUB 1111111 TREE c::::J TOTAL 50 40 30 20 10 o~-high intensity low intensity plant removal gap control Treatment Figure 5-9 Average species richness (number of species) for grasses, herbs lianas, shrubs and trees in each gap treatment 9 months following burns (n = 16) Error bars represent the standard error of total species richness per plot. 136

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137 Table 5-5 Similarity of species recorded in the four gap treatments The first matrix is the tot number of species in common between two treatments The second matrix is Sorenson's similarity index (Sc) or percent similarity between two treatments Sc= 2c /(a+ b) where c is the number of species in common between two treatments, and a and b are the total number of species found in each of the two treatments Gap control Plant removal Low itensity High intensity Gap control Plant removal Low itensity High intensity Gap control Plant removal 138 Gap control Plant removal 0.79 Low intensity bum 137 161 Low intensity burn 0 75 0.82 High intensit y bum 124 137 145 High intensity bum 0 74 0 75 0.77

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138 (683 C) and at 3 cm depth (185 C) were well above the temperature range required to kill seeds. Brinkmann and Vieira ( 1971) found that seeds of 31 Amazon rainforest tree species were killed during fires where soil temperatures at 2 cm depth ranged 95-125 C. Similarly Uhl et al. (1981) reported that seeds of 7 successional species were killed at temperature> 100 C. Even the most tolerant hard-seeded species from fire-prone systems are killed at temperatures above 150 C (Probert 1992) Therefore the few seeds that did survive high intensity bums likely escaped these high temperatures in unburned patches Average soil surface temperatures during low-intensity bums ( 160 C) were also sufficient to kill seeds In this treatment however the density of viable seed was only reduced by approximately half that of gap controls ( 15 vs. 32 seeds m 2 ). This seed survival may reflect the greater heterogeneity of low-intensity bums or it may reflect survival of seeds buried below the soil surface Average soil temperatures at 3 cm depth during low-intensity bums (54 C) ma y have been below the range lethal to seeds Similarly Uhl et al (1981) found that seed banks were reduced by cutting and burning of Amazonian rainforest but that a substantial number of buried seeds survived. The number of seeds surviving low-intensity bums may also reflect the resistance of some seeds to low-intensity fires For example of 80 Copaifera chodatiana seeds placed on the soil surface before low-intensity bums, 35% germinated (Chapter 4). The high mortality of seeds caused by high-intensity bums implies that seedling regeneration in these plots was mostly limited to seeds dispersed after bums (approximately 24 seeds m2 in the first 9 weeks following bums). Common species establishing after high-intensity bums from seed Anadenanthera colubrina, Astronium urundueva and Thiloa paraguariensis were species frequently caught in gap seed traps. Astronium and Thiloa are wind dispersed, and compared with their total seed rain a disproportionate percentage their seeds were trapped in gaps This observation agrees with other studies conducted in the neotropics that noted seed rain of wind dispersed seeds is greater in gaps than forest understories (Augsperger and Franson 1988, Loiselle

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139 et al 1996). Higher survival of seeds in low-intensity bums implies that seedlings establishing in these plots may have originated either from seed banks or seeds dispersed following bums The density of buried seeds in gap control plots (5-164 seeds m 2 ) was substantiall y lower than densities reported in the literature for other neotropical forests (344-682 seeds m2 Guevera and Gomez-Pompa 1972 752 seeds m2 Uhl et al. 1981 177 seeds m2 Uhl et al 1982 742 seeds nf Putz 1983 7000 seeds m Young et al. 198 7 4535 seeds m 2 Butler and Chazdon 1998) The low seed density I observed likely does not reflect an actual difference between the Las Trancas forest and other forest types in seed bank densities, as Vaca (1999) also reported higher seed bank densities in treefall gaps from Las Trancas '95 of 477-774 seeds m 2 (sampled to 4 cm depth). This disparity between my results and Vaca's is likely due to difficulties keeping seed flats moist due to a drought that persisted during the first 2 months of my germination trial. Total plant cover: differences in rates of recovery among treatments Total plant cover was reduced to zero immediately following bums and plant removal and rates of plant colonization in the following months differed greatly among these treatments Increase in plant cover was slowest after high-intensity bums, likely due to mortality during bums of both advanced regeneration and seeds stored in the soil. Seeds dispersed in the second year appeared to be an important component of regeneration in high-intensity bum plots ; total cover in high-intensity bum plots increased more during the first 6 months of the second year (from 25 to 60%) than during the entire first year (0 to 25%) By comparison, increases in plant cover after low-intensity bums and plant removal were rapid; after only 6 months colonizing plants covered 60% and 75% of these treatments respectively. Again the slightly lower cover in the low-intensity bum treatment was likely due to partial sterilization of seed banks and some death of adva...riced regeneration caused by the bums Although canopy cover> 2 m did not differ among gap treatments (Chapter 2) differences in total plant cover< 2 m will create notably different opportunities for establishing tree seedlings.

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140 Due to the abundance of advance regeneration and rapid increase in total cover gap control plots provide the least suitable establishment sites for new tree seedlings particularl y those of shade intolerant species And although the low-intensity bum and plant removal plots provided ample opportunity for seedling establishment immediately following treatments this space was re colonized by competing plants within 6 months The high-intensity bum treatment pro v id e d the most space for the longest period for establishing tree seedlings These rapid increases in plant cover stress the importance of early colonization for the successful establishment of tree seedlings Seedlings that colonized immediately following bums plant removal or canopy gap formation would have experienced significantly less competition than those colonizing several months later. The importance of early establishment in determining later success of tree seedlings is frequently reported For example Brokaw ( 1985a) found that growth rates were highest and mortality lowest for trees establishing before or soon after canopy gap formation in Panama Dominant life forms and species in gap treatments Gap treatments altered the relative cover of plant life forms from that of forest understories For example shrubs which were a dominant life form in forest understory plots formed a low percentage of cover relative to herbs vines and trees in all four gap treatments Also, percent cover of ground bromeliads was lower in the plant removal and burned treatments relative to the forest understory and gap control plots. Within gap treatments the relative proportions of herb vine and tree cover were fairly similar each contributing approximately 2030% of total plant cover. Yet within each life form particular species tended to dominate in different gap treatments Herbs. Herb cover in gap control plant removal and low-intensity bum plots was dominated by several species in the genus Ruellia perennial herbs in the Acanthaceae Ruellia spp were the most common gerrninants in seed banks sampled from these treatments These species

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141 appear to respond rapidly to canopy opening and/or soil disturbance ; while the y ma y persist in forest understories they form dense hedges along logging roads and skid trails Also Ruellia was the onl y genus of an y life form that alone covered 100% of some subplots Due to its tendenc y to form dense patches Ru e llia spp ma y be the greatest competitors of establishing tree seed lin gs in gap control plant removal and low-intensity bum treatments Ruellia spp comprised a much smaller fraction of the herb cover of high-intensity bum plots, likely due to seed mortality during bums Herb cover in high-intensity bum plots after 18 months was largely composed of annual herbs particularly species of Asteraceae. Although present in the first year following bums cover of these herbs increased greatly in the second y ear presumably from newl y dispersed seeds. Two other herbaceous species Co mmelina sp and an unknown Euphorbiaceae appeared to establish preferentiall y in burned areas as well Th e distribution of these two species suggest the y were present as seeds or bulbs before bums and were stimulated to germinate or sprout b y medium or low-intensity heat. Vines. The dominant woody vine species in the gap control plant removal and low intensity bum treatments was Mimosa sp. perhaps more correctly called a scrambling shrub because it is free standing up to heights of 5 m and only then begins to support i ts limbs on branches of adjacent trees. This species was most frequentl y the tallest individual in the 4 m 2 subplots and was the second most common species > 2.5 m tall In this stud y, Mimosa regenerated exclusively by sprouting from the stem and root collar. These results agree with those of Sampiao et al. (1993) who observed higher dominance of Mimosa sp following bums in the Brazilian caatinga which they attributed to its ability to survive fires by sprouting It appears that Mimosa sp. does not survive fires of high-intensity as it was infrequent in high-intensity bum plots in m y study Vine cover in high-intensity bum plots was dominated by a liana species that regenerated primarily from seed Thiloa paraguariensis (Combretaceae) The wind dispersed seeds of Thiloa

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142 were the most abundant of any vine species caught in seed traps following bums Although Thiloa did not commonly reach heights > 2.5 m in the treatment plots during the 18 month observation period, it was frequently the tallest individual in high-intensity bum subplots This species was also abundant in an area of Las Trancas previously burned by a wildfire (Mostacedo et al.1999) Herbaceous vine cover was patchy ; where herbaceous vines were present the y tended to dominate all treatments within a block. For example one block was dominated by two unknown herbaceous vine species (Leguminosae and Euphorbiaceae) and t\vo other blocks were each dominated by Echinopepon sp ., a herbaceous vine in the Cucurbitaceae It is noteworthy that herbaceous vines were dominant in only 3 of 16 blocks Herbaceous vines are often observed proliferating in disturbed forests For example, in a sub-humid forest in Bolivia Pinard et al. (1998) found a density of 21 000 herbaceous vines per ha the first y ear following a wildfire In disturbed dry forest in Ecuador Gentry ( 1995) observed that small vines literally blanketed the remaining forest. Other evidence from Las Trancas suggests that vine proliferation in this dry forest may not be as extensive as in other forest sites Mostacedo et al (1998) found that the sub humid forest used in the study by Pinard et al. (1998) had a greater increase in post-fire vine infestation than an area of Las Trancas damaged by wildfire. Trees. Although percent tree cover was similar to that of herbs and lianas the majority of tall individuals were trees For example trees comprised 86% of individuals> 2.5 m tall and 53% of the tallest individuals in subplots Among this group of dominant trees the majority were non commercial species For example, non-commercial species comprised 60% of trees> 2.5 m tall and 70% of the trees that were the tallest individuals in subplots. Most of these non-commercial trees, which are shade tolerant and relatively abundant in forest understories (Killeen et al. 1998) regenerated primarily by sprouting from stems present before gap formation Therefore where advance regeneration is abundant regeneration of commercial tree species ma y be dominated by non-commercial tree species

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143 This pattern did not occur in high-intensity burn plots Most advanced regeneration was killed during high-intensity burns, allowing individuals of shade-intolerant commercial tree species to establish and dominate For example among the commercial trees> 2 5 m tall and among those tallest in subplots approximately 50% were located in high-intensity burn plots the remaining half distributed among the other 3 treatments. Commercial tree species that did dominate in the gap control, plant removal and low-intensity burn treatments mostly arose from pre-existing stems or roots. For example, Centrolobium microchaete the most common commercial tree species> 2 5 m tall. regenerated as suckers from the root system of mature trees Tall individuals of Caesalpinia pluviosa a moderately shade-tolerant species sprouted from cut or broken stems in gap controls The future value of these sprouting commercial trees is questionable particularly those sprouting from stems Due to the susceptibility of cut or damaged stems to fungal pathogens mature trees that arose as sprouts may not be harvestable Although root sprouts of temperate species can produce healthy stems (e g ., Blake 1983), little is known about potential value of root sprouts of tropical species. Bromeliads. Although several bromeliad species are present at this site only one species Pseudoananas saginarius, is typical of forest understorys Pseudoananas occurs at high densities in Lomerio : approximately 81 % of the forest is occupied by these bromeliads and dense patches of roughly 8 000 ramets per hectare may occur (MacDonald et al. 1998) Removal of Pseudoananas manually or by burning potentially has tremendous consequences on the establishment of other species In a study specifically addressing the effects of Pseudoananas on tree seedling regeneration in Las Trancas Fredericksen (1998) found that tree seedling establishment and growth was enhanced more by the removal of bromeliads than by the removal of other understory vegetation. Therefore the rapid colonization witnessed in the plant removal and low-intensity burn treatments may have been in large part due to the removal of Pseudoananas

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144 Dominance of tree and liana sprouts Tree and liana sprouts were taller had larger crown areas larger basal diameters and more stems than tree and liana seedlings in all gap treatments Very similar results were r e ported b y Miller and Kaufmann ( 1998b) who compared the size of tree seedlings and sprouts after slashing and burning of a tropical deciduous forest in Mexico In their stud y the y found mean sprout height crown area stem diameter and number of stems were significantl y greater than for seedlings Miller and Kaufmann ( 1998b) attributed the larger sizes of sprouts than seedlings to their larger carbohydrate reserves Sprouts which draw carboh y drates from pre-established root systems, have much larger carbohydrate reserves than seedlings which are limited to se e d reserves Even root systems of seedlings ma y provide larger carbohydrate reserves for sprouting shoots than that provided b y seed reserves particularly in dry forests where most seeds are wind dispersed and necessarily small (Gentry 1995) and root systems are proportionatel y larger than in wetter forests (Cuevas 1995) In Las Trancas for example storage tap roots were observed in 3 month-old Anadenanthera seedlings Sprouts may still have an advantage over seedlings even after both have depleted their carbohydrate reserves The larger root system of sprouts would offer more surface area for water and nutrient uptake and likely extend deeper into the soil than seedling root s y stems This difference would confer an advantage to sprouts particularly in dry forests where water is seasonally limiting Root systems of sprouts may also provide an additional source of mineral nutrients for developing shoots For example Coutinho ( 1990) demonstrated that ash contents in the underground organs of woody plants increased and remained elevated for 5 months following ground fires in the Brazilian cerrado The high carbon cost of maintaining an extensive root system is one possible disadvantage of sprouts. Sprouts have to export more carbon below-ground to support larger roots systems

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145 while seedlings with balanced root/shoot ratios may be able to allocate relatively more resources to aboveground photosynthetic tissue. Also, sprouts often divide height growth among multiple stems whereas seedlings usually support only one stem. These reasons may explain why seedlings and sprouts of several tree species had similar relative growth rates during the second year following bums (Chapter 4). Similarly, Miller and Kaufinann (1998b) reported similar relative growth rates among seedlings and sprouts in Mexican dry forest. Changes in relative dominance of sprouts among treatments Although sprouts were larger than seedlings in all treatments the magnitude of this difference was dependent on treatment intensity For example, while the average height of sprouts was roughly 3 times the average height of seedlings in the unburned treatments, sprouts were only slightly taller than seedlings in the high-intensity bum treatment (Figure 5-6) This general pattern is also apparent for crown area and stem diameter. The shift in relative dominance among sprouts and seedlings in high-intensity bum plots is particularly apparent among individuals> 2 5 m tall. In high-intensity bum plots, seedlings comprised 40% of individuals > 2 5 m tall, whereas in the gap control, plant removal, and low-intensity bum treatments only 10% of individuals> 2.5 m tall were seedlings. The decrease in dominance of sprouts in high-intensity bum plots was partly due to lower sprout densities and smaller sprouts in this treatment. Plant cell death during fires depends on both the duration of heating and maximum temperature ; fire intensity therefore, is an important determinate of post-bum sprouting. Often, a greater number of plants of increasingly larger size are killed at higher fire intensities (Moreno and Oechel 1994) Intense fires may also hinder growth of sprouts by killing part of the root system A decrease in sprout size and density following intense fires has been reported in several studies conducted in tropical forests For example Sampaio et al. ( 1993) studied the effects of fire intensity on coppicing of caatinga vegetation in Brazil and found sprouts were both smaller and less abundant in areas that experienced more

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intense fires. Similarl y, Kaufmann ( 1991) found decreasing frequency of sprouting and slower sprout growth in areas of greater fire intensity in moist forests in the Brazilian Amazon 146 The decrease in dominance of sprouts is also partially due to the greater size of seedlings in high-intensity burn plots. More seedlings reached heights> 2.5 min high-intensity burn plots than in other treatments Most of these seedlings> 2.5 m were shade-intolerant species establishing from seeds dispersed after the experimental burns i e Anadenanthera and Astronium. As discussed in Chapters 3 and 4 relative growth rates of these species were greatest in high-intensity burn plots presumably due to both increased soil nutrients and decreased competition Treatment effects on species richness and composition Although high-intensity burns caused substantial mortality of advanced regeneration and buried seeds, total species richness in this treatment was higher than in the least disturbed treatment the gap control. This relatively high species richness in the high-intensity burn treatment was likely due to the presence of several colonizing species that were absent from gap control plots. However due to the sparse vegetation after high-intensity bums diversity indices that incorporate abundance of species would likely be much lower in high-intensity burn plots Therefore the high species richness found in this study following high-intensity burns should be interpreted with caution According to the intermediate disturbance hypothesis (Connell 1978) frequent large or intense disturbances decrease species diversity Several studies from tropical forests support this hypothesis. For example, Uhl et al. ( 1988a) found that tree species richness 8 years after abandonment of light-use pastures in the Amazon was relativel y high at 20 tree species per 100 m 2 In contrast they found that after intense use pastures recover substantiall y fewer tree species ; they found only l tree species per l 00 m 2 8 y ears after abandonment of intensel y used pastures In tropical dry forest in Mexico Miller and Kaufmann ( 1998a) similarly observed that

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species diversity after slash-and-bum was similar to undisturbed forest but after 3 successive burns diversity of woody species declined by 25% 147 Species composition was in general very similar among gap treatments ; similarity indices ranged from 74-82% The high degree of sprouting ma y account for this relative! stable species composition M i ll e r and Kaufman ( 1998a) attribut e d the maintenance of di vers i ty following slash and-burn to the high sprouting capacity of tree species at the Mexican dry forest site Similarly in an Australian rain forest Stocker ( 1981) observed high floristic stability following felling and burning which he also attributed to the high percentage of trees regenerating b y sprouting Conclusions Dominance of species life forms and modes of regeneration were all altered b y gap treatments While cover of ground bromeliads and shrubs decreased cover of trees herbs and lianas increased in gap treatments relative to forest plots Several species also tended to domin c' t -: in gap treatments although species that were dominant in gap control plant removal and low intensity bum treatments differed from species dominant in high-intensity burn treatments Tree and liana sprouts we r e the largest individuals in the gap control plant removal and low-intensity burn treatments Seedlings were comparatively more dominant in high-intensity bum plots Although total species richness was not reduced by high-intensity bums species compositior. in this treatment was least similar to gap controls. Total plant cover increased rapidly in all treatments except after high-intensity burns Due to this rapid increase and the abundance of advance regeneration, opportunities for establishing tree seedlings were short-lived in all gap treatments but high-intensity bums In the gap control plant removal and low-intensity bum treatments commercial trees were dominated b y sprouts of non-commercial trees present before gap formation Onl y in the high-intensity bum treatment were individuals of commercial tree species dominant. These results have serious implications for silvicultural treatments that aim to improve regeneration from seed of commercial tre e sp e cies In

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148 this study the most intense treatment resulted in the most vigorous regeneration of commercial tree seedlings yet this treatment also altered communit y structure and composition mor e drasticall y than the other treatments

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CHAPTER6 COMMERCIAL TREE REGENERATION FOLLOWING AGRICULTURAL ABANDONMENT IN BOLIVIAN DRY FORESTS Introduction Forest management techniques are strengthened by knowledge of the autoecology of desirable species and knowledge of the forest disturbance regime. Results of the studies presented in Chapters 2-5 indicate that many of the shade-intolerant commercial species in Lomerio require disturbances more intense than canopy gap formation for their regeneration However these studies only examined regeneration over an 18-month period Clearly knowledge of the autoecology of these species would benefit from studies that examine patterns of regeneration over longer time scales In Lomerio shifting agriculture is the principle occupation of the Chiquitanos and therefore there is an abundance of fallow agricultural fields abandoned at various times in the past. Examining population structures of individual species in such fields can increase understanding of the regeneration strategies of these species Furthermore comparing structural features of secondary forests to mature forests may provide clues to past disturbance regimes or events that structured present-day mature forest. I examined tree regeneration and stand structure in a 50-year chronosequence of secondary succession after shifting cultivation in Lomerio My specific objectives were : 1) to characterize population structures of commercially valuable tree species in secondary forests of different ages ; 2) to characterize forest structure and diversity in secondary forests of different ages and compare 149

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these with mature forest and, 3) to discuss the role large disturbances may have played in the formation of mature forests in Lomerio. Methods Study site and background 150 The forest fallows chosen for this study were located in lands communally owned by the community of San Lorenzo (Figure 2-1). All fallows were abandoned slash-andburn agricultural clearings. In Lomerio forest clearing for slash-and-bum agriculture begins at the beginning of the dry season (May-June) when understories of selected forest are slashed with machetes and then trees are felled by ax or chainsaw Slash is allowed to dry for at least 2 months until the end of the dry season (August-September), when fields are burned Fields are not plowed. Crops are planted after the start of the rainy season (October-November). Each field that is opened is used for three to five years, depending on soil fertility The general sequence of crop rotation for fields on more fertile soils is rice, corn and occasionally beans the first year manioc and plantains the second year, corn and rice again in the third year and manioc or sugar cane the fourth year (McDaniel unpublished) While the site is being actively farmed it must be weeded 2-3 times a year to prevent second growth vegetation from taking over. Abandoned fields are sometimes cultivated again after 15-20 years although many fields are left for much longer periods The size of agricultural fields varies from 1 to 7 ha although most fields are about 2 ha (McDaniel unpublished) Most fields are surrounded by mature forest or older fallows so seed sources are locall y available Study plots Fourteen fallows representing 12 different ages from 1 to 50 years were located with the help oflocal Chiquitano farmers. Each age had one replicate except for the I-yr-old fallow which had 3 replicates The replicates for the 1-yr-old fallow were used to examine variation among fallows of same age but were averaged for descriptive statistics. Fallow ages were estimated by

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151 talking with the owners of the original fields These ages are only estimates and confidence of ages decreases with fallow age However I am confident that the chronological order of field abandonment is correct. The mature forest stand used for comparison was Las Trancas 9 5 the site of the studies described in Chapters 2-5 Data from an intensive floristic inventory conducted in this forest in 1995 prior to the initiation of logging activities are used for the mature forest values (Killeen et al 1998). Data collection Tree inventories were conducted during the wet season (February and March) of 1998 Sampling design follows that of Killeen et al. (1998) Within each fallow six 50x20 m plots were randoml y located along transects with plots separated b y 50-100 m Each plot was composed of nested quadrats with smaller quadrats situated within larger ones (Figure 6-1) In each quadrat all trees within the targeted size class were identified to species their diameter at breast height measured (1.3 m dbh) and total height estimated visually. In each plot total plant cover< 2 m and percent cover by lifeform (grasses herbs bromeliads shrubs lianas and trees) was estimated visually for four 1 m 2 plots located at the first established comer. Due to very abundant seed production of Anadenanthera the year of this study newly germinated seedlings of this species were also counted in the 1 m 2 plots rather than the 16 m 2 quadrats in order to distinguish them from seedlings > 1 year-old. Canop y cover was estimated with a spherical densiometer at one comer of each plot. Voucher specimens were collected and identified at the Museo de Historia Natural de Noel Kempff Mercado Santa Cruz. Mature forest values of basal area stem density and canopy height were obtained from Killeen et al. ( 1998) who sampled 100 plots in a 300 ha area of forest in Las Trancas 95 However as species richness varies as a function of area sampled tree species richness was estimated from a subsample of 6 plots randomly selected from Killeen s data set. Plant cover

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A 50m B C 4m 1 Sm .--... D I 20m L 10m 152 Quadrat A: Size class 6 ( dbh > 40 cm ) Quadrat B: Size class 5 ( dbh 20 40 cm ) Quadrat C: Size class 4 ( dbh 10 20 cm ) Quadrat D: Size class 3 ( dbh 5 10 cm ) Quadrat E: Size class 2 ( dbh < 5 cm ,> 2 m tall) Quadrat E: Size class 1 ( dbh < S cm ,<2 m tall) Figure 6-1 Layout of sample plots used to measure tree structure and diversity in abandoned shifting agricultural fields Each plot had nested quadrats of 5 sizes ( A-E) with a different tree size class (based on dbh ) targeted for each quadrat.

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< 2 m and canop y cover of the mature forest were measured in the forest understory plots described in Chapter 2 153 Importance values for each species were calculated as the average of relative dominance relative abundance, and relative frequency. In this chapter dominance is the proportion of total basal area of each species Results Regeneration ecology of selected species The combined importance values of the commercial tree species listed in Table 6-1 decline with increasing fallow age In this section I highlight the population structures of 4 species with the highest total importance values : Centrolobium microchaete, Anadenanthera colubrina Astronium urundueva and Acosmium cardenasii. All four of these species are widely distributed throughout South American dry forests (Prado and Gibbs 1993). All except Acosmium are commercially valuable in Bolivia Additionally, all except Acosmium were used in the seed addition treatment of the study described in Chapter 4 Astronium urundueva (Figure 6-2). Astronium seedlings were abundant in the 1 and 2yr-old stands, but scarce in stands >20-yr-old Population structures in 8-, 20-, 23, 40, and 50yr-old stands resemble an even-aged or single-cohort stand. Density was markedly lower in the mature forest stand than the secondary stands. Anadenanthera colubrina (Figure 6-3). Similar to Astronium Anadenanthera seedlings (> 1-yr-old) and saplings were abundant in young stands but seedling density (> I-yr-old) was low in the 15-, 30, 40, and 50-yr-old stands One-year-old Anadenanthera seedlings were abundant in all stands ; densities ranged from 0.5-22 seedlings m2 It cannot be stated with certainty that the abundance of the smallest size class in the mature stand was due to 1-yr-old seedlings, as newly germinated seedlings were not distinguished from older seedlings when the mature forest was surveyed in 1995. It is likely this was the case, however because two y ears

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154 Table 6-1 Importance values (relative abundance+ relative dominance+ relative frequency) of individual tree species in each of 12 differently aged stands in a 50y ear chronosequence ol abandoned agricultural fields. Only species with total importance values for all stand ages greater than 5% are reported Species are in descending order of summed importance values Commercial tree species are in bold print. S~ecies 1 2 3 5 7 10 15 20 23 30 40 50 M Centrolobium microchaete 27 10 32 2 9 l 5 15 24 13 7 4 1 Anadenanthera colubrina 9 8 11 13 11 15 11 7 15 11 11 9 13 Astronium urundueva 9 22 6 18 10 10 8 5 2 1 5 7 2 Acosmium cardenasii 9 4 1 5 3 4 6 14 13 30 Phyllostylon rhamnoides 2 2 2 9 2 9 6 6 14 14 1 Casearia gossypiosperma 6 5 2 4 5 3 4 5 4 4 4 3 2 Caesalpinia pluviosa 9 4 3 4 4 5 3 3 2 4 6 3 Cordia alliodora 8 5 7 6 0.4 4 2 3 4 0 5 1 2 0 3 Aspidosperma rigidum 7 2 3 3 2 1 2 1 3 5 Chorisia speciosa 1 1 3 3 3 8 4 4 2 Machaerium acutifolium 1 2 0.4 2 3 1 1 3 4 4 2 3 Galipea trifoliata 0.4 1 2 2 2 2 2 3 5 4 3 Rhamnidium elaeocarpum 1 4 3 3 1 4 3 5 Simira rube sce n s 0.4 1 2 I 2 2 2 l 5 I 4 Allophylus pauciflorius l 1 7 1 1 4 1 1 3 2 Spondias mombin 6 5 2 l 3 1 1 2 1 Guazuma ulmifolia 4 1 4 1 7 2 l l Combretum leprosum 5 2 4 1 4 0.4 2 0 1 Acacia lorentensis 1 1 2 10 1 l 3 Machaerium scleroxylon 0.4 5 1 3 2 1 1 l 1 2 Cecropia concolor 2 3 5 5 3 1 0 Bougainvillea modesta 1 2 5 2 4 2 Cariniana estrellensis 0 4 1 4 1 6 1 Sterculia apetala 3 1 1 3 1 2 2 1 Pterogyne nitens 3 2 1 1 2 1 2 2 1 Cedrela fissilis 6 4 1 0 1 Talesia esculenta 1 2 2 3 1 1 1 1 0 1 Neea hermaphrodites 0 3 0 3 1 0.4 1 2 5 Cybistax antisyphilitica 0 3 1 1 2 2 0.4 l 1 0.2 1 Rollinia herzogii 2 1 1 1 2 1 Ce/tis pubescens 1 2 1 0 5 1 1 l l 1 0.2 Pogonopus tubulosus 0.4 1 5 1 l Eugenia flavescens 0.3 1 2 1 3 1 0.3 Samanea saman l l 2 1 1 1 1 0.1 Euterolobium coutortisiliquum l l 2 3 0 1 Unknown sp 0 1 0 3 2 Trema micrantha 5

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1500 1000 500 0 1000 500 1 100 75 I 50 25 0 123456 123456 1 2 3 4 5 6 2 3 4 5 6 155 Astronium urundueva 1500 1000 500 0 1 2 3 4 5 6 1 2 3 4 5 6 1500 1500 1000 1000 500 500 0 0 123456 123456 123456 2 3 4 5 6 1 2 3 4 5 6 1 2 3 4 5 6 Figure 6-2 Population structures of Astronium urundue v a in agricultural fallows 1 to 50 years after abandonment. Each graph represents the density (individuals/ha y -axis) of each size class (x-axis) Size classes follow those explained in Figure 6-1 Numbers in upper right comers of each graph represent approximate fallow age Note changes of scale in the y-axis

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123456 1000 500 0 1 2 3 4 5 6 400 I 350 1 300 250 200 150 100 50 0 1 2 3 4 5 6 400 300 200 100 0 1 2 3 4 5 6 Mature forest 400 ,,..~ ===== ~ I 200 : ,:, :, :, :,:,:,:,:::::::::::::::,:::::::::::;;,:::,:::;:;:::::::::::::: I 100 I O +--,--.:;-,ao.:;,-;--.:;-,-i 123456 I Anadenanthera colubrina 1000 500 0 1 2 3 4 5 6 1000 500 0 1 2 3 4 5 6 400 300 200 100 0 1 2 3 4 5 6 2 3 4 5 6 1 2 3 4 5 6 0 1 2 3 4 5 6 400 300 200 100 0 2 3 4 5 6 400 -======== 350 300 250 200 150 100 50 0 ~11:.:4"'""'+-.-~""'f"'-i 1 2 3 4 5 6 156 Figure 6-3 Population structures of Anadenanthera colubrina in agricultural fallows 1 to 50 years after abandonment. Each graph represents the density (individuals/ha, y-axis) of each size class (x-axis). Size classes follow those explained in Figure 6-1. Numbers in upper right comers of each graph represent approximate fallow age Note changes of scale in the y -axis

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157 after the 1995 survey I did not find a single Anadenanthera seedling in 32 plots ( 4 m 2 ) sampled in the mature forest understory (Chapter 4) Population structures of the 15, 40, and 50-yr-old stands resemble even-aged stands. Excluding the smallest size class (class 1) the population structure in the mature forest also resembles an even-aged stand Centrolobium microchaete (Figure 6-4). Patterns in population structures of Centrolobium throughout the 50-yr chronosequence are less clear than those of Anadenanthera or Astronium Centrolobium individuals were more abundant than all other species in the 1yr -old stand However density in older stands was sporadic ; fallows ages with small populations (5-, 10, 15, 40, 50-yr-old and mature stands) were interspersed among fallow ages with larger populations (8, 20-, 23-, 30-yr-old stands). Population structures in the older stands (30, 40, 50-yr-old, and mature) appear even-aged Acosmium cardenasii (Figure 6-5). The densest seedling bank of Acosmium was found in mature forest (7 125 per ha) AlthoughAcosmium was established in moderatel y high densities in the 1-yr-old stand, seedlings were mostly sparse or absent in the 2, 3, 5, 10, and 15y r-old stands Seedlings and saplings were abundant in stands 23-yr-old and older. Population structures in the 30-, 40, 50-yr-old and mature stands resemble a reverse J-shaped structure Dominance of regeneration guilds throughout the chronosequence Figure 6-6 depicts the dominance of four regeneration guilds (short-lived pioneers long lived pioneers partially shade-tolerant and shade-tolerant) throughout the 50 yr chronosequence Classifications are based on species shade tolerance obtained from Pinard et al. (1999) Long lived pioneers dominate all stands throughout the chronosequence Shade-tolerant species become more dominant in older stands but do not dominate long-lived pioneers even in the mature stand Short-lived pioneers and partially shade-tolerant species did not dominate the successional stands at any point in the chronosequence.

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1 2 3 4 5 6 1 2 3 4 5 6 1 2 3 4 5 6 1000 -======= 800 600 400 200 0+-"f'"l'9"'1~-.,i:~~ 1 2 3 4 5 6 Mature forest Centrolobium microchaete 1000 800 600 400 200 0 1000 800 600 400 200 0 123456 123456 123456 123456 1000 800 600 400 200 0 1000 800 600 400 200 0 123456 2 3 4 5 6 123456 1 2 3 4 5 6 Figure 6-4 Population structures of C. micro chaete 158 in agricultural fallows 1 to 50 years after abandonment. Each graph represents the density (individuals/ha y -axis) of each size class (x-axis) Size classes follow those explained in Figure 6-1 Numbers in upper right corners of each graph represent approximate fallow age Note changes of scale in the y -axis

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500 .,,.t':':':11..,,,,,==== : 1 i_! !_!_!_!_!_!_!_!_:_ 1_:_ :_:_l_l_!_l_!_l_!_! i_! !_!_~ 200 1111 .. . .... ~ I~ 1 2 3 4 5 6 1 2 3 4 5 6 1 2 3 4 5 6 1 2 3 4 5 6 500 -i11i-m,.,,= ===="' ::.: : : .: ::. x~.::::.:::~: : ::'.: :.:::..<.<.::::.:::. 0 1 2 3 4 5 6 Acosmium cardenasii 500 ....,,,,======......, 400 300 200 100 0 +-"4"-"""r'""""!=f""""'f"'~ 2 3 4 5 6 500 400 300 200 100 0 1 2 3 4 5 6 2 3 4 5 6 2 3 4 5 6 500 ..,,,,....,,,,.,,,,,,,.,,,.,,,"""""""""""'""" 400 300 200 100 0 2 3 4 5 6 2 3 4 5 6 2 3 4 5 6 1 2 3 4 5 6 Figure 6-5 Population structures of Acosmium cardena sii in agricultural fallows 1 to 50 years after abandonment. Each graph represents the density {individuals/ha y -axis) of each size class ( x -axis) Size classes follow those explained in Figure 6-l Numbers in upper right comers of each graph represent approximate fallow age Note changes of scale in the y -a xis. 15 9

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1.0 ,----------------------,/ cd 0 8 "@ "' 0 6 123 5 8 10 15 20 23 30 short-lived pioneers long-lived pioneers partially shade tolerant shade tolerant -40 50 Stand age (years) Figure 6-6 Proportion of total basal area of four different regeneration guilds (short-lived pioneers long-lived pioneers partially shade-tolerant and shade-tolerant) over the 50-year chronosequence of forest fallows following agricultural abandonment. Classification of species into regeneration guilds follows Pinard e I al. 1999 The mature forest is rep resented by"M ."

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161 Species richness and forest structure Tree species richness reached 75% of matur e forest richness within 5 years after agricultural abandonment (Figure 6-7 A) Canopy height and basal area were slower to recover ; both of these indices were 75% of the mature forest values in the 23-yr-old stand (Figure 6-7B and C) Total stem density (stems > 2 m height) was highly variable in stands up to the 30-yr-old stand (Figure 6-7D) ; total stem density of the 50-yr-old stand was almost twice that of the mature forest stand. Changes in stem density and basal area by size class reveal important developments in forest structure (Figure 6-8). Most of the variation in total stem density among stand ages was due to sterns< 5 cm DBH ; the difference in total stem density between the 50y r-old stand and mature forest is largely due to the greater abundance of these small stems in the y ounger stand In contrast, the variation in total basal area was due more to large stems > 20 cm dbh reflecting the contribution of large trees Basal areas of the 5 size classes were similar in the 50y r-old and mature forest stand although the mature forest stand had comparatively more basal area in the > 20 cm size classes and less in the < 20 cm size classes Canopy cover and understory cover Canopy cover recovered to 75% of mature forest richness within 8 years after agricultural abandonment (Figure 6-9) Canopy cover of the 50-yr-old stand and the mature forest were similar both with 80% cover Grasses and herbs were a dominant part of plant cover in the young stands but their cover declined in older stands Shrubs and the common ground bromeliad Pseudananas sagenarius were absent in young stands but were more abundant in older stands Tree regeneration lost dominance in the understory in stands> 15y r old likel y due to saplings recruiting into larger size classes Vine cover varied from 5 to 18% throughout the stands of the chronosequence wi th no distinct pattern in cover with stand age

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162 n 60 n 50 c A Tree species richness .J 40 n \) 30 '3 \) 20 0.. n \) 10 \) -... 0 E ..... 30 ..c: 00 25 ..... Q) B Canopy height ..c: >, 0.. 20 0 15 c:: d 0 10 E ;:3 5 E x 0 35 30 d .e 25 N C Total basal area E 20 d 15 Q) a 10 5 V) d :Q 0 10000 d .e V) 8000 E D Total stem density Q) ..... V) 6000 .:;, ;:;; 4000 c:: Q) -c:, 2000 E ---. Q) ..... 0 (/) 123 5 810 15 20 23 30 40 50 Stand age (years) Figure 6-7. Tree species richness, canopy height, total basal area, and total stem density of agricultural fallows I to 50 years after abandonment. Dotted lines represent 75% of mature forest values for each index. Tree species richness is the total number of tree species found in each fallow. Canopy height, total basal area, and total stem density are averages (with standard errors) of the 6 sampling plots used in each fallow. Mature forest values are taken from Killeen et al. (1998). Stem density does not include stems< 2 m tall.

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163 8000 --,------------------------7 7000 ,-._ 6000 ME ---5000 VJ E Q) ..... VJ 4000 .__, ;,-.. ..... cii 3000 i::: Q) "'d E 2000 Q) ..... {/) 1000 0 12 11 10 9 ,-._ 8 ro 7 M E .__, ro 6 Q) 5 a 4 VJ ro a:l 3 2 0 -1 A. Stem density 0 .0 123 5 810 15 20 23 30 40 50 B Basal area J'. V \ / \ .. ~ Z / "' ---< 5 cm O 5 10 cm ---'9'10 20 cm -v20-40cm -a> 40cm =..:..:~ ----' I 0 \ :r--"1 ;JI ...__ b "~! / -r '0 I --._..., / . / v ... -. t 1 / l23 5 810 15 2023 30 40 Stand age @ .. .... o 50 Figure 6-8. Stem density (A) and basal area (B) by diameter size class of agricultural fallows 1 to 50 years after abandonment. The mature forest stand is designated on the x-axis by a "M"

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-100 0 "-' I-. 0 (.) >. 60 0. 0 C: 40 ro (.) .... 20 C: 0 (.) 20 0 164 A. Percent canopy cover ... .... . B. Percent understory cover < 2 m 2 3 5 8 al grass bromeliads k lt@M d herbs .. shrubs lianas trees 10 15 20 23 30 40 50 I Ml Stand age (years) Figure 6-9 A Percent canopy cover> 1.3 m measured with a spherical densiometer and B percent understory cover <2 m by lifeform The mature forest stand is designated on the x-axis by a "M"

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165 Discussion Tree species regeneration strategies The four species I have highlighted show considerable variation in their regeneration strategies. Based on their size-class distributions in the chronosequence as well as results from experimental studies (Chapter 4 and 5) these species span the continuum from shade-intolerant to shade-tolerant. Acosmium, with its densest seedling bank in mature forest appears to regenerate well in the shade and is the most shade-tolerant of the four species The remaining three species all exhibit characteristics of shade-intolerant regeneration but appear to require different degrees of disturbance for their regeneration. Of the four species, Astronium is perhaps the most shade-intolerant. Based on the absence of seedlings in older fallows Astronium appears to have little ability to establish in shade However canopy opening alone does not appear to be sufficient for its regeneration Even with seed addition A s tronium seedlings did not establish under canopy gaps but did establish following plant removal and controlled bums (Chapter 4) suggesting this species may need soil disturbance for its establishment. Unlike Astonium Anadenanthera has the ability to establish in forest understories ; newly germinated seedlings were abundant in all stands in the chronosequence. However this seedling bank appears to be emphemeral. Mortality of Anadenanthera seedlings in forest understories was high (Chapter 4) and few are expected to recruit into larger size classes The low density of individuals in the second size class in the mature stand supports this claim Therefore Anadenanthera appears to be less shade-intolerant than Astronium because it can establish in shade but still requires disturbance to survive beyond the seedling stage and recruit into larger size classes. The sporadic abundance of Centrolobium among the differently aged stands ma be due in part, to its mode of regeneration. Centrolobium primarily regenerates vegetativel y as root

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166 sprouts from the root systems of mature individuals (Chapter 4 and 5) ; Centrolobium seeds have extremely low viability. This mode of regeneration would contribute to a clumped distribution as regeneration would be more abundant near mature individuals A clumped distribution would explain why Centrolobium was abundant in some fields and almost absent in others Rather than light availability alone Centrolobium may be dependent on damage to a mature individual for regeneration Despite their different regeneration strategies Anadenanthera Astronium and Centrolobium can be classified as long-lived pioneers (Finegan 1996) ; they all exhibited shade intolerant regeneration become emergents at maturity and produce dense wood (1.00 l. 10 and 0.75 g/cm 3 respectively) Approximately 13 commerical species in Lomerio also show characteristics of long-lived pioneers. As with Anadenanthera Astronium and Centrolob i um these other shade-intolerant commercial species were more dominant in forest fallow s than the mature stand Although long-lived pioneers dominated all successional stands with increasing stand age their dominance declined as that of more shade-tolerant species increased Without introducing new disturbance to the mature stand populations of long-lived pioneers will senesce and the abundance of shade-intolerant commercial species will likely decline Successional patterns and recovery of forest structure The patterns of change in forest structure observed among the variousl y aged abandoned fields in this study follow descriptions of tropical secondary forest dynamics reviewed by Brown and Lugo ( 1990) specifically total stem density decreases during stand development as density of trees > 10 cm dbh canopy height, and basal area increase. These structural changes occurred relatively quickly in this chronosequence : tree species richness reached 75% of mature forest richness in the 5-yr-old stand and both basal area and maximum canopy height reached 75% that of mature forest in the 23y r-old stand

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167 These rates of recovery relative to mature forest values are among the higher rates documented in other studies of tropical forest regeneration, particularly for species richness (Table 6-2) It is important to note however that species richness and species composition vary independently and although many studies report quick recoveries of species richness species composition is almost always slower to recover (Finegan 1996). The fast relative recovery of the Bolivian site compared to the other forest sites ma y be due to several factors First a high percentage of tree species in these forests regenerate b y sprouting Communities with a high percentage of sprouting species are thought to be more resilient to disturbance (Janzen 1975 Ewe! 1980 Cortlett 1981 Nyerges 1989) as growth rates of sprouts are generally higher than that of seedlings, allowing sprouts to form taller and larger crowns soon after disturbance (Miller and Kauffinan 1998b Chapters 4 and 5). And although not quantified seed rain of trees into abandoned fields was potentially high due to the large percentage of wind dispersed tree species in this forest (Justiniano 1997) and the relatively small size of forest clearings (usually ~2 ha) Notably, short-lived pioneers did not dominate the successional stands early in the chronosequence (Figure 6-6) Instead, long-lived pioneers begin to dominate the stands almost immediately This pattern contrasts with succession in tropical moist forests where short-lived pioneers typically dominate successional stands for up to l 0 years (Uhl et al 1981 Finegan 1996 Howlett 1998) This pattern may be related to differences among dry and moist forest tree species in the ability of their seeds to establish and survive in abandoned fields As Anderson ( 1990) explains the rnicroclimatic gradient from forested to non-forested sites is notably steep in the moist tropics This gradient is more gradual in dry forests due to a more open mature forest canopy Therefore, while many species common in mature dry forests are adapted to growing in high light conditions non-forested sites may be beyond the physiological tolerance of many forest trees in the moist tropics (Anderson 1990)

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Table 6-2 Comparison of rates of recovery of basal area tree species richness and canopy height reported b y other studies of forest succession in the tropics and subtropics For each index the age b y which secondary stands reach 75% of the mature forest stand and the mature stand values are reported. Basal area (m 2 /ha) Canopy height Tree species richness Age @ Mature Age @ Mature Age @ Mature Forest type 75% (m 2 /ha) 75% (m) 75% (#) Source montane > >30 60 3 > >30 36 15 20 Kappelle et al 1996 lower montane > 35 62 8 11 21 > 35 105 3 Kuzee et al 1994 subtropical wet 20 33 8 20 24 > 21 37 Lugo 1990 subtropical wet 25 20 Aide et al 1995 tropical moist > 80 34 8 > 35 25-35 60 67 Saldariaggaetal 1986 tropical moist > 80 35 6 5 66 Saldariagga et al 1988 tropical wet 17 33 > 17 70 Guariguata et al. 1997 tropical dry > 10 I 0.4 l 0 25 Aweto 1981 a b tropical dry 23 25 23 25 5 3 7 This study > > signifies that the mature stand value was more than twice that of the oldest secondary stand reported a Total species richness

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169 While it is likely that differences in climate or soil fertility may be partl y responsible for the differences in regeneration rates among the forests types listed in Table 6-2 (Lugo and Brown 1990 Guariguata et al 1997) the type and intensity of previous use also influences rate of recovery For example Aide et al. (1995) found forest growth to be much slower after pasture abandonment than following less intense forms of disturbance in wet forests of Puerto Rico In Mexican humid forests above-ground biomass accumulation was inversel y related to the duration of prior land use (Hughes et al 1999) Kappelle et al. ( 1996) found less abundant oak reg e neration in areas of Costa Rican montane forest that had experienced high-intensity burns In the Venezuelan Amazon tree species richness was 3 to 66 times greater in land that was immediatel y abandoned after slash-and-burn than land that was cultivated before abandonment (Uhl e t al.1988a) Therefore the rates of recovery of these different forests should be interpreted with caution as the intensit y of disturbance likely differed among the studies Although a distinct pattern emerged from the chronosequence there was notabl e variation in structural traits among the different stands not explained b y differences in ag e For example the 23y r-old stand had higher stem densities and canopy height than expected This ma y be du e to a number of factors that likely varied between fields formed at different times and in different locations including differences in the length of cultivation soil type size of clearing availability of propagules and rainfall during the first years of development or disturbance b y fire cattle grazing or fuel wood collection in the years after abandonment. Substantial differences among the three stands formed in the same year suggest that variation among factors operating very early in succession such as initial site conditions seed dispersal germination and predation could account for much of the variation among stands The three 1-yr-old stands had ranges of tree spec i es richness of 11-18 species basal area 0 84-2 28 m 2 /ha and stem density of 9 880-42 800 stems/ha < 5 cm dbh

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170 History of the mature stand The dominance of long-lived pioneers and their even-aged size class structures in the mature stand indicate the present forest formed after a severe disturbance. Several additional findings suggest that this disturbance was most likel y fire Dendrochronology of Cedrela ji silis one of the few species that creates annual rings indicated that intense fires have occurred roughl y every 30-60 years in Las Trancas (J Huffman pers comm) In the course of the soil studies conducted in Chapter 3 I found charcoal fragments at various locations and soil depths in both Las Trancas '94 and 95 Although the cause of these past fires are unknown pottery shards (found in soil sampled from 6 different sites within Las Trancas) suggests that at least some fires were anthropogenic in origin How old is the mature forest?" While this stand s structural features such as stem density and total basal area are very similar to those of the 50-yr-old stand (Figure 6-8) more information on features that may change less dramaticall y, such as the composition of understory species or size of the largest individuals may show that it is much older. For example the largest tree in the mature stand was 115 cm dbh (Schinopsis brasilensis) in the 50-yr-old stand it was only 78 cm dbh (Phyllostylon rhamnoides) Presumably the mature stand is as old as its even aged population of canopy trees According to estimates from other neotropical forests the life span of long-lived pioneers ranges from 75 to 150 years (Lieberman and Lieberman 1987 Finegan 1996). The role of disturbances in the dry forests of Lomerio There is ample evidence that disturbance of both natural and anthropogenic origins has been an e>..1ensive force throughout the Amazon basin for rnillenia (Clark and Uhl 1987 Goldammer 1990 1992 1993 Schule 1990, Kershaw 1997) Much of what was once considered virgin forest is now recognized as regrowth (e g. Budowski 1965 Dencvan 1976 Denevan et al. 1989). This study provides additional evidence of the pervasive if not frequent role of

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disturbance in the neotropics And I will argue that tropical dry forests have likel y been more prone to both natural and human disturbance than wetter regions of the Amazon basin. 171 Previously major fires were thought to occur infrequently in tropical dry forests (Malaisse 1978 Janzen 1988 Murphy and Lugo 1986) presumably due to the sparcity of combustible material on forest floors (Hopkins 1983) However, natural fires caused by dry thunderstorms are likely to occur during the transition from the dry period to the rainy season and have been witnessed in dry tropical forests (Middleton et al 1997) The susceptibility of this vegetation type to natural fire is also supported by the charcoal record Radiocarbon dates of charcoal throughout the Amazon correspond to dry episodes during glacial periods when dry forests expanded and more humid forests contracted (Goldhammer 1993 Clark and Uhl 1987 Saldriagga et al 1986) Janzen ( l Q75) hypothesiz e d that the dominance of sprouting species in the dry tropics. as well as the comparatively simple forest structure may be the consequence of their more extensive disturbance regime Evidence from lowland Bolivia suggests that dry forests suffered less mortality from wildfire than moister forests (Mostacedo et al. 1998) which may also be indicative of the pervasive role fires have played in dry forests In the dry forests of Bolivia anthropogenic disturbance also has likely been more extensive than in wetter parts of the Amazon Denevan (1976) estimated that the pre-European contact population density of the region that includes Chiquitania was nine times higher than other Amazonian lowland forests due to its more fertile soil. 1n fact present population densities in Lomerio ( ~2.0 people km 2 ) are only slightly higher than the pre-contact density estimated b y Denevan (0 6-1 8 people km ). Indicative of this high indigenous population Chiquitania was one of the centers of Jesuit missions in South America Within Chiquitania Lomerio has historically been a refuge of indigenous people fleeing the Jesuit missions in the 1700s and later white land owners and rubber barons attempting to enslave them in the 1800s and early 1900s (Krekeler

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172 1985). The even-aged population structure of many of the shade-intolerant tree species ma y in fact be the legacy of centuries of slash-and-bum cultivation Implications: Management potential of secondary forests in Lomerio Management potential of an y forest stand is dependent on the harvestable vol ume of trees In Lom e rio approxin1at e l y half of the commercial timber tree species are lon g -l ive d pioneers and are found in abundance in secondary forests The 50-yr-old stand in this stud y averaged 30 commercial trees of harvestable size(> 40 cm DBH) per hectare comparable to the number in mature forest (33 trees/ha). The abundance of shade-intolerant commercial tree species in secondary forests and the declin e in their populations in maturing stands has important implications for the future management of these species in Lomerio. Managing secondary forests for secondary forest species I (i e., long-lived pioneers) is likely a more viable strategy than the current practice in Lomerio : managing mature forests for secondary forest species Managing mature forests for species that naturally regenerate in even-aged populations logically entails heavy intervention (Dawkins 1958) As such silvicultural methods intensive enough to enhance regeneration of long-lived pioneers are also likely be very expensive In Bolivia preliminary cost estimates of using prescribed bums to enhance regeneration of the species highlighted in this study were higher than the expected benefits resulting from increased seedling densities ($8-14 per gap ; Ramirez 1998) Use of s e condary forests in Lomerio could offset these high silvicultural costs by taking advantage of the abundant regeneration of commercial species in agricultural fallows Additionally intensive silvicultural treatments ma y compromise goals of maintaining forest integrity and biodiversity In Lomerio the objecti ve of the current management plan is to sustainably produce timber while minimizing negative impacts on the other biological and ph ysi cal resources in the forest (Pinard e t al. 1999) Although the dry forests in Lomerio ma y be more

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resilient to disturbance than wetter tropical forests intensive silvicultural intervention ( i e ., prescribed burning) ma y alter species composition if applied on a large scale (Chapter 5 ) Management of secondary forests for timber is certainl y not a new idea ( e. g ., Holdridge 17 3 I 9 57 E w e! 197 9 Wadsworth 1983 Budowski 1985 Lamprecht 1989 Dubois 1990. Fine g an 1992). High concentrations of economically important tree species (i e Did y mo p an ax V o c hysi a Cordia Simarouba Goupia Laetia Cedrela) have been documented elsewhere in s e condary neotropical forests (reviewed by Finegan 1992) In fact one of the earliest tropical s1lvicultur e systems was based on the management potential of abandoned slash-and-bum fi e lds Dev e loped by Dutch foresters in Burma in the mid 1800s the taungya or tumpang sari s y stem leased ro y al forest lands to shifting cultivators who planted teak between their usual crops of rice and cotton After several y ears of cultivation fields were left fallow and managed for teak (Peluso 199 2 Palmer and Dawkins 1993 Dawkins and Philip 1999) A more recent example i s the Trin i dad Shelterwood system initiated in the 1950s in which secondary forests in Trinidad w ere managed for fast-gro w ing li ght-demanding species such as Did y mopanax B y ronsomium and L a etia (Finegan 1992) The management of secondary forest in generall y is onl y economicall y possible w h e re the light demanding trees are commercial as opposed to merely utilizable (Finegan 1992) ; this condition which exists in Bolivia (Tabel 1-1 ) However use of secondary forests in Lom e rio would not prevent harvesting of mature forest altogether. There was a notable lack of shad e tolerant species in secondary stands and therefore a lower diversity of commercial species In order to maintain current markets for these species it would be necessary to manage matur e forest for shade-tolerant species The silvicultural treatments required to insure the reg e neration of shade-tolerant species in ma~re forest require comparativel y little intervention and are therefore not likel y to be as costl y or detrimental to biodiversity.

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CHAPTER6 SUMMARY AND CONCLUSIONS Summary of Study Results Due to the diversity of regeneration strategies among commercial tree species in Lomerio canopy gap formation, plant removal and low and high intensity controlled burns had variable effects on their regeneration Regeneration of shade-intolerant species that lacked seedling banks was dependent on the creation of sites suitable for their seedling establishment, growth, and survival. Reduction of competing vegetation by controlled burning and plant removal benefited seedling establishment of shade-intolerant species. Subsequent seedling survival and growth of these species was greatest after high intensity burns likely due to the slow recovery of competing vegetation and the dramatic although short-lived increase in available nutrients Shade-tolerant species had seedling banks in mature forest understories therefore the success of these species was dependent on the survival of this advanced regeneration and the creation of microsites that enhanced growth Although survival of advanced regeneration was high in canopy gaps growth of this regeneration will likely be slow due to the abundance of competing vegetation Sprouting of shade-tolerant species was a common means of surviving plant removal and low intensity burns. Sprouts exhibited vigorous growth and dominated con specific seedlings in these treatments. Most advanced regeneration of shade-tolerant species was killed during high intensity burns At least one commercial tree species in Lomerio regenerates predominately as root sprouts Root sprouts of this species C entrolobium micro c haete were most abundant in gaps 174

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17 5 where a mature individual was harvested. Within these gaps root sprouts were most abundant in treatments where damage but not mortality of roots was likely (plant removal and low intensity burn treatments) Overall sprouting proved to be an important pathway for regeneration following treatments. Sprouts from roots root collars or sterns dominated canopy gap, plant removal and low intensity burn treatment plots throughout the study In treatment plots not dominated by root sprouts of C entrolobium microchaete the largest sprouts were of non-commercial species Commercial tree seedlings were among the largest individuals only on high intensity burn plots. The importance of sprouting in this dry forest likely contributed to a relatively high species similarity among treatments Only where the frequency of sprouting was reduced by high intensity burns was species composition altered most from the other treatment plots. Regeneration of shade-intolerant tree species was also abundant in recently-abandoned slash-and-bum fallows in Lomerio Older fallows ( 40-50 yrs) had similar densities of commercial trees as managed mature forest. The mostly even-aged stands of shade-intolerant canopy species in the mature forest suggest that it developed after some type of large-scale disturbance Pottery shards and charcoal fragments found in soil sampled from this mature forest suggests it may have once been under cultivation. Implications for management In Lomerio tree species spanning the continuum from shade-intolerant to shade-tolerant are commercially valuable and are harvested from mature forests. Due to the well-planned labor intensive harvesting operations and the relatively low volume of timber extracted only modest disturbance to the remaining forest occurs. This management scheme may be suitable for one group of commercial species : shade-tolerant species that may only require small canopy openings to release their advanced regeneration. As is commonly recommended for managed natural forests in the tropics managing advance-regenerating species would require little damage to the

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176 residual forest during harvesting operations as apparently is happening in the managed forests of Lomerio Due to the abundance and vigor of competing vegetation in harvesting gaps additional treatments such as weeding, release or thinning may be necessary at various intervals after harvesting. Alternatively limiting the sizes of harvesting gaps to discourage dense regeneration of more light demanding non-commercial species may also aid the growth and survival of advanced regeneration of commercial species In contrast, harvesting activities alone do not provide appropriate microsites for the regeneration of many commercial species in Lomerio the majority of which are shade-intolerant. Most of these species lack seedling banks in mature forests and therefore must rely on regeneration from seed following harvest. The results of these studies suggest that even large multiple-tree harvesting gaps do not provide sufficient light for seedling establishment of many of these species, again due to the abundance of competing vegetation. By reducing this competition exposing mineral soils and increasing available forms of nutrients controlled bums can enhance seedling establishment growth, and survival of these shade-intolerant species. Although this evidence strengthens the promise of controlled bums as a silvicultural tool in Lomerio several points stilJ need to be addressed before bums should be prescribed on a management scale Fire intensity. Controlled bums can be manipulated in various ways to achieve different results. The 'success of controlled burns depends to a great extent on if their results meet specified management objectives. Fire intensity is an important determinate of the extent and duration of competition removal the amount of mineral soil exposed, and alterations of nutrient availability While low intensity burns control competing vegetation little more than manual plant removal high intensity bums prevent regeneration of all but the most fire-tolerant sprouting species or those colonizing from dispersed seeds. Also, high intensity burns can greatly increase available nutrients but often at the expense of total nutrient stores and soil structure Low intensity burns increase avaiJable nutrients to a lesser extent than high intensity bums but also are not as

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177 damaging to soil physical properties. Forest managers should control as much as is possible the intensity of burns in accordance with their management objectives. Seed availability. Improving microsite conditions for establishment of shade-intolerant species through controlled burning will only meet management objectives if seeds are available Due to the irregular inter-annual reproduction of many of tree species in Lomerio seed limitation will likely be a problem for at least some commercial species in any given year. Also seed trees of most commercial species are infrequent, and therefore their seeds are patchily dispersed. Therefore augmenting naturally dispersed seeds should be done in combination with controlled burns to ensure sufficient seedling establishment. Although low seed longevity may prevent seed storage from year to year manually redistributing collected seeds into burned areas should increase seedling abundance of many species. Also as few seeds on soil surfaces or buried in the soil survive burns, the timing of burns can also be controlled to maximize seed availability Seed survival is most easily ensured by timing controlled burns before peak seed fall of targeted commercial species. Deciding when and where controlled bums are appropriate. Intensive silvicultural treatments such as controlled burns designed to enhance regeneration of shade-intolerant species will also damage or kill advanced regeneration of shade-tolerant species 1bis conflict requires that species guilds be managed either in different areas or at different times For example if advanced regeneration of shade-tolerant species is sufficient in forest understories, then low impact han esting methods followed by weeding or thinning should be prescribed rather than more intensive treatments such as controlled burns. Alternatively these two guilds can be managed within the same forest at different cutting cycles as in the Trinidad Shelterwood System (Finegan 1992) The TSS was projected as a polycyclic system with a 60-year rotation and two 30-year cutting cycles The first harvest focused on the even-aged stand of fast growing shade-intolerant species and the second harvest

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178 was to have exploited the more shade-tolerant species which regenerated continuously in the understory In order to maintain current markets for Bolivian timber species management units should be at different cycles in the rotation to ensure supplies of both secondary and more shade tolerant species. However even in this polycylic system the choice of opting for intensive silvicultural treatments over less intensive treatments will still arise at the start of each new rotation A second alternative is to reserve management of mature forests for more shade-tolerant species while focusing management of shade-intolerant species where their regeneration is abundant, such as in secondary forests that regenerate following abandonment of slash-and-bum fields. Commercial management of agricultural fallows for shade intolerant species was a common practice in Burma in the late 1800s and early 1900s (Dawkins and Philip 1998) and more recently has been attempted with promising results in Mexico (L. Snook, pers. comm.) One advantage of harvesting shade-intolerant species from abandoned slash-and-bum fields is that the costs of these intensive silvicultural treatments are incorporated into the traditional agriculture system of the local population However this system has several disadvantages including the small size of clearing made by farmers as well as maintaining fallows for longer rotation periods than local farmers are accustomed Also, farmers may not wish to travel far from their communities causing difficulties in the distribution of harvest units over the entire forest. Gaps in knowledge of controlled bums in managed tropical dry forests Effects of scale. Although many of the changes brought about by bums are short lived (e.g. nutrient availability) other changes may last for longer periods (e g ., altered soil physical properties) This study examines only a short duration of forest regeneration following controlled bums Furthermore the controlled burns in this study were small in scale. As area : edge ratios are likely to affect regeneration (i e ., through their affects on dispersal, predation, duration of insolation etc ) the effects of controlled burns are expected to vary with burn size. As such the

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179 ecological effects of larger burns as well as long-term effects of small and large burns should be documented to provide a more complete knowledge base for forest managers. Economic viability. The studies presented in this dissertation only document ecological effects of controlled burning. Before bums are prescribed on a management scale the economic viability of controlled burning in Bolivian forests should be assessed as well. Preliminary costs estimates by BOLFOR suggest that due to the labor needed to construct fire breaks and guard burns burning relatively small areas such as harvesting gaps i s not cost-effective. Burning large areas may be more cost-effective as cost per unit area decreases with increasing bwn size However, conducting large bums requires considerably more skill and knowledge of fire behavior than smaller burns and therefore education and training of bum crews will become more important, and possibly more expensive Interpretations of any economic analysis must be based on what factors are accounted for in the analysis According to some authors present investments in silvicultural treatments for a tree crop that will be harvested in 30 years are likely to be financially unattractive due to the high discount rates typical of many tropical countries (e g ., Rice et al 1997). However this argument is based on no consideration of externalities (i e. costs to biodiversity watersheds air quality etc ) Improving regeneration through controlled bums may indirectly benefit the conservation of biodiversity and maintain water and air quality by slowing the c onversion of managed forests to non-forest uses. On the other hand, controlled bums may also alter biodiversity as well as lower air and water quality through the release of particulates Economic analyses of prescribed burning will also be complicated by the contribution of prescribed burns to carbon emissions. The potential for carbon release from forest fires can be magnitudes greater than that caused by logging (Nepstad et al 1999) therefore carbon emissions from prescribed burns should be accounted for in economic analyses where externalities are considered The danger of fire escapes from prescribed burns and the resulting carbon emissions

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180 are potentially costly as well For example in the dry season of 1999 wildfire damaged over 1.6 million hectares of Bolivian forests liberating an estimated 17 3 million tons of carbon into tl1e atmosphere (W. Cordero pers comm.) This estimate of carbon emissions from only one wildfire season, is roughly equivalent to the 30-year goal for carbon sequestration of a 2 1 million hectare addition to the Noel KempffMercado National Park in eastern Bolivia, a purchase that was funded through a carbon-offset program (Environmental News Network 1998) Although introducing more fire into a system through prescribed burning may be viewed as increasing the danger of fire escape sound fire management can actually decrease the likelihood of wildfire Fire management policy. If controlled burning is adopted as a forest management tool in Bolivia, it should be as a part of a larger integrated fire management system In addition to using controlled bums to meet management goals, integrated fire management requires the capability to actively manage all fire situations including preventing and/or suppressing undesirable fires (Goldammer 1992). In addition to the relevant ecological and economic knowledge, an integrated fire management system requires substantial infrastructure and trained personnel. Many Bolivian forest managers express a doubt that fire management will be conducted in Bolivian forests in the near future, citing the inaccessibility of forests and the weak institutional capacity of organizations that might eventually conduct fire management as two important barriers to implementation (K. Gould pers. comm.). Although the studies described in this dissertation suggest that controlled bums may benefit the regeneration of commercial trees in Bolivian dry forests, it is likely that institutional and economic factors will ultimately determine whether prescribed burning is integrated into forest management.

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APPENDIX Appendix. Species recorded in four gap treaments 9 months following burns. The frequency of each species in the 16 plots of each treatment is listed for gap controls (GC) plant removal (PR), low intensity burn (LI) and high intensity burn (HI). "ALL" is the frequency of species in all 64 treatment plots. Frequency per treatment (%) Famill'. S~ecies GC PR LI HI All Bromeliads Bromeliaceae Psuedoananas saginarius (Arruda) 100 100 100 69 92 Cacti Cactaceae Pereskia sacharosa Griseb. Cactaceae C ereus tacuara/ensis Cardenas 6 6 6 5 Cactaceae Opuntia brasiliensis (Willd ) 31 19 6 14 Fern Schizeaeceae Anemia rotundifolia Schrader 25 19 ,., 1-' 14 Grass Cyperaceae Cyperus sp. 19 13 13 11 Poaceae Chusquea ramosissima Lindrn 6 6 3 Poaceae Panicum sp. 6 6 19 8 Poaceae Lasiacis sorgoides (Desv.) Hitchc. 19 31 31 25 27 Poaceae O/yza /atifolia Desv. 44 31 19 19 28 Poaceae Poaceae sp. 1 19 19 25 19 20 Poaceae Pharus lappulaceus Aubl. 56 50 31 13 38 Poaceae Poaceae sp 2 19 13 25 25 20 herb Acanthaceae Anisacanthus boliviensis (Wees ) 6 6 3 Acanthaceae Justicia velascana Lindau 6 2 Acanthaceae Ruellia brevifolia (Pohl) 6 2 Acanthaceae Ruel/ia sp 1 88 94 88 50 80 Acanthaceae Ruellia sp. 2 88 94 88 31 75 Acanthaceae Rue/liasp. 3 50 50 56 13 42 Acanthaceae Ruellia sp. 4 75 88 88 38 72 Amaranthaceae Chamissoa acuminata Mart. 6 13 6 6 Amaranthaceae Jresine diffusa H.& B ex Willd. 25 6 19 6 14 Araceae Anthurium plowmanni Croat 6 25 13 11 Araceae Anthurium sp 6 2 Araceae Philodendron camposportoanum G M Burrose 19 13 13 13 14 Araceae Philodendron tweedieanum Schott 88 69 75 31 66 Asteraceae Ageratum sp. 38 50 31 30 Asteraceae Aster sp 1 6 2 Asteraceae Aster sp. 2 6 2 Asteraceae Aster sp. 3 6 2 Asteraceae Aster sp. 4 6 2 Asteraceae Aster sp. 5 6 6 13 6 181

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182 Appendix A. ( continued) Frequency(%) Famil~ S!!ecies GC PR LI m All Asteraceae Aster sp. 6 19 50 17 Asteraceae C hromolaena extensa 19 13 6 9 Asteraceae E rechtites hieracifolia 6 2 Commelinaceae Co mmelina sp. 1 38 44 38 38 39 Commelinaceae Co mmelina sp. 2 13 6 6 6 Commelinaceae Co mme/ina sp. 3 19 31 38 38 31 Commelinaceae Co mmelina sp. 4 Commelinaceae Dichorisandra sp 6 6 3 Costaceae C ostus arabicus L. 31 25 38 19 28 Euphorbiaceae Aca/ypha multicoulis Mull. 56 75 81 50 66 Euphorbiaceae Cro ton sp 6 13 6 13 9 Euphorbiaceae Manihot sp. 1 6 6 3 Euphorbiaceae Manihot sp 2 6 13 19 13 13 Euphorbiaceae Manihot sp. 3 Euphorbiaceae Sebastiana sp. 6 2 Euphorbiaceae Euphorb sp. l 38 19 25 38 30 Euphorbiaceae Euphorb sp 2 13 13 6 8 Leguminosae Legume sp. 1 6 2 Leguminosae Legume sp 2 6 2 Malvaceae Abuti/on benense (Britlon) 6 6 3 Malvaceae Malvaceae sp. 1 6 2 Malvaceae Malvaceae sp. 2 6 6 3 Malvaceae Sida glabra Mill. 25 44 69 50 47 Marantaceae Calathea villosa Lindl. 25 44 63 25 39 Musaceae Heliconia sp 6 6 3 Orchidaceae Orchid sp 1 6 2 Orchidaceae Orchid sp 2 0 6 6 3 Orchidaceae Orchid sp. 3 6 2 Phytolacaceae Hilleria latifolia (Lam.) 6 2 Phytolacaceae Peliveria al/iaceae L. 6 13 6 6 Phytolacaceae Phytolacaceae sp 13 ,, ., Phytolacaceae Rivinia humilis (L.) 6 13 6 6 Piperaceae Pipersp. 6 13 5 Portulacca Talinum trianulare (Jacq ) 13 13 6 Rubiaceae Rubiaceae sp. l 6 2 Rutaceae Moniera trifolium L. 6 6 3 Solanaceae C apsicum chacoense Hunziker 13 6 5 Solanaceae Solanum apaense 13 3 Solanaceae Solanum riparium Pers. 13 25 38 19 unknown sp. 1 6 13 5 unknown sp 2 13 3 unknown sp. 3 6 6 3 liana Acanthaceae Mendoncia sp 44 19 50 13 31 Apocynaceae Forsteronia pubescens 6 13 13 8 Apocynaceae Prestonia acutifo/ia (Benth ex Muell Arg ) 6 6 13 6 Apocynaceae Apocynaceae sp 31 13 6 31 20 Asclepiadaceae Fischeria stellata (Yell.) E. Foum 19 13 13 11

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183 Appendix A. ( continued) Frequency (%) Familr Sl!ecies GC PR LI m All Asclep1adaceae Gonolobus denticulatus (Vofil) Steven l3 6 5 Asteraceae Mikaniasp 6 2 Bignoniaceae Adenocalymma bracteatum (Chamisso) DC. 6 2 Bignoniaceae Adenocalymma bracteolatum 6 6 13 6 8 Bignoniaceae Adenocalymma purpurascens 6 2 Bignoniaceae Arrabidaea brachypoda 6 2 Bignoniaceae Arrabidaea fagoides 88 94 94 56 83 Bignoniaceae Arrabidaea sp. 1 13 6 6 6 Bignoniaceae Arrabidaea sp. 3 50 38 38 31 39 Bignoniaceae Arrabidaea sp 2 6 2 Bignoniaceae Bignon. sp 1 6 13 13 19 13 Bignoniaceae Bignon sp 2 25 38 44 69 44 Bignoniaceae Bignon. sp 3 19 13 19 13 16 Bignoniaceae Bignon sp 4 6 13 6 6 8 Bignoniaceae Bignon sp. 5 6 13 5 Bignoniaceae Bignon. sp. 6 6 6 3 Bignoniaceae Bignon. sp. 7 6 6 6 5 Bignoniaceae Bignon sp 8 6 2 Bignoniaceae Bignon. sp 9 6 2 Bignoniaceae Bignon. sp. 10 13 3 Bignoniaceae Bignon sp 11 6 2 Bignoniaceae Bignon. sp. 12 6 2 Bignoniaceae Bignon. sp. 13 6 6 3 Bignoniaceae Clytostoma binatum (Thunb.) 13 6 6 6 Bignoniaceae C ydista decora (Spencer Moore) A. Gentry 88 75 56 31 63 Bignoniaceae Macfadyena sp. 56 56 63 63 59 Bignoniaceae Macfadyena uncata 13 3 Bignoniaceae Macfadyena unguis-cati (L.) A. Gentry 6 2 Bignoniaceae Manaosella ajf cordifolia (A. DC.) Gentry Bignoniaceae Perianthomega vellozoii 56 50 38 31 44 Bignoniaceae Pithecoctenium crucigerum (L.) A. Gentry 6 2 Caricaceae Carica sp 6 6 3 Combretaceae Thiloa paraguariensis Eichler 31 50 75 88 61 Convolvulaceae Convolvulaceae sp 1 13 13 13 19 14 Convolvulaceae Merremia sp. 13 13 6 Cucurbitaceae Cucurbit sp. 1 6 6 25 19 14 Cucurbitaceae Cucurbit sp. 2 6 6 6 13 8 Cucurbitaceae Cucurbit sp 3 6 2 Cucurbitaceae Echinopepon racemosus (Stued ) 19 19 19 13 17 Cucurbitaceae Fevillea sp 13 6 13 19 13 Cucurbitaceae Psiguria ternata (Roem ) C. Jeffrey 6 6 3 Dioscoreaceae Dioscorea sp. J 13 3 Dioscoreaceae Dioscorea sp. 2 6 2 Dioscoreaceae Dioscorea sp. 3 6 13 13 13 11 Euphorbiaceae Omphaleadiandra L. 25 56 19 13 28 Euphorbiaceae Tragia volubi/is L. 13 31 6 13 Hippocrateaceae Hippocratea volubilis L. 81 75 75 50 70 Leguminosae Coursetia hassleri Chodat

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18 4 Appendix A. ( continued) Frequency (%) Famill:'. Seecies GC PR LI HI All Legununosae Legume sp. 1 6 6 6 5 Leguminosae Legume sp. 2 6 2 Leguminosae Legume sp. 3 25 50 13 31 30 Leguminosae Legume sp. 4 6 13 19 13 13 Leguminosae Legume sp. 5 13 13 13 9 Leguminosae Legume sp. 6 6 6 "' ., Leguminosae Legume sp. 7 6 6 6 5 Leguminosae Legume sp. 8 38 25 38 44 36 Leguminosae Machaerium amp/um Benth 6 19 6 Leguminosae Mimosasp 88 63 69 13 58 Leguminosae Phaseolus sp. 6 13 13 6 9 Leguminosae Rhynchosia rojasii Hassl. 13 6 13 6 9 Liliaceae Herreria montevidenss Klotzsch 6 13 13 8 Malpighiaceae Stigmaphyllon sp. 6 6 3 Malpigiaceae Malpighiaceae sp. 1 50 63 44 6 41 Malpigiaceae Malpighiaceae sp. 2 6 13 5 Malpighiaceae Malpighiaceae sp 3 6 6 6 6 6 Malpigiaceae Malpighiaceae sp. 4 19 6 25 6 14 Malpigiaceae Malpighiaceae sp. 5 6 2 Malpigiaceae Malpighiaceae sp. 6 6 6 3 Malpigiaceae Malpighiaceae sp. 7 25 19 19 16 Malpigiaceae Malpighiaceae sp 8 6 2 Menispermaceae Cissampelos tropaeolifolia DC 6 2 Passifloraceae Passiflora amethystina 6 13 13 6 9 Sapindaceae Sapindaceae sp 1 6 2 Sapindaceae Sapindaceae sp. 2 6 6 3 Sapindaceae Sapindaceae sp 3 6 2 Sapindaceae Sapindaceae sp 4 6 2 Sapindaceae Serjania hebecarpa 13 13 13 9 Sapindaceae Serjania marginata 75 75 56 31 59 Sapindaceae Serjania reticulata 81 69 56 75 70 Sapindaceae Serjania sp 75 69 81 38 66 Sapindaceae Serjania tripleuria 6 6 3 Sapindaceae Thinouia cf paraguayensis 6 6 3 Sapindaceae Thinouia herbecarpa 44 44 44 19 38 Sapindaceae Thinouia sp. 6 6 3 Sapindaceae Urvillea sp 6 2 Trigoniaceae Trigonia sp. 63 63 75 69 67 Ulmaceae Ulmaceae sp 6 2 Vitaceae Cissus sp. 1 6 19 19 11 Vitaceae C issus sp. 2 44 44 69 56 53 Unknown sp 1 6 2 Unknown sp 2 13 "' ., Unknown sp 3 13 6 13 8 Unknown sp. 4 6 2 Unknown sp. 5 13 6 5 palm Palmae Atta/ea phalerata Mart. ex Spreng 13 13 6

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185 Appendix A. ( continued) Frequency (%) Familr SJ:!ecies GC PR LI m All Palmae Syagrus sacona Karsten 13 6 19 l3 13 shrub Apocynaceae Tabernaemontana cymosa Jacq. 6 2 Asteraceae Daciphyllum brasi/iensis (Spr.) 50 31 31 19 33 Caesalpiniodeae Bauhinia longicuspis 6 6 6 6 6 Capparaceae C apparis sp. 19 13 6 6 11 Erythroxylaceae E rythroxylum sp 13 44 19 19 23 Flacortiaceae Prockia crucis P. Browne ex L. 38 25 50 19 33 Myrtaceae C alyptranthes sp. 13 13 6 8 Myrtaceae Eugenia ligustrina Kiaersk 81 81 75 63 75 Myrtaceae Myrcieria cauliflora (Mart ) 13 13 6 6 9 Nyctaginaceae Bougainvillea modesta Heimerd. 6 6 6 5 Nyctaginaceae N y ctaginaceae sp 1 6 6 3 Rhamnaceae Ziziphus sp 25 19 38 19 25 Rubiaceae Rhandia sp. 1 19 13 6 6 11 Rubiaceae Rhandia annata 6 13 56 13 22 Rubiaceae Rhandia sp 2 6 6 6 5 Rubiaceae S imira rubescens (Benth). Bremek ex 13 13 13 6 11 Rutaceae Es enbeckia a/mawillia Kaastra 25 31 38 38 33 Violaceae Hybanthus comunis (St. Hil.) Taub. 81 100 88 69 84 Unknown sp 1 6 2 Unknown sp 2. 6 2 Unknown sp 3 13 13 13 13 13 Unknown sp 4 13 13 6 Unknown sp 5 31 56 50 19 39 Unknown sp 6 6 2 U nknown sp 7. 6 2 commercial tree species Anacardiaceae Astronium urundueva (Allemao) Engl. 13 38 56 94 50 Anacardiaceae Schinopsis brasiliensis Engl. 19 25 11 Apocynaceae A spidosperma cylindrocarpon Muell. 13 13 13 9 Apocynaceae Aspidosperma pyrifolium Mart. 13 6 13 8 Apocynaceae Aspidosperma rigidum Rushy 81 75 69 44 67 Bignoniaceae Tabebuia impetiginosa (Mart ex DC .) 13 13 38 44 27 Boraginaceae C ordia s p. 6 6 .., .) Boraginaceae C ordiasp. 19 13 44 31 27 Caesalpiniodeae C aesa/pinia floribunda Tul. 63 50 56 50 55 Caesalpiniodeae C opaifera chodatiana Hassler 75 69 69 56 67 Meliaceae C edrela fissilis Veil 13 6 6 6 Mimosoideae Anadenanthera colubrina (Veil Cone.) 100 100 100 94 98 Mimosoideae Pithecel/obium sp 25 38 25 19 27 Papilionoideae Amburana cearensis (Allemao) 6 2 Papilionoideae C entrolobium microcha e te (C Marius ex Benth ) 38 56 6 3 69 56 Papilionoideae Machaerium scleroxylon Tul 0 19 6 19 11 Papilionoideae Platymiscium ulei Harms 19 6 31 6 1 6 Rubiaceae C al yc ophyllum multiflorium Griseb 6 6 6 5 Ulmaceae Phyllosty/on rhamnoides (Poisson) 6 6 3 Anacardiaceae Spondias mombin L. 13 13 31 25 20

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186 Appendix A. ( continued) Frequency (%) Familr S~ecies GC PR LI HI All non-commerc1al tree species Annonaceae Annona cf jahnii Saff. 6 19 19 6 13 Bignoniaceae Zeyheria tuberculosa (Veil.) 6 2 Bombacaceae Ceiba samauma (Mart.) Bombacaceae C horisia speciosa St. Hilaire 31 31 25 13 25 Bombacaceae Eriotheca roseorum (Cuatrec.) 6 19 25 6 14 Bombacaceae Psuedobombax marginatum (St. Hilaire) Caesalpiniodeae Bauninia angulata L. 6 2 Caesalpiniodeae Poeppigia procera C. 94 100 100 69 91 Capparaceae C apparis prisca J F Macbr 13 25 13 13 Caricaceae Ca rica sp. 6 6 6 5 Cochlospermaceae C ochlospermum vitifolium (Willd.) 6 6 6 5 Combretaceae C ambre/um leprosum (Mart.) 13 13 13 9 Flacortiaceae C asaeria acu/eata Jacq 13 44 19 13 22 Flacortiaceae C asaeria arborea (Rich .) 56 63 63 38 55 Flacortiaceae C asaeria gossypiosperma Briq 94 88 88 81 88 Flacortiaceae Prokia crucis P Browne ex L. Humiriaceae Sacoglottis mattogrossensis Malme. 6 6 3 Lecithidaceae C ariniana estre/lensis (Raddi) Kuntze 13 6 13 6 9 Lecithidaceae Cariniana sp. 6 2 Meliaceae Trichi/ia elegans A. Juss. 63 69 81 13 56 Mimosoideae Acacia loretensis J.F Macbr. 31 38 50 31 38 Mimosoideae Ac acia polyphylla DC. 38 63 56 38 48 Mimosoideae Enterolobium contortisiliquum (Veil Cone.) ) 19 19 13 13 Moraceae Cec ropia concolor Willd 6 13 31 38 22 Nyctaginaceae Neea hermaphrodita S. Moore Uel 94 100 94 63 88 Nyctaginaceae Nyctaginaceae sp 13 3 Papilionoideae Platypodium elegans 63 75 63 50 63 Phytolacaceae Gallesia integrefolia (Sprengel) 19 19 13 25 19 Rubiaceae Simira rubescens (Benth). Bremek ex Steryerm 81 63 75 44 66 Rutaceae Ga lipea trifoliata Aublet. 44 69 88 50 63 Rutaceae Zanthoxylum sp. 44 56 50 13 41 Sapindaceae Al/ophyl/us pauciflorus Radlk 63 69 50 19 50 Sapindaceae Talesia esculenta (St. Hil.) 13 6 13 13 11 Sapotaceae C hry so phyllum gonocarpum (Martius & Eichler) 6 6 6 5 Tiliaceae Heliocarpus americanus L. 6 13 6 6 Tiliaceae Luehea paniculata Martins 6 6 13 6 Ulmaceae Ce /tis iguanaea (Jacq.) 25 25 19 6 19 Ulmaceae Trema micrantha (L.) 44 44 25 28 Ulmaceae Ulmaceae sp. 13 6 6 6 8 Urticaceae U rera baccifera (L.) 38 56 38 19 38 V erbenaceae Lippea sp. 6 2 Unknown sp 1 6 2 Unknown sp. 2 6 6 .., .,

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203 Singh, K P. 1989. Mineral nutrients in tropical dry deciduous forest and savanna ecosystems in India In J Proctor (ed.) Mineral nutrients in tropical forest and savanna ecosystems pp 153-168 Oxford : Blackwell Scientific Publications Smith C. K., H L. Gholz and F De Assis Oliveira 1998 Soil nitrogen dynamics and plant induced soil changes under plantations and primary forest in lowland Amazonia Brazil. Plant and Soil 200 : 193-204 Snook, L. K. 1996 Catastrophic disturbance logging and the ecology of mahogany (Swietenia macrophylla King): Grounds for listing a major tropical timber species in CITES Botanical Journal of the Linnean Society 122 : 35-46. Sokal and Rolf 1981. Biometry 2 nd Edition New York: Freeman. SPSS for Windows 1997 Standard Version Release 8 0 0 Copyright SPSS Inc Stanley, S A 1995. Report on the methodology utilized in the study : Use of prescribed burning to foment the regeneration of commercial tree species in Lomerio Santa Cruz Bolivia BOLFOR Technical Document, Santa Cruz Bolivia Stanley S A. 1999. Prescribed Fire to Augment the Regeneration of Mahogany (Swietenia macrophylla) and Spanish Cedar (Cedrela odorata) in the Maya Biosphere Reserve Guatemala Masters Thesis North Carolina State University Raleigh NC. Stocker G. C. 1981 Regeneration of a North Queensland rain forest following felling and burning Biotropica 13 : 86-92 Stromgaard P 1992 Immediate and long-term effects of fire and ash-fertilization on a Zambian miombo woodland soil. Agriculture Ecosystems and Environment 41: 19-37 Swain M D 1996 The Ecology of Tropical Forest Tree Seedlings UNESCO Casterton UK : Parthenon Publishing Group Limited. Swaine, M D ., and T. C Whitmore 1988 On the definition ofccological species groups in tropical rainforests. Vegetatio 75 : 81-86 Thompson, J J Proctor, D A. Scott P J. Fraser R.H. Marrs, R. P Miller V Viana 1998 Rain Forest in Maraca Island, Rorairna Brazil: Artificial gaps and plant response to them Forest Ecology and Management 102 : 305-321. Thompson, K. 1992. The functional ecology of seed banks. In Seeds : M Fenner (ed ) The Ecology of Regeneration in Plant Communities, pp. 231-258 Wallingford UK : C.A.B International. Uhl C ., and R Buschbacher. 1985 A disturbing synergism between cattle ranch burning practices and selective tree harvesting in the eastern Amazon Biotropica 17 : 265-268. Uhl C. R. Buschbacher and E A. S Serrao. 1988a Abandoned pastures in eastern Amazonia I. Patterns of plant succession. Journal of Ecology 76 : 663-68 l.

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205 Weast R. C. 1988 Handbook of Chemistry and Physics Boca Raton FL: CRC Press Wells C. G ., R E Campbell L. F DeBano C. E Lewis R I. Fredericksen E C. Franklin R C. Froelich and P.H. Dunn 1979. Effects of fire on soil : A state-of-the-art review USDA Forest Serevice General Technical Report WO-7 Washington D.C. Wenger K. I. 19 53. The sprouting of sweetgum in relation to season of cutting and carboh y drate content. Plant Ph y siolog y 28 : 35-49. Weston C J and P M Attiwill 1996. Clearfelling and burning effects on nitrogen mineralization and leaching in soil of old-age Eucalyptus regnans forests. Forest Ecology and Management 89 : 13-24 Whelan J 1994 Ecology of Fire San Diego : Academic Press Whitman, A. A. N V L. Brokaw J M Hagan 1997 Forest damage caused by selection logging of mahogan y (Swietenia macrophylla) in northern Belize Forest Ecology and Management 92 : 87-96 Whitmore T.C. 1984. Tropical Forests of the Far East 2 nd Edition Oxford : Clarendon. Whitmore T. C. 1989 Canopy gaps and the two major groups of forest trees Ecology 70 : 536538 Wright H. A. and AW Bailey 1982 Fire Ecology : United States and Southern Canada New York: John Wil ey and Sons Wyatt-Smith J 1987 The Management of Tropical Moist Forest for the Sustained Production of Timber : Some Issues IUCN/IIED Tropical Forest Policy Paper 4. IUCN Gland Switzerland Young K. R. J J Ewel and B. J. Brown 1987 Seed dynanlics during forest succession in Costa Rica. Vegetatio 71: 157-173. Zedler P H ., C.R Gautier and G S McMaster. 1983 Vegetation change in response to extreme events : The effect of a short interval between fires in California chaparral and coastal scrub Ecolog y 64: 809-818

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BIOGRAPHICAL SKETCH Deborah Karen Kennard was born in Dickinson Texas in 1969 She attended junior and senior high school in Tokyo Japan She earned a bachelor of arts degree in biology with a minor in fine arts from Trinity University in San Antonio Texas in 1991 Before entering University of Florida s master s program in botany in 1994 she taught English in Japan and volunteered on various conservation and management projects in both Florida and Borneo She was awarded a master of science degree in botany in 1996 and a doctorate in 2000. Deborah plans to continue research on the ecological effects of forest management. 206

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I certify that I have read this study and that in my opinion it conforms to acceptable standards of scholarly presentation and is fully adequate in scope and quality a dis tio the degree of Doctor of Philosophy F E Putz, Chair ofessor of Botany I certify that J have read this study and that in my opinion it conforms o acceptable standards of scholarly presentation and is fully adequate in scope and quality as ssertation ti of Doctor of Philosophy I certify that I have read this study and that in my opinion it conforms to acceptable standards of scholarly presentation and is fully adequate in scope and quality as a dissertation for the degree of Doctor of Philosophy oru taJlllla Assistant Professor of Botany I certify that I have read this study and that in my opinion it conforms to acceptable standards of scholarly presentation and is fully adequate in scope and quality as a dissertation for the degree ofDoctor of Philosophy. / J,, ,;;:,_, Earl Stone ssor Emeritus of Soil and Water Science I certify that I have read this study and that in my opinion it conforms to acceptable standards of scholarly presentation and is fully adequate in scope and quality as a di rtation for the degree of Doctor of Philosophy. This dissertation was submitted to the Graduate Faculty of the College of Agricultural and Life Sciences and to the Graduate School and was accepted as partial fulfillment of the requirements for the degree of Doctor of Philosophy ~I 1 / May 2000 Dean, College of Agricultural and Life Sciences Dean, Graduate School

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00 UNIVERSITY OF FLORIDA II I II IIIII I Ill I l l ll l ll l l lll I I I II I II I I II I I II III I II II II I III I I Ill I I 3 1262 08555 169 4


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