Regression of 4-year survivorship percentage on the coefficient of variation of SLM of 23 non-pioneer tropical tree species growing in an experimental planting ...
SELECTING LATE-SUCCESSIONAL TREES FOR TROPICAL FOREST RESTORATION
BY CRISTINA MARTÍNEZ-GARZA B.S., National University of Mexico (UNAM), 1996
THESIS Submitted as partial fulfillment of the requirements for the degree of Doctor of Philosophy in Liberal Arts and Sciences in the Graduate College of the University of Illinois at Chicago, 2003 Chicago, Illinois
iii This thesis is dedicated to the Martínez and Garza families, with special thanks to the Martínez-Garza family, José Martínez, Magdalena Garza, José de Jesus, Silvia, Miguel Angel, Clara, Alejandro, Carlos, and to my husband Raúl Díaz-Heredia.
iv ACKNOWLEDGMENTS I would like to thank my thesis committee, Hormoz BassiriRad, Joel S. Brown, Robin B. Foster, Henry F. Howe and Martin Ricker, for their support and assistance. I especially thank Hank Howe for his advice and friendship. I also thank my colleagues from UIC, especially Norbert Cordeiro, Elaine Hooper, Malu Jorge, Maria Miriti, Gaby Nuñez, Manuel Pacheco, Sonali Saha (and Amartya), Amy Sullivan, Barbara Zorn-Arnold and, in Mexico, Reyna A. Castillo, Cecilia Sánchez, Juan Fornoni and Fernando Pacheco, all of them were a constant source of technical and emotional support. I thank all my brothers and sisters, especially Clara (and Gil), Alejandro, Miguel (and Frederike) and Silvia for all their help in the fieldwork. I am grateful to my husband Raul for his help through everything, life, the fieldwork and the writing of the thesis. I infinitely thank my parents and my grandmother (Esperanza Portillo) for their love and support that always accompany me. Finally, I thank the staff of the Los Tuxtlas Biological Station where the fieldwork was conducted. Financial support was provided by the National Council of Science and Technology of Mexico (CONACyT) and The Lincoln Park Zoo Neotropic Foundation of Chicago.
CMG
v TABLE OF CONTENTS CHAPTER 1.
PAGE GENERAL INTRODUCTION 1.1 Introduction…………………………………………………… 1.2 Restoration objectives………………………………………… 1.3 Conceptual Framework……………………………………….. 1.3.1 Leaf Traits of Tropical tree species……………….. 1.4 Literature Cited……………………………………..………….
2.
RESTORING TROPICAL DIVERSITY: BEATING THE TIME TAX ON SPECIES LOSS 2.1 Abstract……………………………………………………….. 2.2 Introduction…………………………………………………… 2.3 Natural succession……………………………………………. 2.4 Arriving and surviving in unnatural landscapes……………… 2.5 Attrition of species from fragments in alien matrices………... 2.6 Restoring the matrix: beating the time tax in diversity………. 2.7 Synthesis and Applications…………………………………… 2.8 Acknowledgment……………………………………………... 2.9 Literature Cited………………………………………………..
3.
1 2 3 4 7
11 12 13 14 17 19 24 25 26
ONTOGENY OF SUN AND SHADE LEAVES OF EIGHT NONPIONEER TROPICAL TREE SPECIES 3.1 Abstract………………………………………………………. 3.2 Introduction…………………………………………………... 3.3 Material and methods………………………………………… 3.3.1 Study site…………………………………………. 3.3.2 Sampling…………………………………………. 3.3.3 Data Analysis…………………………………….. 3.4 Results………………………………………………………… 3.4.1 Effects of species and ontogenetic stage on leaf traits……………………………………………….. 3.4.1.1.Leaf State…………………………………. 3.4.1.2.Leaf Flexibility…………………………… 3.4.2 Ontogenetic leaf variability and maximal height species……………………………………………. 3.5 Discussion……………………………………………………. 3.6 Acknowledgment…………………………………………….. 3.7 Literature Cited………………………………………………..
31 32 36 36 36 40 41 41 41 44 46 51 53 54
vi TABLE OF CONTENTS (continued) CHAPTER 4.
PAGE RESTORING BIODIVERSITY: PREDICTING SURVIVAL AND GROWTH OF LATE-SUCCESSIONAL TREE SPECIES IN EARLY SUCCESSIONAL ECOSYSTEMS 4.1 Abstract……………………………………………………….. 4.2 Introduction…………………………………………………… 4.3 Methods………………………………………………………. 4.3.1 Study site…………………………………………. 4.3.2 Land use history and Experimental planting…….. 4.3.3 Data analysis……………………………………… 4.4 Results………………………………………………………… 4.4.1 CV of Leaf traits and Survival……………………. 4.4.2 CV of Leaf traits and Increments in height and Diameter………………………………………….. 4.4.3 Maximal tree height…….…………………………. 4.5 Discussion……………………………………………………. 4.6 Conclusions…………………………………………………… 4.7 Acknowledgment……………………………………………… 4.8 Literature Cited…………………………………………………
5.
58 59 62 62 62 66 68 68 68 73 73 76 77 77
PREDICTING SURVIVAL AND GROWTH RATES OF LATESUCCESSIONAL TREE SPECIES IN ABANDONED PASTURES 5.1 Abstract………………………………………………………. 5.2 Introduction………………………………………………….. 5.3 Methods……………………………………………………… 5.3.1 Study Site………………………………………… 5.3.2 Land use and Experimental Plantation…………… 5.3.3 Data analysis……………………………………… 5.4 Results………………………………………………………… 5.4.1 Survival…………………………………………… 5.4.2 Increments in Height and Diameter………………. 5.4.3 Leaf Traits………………………………………… 5.4.3.1. Leaf size………………………………….. 5.4.3.2. SLM……………………………………… 5.4.3.3. Leaf Water Content……………………… 5.4.3.4. Leaf Density……………………………… 5.4.4 Leaf traits and performance in secondary forest and Pastures……………………………………………
82 83 86 86 86 89 91 91 91 92 93 93 95 95 96
vii TABLE OF CONTENTS (continued) CHAPTER
PAGE 5.4.5
6.
CV of leaf traits and performance in secondary forest and Pastures………………………………… 5.5 Discussion…………………………………………………….. 5.6 Conclusions…………………………………………………… 5.7 Literature Cited………………………………………………..
103 108 110 110
SYNTHESIS AND IMPLICATIONS……………………………. 6.1. Literature Cited……………………………………………….
113 123
APPENDICES Appendix A………………………………………………………. Appendix B………………………………………………………. Appendix C………………………………………………………. Appendix D……….……………………………………………… Appendix E………………………………………………………. Appendix F……………………………………….……………… Appendix G……………………………………………………… Appendix H……………………………………………………… Appendix I………………………………………………………. Appendix J………………………………………………………. Appendix K……………………………………………………… Appendix L……………………………………………………… Appendix M……………………………………………………… Appendix N………………………………………………………. Appendix O………………………………………………………. Appendix P………………………………………………………..
128 129 130 131 132 133 134 136 137 138 139 140 141 142 143 144
VITA……………………………………………………………….. ……
145
viii LIST OF TABLES TABLE I.
PAGE VARIABLES ASSESSED IN EIGHT TROPICAL NON-PIONEER TREE SPECIES……………………………….
35
SPECIES, AUTHORITIES AND RANGE OF STATURE OF EIGHT TROPICAL NON-PIONEER TREE SPECIES…….
37
LEAF TRAITS MEASURED IN EIGHT TROPICAL TREE SPECIES IN LOS TUXTLAS, VERACRUZ, MEXICO.............
39
SUMMARY OF MULTIVARIATE ANALYSIS OF VARIANCE OF LEAF TRAITS MEASURED IN EIGHT TROPICAL TREE SPECIES IN LOS TUXTLAS, VERACRUZ, MEXICO…………
41
UNIVARIATE ANALYSIS OF VARIANCE OF LEAF TRAITS AT EACH ONTOGENETIC STAGE OF EIGHT TROPICAL TREE SPECIES IN LOS TUXTLAS, VERACRUZ, MEXICO…
43
UNIVARIATE ANALYSIS OF VARIANCE OF LEAF TARITS FOR EACH LEAF ENVIRONMENT OF EIGHT TROPICAL TREE SPECIES IN LOS TUXTLAS, VERACRUZ, MEXICO…
43
SUMMARY OF MULTIVARIATE ANALYSIS OF VARIANCE OF LEAF FLEXIBILITIES MEASURED IN EIGHT TROPICAL TREE SPECIES IN LOS TUXTLAS, VERACRUZ, MEXICO…
45
UNIVARIATE ANALYSIS OF VARIANCE OF LEAF FLEXIBILITY FOR EACH ONTOGENETIC STAGE OF EIGHT TROPICAL TREE SPECIES IN LOS TUXTLAS, VERACRUZ, MEXICO…………………………………………..
46
PHOTON FLUX DENSITY (PFD), GRAVIMETRIC SOIL MOISTURE AND BULK DENSITY AT DIFFERENT LEVELS OF CROWN ILLUMINATION INDEX…………………………
64
FAMILY, FRUIT TYPE AND RANGE OF HEIGHT OF 23 LATE-SUCCESSIONAL TROPICAL TREE SPECIES…………
67
XI.
CORRELATIONS AMONG LEAF TRAITS……………………
67
XII.
PHOTON FLUX DENSITY (PFD), GRAVIMETRIC SOIL MOISTURE AND BULK DENSITY IN SECONDARY FOREST AND PASTURES…………………………………………………
89
II. III. IV.
V.
VI.
VII.
VIII.
IX.
X.
ix LIST OF TABLES (continued) TABLE XIII.
PAGE FAMILY, FRUIT TYPE AND RANGE OF MAXIMAL TREE HEIGHT OF 12 LATE-SUCCESSIONAL TROPICAL TREE SPECIES………………………………………………………….
XIV. SUMMARY OF MULTIVARIATE ANALYSIS OF VARIANCE OF INCREMENTS IN HEIGHT AND DIAMETER MEASURED IN 12 TROPICAL TREE SPECIES IN LOS TUXTLAS, VERACRUZ, MEXICO.............…………………………………. XV.
90
92
SUMMARY OF MULTIVARIATE ANALYSIS OF VARIANCE OF LEAF TRAITS MEASURED IN 12 TROPICAL TREE SPECIES IN LOS TUXTLAS, VERACRUZ, MEXICO…
95
XVI. LEAF SIZE MEASURED IN EIGHT TROPICAL TREE SPECIES AT THREE ONTOGENETIC STAGES IN LOS TUXTLAS, VERACRUZ, MEXICO..............................................
127
XVII. SLM MEASURED IN EIGHT TROPICAL TREE SPECIES AT THREE ONTOGENETIC STAGES IN LOS TUXTLAS, VERACRUZ, MEXICO…………………………………………..
128
XVIII. LEAF DENSITY MEASURED IN EIGHT TROPICAL TREE SPECIES AT THREE ONTOGENETIC STAGES IN LOS TUXTLAS, VERACRUZ, MEXICO.............................................
129
XIX. LEAF TOUGHNESS MEASURED IN EIGHT TROPICAL TREE SPECIES AT THREE ONTOGENETIC STAGES IN LOS TUXTLAS, VERACRUZ, MEXICO.............................................
130
XX.
LEAF WATER CONTENT MEASURED IN EIGHT TROPICAL TREE SPECIES AT THREE ONTOGENETIC STAGES IN LOS TUXTLAS, VERACRUZ, MEXICO..............................................
131
XXI. UNIVARIATE ANALYSIS OF VARIANCE OF FOLIAR FLEXIBILITY OF EIGHT TROPICAL NON-PIONEER TREE SPECIES IN LOS TUXTLAS, VERACRUZ, MEXICO………...
132
XXII. FAMILY, COEFFICIENT OF VARIATION OF LEAF TRAITS AND SAMPLE SIZE OF 23 LATE-SUCCESSIONAL TROPICAL TREE SPECIES…………………………………………………...
133
XXIII. SURVIVAL OF 12 LATE-SUCCESSIONAL TROPICAL TREE SPECIES IN PASTURES AND SECONDARY FORESTS……..
136
x LIST OF TABLES (continued) TABLE
PAGE
XXIV. LEAF TRAITS AND TOTAL NUMBER OF INDIVIDUALS SAMPLED OF 12 LATE-SUCCESSIONAL TROPICAL TREE SPECIES………………………………………………………….
137
XXV. PEARSON CORRELATION AMONG LEAF TRAITS MEASURED IN PASTURES AND SECONDARY FOREST FOR 12 TREE SPECIES AT LOS TUXTLAS, MEXICO………
138
XXVI. SUMMARY OF UNIVARIATE ANALYSIS OF VARIANCE OF INCREMENTS IN HEGHT FOR BROSIMUM ALICASTRUM AND POUTERIA CAMPECHIANA FROM A GREENHOUS EEXPERIMENT…………………………………
139
XXVII. SUMMARY OF MULTIVARIATE ANALYSIS OF VARIANCE OF LEAF TRAITS OF BROSIMUM ALICASTRUM AND POUTERIA CAMPECHIANA GROWING IN A GREENHOUSE EXPERIMENT AT THREE LIGHT LEVELS AND TWO WATER LEVELS…………………………………………………………..
141
XXVIII. LEAF TRAITS OF BROSIMUM ALICASTRUM AND POUTERIA CAMPECHIANA GROWING IN A GREENHOUSE EXPERIMENT AT THREE LIGHT LEVELS AND TWO WATER LEVELS………………………………………………………….. 142
xi LIST OF FIGURES FIGURE 1.
2.
3.
4.
5.
6.
PAGE Correlation between ontogenetic leaf variability and the maximal height of species for eight non-pioneer tropical tree species from Los Tuxtlas, Veracruz, Mexico. A) Leaf size from juveniles to adults, B) SLM from seedlings to juveniles, C) SLM from seedlings to adults. The key for the species follows Table II……………………………………….
48
Regression of 4-year survivorship percentage on the coefficient of variation of SLM of 23 non-pioneer tropical tree species growing in an experimental planting in the Cooperative of Lázaro Cárdenas, Los Tuxtlas, Mexico. Open circles correspond to canopy and emergent species (> 25 m) and filled circles to understory species. The key for the species follows Table X……………………………………………………………
69
Regression of the Incremental growth in basal diameter on the coefficient of variation of SLM of 23 non-pioneer tropical tree species growing in an experimental planting in the Cooperative of Lazaro Cárdenas, Los Tuxtlas, Mexico. Open circles correspond to canopy and emergent species (> 25 m) and filled circles to understory species. The key for the species follows Table X…………………………………………………..
71
Regression of the Incremental growth in height on the coefficient of variation of SLM of 23 non-pioneer tropical tree species growing in an experimental planting in the Cooperative of Lazaro Cárdenas, Los Tuxtlas, Mexico. Open circles correspond to canopy and emergent species (> 25 m) and filled circles to understory species. The key for the species follows Table X……………………………………………………………
72
Regression of the Coefficient of variation of SLM on the maximal tree height of 23 non-pioneer tropical tree species growing in an experimental planting in the Cooperative of Lazaro Cárdenas, Los Tuxtlas, Mexico. The key for the species follows Table X……………………………………………………………
74
Log Increment of diameter of 12 late-successional tropical tree species Key for species follows Table XIII. Asterisks over bars represent significant differences between secondary forest and pasture tested with Pos Hoc Tuke Test…………………………...
94
xii
LIST OF FIGURES (continued) FIGURE 7.
8.
9.
10.
11.
12.
13.
PAGE Regression of 17-month survivorship percentage on the Leaf size of 11 non-pioneer tropical tree species growing in an experimental planting in the Cooperative of Lázaro Cárdenas, Los Tuxtlas, Mexico. The key for the species follows Table XIII.
97
Regression of Log of monthly increment in diameter of individuals growing in pasture on log of SLM of conspecific growing in the secondary forest for 12 late-successional tropical tree species in the Cooperative of Lázaro Cárdenas, Los Tuxtlas, Mexico. The key for the species follows Table XIII……………..
99
Regression of Log Increment in diameter on log of SLM of individuals growing in the pasture for 12 late-successional tropical tree species in the Cooperative of Lázaro Cárdenas, Los Tuxtlas, Mexico. The key for the species follows Table XIII.
100
Regression of Log Increment in diameter of individuals growing in pasture on Log of Leaf Density of conspecific growing in the secondary forest for 12 late-successional tropical tree species in the Cooperative of Lázaro Cárdenas, Los Tuxtlas, Mexico. The key for the species follows Table XIII…..………………………..
101
Regression of 17-months survivorship of individuals growing in pastures on Log Increment in height for 12 late-successional tropical tree species in the Cooperative of Lázaro Cárdenas, Los Tuxtlas, Mexico. The key for the species follows Table XIII.
102
Regression of 17-month survivorship in secondary forests on the coefficient of variation of Water Content of 12 non-pioneer tropical tree species growing in an experimental planting in the Cooperative of Lázaro Cárdenas, Los Tuxtlas, Mexico. The key for the species follows Table XXIII……………………………….
104
Regression of the increments in height in secondary forests on the Coefficient of variation of Water Content of 12 non-pioneer tropical tree species growing in an experimental planting in the Cooperative of Lázaro Cárdenas, Los Tuxtlas, Mexico. The key for the species follows Table XXIII………………………………
105
xiii LIST OF FIGURES (continued) FIGURE 14.
15.
16.
17.
18.
PAGE Regression of the increments in height in secondary forests on the Coefficient of variation of Leaf Density of 12 non-pioneer tropical tree species growing in an experimental planting in the Cooperative of Lázaro Cárdenas, Los Tuxtlas, Mexico. The key for the species follows Table XXIII………………………….……
106
Regression of the increments in height in pastures on the Coefficient of variation of Leaf Density of 12 non-pioneer tropical tree species growing in an experimental planting in the Cooperative of Lázaro Cárdenas, Los Tuxtlas, Mexico. The key for the species follows Table XXIII……………………………....
107
Regressions of Log of monthly Increment in height on the coefficient of variation of SLM for canopy and understory species growing in an experimental planting in the Cooperative of Lazaro Cárdenas, Los Tuxtlas, Mexico……………………………………
135
Log increments in Height measured in Brosimum alicastrum and Pouteria campechiana individuals growing in a greenhouse experiment at three light levels and two water levels…………….
140
Regression of Log increment in height and Log of SLM for Brosimum alicastrum (bold letters) and Pouteria campechiana. Each point represent mean Log increment in height and mean Log of SLM for a combination of light levels (HL, high light levels [100 %], ML, medium [50 %] and LL, low light [30 %], and water availability (HW, high water availability [70 %], LW, low water [50 %])…………………………………………………
143
xiv LIST OF ABBREVIATIONS ANOVA
Analysis of Variance
CV
Coefficient of Variation
Env
Environment
LTBS
Los Tuxtlas Biological Station
Leaf E
Leaf Environment
MANOVA
Multivariate Analysis of Variance
Ontogenetic S
Ontogenetic Stage
P
Pasture
PFD
Photon Flux Density (µmol /m2 s)
SLM
Specific Leaf Mass
SF
Secondary Forest
Sec For
Secondary Forest
WC
Leaf Water Content
xv SUMMARY Planting seedlings of interior forest species after land abandonment could sharply accelerate the process of re-vegetation of complex communities. Pioneer stands or monocultural plantations may be enriched with seedlings of late-successional animaldispersed trees, or initial plantings could be mixes of late-successional and pioneer species. This thesis sets criteria for selecting species for enrichments and in some cases for overstories. My main hypothesis is that those species with leaf traits associated to high light levels and/or higher ability to adjust leaf traits to different light levels and water availability, will survive and grow better in environmental conditions typical of early successional environments (i.e., abandoned pastures, secondary forest). Further, tall species (canopy and emergent) experience a vertical gradient in light levels and water availability from their seedling stage in the shade understory to the bright canopy. Therefore, I expect tall species to show leaf traits related to high light levels and higher ability to adjust leaf traits to different environmental conditions, reflected by higher intraspecific variation in leaf traits measured across various microhabitats. From 23 late-successional tree species evaluated, those with high intraspecific variability in leaf mass per unit leaf area (specific leaf mass; SLM) showed higher survival and growth rates across different microhabitats of early successional environments than species with low variability in SLM, while for thwelve latesuccessional tree species evaluated, those species with low mean SLM measured in sun (pastures) or shade (secondary forest) leaves showed higher growth rates in height when growing in abandoned pastures with long-lasting high light conditions. Maximal tree
xvi SUMMARY (continued) height was related to variation in SLM but it was not correlated to survival and growth rates in early successional ecosystems and neither in pastures. Leaf traits offer easily measurable variables that may provide criteria for selection of species for planting mixed species stands. This alleviates the need to individually screen large numbers of latesuccessional species for performance in different scenarios of restoration (i.e. early successional ecosystems, pastures).
1. GENERAL INTRODUCTION
1.1
Introduction It is widely recognized that deforestation of tropical habitats is occurring at a
rapid and in many places accelerating pace (Bawa and Dayanandan, 1997); what is less appreciated is that tens of thousands of hectares of denuded land are abandoned annually to succession back to forest (Moran et al., 1994). The normal process of secondary succession in neotropical habitats involves initial colonization by early successional fastgrowing species, which occupy sites for up to 15 years or more before a mixture of them and a species-poor cohort of long-lived tree species is established for up to several decades (e.g., Denslow, 1985; Uhl et al., 1988; Guariguata et al., 1997). This early to mid-successional forest may not resemble the original forest for a long time, if ever. Restoration efforts frequently aim to recover ecosystem function as soon as possible; therefore, usually exotic or native fast growing species are used when planting is necessary. If exotic fast growing species are used, they may invade other natural areas and displace native species (Fine, 2002). Native fast growing species, “pioneers”, represent only 20 % of the diversity in the tropical forest (Gómez-Popma and VázquezYanes, 1981). This and both, high initial mortality and establishment of low-diversity stands for decades reduce the desirability of reforestation with pioneers alone. The objective of this thesis is to presents new criteria to choose late-successional tropical tree species to restore biodiversity in early successional environments, based
1
2 on vegetative traits that predict species survival and growth rates. The restoration of biodiversity and the subsequent recovery of many plant-animal interactions has been one overlooked objectives of restoration. In the next section I present some of the general objectives of restoration:
1.2
Restoration objectives Restoration projects aim to start or accelerate successional process to achieve
ecological fidelity by combining i) structural replication, ii) restoration of ecosystem function, and iii) durability in time (Higgs, 1997; Lamb, 1998). From this it follows that the most simple restoration effort is based on the removal of the perturbation (i.e. cattle, agriculture, logging) and the hope that natural succession progresses to create a forest similar to the original one. Many barriers to natural regeneration have been identified and procedures to overcome them have been suggested (Nepstad et al., 1990; Holl, 1999; Ashton et al., 2001). Among these barriers, dispersal limitation of deep forest species in early successional environments is a key phenomenon that deserves more attention. A general consequence of dispersal limitation is that secondary forest lack deep forest species for many years. Here, “letting nature takes its course” encourages continuing random species loss from primary forest remnants during what could be recovery of floristic and faunistic diversity in and near forest fragments (see Janzen, 1988). A second general consequence is that this usual process of secondary succession precludes recovery of biotic diversity representative of either forests prior to deforestation, or of remaining fragments, for decades.
3 Managed succession may have a variety of objectives (Parker, 1997; also Howe 1994, 1999). Restoration of gross ecosystem functions, such as material retention and loss, energy inputs and outputs, and the outlines of food-web structure, may require only a small subset of the 30 to 100 tree species/ ha representative of forest structure (see Ehrenfeld and Toth, 1997; Palmer et al., 1997), but establishing methods for successfully accelerating the recovery of diversity of mid-and late-successional trees will be necessary for slowing the loss of diversity from remnants and restoring floristic and faunistic interactions and dependencies. How should this be done, with passive secondary succession, somewhat facilitated natural succession, consciously enriched succession, or manipulated succession under exotic monocultures? These issues are reviewed in Chapter 2. Restoration efforts benefit from the knowledge of the natural dynamics in mature forest, to study this dynamic, tropical tree species have been classified in functional groups. This is reviewed in the following section.
1.3
Conceptual Framework Rain forest tree species are classified into two big groups pioneer and non-
pioneers using as criterion demography as well as the requirements of seed and seedling (Martínez-Ramos, 1985; Swaine and Whitmore, 1988). Early successional or pioneer species have small seeds, long-distance dispersal capacity, long persistence in the seed bank, and ability to colonize gaps within the forest. These canopy gaps receive higher light levels than the understory, have lower humidity at the soil surface, and in some cases a temporal release of nutrients (Bazzaz and Wayne 1994). Pioneer species are known to respond immediately to environmental factors of canopy gaps that give them an
4 advantage over non-pioneer species in this particular habitat. These species have short life spans (Martínez-Ramos, 1985), often senescing or dying in 15-30 years. Non-pioneer species represent the remaining 80 % of tropical tree species (Martínez-Ramos, 1985). Non-pioneers may germinate and establish inside the forest in shade (Swaine and Whitmore, 1988). For this study, non-pioneer species and long-lived species are considered as the same group of late-successional species, although there are many exceptions in more species-rich tropical forests (Martínez-Ramos, 1985; Foster et al., 1986). The late-successional group may be further subdivided on the basis of maximal height of adults as emergent and canopy (tall species) and understory categories (short species). Certain species may be exposed to less environmental heterogeneity during most of their life cycle. For example pioneer species may germinate, establish and reach the canopy always experiencing high light level while short statured tree species may germinate and inhabite always the dark forest understory. On the other hand, emergent and canopy late-sucessional species depend on eventual opening of gaps to reach the canopy, therefore they are exposed to higher environmental heterogeneity during their lives. The environmental conditions that species experience for the most part of their lives is related to morphological and functional plant characteristics, for example, leaf traits, which will be reviewed in the next section.
1.3.1. Leaf traits of tropical tree species Late-successional tree species show substantial phenotypic plasticity (sensu Bradshaw, 1965): they can function in a range of environments. As adults, they have sun leaves at the top of their crown and shade leaves at the bottom of the crown. Shade
5 leaves have lower specific leaf mass (SLM; leaf dry weight/ leaf area, g m -2), lower leaf water content (WC: fresh leaf weight-dry leaf weight/ leaf area, g m-2) and thinner leaves, and higher stomatal density in comparison with sun leaves within an individual crown (Popma et al., 1992; Oberbauer and Strain, 1986, Pearcy and Sims, 1994). These traits increase whole-plant productivity by maximizing photosynthesis and minimizing loss of water (Björkman, 1981; Bongers and Popma, 1988; Popma et al., 1992). Reduction in leaf area and increase in leaf thickness in sun conditions alleviates heat load and decreases loss of water (Chiariello,1984). On the other hand, shade leaves have larger leaves for light capture and maximization of sunfleck use. How light heterogeneity, that different species experiences through their ontogeny, affects their leaf traits is shown in Chapter 3 for eight late-successional species with a maximal tree height ranging from 8 to 40 m. Leaf area represents the available photosynthetic area for light harvesting events. A minimum total leaf area on a per plant basis should be maintained at any set of conditions to ensure survival (Kohyama, 1987). This may be accomplished through leaf production and leaf life span (King, 1994). Short leaf life span is considered an adaptation for rapid growth rate and drought-avoidance (Reich et al., 1991), generally attributed to pioneer species (Coley, 1988). Under shade conditions, pioneer species are unable to maintain their leaf area, due to short life span and lack of resources for high leaf production. In contrast, at similar light levels, emergent and canopy species have long leaf life span and lower leaf production (King, 1994). In Panama, survival of two herb species with different dependence on gaps was positively related to intrinsic leaf longevity (Mulkey et al., 1991). For 142 tropical tree species at different DBH classes
6 survival was negatively correlated with growth rates in the forest (Condit et al., 1995). It appears that the ability to function in a range of environments might be at least as well predicted by functional and demographic features than by physiological traits alone (Mulkey et al., 1991), particularly in field situations where physiological work is difficult. Determinations of which traits predict species survival and growth in restoration planting have been done for grassland (Pywell et al., 2003), whilet this essential information is not available for tropical forest. Chapter 4 shows a study of the variation in leaf traits that predict survival and growth rates of 23 late-successional species growing in the different microhabitats of early successional ecosystems. In this plantation the natural arrival of pioneer species lead to rapid changes in microenvironmental conditions as it may happen in restoration projects that are located close to sources of pioneer seeds. However, not all restoration projects will be carried out in such environments. Chapter 5 reports results from the first 17 months of 12 mid-successional and late- successional species growing in a plantation that mimics the environmental conditions of a pasture without natural arrival of pioneer species. In this plantation leaf traits that predict performance were also evaluated. Restoration projects aim to recover former forests, including the high biodiversity typical of mature forests; consequently the planting of late-successional species is necessary. However the lack of information about species establishment requirements makes it easier to use few species with high growth rates and survival in early successional environments. Determination of which trait(s) best predict(s) latesuccessional species survival in different scenarios of restoration (i.e. early successional
7 ecosystems, pastures) will obviate the need to test all tropical tree species (about 300 species at the Los Tuxtlas) individually to decide which of them to plant. Maximizing survival per planted seedling reduces the cost of nursery maintenance and replanting in restoration projects. By resolving which traits are better predictors of species growth rates in adverse conditions of tropical pastures, we will minimize the time needed to regenerate a forest of high structural and species diversity.
1.4
Literature Cited
Ashton, M. S., C. V. S. Gunatilleke, B. M. P. Singhakumara and I. Gunatilleke. 2001. Restoration pathways for rain forest in southwest Sri Lanka: a review of concepts and models. Forest Ecology and Management, 154, 409-430 Bawa, K. S. and Dayanandan, S. 1997. Socioeconomic factors and tropical deforestation. Nature, 386, 562-563. Bazzaz, F. A. and Wayne, P. M. 1994. Coping with environmental heterogeneity: The physiological Ecology of tree seedling regeneration across the gap-understory continuum. Exploitation of environmental heterogeneity by plants (eds. M. M. Caldwell and R. W. Pearcy), 349-390. Academic Press, USA. Björkman, O. and Demmig-Adams, B. 1994. Regulation of Photosynthetic Light Energy Capture, Conversion, and Dissipation in Leaves of Higher Plants. Ecophysiology of Photosynthesis (eds. E.-D. Schulze and M. M. Caldwell), 17-47. Spring-Verlag, New York, USA. Bongers, F. and J.Popma. 1988. Is exposure-related variation in leaf characteristic of tropical rain forest adaptive? Plant form and vegetation structure (eds. M. J. A. Werger, P. J. M. van der Aart, H. J. During and J. T. A. Verhoeven), 191-200. Academic Publishing, The Netherlands. Bradshaw, A. D. 1965. Evolutionary significance of phenotypic plasticity in plants. Advances in Genetics (eds. E. W. Caspari and J.M. Thoday), 115-155. Academic Press, New York, USA. Chiariello, N. 1984. Leaf energy balance in the wet lowland tropics. Physiological ecology of plants of the wet tropics (eds. E. Medina, H. A. Mooney and C. Vázquez-Yanes), 85-98. Dr. Junk Publishers, The Netherlands.
8 Coley, P. D. 1988. Herbivory and defensive characteristics of tree species in a lowland tropical forest. Ecological Monographs, 53, 209-229. Condit, R., Hubbell, S. P. and Foster, R. B. 1995. Mortality-Rates of 205 Neotropical Tree and Shrub Species and the Impact of a Severe Drought. Ecological Monographs, 65, 419-439. Denslow, J. S. 1985. Disturbance-mediated coexistence of species. The Ecology of Natural Disturbance and patch dynamics (eds. S. T. A. Pickett and P. S. White), 307-323. Academic Press, Inc, USA. Ehrenfeld, J. G. and Toth, L. A. 1997. Restoration ecology and the ecosystem perspective. Restoration Ecology, 5, 307-317. Fine, P. V. A. 2002. The invasibility of tropical forests by exotic plants. Journal of Tropical Ecology, 18, 687-705. Foster, R. B., J. Arce B., and T. Wachter.1986. Dispersal and the sequential plant communities in Amazonian Peru floodplain. Frugivores and Seed Dispersal (eds. A. Estrada and T.H. Fleming), 357-370. W. Junk Publishers, Dordrecht, Netherlands. Gómez-Pompa, A. and C. Vázquez-Yanes. 1981. Sucessional Studies of a rain forest in México. Forest Sucession Concepts and Application (eds. D. C. West, H. H. Shugart and D. B. Butkin), 246-266. Springer-Verlag, New York, USA. Guariguata, M. R., Chazdon, R. L., Denslow, J. S., Dupuy, J. M. and Anderson, L. 1997. Structure and floristics of secondary and old-growth forest stands in lowland Costa Rica. Vegetation, 132, 107-120. Higgs, E. S. 1997. What is good ecological restoration? Conservation Biology, 11, 338 -348. Holl, K. D. 1999. Factors limiting tropical rain forest regeneration in abandoned pasture: Seed rain, seed germination, microclimate, and soil. Biotropica, 31, 229-242. Howe, H. F. 1994. Managing Species-Diversity in Tallgrass Prairie - Assumptions and Implications. Conservation Biology, 8, 691-704. Howe, H. F. 1999. Dominance, Diversity and Grazing in Tallgrass Restoration. Ecological Restoration, 17, 59-66. Itoh, A., Yamakura, T., Ogino, K., lee, H. S. and Ashton, P. S. 1997. Spatial distribution patterns of two predominant emergent trees in a tropical rainforest in Sarawak, Malaysia. Vegetation, 132, 121-136.
9
Janzen, D. H. 1988. Tropical Ecological and Biocultural Restoration. Science, 239, 243244. King, D. A. 1994. Influence of light on the growth and morphology of saplings in a Panamanian forest. American Journal of Botany, 81, 948-957. Kohyama, T. 1987. Significance of architecture and allometry in saplings. Functional Ecology, 1, 399-404. Lamb, D. 1998. Large-scale ecological restoration of degraded tropical forest lands: The potential role of timber plantations. Restoration Ecology, 6, 271-279. Lovelock, C. E., Jebb, M. and Osmond, C. B. 1994. Photoinhibition and Recovery in Tropical Plant-Species - Response to Disturbance. Oecologia, 97, 297-307. Lugo, A. E. 1997. The apparent paradox of reestablishing species richness on degraded lands with tree monocultures. Forest Ecology and Management, 99, 9-19. Martínez-Ramos, M. 1985. Claros, ciclos vitales de los arboles tropicales y regeneración natural de las selvas altas perennifolias. Investigaciones sobre la Regeneración de selvas altas en Veracruz, México. (eds. A. Gómez-Pompa and S. Del Amo), 191-240. Alhambra Mexicana, Mexico. Martínez-Ramos, M. 1994. Regeneración natural y diversidad de especies arboreas en selvas humedas. Boletin de la Sociedad Botánica de México, 54, 179-224. Moran, E. F., Brondizio, E., Mausel, P. and Wu, Y. 1994. Integrating Amazonian Vegetation, Land-Use, and Satellite Data. Bioscience, 44, 329-338. Mulkey, S. S., Smith, A. P. and Wright, S. J. 1991. Comparative Life-History and Physiology of 2 Understory Neotropical Herbs. Oecologia, 88, 263-273. Nepstad, D., Ch.Uhl and A.E.Serrao. 1990. Surmounting barriers to forest regeneration in abandoned highly degraded pastures: a case study from Paragominas, Pará, Brazil. Alternatives to deforestation: steps toward sustainable use of the amazon rain forest (ed. A. B. Anderson), 215-229. Columbia University Press, New York, USA. Oberbauer, S. F. and Strain, B. R. 1986. Effect of canopy position and irradiance on the leaf physiology and morphology on Pentaclethra macroloba (Mimosaceae). American Journal of Botany, 73, 409-416. Palmer, M. A., Ambrose, R. F. and Poff, N. L. 1997. Ecological theory and community restoration ecology. Restoration Ecology, 5, 291-300.
10 Parker, V. T. 1997. The scale of successional models and restoration objectives. Restoration Ecology, 5, 301-306. Pearcy, R. W. and Sims, D. A. 1994. Photosynthetic Acclimation to Changing Light Environments: Scaling from the Leaf to the Whole Plant. Exploitations of Environmental Heterogeneity by Plants. Ecophysiological Processes Above-and Belowground (eds. M. M. Caldwell and Pearcy, R.W.), 145-173. Academic Press: Harcourt Brace & Company, USA. Poorter, L. 1999. Growth responses of 15 rain-forest tree species to a light gradient: the relative importance of morphological and physiological traits. Functional Ecology, 13, 396-410. Popma, J. and Bongers, F. 1991. Acclimation of seedlings of three Mexican tropical rain forest tree species to a change in light availability. Journal of Tropical Ecology, 7, 85-97. Popma, J., Bongers, F. and Werger, J. A. 1992. Gap-dependence and leaf characteristics of trees in a tropical lowloand rain forest in México. Oikos, 63, 207-214. Pywell, R. F., J.M. Bullock, D.B. Roy, LIZ Warman, K.J. Walker and P. Rothery. 2003. Plant traits as predictors of performance in ecological restoration. Journal of Applied Ecology, 40, 65-77. Reich, P. B., Uhl, C., Walters, M. B. and Ellsworth, D. S. 1991. Leaf Life-Span as a Determinant of Leaf Structure and Function among 23 Amazonian Tree Species. Oecologia, 86, 16-24. Ricker, M. 1998. Enriching the tropical rain forest with native fruit trees: a biological and economic analysis in Los Tuxtlas (Veracruz, Mexico). PhD. Dissertation. 262 pp. Yale University, USA. Swaine, M. D. and Whitmore, T. C. 1988. On the Definition of Ecological Species Groups in Tropical Rain Forests. Vegetatio, 75, 81-86. Uhl, C., Buschbacher, R. and Serrao, E. A. S. 1988. Abandoned Pastures in Eastern Amazonia .1. Patterns of Plant Succession. Journal of Ecology, 76, 663-681. Welden, C. W., Hewett, S. W., Hubbell, S. P. and Foster, R. B. 1991. Sapling Survival, Growth, and Recruitment - Relationship to Canopy Height in a Neotropical Forest. Ecology, 72, 35-50.
2. RESTORING TROPICAL DIVERSITY: BEATING THE TIME TAX ON SPECIES LOSS
2. 1
Abstract Fragmentation of tropical forest is accelerating, at the same time that already
cleared land reverts to second growth. Fragments inexorably lose deep forest species to local extinction while embedded in low-diversity stands of early successional pioneer trees. Pasture matrices undergoing passive secondary succession become a “pioneer desert” from the vantage of remnant immigration, imposing a “time tax” of loss of deepforest plants from forest fragments. If seeds of deep-forest trees find pastures, or seedlings are planted there, many prosper. Bypassing early domination of pioneer trees in regenerating matrices, or enriching matrices with animal-dispersed forest trees, may stem loss of species from forest fragments and accelerate succession far from edges of old forest. Planting disperser-limited trees that establish in open ground may bypass 3070 years of species attrition in isolated remnants by attracting animals that encourage normal processes of seed dispersal into and out of the fragments. Development of criteria for selection of persistent, reasonably rapidly growing animal-dispersed species that are mixed with planted or naturally-arriving pioneers will be an important component of enrichment planting.
11
12 2. 2
Introduction In vast areas of the tropics, forest has been displaced by crops and commercial
cattle grazing, leaving land devoid of natural vegetation and the soil seed banks, seedling cohorts, and suppressed saplings of mature forest trees that might restore it (QuintanaAscencio et al., 1996; Miller, 1999). Often forest nuclei remain, but as fragments of a few hectares that inexorably lose species to local extinction (Turner, 1996). Where intensive crop production or cattle ranching is unprofitable, restoration of biological diversity in regenerating matrices between fragments has the potential to stem the loss of species from relict forests, and may actually facilitate restoration of natural metapopulation dynamics. How should this be done, with passive secondary succession, somewhat facilitated natural succession, consciously enriched succession, or manipulated succession under exotic monocultures? I argue that in restoration of biological diversity between forest remnants, the vegetation matrix matters. The matrix influences rates of accumulation of organic matter and nutrients (Lugo, 1997), and the population and community dynamics of species in the remnants which the matrix surrounds (Vandermeer and Carvajal, 2001). Almost any vegetation adds organic matter and nutrients, and retains water, to a greater extend than barren land. However, loss of species from remnants may not be stemmed if land is quickly occupied by one or a few early successional species, followed by what I call the “pioneer desert” of early and late pioneers that retard influx of disperser-limited deepforest trees for a century or more (Finegan, 1996). If intensive land use imposes a “time tax” of soil degradation (Lugo, 1988), the 100+ year pioneer matrix imposes another
13 “time tax” of species loss from remnants, a lost opportunity cost which may be irredeemable. Loss of mutualists and herbivores may accelerate plant extinctions in forest fragments. A disproportioante share of tropical trees require insects, bats or birds for pollination, and bats, birds, or terrestrial or arboreal mammals for seed dissemination (Howe and Westley, 1997), and the diversity of recruiting cohorts of seedlings is maintained by thinning by ground-foraging mammals (Dirzo and Miranda, 1991). An immense array of ecological disruptions occurs if herbivores, predators, dispersal agents, and pollinators lack access to forest patches for long intervals (Cordeiro and Howe, 2001; Terborgh et al., 2001; Wright and Duber, 2001; Tewksbury et al., 2002). A matrix that accelerates the influx of frugivores and herbivores is an overlooked but needed component of forest regeneration. I argue that loss of tree species continues unabated from forest remnants embedded in a matrix of pioneer species, imposing a time tax of local extinction that will impede future restoration efforts. A remedy is the creation of matrices enriched with large-seeded, animal-dispersed trees which attract dispersal agents that accelerate succession to mature forest, mediate immigration of forest species back into remnant fragments, and augment emigration of trees of relict forests into regenerating matrices.
2. 3
Natural succession In rainforests, natural disturbances vary in scale, intensity, and frequency, and
trigger a variety of successional processes. Most disturbances are small branchfalls and treefalls that form canopy gaps of 50-1000 m2 and close within a few years (e.g. Fig. 1,
14 Howe, 1990), usually filled by deep-forest species already present as seeds, seedlings, saplings, re-sprouts, or overhanging canopy trees (Brokaw and Scheiner, 1989). Even extensive hurricane blow-downs yield to tree-species already present as suppressed plants in the understorey or stumps capable of re-sprouting after damage (Boucher et al., 2001). Most natural regeneration in rainforests free from human disturbance comes from established plants, the species composition of which roughly reflects the pre-disturbance community. Nature can favour immense stands of early pioneer trees, but the circumstances are limited and rare. Massive mudslides following earthquakes or volcanism are closed by small-seeded pioneer species with high dispersal capacity (e.g. Thebaud and Strasberg, 1997). For instance, Garwood and colleagues (1979) reported that an earthquake in eastern Panama sloughed an entire watershed into the sea. Much like a large anthropogenic disturbance, the exposed earth was quickly choked with saplings of invasive bird-dispersed Trema (Ulmaceae), a common early successional tree that produced fruits within four years. Large natural disturbances that remove vegetation over several to hundreds of hectares involve quite different processes of succession than simple closure of canopy gaps, or even re-vegetation after damage by wind or fire that leaves many seeds, seedlings, and older plants alive and ready to respond to increased light and soil resources.
2.4
Arriving and surviving in unnatural landscapes Landscape features influence the arrival and survival of tree species after major
disturbances (e.g. Tewksbury et al., 2002). In the tropics, abandoned croplands and
15 pastures do not have seed banks of forest species, nor are there seedling or sapling cohorts waiting to respond to opportunity. Low rates of colonization and emergence, and high mortality of seeds and seedlings, result in low densities and diversities of tree juveniles and saplings (Uhl et al., 1988; Nepstad et al., 1996; Zimmerman et al., 2000). Under such conditions, dispersal becomes a limiting factor in the rate of forest regeneration (Quintana-Ascencio et al., 1996; Holl, 1999; Miller, 1999), and the size and distribution of remnant sources of forest seeds becomes an important determinant of seed dispersal and seedling recruitment. Even when small pastures experience substantial seed rain from late-successional trees in nearby mature forest, they always receive different seed input than small gaps in within the mature rainforest (see Purata, 1986; MartínezGarza and Montagut, 1999; 2002). It is not realistic to expect early large-scale colonization of open land by large-seeded bird, bat, or primate-dispersed tree species even tens of meters, much less hundreds of meters, from a forest edge (e.g. see Hooper et al., 2002; Ingle, 2003). Unsuccessful intensive farming and ranching have made such large tracts of thoroughly disturbed land commonplace, and have coincidently made the dispersal limitation that is characteristic of such situations the rule rather than the exception throughout much of the Neo- and Paleotropics. Only part of the problem is what can establish and grow under the canopy of such pioneers; much of the problem is what can get there. Examples from the Neotropics illustrate the issues. A few early pioneers such as Cecropia (Moraceae [Cecropiaceae]), Croton (Euphorbiaceae), Trema (Ulmaceae), and Vismia (Guttiferae [Clusiaceae]) hold ground for 20 - 30 years, to be replaced by such longer-lived late pioneers as Cordia (Boraginaceae), Goupia (Goupiaceae
16 [Celastraceae]), Guazuma (Sterculiaceae), Laetia (Flacourtiaceae), Rollinia (Annonaceae), Spondias (Anacardiaceae), Stryphnodendron (Fabaceae) and Vochysia (Vochysiaceae) (Uhl and Jordan, 1984; Finegan, 1996; Mesquita et al., 2001). In a mix of successional forests, species richness and diversity may approach that of mature forest in 100 years, but species composition does not resemble mature forest; dominants of primary forest, if present, are rare (Finegan, 1996; Aide et al., 2000). A century of revegetation after intensive grazing or farming produces forests that restore most ecosystem function, but the new forests do not have - for many years, if ever - the same species of trees that were present before the cow or plough. Such a “pioneer desert” does not supply stands of remnant mature forest with propagules of deep forest species, nor provide genetic diversity from pollen for isolated relict trees, for at least several decades. Ecosystem function may return, but much of the biological function of a mature forest does not. A useful irony is that many disperser-limited deep forest species do well in open pastures, if they get there. In Panama, Hooper and colleagues (2002) find > 70% survival of large-seeded, shade-tolerant tree seedlings in pastures, where per-capita survival of seedlings of pioneers is < 0.1%. In Costa Rica, seedlings of large-seeded Octoea glaucosericea, O. whitei (Lauraceae), and Calophyllum brasiliense Cambess. (Guttiferae [Clusiaceae]) have greater than 50% survival when planted in pastures (Loik and Holl, 1999), while this Calophyllum and deep-forest trees Virola koshnyi Warb. (Myristicaceae), Dipteryx panamensis (Pittier) Record & Mell [D. oleifera Benth], Pithecellobium elegans ([Albizia elegans]; Fabaceae), and Genipa americana L. (Rubiaceae) experience 60-98 % survival when planted as seedlings in other abandoned
17 pastures (Montagnini et al., 1995). Evidence is mounting that many deep-forest species survive in pastures. Similar success occurs in the Old World. Hardwick and her colleagues (1997) find that in abandoned agricultural land in northern Thailand, seedlings of animal-dispersed Beilschmiedia spp. (Lauraceae) with large seeds (6-7 g) and Prunus cerasoides D. Don (Rosaceae) with seeds of moderate size (0.23 g) have survival in weedy pastures of 93% and 70%, respectively. Pioneers colonize pastures, but only because small (< 10 mg) seeds are present in immense numbers. Far from a forest edge, even pioneer arrival may be very slow. However, if deep-forest trees with large seeds find pastures, or are planted there, seedlings of many species prosper.
2.5
Attrition of species from fragments in alien matrices Forest fragments excised from continuous forest and imbedded in alien matrices
lose many species of animals and plants over 20 - 100 years (Turner, 1996). Proximity of fragments to sources of dispersal agents influence which species are vulnerable to local extinction in remnants. The landscape of sources, and the quality of surrounding matrices, will likely determine which species are most vulnerable in a given area. Very low densities of most species make tropical forest fragments especially vulnerable to high rates of local extinction (e.g. Maina and Howe, 2000). Species number is positively associated with size of habitat patches, while population densities of all but the dominants are inversely correlated with richness (Preston, 1948; MacArthur, 1972). Recent studies show that seedlings and juveniles of many rainforest species suffer declines in fragments (Cordeiro and Howe, 2001; Githiru et al., 2002), and some island
18 communities show dramatic losses through older juvenile and sapling stages (see Leigh et al., 1993). With adult tropical trees dying at a rate of roughly 1% per year in mature forest (Brokaw, 1985; Sukumar et al., 1998), the less representative the matrix is of mature forest, the more rapidly small populations of dispersal agents and trees with reduced recruitment will be lost from remnants. Maina and Howe (2000) argue that tree species most vulnerable to local extinction from small forest remnants are those at the moderately abundant to rare end of the species abundance distribution that lose pollinators or dispersal agents in small fragments, while those most likely to persist are highly vagile weeds, augmented by successional pioneer deserts, and “always rare” species that function as successful metapopulations. Records from old fragments in East Africa have borne out these expectations (Cordeiro and Howe, 2001). In the East Usambara Mountains of Tanzania, a rainforest of largely animal-dispersed trees (23% endemic) was fragmented into 30 ha) than small (< 10 ha) forest fragments, whereas recruitment of eight wind- and gravity-dispersed trees of the forest interior was unaffected. Recruitment of 10 endemic, animal-dispersed tree species was 40 times lower in small fragments than in continuous forest or large fragments. Even in this system, with far fewer species than many forests of South America or East Asia, many tree species were represented by 1-5
19 adult individuals per fragment, far fewer than necessary to maintain viable populations. In such situations, restoration of the matrix is a race.
2.6
Restoring the matrix: beating the time tax on diversity Restoring matrix diversity encourages natural processes of immigration and
integration among nuclei of forest remnants. One measure is planting buffers, corridors, and stepping stone stands around and between remnants (Janzen, 1988; Lamb et al., 1997; Tewksbury et al., 2002). Another measure after release from intensive agriculture is to upgrade or eliminate the 100 year pioneer desert by encouraging late-successional trees long before they would passively arrive of their own accord. This in turn should attract vertebrate dispersal agents that accelerate the process of seed dispersal into and out of forest fragments. A question is, are some means of encouragement more effective than others? In general, in forest regeneration of large areas of abandoned agricultural land or pasture, passive succession cannot stem the loss of species from forest remnants. Isolated trees that attract dispersal agents or artificially positioned perches that do the same increase seed rain of animal-dispersed species (Guevara et al., 1991; Miriti, 1998), but recruitment is slow and remains unrepresentative of mature forest (Holl, 1999). I doubt that passive succession, or succession encouraged with perches, will effectively counter the time tax of loss from remnants, except for edges close to mature forest. Encouraging disperser activity in this minimal way may be more appropriate for buffering edges of remnants (e.g., Janzen, 1988) than for bypassing the pioneer desert.
20 I suggest that beating the time tax on biodiversity is possible if successions of pioneers are actively enriched with plantings of late-successional and deep forest animaldispersed tree species. Pioneers arrive of their own accord close to forest edges, or may be planted far from edges in the largest deforestations. In some cases valuable timber species may be mixed in plantings for later commercial harvest, a conservation measure that as has been attempted in Australia (Lamb and Lawrence, 1993). I suggest that criteria should include a mix capable of growing in xeric conditions of open ground, capable of growing under exotic grasses, capable of establishing under the inevitable pioneer canopy, and when mature capable of attracting animate dispersal agents that accelerate the process. Tucker and Murphy (1997) demonstrate enhanced regeneration of animal-dispersed species in enriched successions in tropical Queensland, where dispersal agents and the trees that they use are far more common with plantings of latesuccessional trees than in control plots of the pioneer matrix. Criteria for selection of those late-successional species most likely to accelerate succession to complex forests should become an important facet of enrichment plantings. At some sites, exotic plantation monocultures may admit deep forest species native to surrounding forests. Lugo (1997) makes the interesting point that plantation monocultures of some short-lived exotics cast enough shade to suppress pioneers, but admit invasion by deep-forest species at least as well as pioneers and better than other plantation crops. This variation on passive succession would permit a cash crop, and cost little. Whether such a method would create unintended consequences, be a superior strategy to enrichment of early and late pioneer stands, or simply be the only option in regions used for plantation crops, should be a matter of debate.
21 The evidence indicates that the feasibility of using plantations of exotic monocultures to promote late-successional indigenous species depends on the species used and the location. In Brasil, Leucaena leucocephala (Lam.) De Wit (Fabaceae) plantations admitted more native forest species than Casuarina plantations (Parrotta 1995). In Hawaii, plantations of Eucalyptus saligna Sm. (Myrtaceae) and Flindersia brayleyana F. V. Muell. (Rutaceae) from Australia and Fraxinus uhdei (Wenzig) Lingelsh (Oleaceae) from Mexico established in the 1950s and 1960s fostered very different regeneration pathways (Harrington and Ewel, 1997). E. saligna strongly favored exotics and F. braylenana replaced itself, while F. uhdei favoured two dominants of surrounding mature forest, Cibotium glaucum (Sm.) Hook. and Arnott (Dicksoniaceae) and Metrosideros polymorpha Gaud. (Myrtaceae). Various pines and eucalypts in South Africa admitted a few common animal- and wind-dispersed forest canopy species, which appear to represent late pioneer species (see Geldenhuys, 1997). In the Congo, eucalypt plantations admit a number of native forest species, especially close (< 50 m) to the forest edge, with strong representation by wind-dispersed species that show especially rapid regrowth after clear-cutting of the plantation crop (Loumeto and Huttel, 1997). With some exceptions, such as Leucanea in Brazil and Fraxinus in Hawaii, forests developing under plantation monocultures resemble pioneer desert produced by passive regeneration of unnaturally extensive abandoned agricultural land, or admit even fewer species than natural successions if exotics replace themselves. Another option might be the planting of native monocultures or mixed stands (Guariguata et al., 1995; Haggar et al., 1997; Murcia, 1997). Monocultures or mixed stands of native species admit the invasion of other native species, but success also
22 depends on which species are planted and the objectives of restoration. For instance in Costa Rica, establishment of forest trees is favored under monocultures of Vochysia guatemalensis (Donn.Sm.) Standl. (Vochysiaseae), but plant growth is slower than under Jacaranda copaia (Aubl.) D. Donn. (Bignoniaceae) monocultures (Guariguata et al., 1995). In the Colombian Andes, recruitment of forest species is lower in plantations of Alnus acuminata (Kunth) Kuntze (Betulaceae) than in the natural regenerated forest (Murcia, 1997). As in natural successions, few of the species that recruit under the plantations belong to the mature forest (Murcia, 1997; Haggar et al., 1997). These planted overstorey trees are wind-dispersed, which would not draw dispersal agents as effectively as animal-dispersed species In summary, enrichment of successional plots with seedlings of a variety of animal-dispersed tree species would appear to be the most feasible means of beating the time tax on species loss from remnants embedded in secondary successions (see Tucker and Murphy, 1997). Variations on passive succession, unattended or under overstories of exotic or native trees, may offer a means of restoring ecosystem function, but they do not avoid the equivalent of the pioneer desert: forest remnants remain surrounded by low-diversity matrix. Adding perches may or may not speed succession near forest edges, and planting animal-dispersed monocultures might accelerate attraction of birds and mammals that eat fruits and disperse other seeds in the process. The risk in any variation of passive succession is embedding remnant forests in alien matrices for unnecessarily long times. In our view, useful factors to consider as criteria for enrichment plantings of late successional trees include both attributes of growth and survival of the tree species, and their role in dispersal dynamics.
23 Not all deep-forest species can survive in xeric pasture conditions, or prosper under thick cover of exotic grasses or exotic or native weeds. Some deep forest species are more likely to survive in the open than others, and some species grow more rapidly than others (e.g., Hooper et al., 2002). High mortality and very slow growth could admit weedy herbaceous grasses and forbs, thereby setting back succession. Pioneer trees will arrive of their own accord near forest edges (e.g., Martínez-Garza and Montagut, 1999; 2002; Ingle, 2002), but animal-dispersed pioneers such as Cecropia, Cordia, Ficus, and Trema could be gainfully planted far (> 100 m) from source forests edges to secure soil, provide shade, and initiate the process of frugivore assembly. Maury-Lechon (1993) has found that growth plasticity of juvenile stages of trees is a predictor of growth rate in secondary succession. I found that variation (coefficient of variation) in an easilymeasured leaf character, specific leaf weight, is a reasonable predictor of growth rate and survival of deep forest trees in one five-year-old planting in southern Mexico (see Chapter 4). Given a pool of dozens to hundreds of moderate- to large-seeded species in the vicinity of almost any tropical forest, the high survival rate of those mature forest species that have been planted in open pastures, and the evolving understanding of easily detected attributes that are correlated with high growth rates and persistence, I am confident that plant functioning will not limit enrichment. Besides capacity for survival, species used in enrichment plantings should also serve other purposes. Species chosen might represent rare species logged out of many remnants, and therefore serve a conservation function (e.g., Lamb and Lawrence, 1993). A more general tactic is to select those species that, once mature, will attract birds, primates, bats, or other fruit-eating animals that accelerate seed dispersal of late-
24 successional species into and out of forest remnants. Because the greater majority of tree species in many tropical forests bear fruits adapted for animal dispersal (Howe and Westley, 1997), a wide range of species appealing to generalist and specialist species is likely to be available, and otherwise suitable for planting as seedlings. With refined criteria for inclusion of components of the mature forest community, Tucker and Murphy’s (1997) approach could be the basis for a model for effective re-vegetation in most tropical regions of the world.
2.7
Synthesis and Applications Planting seedlings of interior forest species after land abandonment should sharply
accelerate the process of re-vegetation of complex communities. Two to 10 years may be lost in slower growth of some trees, but decades saved in total community recovery. Bypassing acute dispersal limitation with plantings that accelerate re-occupation of land far from remnants by herbivorous and fruit-eating animals may subtract 30-100 years of the time tax on diversity imposed by a low-diversity pioneer matrices. Effects of enrichment between forest remnants will be strongest where matrices are dominated by wind-dispersed trees, where plantation exotics self-seed easily, and in deforested areas far (> 100 m) from forest edges. Pioneer stands or plantation monocultures may be enriched with seeds or better with seedlings of late-successional animal-dispersed trees, or initial plantings could be mixes of late-successional and pioneer species. Whatever enrichment is used will almost certainly do more to beat the time tax on species loss from remnants than passive secondary succession or minimal encouragement with perches.
25 Active enrichment of successions or plantations with a variety of animal-dispersed species appears to be the best method of recruiting a variety of mature forest trees in and from remnants, with some care to selection of trees bearing fruits for consumption and dispersal by different animal guilds. Enrichment of plantations, when possible by animal-dispersed species, has the advantage of adjusting overstorey densities enough to suppress pioneers but not so much as to suppress mature-forest species planted underneath. A challenge will be to establish criteria for selecting candidates for enrichments and in some cases for overstories. The mix should include species that are able to grow in xeric open pastures in some cases, under shade of trees or grasses in others (e.g., Hooper et al., 2002). Species should be animal-dispersed to most effectively harness natural processes of dissemination into and out of remnant forests, and should represent a variety of fruit sizes and attractiveness to the birds, primates, bats, and other arboreal and terrestrial mammals that will mediate the process.
2.8
Acknowledgments I am grateful to N. Cordeiro, P. Fine, B. Finegan, H.F. Howe, N. Ingle, J. Klock,
G. Nuñez, S. Saha, A. Sullivan, B. Zorn-Arnold and one anonymous reviewer for comments on the manuscript, and to the Lincoln Park Zoo Neotropic Foundation, CONACyT of Mexico for support of research leading to these insights.
26 2.9
Literature Cited
Aide, T.M., Zimmerman, J.K., Pascarella, J.B., Rivera, L. and Marcano-Vega, H. 2000. Forest regeneration in a chronosequence of tropical abandoned pasture: Implications for restoration ecology. Restoration Ecology, 8, 328-338. Boucher, D. H., Vandermeer, J. H., Granzoa de la Cerda, I., Mallona, M. A., Perfecto, I. and Zamora, N. 2001. Post-agriculture versus post-hurricane succession in southeastern Nicaraguan rain forest. Plant Ecology, 156, 1131-1137. Brokaw, N. V. L.1985. Gap-phase regeneration in a tropical forest. Ecology, 66, 682-687. Brokaw, N. V. L. and Scheiner, S. M.1989. Species composition in gaps and structure of a tropical forest. Ecology, 70, 538-541. Cordeiro, N. and Howe, H. F. 2001. Low recruitment of trees dispersed by animals in African forest fragments. Conservation Biology, 15, 1733-1741. Dirzo, R. and Miranda, A.1991. Altered patterns of herbivory and diversity in the forest understory: a case study of the possible consequences of contemporary defaunation. Plant-animal Interactions: Evolutionary Ecology in Tropical and Temperate Regions (eds. P. Price, T. Lewingston, G. Fernandes, W. Benson), pp. 273-287. John Wiley and Sons, New York. Finegan, B.1996. Pattern and process in neotropical secondary rain forests: the first 100 years of succession. Trends in Ecology and Evolution, 11, 119-124. Garwood, N. C., Janos, D. and Brokaw, N.1979. Earthquake-caused landslides: a major disturbance to tropical forests. Science, 205, 997-999. Geldenhuys, C.J.1997. Native forest regeneration in pine and eucalypt plantations in Northern Province, South Africa. Forest Ecology and Management, 99, 101-115. Githiru, M., Bennun, L. and Lens, L. 2002. Regeneration patterns among bird-dispersed plants in a fragmented Afrotropical forest, south-east Kenya. Journal of Tropical Ecology, 18,143-149. Guariguata, M.R., Rheingans, R. and Montagnini, F. 1995. Early woody invasion under tree plantations in Costa Rica: Implications for forest restoration. Restoration Ecology, 3, 252-260. Guevara, S., Meave del Castillo, J., Moreno-Casasola, P. and Laborde, J.1991. Floristic composition and structure of vegetation under isolated trees in neotropical pastures. Journal of Vegetation Science, 3, 655-664.
27 Haggar, J., Wightman, K. and Fisher, R.1997. The potential of plantations to foster woody regeneration within a deforested landscape in lowland Costa Rica. Forest Ecology and Management, 99, 55-64. Harrigton, R.A. and Ewel, J.J.1997. Invasibility of tree plantations by native and nonnative indigenous plant species in Hawaii. Forest Ecology and Management, 99, 153-162 Hardwick, K., Healey, J., Elliott, S., Garwood, N. and Anusarnsunthorn, V.1997. Understanding and assisting natural regeneration processes in degraded seasonal evergreen forests in northern Thailand. Forest Ecology and Management, 99, 203-214. Holl, K. H.1999. Factors limiting tropical rain forest regeneration in abandoned pastures: seed rain, seed germination, microclimate and soil. Biotropica, 31, 229-242. Hooper, E., Condit, R. and Legendre, P.2002. Responses of 20 native tree species to reforestation strategies for abandoned pastures in Panama. Ecological Applications, 12, 1626-1641. Howe, H. F.1990. Seedling survival in a Central American tree (Virola surinamensis): Effects of herbivory and canopy closure. Journal of Tropical Ecology, 6, 259-280. Howe, H. F. and Westley, L. C.1997. Ecology of pollination and seed dispersal. Plant Ecology, Second Edition. (ed. M. J. Crawley), pp. 262-283. Blackwell Scientific Publications, London. Ingle, N. R. 2003. Seed dispersal by wind, birds and bats between Philippine montane rainforest and successional vegetation. Oecologia, 134, 251-261. Janzen, D. H.1988. Management of habitat fragments in a tropical dry forest: growth. Annals of the Missouri Botanic Garden, 75, 105-116. Lamb, D. and Lawrence, P.1993. Mixed species plantations using high value rainforest trees in Australia. Restoration of Tropical Forest Ecosystems (eds. H. Lieth and M. Lohmann), pp. 101-108. Kluwer Academic Publishers. Lamb, D., Parrotta, J., Keenan, R. and Tucker, N.1997. Rejoining habitat remnants restoring degraded rainforest land. Tropical Forest Remnants (eds. W.F. Laurance and R.O. Bierregaard, Jr.), pp. 366-385. University of Chicago Press. Leigh, E. G.Jr., Wright, S. J., Herre, E. A. and Putz, F. E.1993. The decline of tree diversity on newly isolated tropical island: a test of a null hypothesis and some implications. Evolutionary Ecology, 7, 76-102
28 Loik, M. E. and Holl, K. D. 1999. Photosynthetic responses to light for rainforest seedlings planted in abandoned pasture, Costa Rica. Restoration Ecology, 7, 382391. Loumeto, J, J. and Huttel, C.1997. Understory vegetation in fast-growing tree plantations on savanna soils in Congo. Forest Ecology and Management, 99, 65-81. Lugo, A. E.1988. The future of the forest: ecosystem rehabilitation in the tropics. Environment, 30, 17-20, 41-45. Lugo, A. E.1997. The apparent paradox of establishing species richness on degraded lands with monocultures. Forest Ecology and Management, 99, 9-19. MacArthur, R.1972. Geographical Ecology. Princeton University Press, Princeton, New Jersey. Maina, G. and Howe, H. F. 2000. Inherent rarity in community restoration. Conservation Biology, 14, 1335-1340. Martínez-Garza, C. and González-Montagut, R.1999. Seed rain from forest fragment into tropical pastures in Los Tuxtlas, Mexico. Plant Ecology, 145, 655-665. Martínez-Garza, C. and González-Montagut, R.2002. Seed rain of fleshy-fruited species in tropical pastures in Los Tuxtlas, Mexico. Journal of Tropical Ecology, 18, 457-462. Maury-Lechon, G.1993. Biological characters and plasticity of juvenile tree stages to restore degraded tropical forests: A systems framework for site analysis and restoration research. Restoration of Tropical Forest Ecosystems (eds. H. Lieth and M. Lohmann), pp. 37-46. Kluwer Academic Publishers. Mesquita, R. C. G., Ickes, K., Ganade, G. and Williamson, G. B. 2001. Alternative successional pathways in the Amazon Basin. Journal of Ecology, 89, 528-537. Miller, P. M.1999. Effects of deforestation in seed banks in a tropical deciduous forest of eastern Mexico. Journal of Tropical Ecology, 15,179-188. Miriti, M. N.1998. Regeneracao florestal em pastagens abandonadas na Amazonia Central: competicao, predação, e dispersao de sementes. Floresta Amazonica: Dinamica, Regeneracao e Manejo. (eds. Gascon, C., Nepstad, D. and Moutinho, P.), pp. 179-191. Instituto Nacional de Pesquisa da Amazonia, Manaus, Brazil. Montagnini, F., González, E., Porras, C. and Rheingans, R.1995. Mixed and pure forest plantations in the humid neotropics: a comparison of early growth, pest damage and establishment costs. Commonwealth Forest Review, 74, 306-314.
29 Murcia, C.1997. Evaluation of Andean alder as a catalyst for the recovery of tropical cloud forest in Colombia. Forest Ecology and Management, 99, 163-170. Nepstad, D. C., Uhl, C., Pereira, C. A., and Cardoso Da Silva, J. M.1996. A comparative study of tree establishment in abandoned pasture and mature forest of eastern Amazonia. Oikos, 76, 25-39. Parrotta, J.A.1995. Influence of overstory composition on understory colonization by native species in plantations on a degraded tropical site. Journal of Vegetation Science, 6, 627-636. Preston, F.1948. The commonness and rarity of species. Ecology, 29, 254-283. Purata, S.1986. Floristic and structural changes during old-field succession in the mexican tropics in relation to site history and species availability. Journal of Tropical Ecology, 2, 257-276. Quintana-Ascencio, P. F., Gonzales-Espinoza, M., Ramírez-Marcial, N., DominguezVázquez, G. and Martínez-Icó, M.1996. Soil seed bank and regeneration of tropical rain forest from milpa fields at the Selva Lacandona, Chiapas, Mexico. Biotropica, 28, 192-209. Sukumar, R., Suresh, S., Dattaraja, H. S. and Joshi, N. V.1998. Dynamics of a tropical deciduous forest: population changes (1988 through 1993) in a 50- ha plot at Mudumalai, southern India. Forest Biodiversity, Research and Modelling. (Eds. S. Dallamier, and J. Comiskey). Pp. 495-506. Smithsonian Research Institution, USA. Terborgh, J.T., Lopez, L., Nuñez, P., Rao, M., Shahabuddin, G., Orihuela, G., Riveros, M., Ascanio, R., Adler, G. H., Lambert, T. D., and Balbas, L.2001. Ecological meltdown in predator-free forest fragments. Science, 294, 1923-1926. Thebaud, C. and Strasberg, D.1997. Plant dispersal in fragmented landscapes: A field study of woody colonization in rain forest remnants of the Mascarenie Archipielago. Tropical Forest Remnants (eds. W.F. Laurance and R.O. Bierregaard, Jr.), pp. 321-332. University of Chicago Press. Tewksbury, J. J., Levey, D. J., Haddad, N. M., Sargent, S., Orrock, J. L., Weldon, A., Danielson, B. J., Brinkerhoff, J., Damschen, E. I. and Townsend, P.2002. Corridors affect aplants, animals, and their interactions in fragmented landscapes. Proceedings of the National Academy of Sciences, USA, 99, 12923-12926. Tucker, N. I. J. and Murphy, T. M.1997. The effects of ecological rehabilitation on vegetation recruitment: some observations from the wet tropics of North Queensland. Forest Ecology and Management, 99, 133-152.
30 Turner, I. M.1996. Species loss in fragments of tropical rain forests: a review of the evidence. Journal of Applied Ecology, 33, 200-209. Uhl, C. and Jordan, C. F.1984. Succession and nutrient dynamics forest cutting and burning in Amazonian. Ecology, 65, 1476-1490. Uhl, C., Buschbacher, R. and Serrao, E. A. S.1988. Abandoned pastures in Eastern Amazonia.1. Patterns of plant succesion. Journal of Ecology, 76, 663-681. Vandermeer, J. and Carvajal, R.2001. Metapopulation dynamics and the quality of the matrix. American Naturalist, 158, 211-220. Wright, S.J. and Duber, H.C.2001. Poachers and forest fragmentation alter seed dispersal, seed survival, and seedling recruitment in the Palm Attalea butyraceae, with implications for tropical seed diversity. Biotropica, 33, 583-595. Zimmerman, J.K., Pascarella, J.B. and Aide,T. M. 2000. Barriers to forest regeneration in an abandoned pasture in Puerto Rico. Restoration Ecology, 8, 350-360.
3. ONTOGENY OF SUN AND SHADE LEAVES OF EIGHT NONPIONEER TROPICAL TREE SPECIES
3.1
Abstract
I asked whether the ability of individual plants to adjust leaf traits during ontogeny reflected an immediate morphological means of responding to current light availability, or was consistent with a strategic change to meet predictable future light conditions. Leaf size, leaf mass per unit area (SLM), leaf density, toughness and water content of eight non-pioneer tropical tree species of different maximal height (8 to 40 m) were determined within three ontogenetic stages (seedling, juvenile and adult). Immediate responses to environmental conditions are reported as the mean values of leaf traits (leaf state) in sun and shade environments and the degree of change from sun to shade environments (leaf flexibility) while strategic changes for future light levels are reported as the correlation between maximal height of species and the direction and magnitude of the change in leaf traits (ontogenetic leaf variability) for the following ontogenetic transitions: from seedlings to adults, from seedlings to juveniles and from juveniles to adults. Seedlings, juveniles and adults of all species showed similar degree of immediate response to environmental conditions (leaf flexibility) irrespective of their maximal height for leaf size, SLM and toughness. Tall species showed greater change in leaf size and SLM from seedling to adult stage than short species as might be expected of species that experience varied light environments in life. These results show that all leaf traits change through
31
32 ontogeny, and that the magnitude and direction of this change (ontogenetic leaf variability) is determined by species maximal height. Still, leaf traits change in respond to current light levels (leaf flexibility) but this response is limited by current ontogenetic stage and the future habitat of individuals as adults.
3.2. Introduction Measurements of a variety of leaf traits of tropical tree species change during ontogenetic development, coinciding with increases in light availability as individuals increase in size (Coleman et al., 1994; Davies et al., 1998). Seedlings and juveniles of the shaded understory undergo morphological and physiological adjustments when gaps created by treefalls or branchfalls suddenly expose them to increased light levels (Bongers et al., 1988a; Agyeman, 1999; Poorter, 2001). Plants in high light conditions have smaller leaves with higher leaf mass per unit area (Specific leaf mass, SLM), dry mass per unit volume (leaf density) and toughness than plants in low-light environments (Parkhurst and Loucks, 1972; Bongers and Popma, 1988; Witkowski and Lamont, 1991; Pearcy and Sims, 1994). Quantitative changes in leaf characters are in part responsible for increased growth rates under higher light levels (King et al., 1997). Few studies, however, explore the changes in leaf traits that result from the interaction of ontogeny, current light availability and potential light availability that species are likely to experience as adults (Thomas and Bazzaz, 1999; Rijkers et al., 2000). An open question is whether changes in dimensions of leaf traits reflect an immediate adjustment to current conditions, or a strategic adjustment that anticipates the future conditions likely to be encountered during the ontogeny of a tree.
33 Plants may adjust leaf traits of photosynthetic importance to the light conditions that individual trees experience as they grow, or changes may be strategic, increasing the chance that individuals can adjust, if necessary, to conditions that they are likely to experience for most of their lives. As an example of a phenotypic response to contemporary conditions, Pearcy and Sims (1994) showed for the tropical herb Alocasia macrorrhiza (L.) G. Don that SLM doubled in contrasting light levels (15 to 700 µmol /m2 s ). In addition, comparative evidence suggests that tree species of different maximal heights modulate leaf characters (magnitude and direction of change) as part of ontogenetic development, what I term ontogenetic leaf variability. For example, in Malaysia, juveniles of tall species (canopy and emergent) have larger leaves than conspecific adults, while juveniles of short species (understory) have smaller leaves than conspecific adults (Thomas and Ickes, 1995; N=51 species). Within this sample, short species show less change in SLM from juvenile to adult than tall species (Thomas and Bazzaz, 1999; N=28 species), perhaps reflecting the more similar light environments shared by juveniles and adults of short species as compared with juveniles and adults of tall species. What remains to be tested is if these and other adjustments are immediate phenotypic responses to light environments of a given ontogenetic stages or strategic changes to face future light environment. Here I ask whether seedlings, juveniles and adults of eight late-successional tree species change their leaf traits in response to immediate light level (sun or shade) or if this response depends in the maximal height of these species that range from 8 to 40 m. Immediate responses to environmental conditions are reported as the mean values of five leaf traits (leaf size, SLM, leaf toughness, leaf density and water content; leaf state) and
34 the degree of change in leaf traits at contrasting leaf environments: sun and shade (leaf flexibility), while strategic changes for future light levels are evaluated as the correlation between maximal height of species and the direction and magnitude of the change in leaf state for three ontogenetic transitions: from seedlings to adults, from seedlings to juveniles and from juveniles to adults (ontogenetic leaf variability, Table I). I reason that all forest tree species regardless of their maximal height encounter variable light levels in their transitions from seedling to adult and are able to adjust their leaf traits to immediate conditions. This hypothesis will be supported if all species show (1) difference in leaf state between sun and shade leaf environments: leaves in sun conditions will show lower leaf size and higher SLM, toughness, leaf density and water content than leaves in shade conditions; (2) similar degree of change between contrasting leaf environments (leaf flexibility) for all leaf traits and, (3) similar magnitude and direction of change in leaf traits (ontogenetic leaf variability) in the three ontogenetic transitions: from seedlings to juveniles, from juveniles to adults and from seedlings to adults, irrespective of maximal height. Adults are expected to develop larger and heavier leaves than seedlings (Oberbauer and Strain, 1986; Kitajima, 1994) and the degree of this change should be similar for all species. Alternatively, leaf state, leaf flexibility and ontogenetic leaf variability may reflect fixed functional strategies of species, which differ by maximal height. In this scenario, (4) ability to show differences in leaf state between leaf environments (sun and shade) for each ontogenetic stages, (5) the degree of change between contrasting leaf environments (leaf flexibility) and, (6) the magnitude and direction of change in the three transitions (ontogenetic leaf variability) will depend on the maximal height of species. Taller species
35 experience a much greater range of light environments though ontogeny that goes from the dark understory to the bright canopy. Therefore, they are expected to show higher leaf flexibility than short species and smaller and heavier leaves at adult stage than at seedlings stage (direction of change) (Thomas and Ickes 1995, Thomas and Bazzaz 1999), in response to increases in light levels. For short species the opposite trend is expected, seedlings are expected to show smaller and lighter leaves than adults (direction of change) and, the magnitude of this change is expected to be smaller than that of tall species. Short species encounter less light variability from seedlings to adult stage, therefore, lower leaf flexibility and less ontogenetic leaf variability are needed or are likely to be expressed.
TABLE I VARIABLES ASSESSED IN EIGHT TROPICAL NON-PIONEER TREE SPECIES Variable
Definition
Leaf State
Mean of the measured leaf characteristic in a given leaf environment (sun or shade in a given ontogenetic stage).
Leaf Flexibility
Ratios of measures of leaf traits in sun as contrasted with shade environments within an ontogenetic stage. Leaf state in sun environment divided by leaf state in shade environment.
Ontogenetic Variability
Ratio of measurements among ontogenetic stages. Three measures include: mean leaf state pooled for both environments (sun and shade) of i) seedlings over adults, ii) seedlings over juveniles, and iii) juveniles over adults
36 3.3
Materials and Methods 3.3.1
Study site
This study was conducted in an experimental plantation located in the Cooperative of Lázaro Cárdenas, and at Los Tuxtlas Biological Station (LTBS) in the State of Veracruz, southeast Mexico (18 º 30’ N and 95º 03’ W). LTBS lies within a reserve of 640 ha of lowland tropical rain forest that constitutes the northernmost rain forest in the Neotropics, the Cooperative of Lázaro Cárdenas is located in the southwest corner of the reserve. Soils are sandy loams, classified as vitric andosols (FAO/UN 1975 in Soto-Esparza, 1976). Mean annual temperature is 27º C and mean annual rainfall is 4900 mm. The dry season extends from March to May, and a rainy season from June to February. The forest has a 35 m tall closed canopy. Nectandra ambigens (Blake) C.K. Allen (Lauraceae) is the most common species in the canopy while Pseudolmedia oxyphyllaria Donn. Sm. (Moraceae) and Astrocaryum mexicanum Liebm. (Palmae) are abundant in the mid-canopy and understory respectively (Bongers et al. 1988b). Field crops, pastures and small forest remnants surround this forest reserve.
3.3.2. Sampling In 1999, eight non-pioneer tree species with different maximal height were chosen from an experimental plantation created in 1997 in former pastures and secondary vegetation close to the Cooperative of Lazaro Cárdenas (Table II). The experimental plantation is located within a 25 ha parcel of former forest that was felled in 1978, used as agricultural land for 4 years and for cattle ranching for 15 years. The secondary forest has a 15 m tall canopy of pioneer species: Cecropia obtusifolia Bertol. (Moraceae),
37 Heliocarpuss appendiculatus Turcz. (Tiliaceae), Bursera simaruba (L.) Sarg. (Burceraceae) and Cedrela odorata L. (Meliaceae). Photon flux density (PFD) was measured in the dry season at midday (Light Meter Li-cor LI-189 quantum sensor) in pastures and secondary vegetation. In the understory of the secondary forest, instantaneous measurement of PFD was 22.53 ± 9.04 µmol/ m2 s (N=75; max 50 and min.10), while in the open pasture PFD was 1947.7 ± 218.78 µmol/ m2 s (N=11; max 2227 and min 1730). Species used in this study were chosen from the experimental plantation depending in the availability of seedlings under contrasting light levels (pasture area and secondary vegetation) that refers to leaf environment.
TABLE II SPECIES, AUTHORITIES AND RANGE OF STATURE OF EIGHT TROPICAL NON-PIONEER TREE SPECIES Species 1
Height (m)1
Family
Key
Nectandra ambigens (S.F. Blake) C.K. Allen
Lauraceae
Nect
20-40
Brosimum alicastrum Sw.
Moraceae
Bros
20-30
Calophyllum brasiliense Cambess.
Clusiaceae
Call
20-30
Pimenta dioica (L.) Merr.
Myrtaceae
Pimi
8-30
Eugenia inirebensis P.E. Sánchez
Myrtaceae
Euge
8-15
Licaria velutina van der Werff
Lauraceae
Lica
10-15
Amphitecna tuxtlensis A.H. Gentry Pouteria rhynchocarpa T.D. Penn. 1
Bignonaceae Amph Sapotaceae
From Ibarra-Manriquez and Sinaca (1995, 1996, 1996a)
Pout
3-10 3- 8
38
Fully expanded undamaged sun and shade leaves of the eight non-pioneer species in three ontogenetic stages (seedling, juvenile and adult) were sampled in November 1999; six individuals at the seedling and juvenile stages and three individuals at the adult stage of each species. Seedlings were < 1.3 cm diameter at the base and < 1.0 m high (except Licaria velutina van der Werff, ≤ 1.21 m height), juveniles were < 3.5 cm DBH and 2.5 m tall except Nectandra ambigens (S.F. Blake) C.K. Allen (≤ 8.9 cm DBH and 4 m tall), while adults were reproductive individuals > 3.5 cm DBH (Diameter at Breast Height; BH=1.5 m). Paucity of smaller seedlings of Licania and Nectandra required the larger cut-offs, which I also expect reflected reality of high growth rates in these species since all individuals in the plantation have the same age. Adults and juveniles of all species were located in the LTBS reserve. Ten to 20 leaves of adult trees were taken from the most exposed part of the canopy (sun leaves) and at the very bottom of the same crown (shade leaves). Three to five leaves were taken from each selected branch. To produce comparable leaf ages among life stages, leaves were sampled just behind the new leaves. Juveniles were located in the understory (shade environment) and in canopy gaps with full vertical illumination (sun environment) inside the reserve. Measures of leaf traits were taken on these different plants to associate leaf characters with low and high light environments (hereafter leaf environments). Ten leaves of each juvenile individual (N=3 for each species) and five leaves of each seedling (N=3 for each species) were sampled in each light environment (sun and shade, total N= 6 per species). Leaves were measured within 2 hr of collection. Leaf fresh mass was measured with a balance Pocket Pro 250-B (readability of 0.1 g) while fresh leaf size was measured
39 by drawing the leaf perimeter on paper and determining area with a Leaf area meter (Li3100, LiCor, Inc. Lincoln Nebraska USA). Leaves were oven dried to constant mass at 100 ° C and weighed to the nearest 0.01 g. With these measurements I calculated the following leaf traits (Table III): Specific Leaf Mass (SLM; inverse of Specific Leaf Area) is the leaf dry mass per unit fresh leaf area. Leaf Water content (WC), that is the difference of fresh leaf mass less leaf dry mass per unit area. I also report the ratio of leaf fresh mass over leaf dry mass often thought of as “leaf density” (Rascio et al., 1990; Garnier and Laurent, 1994). This leaf density reflects the amount of cytoplasm versus hemicellulose and cellulose in plant tissues. A penetrometer was constructed to measure toughness (Choong et al. 1992). Sand could be added to a pan above a punch head of 1 mm diameter to force it down; a reading was grams of sand needed to produce failure (Choong et al. 1992).
TABLE III LEAF TRAITS MEASURED IN EIGHT TROPICAL TREE SPECIES IN LOS TUXTLAS, VERACRUZ, MEXICO. LEAF TRAITS
DEFINITION OR FORMULA
REFERENCES
LEAF SIZE
Fresh leaf lamina projection and petiole
Westoby 1998, Weiher 1999
SLM
Dry mass of the leaf in relation with its area (g m -2)
Reich et al 1992
DENSITY
Dry mass to saturated leaf mass ratio
Garnier and Laurent 1994
TOUGHNESS
The grams of sand needed to make a hole of 1 mm
Choong et al. 1992
diameter through a leaf (g) WATER CONTENT
Fresh mass – dry mass /leaf area (g m -2)
Popma et al. 1992
40
3.3.3. Data analysis To test for the effect of species, leaf environment and ontogenetic stage in leaf traits (leaf size, SLM, leaf density, toughness and water content) multivariate analysis of variance (MANOVA) was used. The mean of each leaf characteristic for 5-20 leaves for each individual was used in all analyses. To calculate leaf flexibility (the ratio of leaf state in sun divided by that in shade) three pairs of individuals for each ontogenetic stage for species were used. To test for the effect of species and ontogenetic stage on leaf flexibility of leaf size, SLM, leaf density, toughness and water content, MANOVA was also used. To test differences within ontogenetic stages for leaf state or flexibility Tukey post hoc comparisons were used. Leaf state and leaf flexibility for all leaf traits were logtransformed to satisfy MANOVA assumptions. The mean leaf state from both leaf environments (sun and shade) of six individuals for each ontogenetic stage and for each species was used in the calculations of ontogenetic variability for three transitions: 1) from seedlings to adults, 2) from seedlings to juveniles, and 3) from juveniles to adults. To test the association of these three ratios with the maximal height of species, Spearman rank correlations were performed. Finally, leaf state from both leaf environments (sun and shade) and leaf flexibility for each species was correlated with the maximal height of species. Species are referred by genus only. Means are accompanied of standard errors trough all. All test were performed with STATISTICA 5.1 (StatSoft, 1998).
41 3.4
Results 3.4.1
Effects of species and ontogenetic stage on leaf traits 3.4.1.1. Leaf state
The mean leaf size for the eight species, at both leaf environment and for the three ontogenetic stages was 51.59 ± 3.1 cm2, mean SLM was 76.49 ± 2.2 g/m2, mean leaf density was 0.46 ± 0.01, mean toughness was 120.4 ± 2.0 g and mean water content was 98.7 ± 3.3 g/m2. MANOVA showed that species, ontogenetic stage and leaf environment had large effects on leaf traits: the interaction of species, ontogenetic stage and leaf environment was also significant (Table IV).
TABLE IV SUMMARY OF MULTIVARIATE ANALYSIS OF VARIANCE OF LEAF TRAITS MEASURED IN EIGHT TROPICAL TREE SPECIES IN LOS TUXTLAS, VERACRUZ, MEXICO MAIN FACTOR
DF 1
P
0.4), and CV of leaf water content (F=0.6, P > 0.4; F=0.8, P > 0.4) were not significant. As predicted, regressions of incremental growth in diameter on CV of SLM (Log Increment in diameter= -2.1 + 0.04 CV of SLM, R2= 0.34, P 0.8) nor CV of leaf water content (F = 0.6, P > 0.5). As expected, maximal tree height predicted the CV of SLM (CV of SLM= 8.64 + 0.31 Maximal Tree Height, R2= 0.18, P < 0.05; Figure 5). Tall species had higher intraspecific variation in SLM than short species. No significant regressions were found of survival on the maximal tree height (F = 0.09, P > 0.7, N=16) neither for the increments in diameter (F= 2.03, P > 0.1) and height (F = 0.27, P > 0.6) on maximal tree height.
4.5
Discussion These results show that an easily-measured leaf traits (CV of SLM) predicts the
performance of late-successional tree species growing in different microhabitats of early successional environments, allowing a criterion for selection of species suitable for planting in abandoned pastures and agricultural fields. Species most capable of adjusting SLM to different microhabitats of early successional environments (high CV of SLM) had higher survival and growth rates in that environment than species with less capability. Planting late-successional species in early successional environments, where pioneer species arrive by unassisted dispersal, helps bypass low-diversity forests (“pioneer desert” of Chapter 2) that may persist for decades.
74
34
CV of SLM = 8.64 + 0.31 Maximal Tree Height 2 R = 0.18, P< 0.05
Coefficient of variation of SLM (%)
30
Pesc
Neam
Coar
Cepe
26 Insi
22
Amtu
Cost Copo
Come Cabr Pidi Poar
Vigu
18 Trhe
14
Live
Gugr
Hitr
Sasa
Porh
10
Rhed Chve
6 2
Amho
Euin
5
10
15
20
25
30
35
40
45
Maximal Tree Height (m)
Figure 5. Regression of the Coefficient of variation of SLM on maximal tree height of 23 non-pioneer tropical tree species growing in an experimental planting in the Cooperative of Lazaro Cárdenas, Los Tuxtlas, Mexico. The key for the species follows Table X.
There are good reasons why variation in SLM is a suitable indicator of growth in varied conditions. A trade-off exists between growth potential and adaptation to adverse conditions (Lambers and Poorter, 1992). Those species adapted to harsh environments grow more slowly than those species growing in high-quality sites. In addition, trade-offs between species leaf traits (leaf area, photosynthetic capacity) and leaf longevity result in
75 different growth rates. The dry leaf weight per unit area (SLM) is a measure of allocation strategy (Wright and Westoby, 1999). In general, those species with low SLM show higher growth rates than species with high SLM when growing under optimal conditions, and this correlation holds for many species of different life forms in various ecosystems (Poorter and Remkes, 1990; Walters et al., 1993; Huante et al., 1995; Cornelissen et al., 1996; Grime et al., 1997; Reich et al., 1998; Westoby, 1998). Further, SLM varies within a species when individuals grow under different abiotic conditions. Plants in high-light conditions have smaller leaves with higher SLM than plants in low-light environments (Parkhurst and Loucks 1972; Bongers and Popma, 1988; Witkowski and Lamont, 1991; Pearcy and Sims, 1994). Given that quantitative changes in leaf traits in response to variable light levels are in part responsible for different growth rates (King et al., 1997) different abilities of species to adjust leaf traits to different microhabitats of an early successional environment predict their growth rates. Intraspecific variation in SLM results from adjusting leaf traits to different light and water availability levels. In the mature forest, seedlings growing in the dark and humid environment of the forest floor adjust their leaf traits to high light levels when canopy gaps open over them. By adjusting their leaf traits (i.e. leaf size) to the new conditions, they optimize light harvesting, thereby increasing their growth rates. However, not all species have the same ability to adjust leaf traits to new abiotic conditions. They still may tolerate the new environmental conditions but they are not able to efficiently use the new resources. Other attributes of species might serve as predictors of growth and survival.
76 Maximal tree height is associated with the plasticity of leaf traits (Niinemets, 1996; Thomas and Bazzaz, 1999). Tall species reach the canopy and develop two kinds of leaves within the same crown to cope with high light levels at the top of the crown and low light levels at the bottom (Popma et al., 1992). These “sun” and “shade” leaves may differ only in quantitative traits, or may differ markedly in structure and shape (Popma et al., 1992). Therefore, tall species have the ability to adjust leaf traits to different light levels and this might provide the functional means of adjusting well to the different microhabitats of early successional environments, resulting in higher survival and growth rates. This study did not support this prediction. Even when intraspecific variation in SLM was positively associated with maximal tree height, and the variation of SLM was related to growth rate and survival, maximal tree height did not predict performance. Selection of tall species from a little-studied sample for planting abandoned agricultural land, without evidence of ability to grow and survive, might not succeed. Of several leaf characters surveyed (Table XXII, Appendix G) all of them highly correlated to each other, coefficient of variation of SLM was the best predictor of growth and survival in early successional habitats.
4.6
Conclusions Late-successional tropical tree species represent > 80% of rain forest tree species
in undisturbed habitats, and provide much of the structural diversity as well as food resources for fruit- and seed-eating animals. For places where natural regeneration takes place, the species that are present in the old growth forest require many years to arrive by themselves (i.e. dispersal limitation exists). The planting of late-successional species is a
77 viable means of restoring diversity much more rapidly than natural regeneration (see Chapter 2). Here I show that tree species with high intraspecific variability in SLM will grow and survive better across different microhabitats of early successional environments than species with low variability in SLM. Use of variation in SLM and potentially other indices that predict performance may alleviate the need to individually screen large numbers of late-successional species for performance in restoration projects, allowing for more time and money to be used to evaluate other criteria for combinations of species that will generate desired restored forest.
4.7
Acknowledgments I thank the staff of the Los Tuxtlas Biological Station, Dr. R. Bye from the
Botanical Garden (UNAM) and Dr. Kumiko Shimada, Institute of Geology (UNAM) for their support, Miguel Sinaca and Jose Luis Paxtian for field assistance. I am grateful to N. Cordeiro, J. Fornoni, M. Jorge, S. Saha, A. Sullivan, and B. Zorn-Arnold for comments on the manuscript. This study was supported by CONACyT of Mexico and the Lincoln Park Zoo of Chicago.
4.8
Literature Cited
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79 Ibarra-Manríquez, G., Sinaca, S.,1996a. Lista florística comentada de la Estación de Biología Tropical “Los Tuxtlas”, Veracruz, México (Violaceae a Zingiberaceae). Revista de Biología Tropical, 44, 427-447. King, D.A., Leigh, E.G., Condit, R., Foster, R.B., Hubbell, S.P., 1997. Relationships between branch spacing, growth rate and light in tropical forest saplings. Functional Ecology, 11, 627-635. Lamb, D., 1998. Large-scale ecological restoration of degraded tropical forest lands: The potential role of timber plantations. Restoration Ecology, 6, 271-279. Lambers, H., Poorter, H., 1992. Inherent Variation in growth rates between higher plants: a search for physiological causes and ecological consequences. Advances in Ecological Research, 23, 187-261. Leopold, A.C., Andrus, R., Finkeldey, A., Knowles, D., 2001. Attempting restoration of wet tropical forests in Costa Rica. Forest Ecology and Management, 142, 243249. Loik, M.E., Holl, K.D., 1999. Photosynthetic responses to light for rainforest seedlings planted in abandoned pasture, Costa Rica. Restoration Ecology, 7, 382-391. Lugo, A.E., 1992. Comparison of tropical tree plantations with secondary forests of similar age. Ecological Monographs, 62, 1-41. Lugo, A.E., 1997. The apparent paradox of reestablishing species richness on degraded lands with tree monocultures. Forest Ecology and Management, 99, 9-19. Martínez-Garza, C., González-Montagut, R., 1999. Seed rain from forest fragment into tropical pastures in Los Tuxtlas, Mexico. Plant Ecology, 145, 655-665. Martínez-Ramos, M., 1994. Regeneracion natural y diversidad de especies arboreas en selvas humedas. Boletín de la Sociedad Botánica de México, 54, 179-224. Montagnini, F., González, E., Porras, C., Rheingans, R., 1995. Mixed and pure forest plantations in the humid Neotropics: a comparison of early growth, pest damage and establishment costs. Common Wealth Forestry Review, 74, 306-314. Moran, E.F., Brondizio, E., Mausel, P., Wu, Y., 1994. Integrating Amazonian Vegetation, Land-Use, and Satellite Data. Bioscience, 44, 329-338. Niinemets, U., 1996. Plant growth-form alters the relationship between foliar morphology and species shade-tolerance ranking in temperate woody taxa. Vegetatio, 124, 145-153.
80
Parkhurst, D.F., Loucks, O.L., 1972. Optimal leaf size in relation to environment. Journal of Ecology, 60, 505-537. Parrotta, J.A., Knowles, O.H., 1999. Restoration of tropical moist forest on bauxite -mined land in the Brazilian Amazon. Restoration Ecology, 7, 103-116. Pearcy, R.W., Sims, D.S., 1994. Photosynthetic acclimation to changing light environments: scaling from the leaf to the whole plant. Exploitation of environmental heterogeneity by plants. (Eds., Caldwell, M.M., Pearcy, R.W.), pp. 145-173, Academic Press, USA. Poorter, H., Remkes, C., 1990. Leaf-area ratio and net assimilation rate of 24 wild species differing in relative growth-rate. Oecologia, 83, 553-559. Popma, J., Bongers, F., Werger, J.A., 1992. Gap-dependence and leaf characteristics of trees in a tropical lowland rain forest in Mexico. Oikos, 63, 207-214. Reich, P. B., Tjoelker, M.G., Walters, M.B., Vanderklein, D.W., Bushena, C., 1998. Close association of RGR, leaf and root morphology, seed mass and shade tolerance in seedlings of nine boreal tree species grown in high and low light. Functional Ecology, 12, 327-338. Reich, P.B., Ellsworth, D.S., Walters, M.B., Vose, J.M., Gresham, C., Volin, J.C., Bowman, W.D., 1999. Generality of leaf trait relationship: a test across six biomes. Ecology, 80, 1955-1969. Ricker, M., Siebe, C., Sánchez, S., Shimada, K., Larson, B.C., Martínez-Ramos, M., Montaginini, F., 2000. Optimising seedling management: Pouteria sapota, Diospyros digyna, and Cedrela odorata in a Mexican rainforest. Forest Ecology and Management, 139, 63-77. Sokal, R.R., Rohlf, F.J., 2003. Biometry: The principles and practice of statistics in Biological Research. Third edition. W. H. Freeman and Company, New York. Soto-Esparza, M., 1976. Algunos aspectos climaticos de la región de Los Tuxtlas. Investigaciones sobre la Regeneración de Selvas altas en Veracruz, México. (Eds., Gómez-Pompa, A., C.Vázquez-Yanes, S., Del Amo, R., Butanda, C. A.), pp. 70-111, CECSA, Mexico. Sterck, F.J., Clark, D.B., Clark, D.A., Bongers, F., 1999. Light fluctuation, crown traits, and response delays for tree saplings in Costa Rica lowland rain forest. Journal of Ecology, 15, 83-95.
81
Thomas, S.C., Bazzaz, F.A., 1999. Asymptotic height as a predictor of photosynthetic characteristic in Malaysian rain forest trees. Ecology, 80, 1607-1622 Turner, I. M., 1996. Species loss in fragments of tropical rain forest: a review of the evidence. Journal of Applied Ecology, 33, 200-209. Walters, M.B., Kruger, E.L., Reich, P.B., 1993. Relative growth-rate in relation to physiological and morphological traits for northern hardwood tree seedlings species, light environment and ontogenic considerations. Oecologia, 96, 219-231. Welden, C.W., Hewett, S.W., Hubbell, S.P., Foster, R.B., 1991. Sapling survival, growth, and recruitment - relationship to canopy height in a Neotropical forest. Ecology, 72, 35-50. Westoby, M., 1998. A leaf-height-seed (LHS) plant ecology strategy scheme. Plant and Soil, 199, 213-227. Whitmore, T.C., 1989. Canopy gaps and the two major groups of forest trees. Ecology, 70, 536-538. Witkowski, E.T.F., Lamont, B.B., 1991. Leaf specific mass confound leaf density and thicknes. Oecologia, 88, 486-493. Wright, I.J., Westoby, M., 1999. Differences in seedling growth behavior among species: trait correlations across species, and trait shifts along nutrient compared to rainfall gradients. Journal of Ecology, 87, 85-97.
5. PREDICTING SURVIVAL AND GROWTH OF LATE-SUCCESSIONAL TREE SPECIES IN ABANDONED PASTURES
5.1
Abstract In the middle of large deforested areas or in sites far from a forest, dispersal of
forest seeds may be low or nil; abandoned lands may remain with no forest cover for a long time. In such cases, erosion degrades land quickly, making future restoration efforts difficult. Here I ask which average or variations in leaf traits predict higher survival and growth rates in sites with no incoming natural regeneration. I evaluate leaf size, leaf mass per unit area (SLM), leaf density and leaf water content and, maximal tree height of 12 tropical tree species grown for 17 months in pastures and secondary forests close to Los Tuxtlas Biological Station in Veracruz, Southeast Mexico. Incoming vegetation was cut once a month in pasture conditions to mimic abandoned pastures with no natural regeneration. Even when survival of latesuccessional species was higher under secondary forest (88 %) than in pastures, overall survival in pastures was high (70 %). All species showed similar increments in height in pasture and secondary forest while only two species (Omphalea and Sideroxylum) showed higher increments in diameter in pasture than in secondary forest. Leaf traits changed in response to contrasting light levels in pastures and secondary forest: individuals growing in pasture showed lower water content and higher SLM and density than individuals growing in the secondary forest. Survival of species in pastures was predicted by increments in height (R2 = 0.28). Mean leaf density or SLM measured in 82
83 individuals growing in secondary forest, and mean SLM measured in individuals growing in pastures predicted the increments in diameter of individuals growing in pastures. Those species with lower mean SLM or leaf density measured in secondary forest showed higher increments in diameter in pasture conditions (R2= 0.39 and R2= 0.45 respectively). By measuring mean SLM or leaf density in individuals growing in understory conditions, it is possible to predict which species will perform better in abandoned pastures with no input of pioneer species that develop a canopy over them. By planting these species it is possible to increase biodiversity in abandoned pastures that otherwise will remain as low diversity environments for long time.
5.2
Introduction After former agricultural or pasturelands are abandoned, the only source of seeds
for forest regeneration comes from dispersal events (Gómez-Pompa and Vázquez-Yanes, 1981; Finegan, 1996). In the middle of large deforested areas or in sites far from the forest, dispersal of forest seeds may be low or nil (Martínez-Garza and GonzálezMontagut, 1999; Hooper et al., 2002; Ingle, 2003). Even the small, highly vagile seeds of early successional species may not arrive. Abandoned lands may remain with no forest cover for years to decades while erosion degrades land quickly, making future restoration efforts difficult. Planting early successional species (pioneers) with high growth rates will only encourage the occupation of large areas with a small proportion (< 20%) of forest biota. Planting late-successional species that are so dispersal-limited that they do not arrive for a long time, if ever, is a preferable reforestation tactic than to create a low-diversity
84 pioneer matrix. Highly diverse mature forest is composed of late-successional species that provide much of the structural diversity and the food resources for fruit- and seedeating animals (see Welden et al., 1991; Martínez-Ramos, 1994). The planting of midand late-successional species, together with pioneer species, is a valuable option to establish a species-rich and structurally diverse forest in places where succession has been arrested. What is unknown is which mid- and late-successional species best endure the harsh conditions of abandoned lands with no woody regeneration. A previous study (Chapter 4) showed that of 23 late-successional species, those with higher variation in leaf mass per unit area (SLM) showed higher survival and growth rates across the microhabitats of early successional ecosystems (secondary forest, edge and open pastures). This high variation corresponds to an adjustment of leaf traits, (i.e., leaf size and weight) to different microhabitats of early successional environments. Here I ask which leaf traits or their variation predict higher survival and growth rates in abandoned pastures where all incoming vegetation was cut once a month to mimic abandoned pastures without natural succession. Further, I compare survival and growth rates of mid- and late-successional species in pasture conditions and in the secondary forest. Even if some species benefit from higher light levels when a gap opens over them in the forest, I expect late-successional species will show lower survival and growth rates in the long-lasting high light levels and low water availability typical of pastures than in the secondary forest (a habitat more similar to the mature forest). Leaf traits change in response to light and water levels. Therefore I expect individuals in long-lasting pasture condition to show leaf traits typical of individuals growing under high light levels and
85 low water availability. This in turn results in smaller leaves and lower water content and higher leaf density and SLM than individuals growing in the shade of secondary forest. Species adjust leaf traits in response to light and water levels, and these changes are in part responsible for higher growth rates. Therefore, mean leaf traits have been found to predict species growth rates under optimal conditions: those species with large leaves and low SLM show higher growth rates (Grime et al., 1997; Reich et al., 1998; Westoby, 1998). In agreement with this, the intraspecific variation in leaf mass per unit area (SLM) predicts species growth rates across microhabitats of early successional environments (Chapter 4). But which leaf traits will predict survival and growth rates in persistant adverse conditions like pastures with no input of pioneer species? I expect that those species with larger leaves and lower SLM show higher survival and growth rates in pasture while the opposite may be true for secondary forest. Previous studies showed that maximal tree height predicts the direction and magnitude of change in leaf traits through ontogeny for eight late-successional tree species (Chapter 3) and also predicted the intraspecific variation in SLM for 23 latesuccessional tree species (Chapter 4). Besides, in a study in Malaysia mean SLM was positively correlated to asymptotic tree height of 28 species (Thomas and Bazzaz, 1999). Tall species experience low light levels as seedlings or juveniles, but high light levels when they reach the canopy. They experience more heterogeneity in light levels during their lives than short species that spend more of their lives in a more homogeneous habitat, the shaded understory. Therefore, seedlings of tall species show larger leaf size and lower SLM than conspecific adults while the pattern is reversed for short species (Chapter 3). I expect tall species to show overall smaller leaves and higher SLM than
86 short species. However even with higher SLM, tall species may show higher survival and growth rates in pastures due to their higher variability in leaf traits.
5.3
Methods 5.3.1
Study site
This study was conducted in a experimental planting in the Cooperative of Lázaro Cárdenas at approximately 6 km from Los Tuxtlas Biological Station (LTBS) in the state of Veracruz, southeast Mexico (18 º 30’ and 18º 40’ N, 95º 03’ and 95º 10’ W). The LTBS lies within a reserve of 640 ha of lowland tropical rain forest that is the northernmost rain forest in the neotropics. The forest has a closed canopy of about 35 m high. Nectandra ambigens (Blake) C.K. Allen (Lauraceae) is the most common species in the canopy, while Pseudolmedia oxyphyllaria Donn. Sm. (Moraceae) and Astrocaryum mexicanum Liebm. (Arecaceae) are abundant in the mid-canopy and understory, respectively (Bongers et al., 1988). Soil is sandy loam classified as vitric andosols (FAU/UN 1975 in Soto-Esparza, 1976). Mean annual temperature and mean annual rainfall are 27º C and 4900 mm, respectively. A dry season extends from March to May, and a rainy season from June to February.
5.3.2
Land use history and Experimental Planting
The study area is located within a 25 ha parcel of former forest land that was deforested in 1968, and used as crop land for 4 years and for cattle ranching for 15 years. It has been undergoing natural regeneration since abandonment in 1987. This early successional environment has three distinct microhabitats: (1) area covered by grasses,
87 (2) secondary forest and, (3) low and broken canopy of 2 to 4 m. From these three, two microhabitats were used: (1) open area of approximately 3 ha covered with the exotic grass Cynodon plectostachyus (K. Schum.) Pilg. up to 1.5 m high in the wet season, and (2) the closest secondary forest with a 15 m tall canopy of pioneer species: Cecropia obtusifolia Bertol. (Moraceae), Heliocarpus appendiculatus Turcz. (Tiliaceae), Bursera simaruba (L.) Sarg. (Burseraceae) and Cedrela odorata L. (Meliaceae). From August to December 2000, seeds of 12 mid and late-successional tree species (Martínez-Ramos, 1985) were collected from the LTBS and germinated (100 individuals per species N= 1200 individuals) under shade in a nursery. Numbers of individuals per species are unequal due to differential seed germination, and seedling survival and mortality during transplantation (Table XXIII, Appendix I). On March 2001, I marked a grid of 1200 squares of 2 X 2 m covering the field, including open pasture and the closest secondary forest. In August 2001, seedlings of approximately 4 months old were planted, one seedling in the middle of each square. Seedlings of all species were systematically randomized in the grid to guarantee interspersion. Light levels (Photon Flux Density, PFD) and soil characteristics (i.e. soil moisture and bulk density) were measured in the pastures and secondary vegetation for a subsample of individuals. Photon-flux density was measured with a Light Meter Li-cor LI189 quantum sensor. Gravimetric soil moisture and bulk density were measured following Chapman (1976). Bulk density estimates soil compaction. Photon-flux density and gravimetric soil moisture were significantly different in the two habitats (MannWhitney U test (1,262) =31.6 P< 0.0001; F(1,52)=4.1, P < 0.05, respectively) while bulk density was not (F(1,54)=0.02, P > 0.9; Table XII).
88 Twelve tree species from nine families were used to represent the variety of maximal tree heights and fruits that attract different frugivorous animals in the area (Table XIII). Maximal tree height is the tallest published stature for adults of a given species in the LTBS (Ibarra-Manríquez and Sinaca, 1995; 1996; 1996a). To calculate growth rates, seedling heights and diameters at the stem base were measured in August 2001, February 2002, April 2002 and January 2003. Grasses and herbs were removed 50 cm around individuals at the time of measurement. To mimic abandoned pastures without natural regeneration, incoming plants growing in the pasture were cut once a month to ensure homogeneous high light conditions. Increments of growth rates in heights (height t2-height t1/
time) and in diameter at the stem base (hereafter diameter) were calculated for
the three periods between censuses. The survival rate of each species was calculated as the number of individuals surviving from 2001 to 2003, divided by the number of individuals present in 2001. One to three fully expanded undamaged leaves were collected from individuals displaying more than 10 leaves (see Table XXIV, Appendix J). Leaves were measured within 2 hr of collection. Leaf fresh weight (FW) was measured with a balance Pocket Pro 250-B (readability of 0.1 g) while fresh leaf size (LS) was measured with a leaf area meter (Ci-202, CID, Inc. Camas, WA, USA). Leaves were oven dried to constant weight at 100 ° C and weighed to the nearest 0.01 g (DW). Specific Leaf Mass (SLM; inverse of Specific Leaf Area) is the leaf dry weight per unit fresh leaf area (=DW/LS). I also report the ratio of leaf fresh weight over leaf dry weight often thought of as “leaf density” (=FW/DW; Rascio et al., 1990; Garnier and Laurent, 1994). This leaf density reflects the amount of cytoplasm versus hemicellulose and cellulose in plant tissues. I also report leaf
89 water content that is the difference between leaf fresh weight and leaf dry weight per unit leaf area (=FW-DW/LS).
TABLE XII PHOTON FLUX DENSITY (PFD), GRAVIMETRIC SOIL MOISTURE AND BULK DENSITY IN SECONDARY FOREST AND PASTURES Habitat Secondary Forest Pasture 1
Median, N=500
2
Mean ± SE, N=120
PFD 1
Soil Moisture 2
BulkDensity2
(µmol -2 s -1)
(mL g -1)
(g cm -3)
20
0.43 ± 0.01
0.67 ± 0.02
1730
0.39 ± 0.01
0.64 ± 0.01
The intraspecific variation in leaf traits is reported as its coefficient of variation (hereafter CV) = Standard Deviation of leaf traits * 100 / Mean leaf traits for the two habitats, and it was corrected for sample size (Sokal and Rohlf, 2003).
5.3.3
Data Analysis
T-tests were used to test for survival differences between habitats overall the 12 species. Multivariate Analysis of Variance (MANOVA) tested the effects of species and habitat (pasture and secondary forest) on increments in height and diameter at the base of the stem. The mean increment in height and diameter for the three periods of time
90 between the censuses during which an individual was alive was used for MANOVA analysis. Individuals that showed a mean negative increment (died back) were not included in MANOVA analysis since increments were log transformed. However, mean individual negative increments are included when means and standard Errors (SE) of increments are shown for each species. Increments and leaf traits were log transformed to satisfy MANOVA assumptions.
TABLE XIII FAMILY, FRUIT TYPE AND RANGE OF MAXIMAL TREE HEIGHT OF 12 LATESUCCESSIONAL TROPICAL TREE SPECIES SPECIES
FAMILY
KEY
FRUIT TYPE 1
HEIGHT (m) 1
Amphitecna tuxtlensis
Bignoniaceae
Amtu
Green berry
3-10
Calophyllum brasiliense
Clusiaceae
Cabr
Green drupes
20-30
Coccoloba hondurensis
Polygonaceae
Coho
Red or black drupes
15-20
Chrysophyllum mexicanum
Sapotaceae
Chme
Red or black berry
10-20
Cymbopetalum baillonii
Annonaceae
Cyba
Capsule with arrillate seeds
8-25
Diospyros digyna
Ebenaceae
Didi
Green berry with black arill
20-25
Guarea grandifolia
Meliaceae
Gugr
Brown capsule
20-30
Nectandra ambigens
Lauraceae
Neam
Black drupe with red cap
20-40
Omphalea oleifera
Euphorbiaceae
Omol
Green berry
15-30
Pouteria campechiana
Sapotaceae
Poca
Brown drupe
15-20
[Garcinia intermedia]
Clusiaceae
Rhed
Yellow berry
5-15
Sideroxylon portoricense
Sapotaceae
Sipo
Brown berry
20-35
Rheedia edulis
1
From Ibarra-Manriquez and Sinaca 1995, 1996, 1996a.
To find out which traits may predict survival and growth rates of mid and latesuccessional species, linear regressions were performed for survival and increments
91 separated for the two microhabitats used on all leaf traits and their variation (CV). The mean increment for species in the three periods of time between censuses and across all habitats was used for regressions. Regressions are used to explore data; for regression with data measured in two habitats causal relationships are not implied. Increments in height and diameter were log transformed to satisfy regression assumptions. All statistical analyses were performed in STATISTICA 4.0
5.4
Results 5.4.1
Survival
Mean 17-month survivorship % for 12 species was significantly higher in the secondary forest (88 ± 1.7 %) than in the pasture (69 ± 4.8 %, t-test (1, 11) =3.22, P< 0.005). Cymbopetalum showed the lowest survival in the secondary forest and pastures (77 and 39 % respectively) while Chrysophyllum showed the highest survival in the secondary forest (97 %), and Coccoloba in the pasture (100 %; Table XXIII, Appendix I).
5.4.2
Increments in height and diameter
Mean increment in height for the twelve species was 1.17 ± 0.05 cm / month. Guarea grandifolia showed the lowest increment in height (0.34 ± 0.01 cm / month), while Calophyllum brasiliense showed the highest (2.86 ± 0.51 cm / month). Nectandra ambigens showed the lowest increment in diameter (0.01 ± 0.005 cm / month), while Cymbopetalum baillonii showed the highest increment in diameter (0.07 ± 0.01 cm /
92 month). MANOVA revealed that overall species and habitat showed differences in increments in height and diameter. The interaction of species and habitat was significant (Table XIV) revealing that only two species (Omphalea oleifera and Sideroxylum portoricense) showed significately higher increments in diameter in open pasture than in secondary forest (Figure 6). All species showed similar increments in height in pasture and secondary forest.
TABLE XIV SUMMARY OF MULTIVARIATE ANALYSIS OF VARIANCE OF INCREMENTS IN HEIGHT AND DIAMETER MEASURED IN 12 TROPICAL TREE SPECIES IN LOS TUXTLAS, VERACRUZ, MEXICO MAIN FACTOR
WILKS ‘
RAO ‘S
DF 1
P
0.2).
4.4.3.2 SLM Species differed in SLM (F= 69.1, P(11, 399) < 0.00001), and habitat had large effect on this leaf trait (F= 174.0, P(1, 399) < 0.00001). Rheedia showed the highest SLM (115.6 ± 4.3 g m-2) while Omphalea showed the lowest SLM (36.4 ± 2.6 g m-2; Table XXIV, Appendix J). Individuals in the pasture showed 1.4 times higher SLM than those in the secondary forest (86 ± 1.6 g/m2 and 62 ± 1.5 g/m2 respectively).
94
Log Increment Diameter (cm / month)
-0.8 *
Secondary Forest Pasture -1.2
* -1.6
-2.0
-2.4
-2.8
Amtu Rhed Chme Coho
Poca
Cyba
Didi
Gugr
Omol
Cabr
Sipo
Neam
Species
Figure 6. Log Increment of diameter at the base of the stem of 12 late-successional tropical tree species Key for species follows Table XIII, species are arranged by increasing maximal tree height from left to right. Asterisks over bars represent significant differences between secondary forest and pasture tested with Post Hoc Tukey Test after Univariate tests were performed.
95 TABLE XV SUMMARY OF MULTIVARIATE ANALYSIS OF VARIANCE OF LEAF TRAITS MEASURED IN 12 TROPICAL TREE SPECIES IN LOS TUXTLAS, VERACRUZ, MEXICO WILKS ‘ LAMBDA
RAO ‘S
DF 1
P
0.1; SLM, F=0.8, P > 0.7; leaf density, F=0.2, P> 0.6) and, neither of them had predictive power for the survival in pastures (leaf size, F=2.43, P > 0.1; SLM, F=0.12, P > 0.7; leaf density, F=1.79, P> 0.2; leaf water content, F=0.05, P > 0.8). The species growth increments in diameter in secondary forest were not predicted by the variation in leaf traits while the increment in diameter in pastures was partially predicted by the coefficient of variation of SLM (F= 4.42, P < 0.06). The species increments in height in the secondary forest were predicted by the CV of Water Content (Log Increment in height = 1.47 - 0.05 CV of WC; R2 = 0.42, P < 0.01; Figure 13) and the CV of leaf density (Log Increment in height = 0.79 - 0.04 CV of Density; R2 = 0.30, P < 0.05; Figure 14), while CV of leaf size (F = 0.38, P > 0.5) and CV of SLM (F=0.18, P > 0.6) did not have predictive power. The increment in height for individuals growing in pastures was predicted only by the Coefficient of Variation in Leaf density (Log Increment in height = 1.01 - 0.03 CV of Density; R2 = 0.27, P < 0.05; Figure 15), the remaining leaf traits did not have predictive power (leaf size, F = 0.5, P > 0.5; SLM, F=0.49, P > 0.5; leaf water content, F = 1.39, P > 0.2).
104
17-month survivorship in Sec Forests (%)
1.02 0.98
17-month survivorship in Sec For = 0.75 + 0.003 * CV of W C 2 R = 0.30, P < 0.05
Neam
0.94
Coho
Chme
Amtu
Poca
Rhed
Omol
0.90 0.86
Didi
Gugr Sipo
Cabr
0.82 0.78 0.74 10
Cyba
15
20
25
30
35
40
45
50
55
Coefficient of Variation of Water Content (%)
Figure 12. Regression of 17-month survivorship in secondary forests on the coefficient of variation of Water Content of 12 non-pioneer tropical tree species growing in an experimental planting in the Cooperative of Lázaro Cárdenas, Los Tuxtlas, Mexico. The key for the species follows Table XIII
Log Increment in Height in Sec For (cm / month)
105
0.6 Didi
Cabr
Neam
0.2
Omol Cyba Chme
-0.2
Poca Coho
-0.6
Amtu
Sipo
-1.0
Gugr
-1.4 -1.8 -2.2 -2.6 10
Rhed
Log Increment in Height in Sec Forest = 1.47 - 0.05 CV of W C 2 R = 0.42, P < 0.01
15
20
25
30
35
40
45
50
55
Coefficient of Variation of Water Content (%)
Figure 13.Regression of the increments in height in secondary forests on the Coefficient of variation of Water Content of 12 non-pioneer tropical tree species growing in an experimental planting in the Cooperative of Lázaro Cárdenas, Los Tuxtlas, Mexico. The key for the species follows Table XIII
Lof of Increment in Height in Sec For (cm / month)
106
0.6 Neam Cabr
Didi
0.2
Omol Cyba Chme
-0.2
Poca
Amtu
Coho
Sipo
-0.6 -1.0
Gugr
-1.4 -1.8
Log of Increment in Height in Sec For = 0.79 - 0.04 CV of Leaf Density 2 R = 0.30 P < 0.05 Rhed
-2.2 -2.6 10
15
20
25
30
35
40
45
50
55
Coefficient of Variation of Leaf Density (%)
Figure 14.Regression of the increments in height in secondary forests on the Coefficient of variation of Leaf Density of 12 non-pioneer tropical tree species growing in an experimental planting in the Cooperative of Lázaro Cárdenas, Los Tuxtlas, Mexico. The key for the species follows Table XIII.
Log of Increment in Height in pastures (cm / month)
107
1.0
0.6
Cabr Amtu
Didi Omol Neam Coho
Chme Poca
0.2
Cyba
-0.2
Sipo
Rhed
-0.6 Log Increment in Height in pastures = 1.01 - 0.03 CV of Density 2 R = 0.27, P < 0.05
-1.0 10
15
20
25
30
35
40
Gugr
45
50
55
Coefficient of Variation of Leaf Density (%)
Figure 15.Regression of the increments in height in pastures on the Coefficient of variation of Leaf Density of 12 non-pioneer tropical tree species growing in an experimental planting in the Cooperative of Lázaro Cárdenas, Los Tuxtlas, Mexico. The key for the species follows Table XIII.
108 5.5
Discussion In deforested areas far from sources of seed to start regeneration, natural
succession may be delayed many years or not start (Finegan 1996). Planting a mix of fast-growing, early-successional species along with mid- and late-successional tree species may fulfill many important restoration goals in such sites. Pioneer species may immediately stop erosion, providing a highly desired ecosystem function, while mid- and late-successional species generate a high diversity forest. However, the high number of species in the mature tropical forest precludes testing all of them in pastures to screen which ones show high survival and growth rates in these environments. Functional leaf traits, like SLM or leaf density are useful for predicting survival and growth rate of late-successional species when planted in adverse conditions (i.e., abandoned pastures). Further, these leaf traits may be measured in pastures or in the shaded conditions of the secondary forest, and still predict growth rates of species growing in pastures. Light conditions in the secondary forest are as low as those prevailing in the mature forest, therefore, by measuring leaf traits under those conditions, it is possible to predict the performance of species in another set of very adverse environmental conditions (pastures). This alleviates the need to individually screen large numbers of late-successional species for restoration projects, allowing for more time and money to be used to evaluate other criteria for combinations of species that will generate desired restored forest. That survival of late-successional species was higher under secondary forest than in pastures is an anticipated result, since these species usually establish in the dark humid understory of the forest. On the other hand, that all species showed similar increments in
109 height in pasture and secondary forest was surprising. Even when late successional species respond to sudden high light conditions caused by formation of gaps in the canopy (Popma and Bongers, 1991), their growth may decrease at extremely high light levels (Ricker, 1998; Poorter, 1999). Especially tall species are known to increase their height when gaps open, which is confirmed by the presence of juveniles and subcanopy size classes preferentially in forest gap or building phases (Clark and Clark, 1992). Two canopy species (Omphalea and Sideroxylum) showed significantly higher increments in diameter in pasture than in the secondary forest, these two species are those with the highest increments in height and in diameter (Figure 6). However, increments in height and diameter were not predicted by maximal tree height at least in the first 17 months; perhaps differences in growth rates will be noticeable later on. In a previous study, survival of late-succesional tree species across various microhabitats of an early successional environment was predicted by the variation in SLM. These different microhabitats result from natural regeneration events, then individuals are exposed to changing light and water levels due to the natural input of pioneer species (Chapter 4). In places where pioneers species do not arrive due to dispersal limitation (long-lasting pasture conditions), the variability in the leaf density that species display in pastures and secondary forest explain their increments in height in pasture (Figure 15). Leaf density represents the amount of cytoplasm relative to the amount of cell wall material. Under water stress, hemicellulose accumulates, thereby increasing leaf density and the water-holding capacity of leaves (Rascio et al., 1990; Garnier and Laurent, 1994). Those mid and late-successional species able to adjust leaf
110 density (water-holding capacity) showed higher increments in height in an environment with very low water availability, as they are pastures.
5.6
Conclusions Late-successional tropical tree species represent > 80% of rain forest tree species
in undisturbed habitats, and provide much of the structural diversity as well as food resources for fruit- and seed-eating animals. For places where natural regeneration does not take place the planting of a mix of early and late-successional species is a viable means of restoring diversity much more rapidly than waiting for natural regeneration to occur. Mean SLM and leaf density measured in the forest understory of mature forest may be used to predict performance of mid and late-successional species in long-lasting harsh conditions (i.e. pastures with not natural regeneration). This may alleviate the need to individually screen large numbers of species for survival and growth rates in restoration projects, allowing for more time and money to be used to evaluate other criteria for combinations of species that will generate desired restored forest.
5.7
Literature Cited
Bongers, F., Popma, J., Meave del Castillo, J., Carabias, J., 1988. Structure and floristic composition on the lowland rain forest of Los Tuxtlas, Mexico. Plant Ecology, 74, 55-80. Chapman, S.B., 1976. Methods in Ecology. Blackwell Science Pub, USA. Finegan, B., 1996. Pattern and process in neotropical secondary rain forest: the first 100 years of succesion. Trends in Ecology and Evolution, 11, 119-124.
111 Gómez-Pompa, A., Vazquez-Yanes, C., 1981. Sucessional studies of a rain forest in Mexico. In: West, D. C., H. H. Shugart, D. B. Butkin (Eds.), Forest Sucession Concepts and Application, Springer-Verlag, USA, pp. 246-266. Grime, J. P., Thompson, K., Hunt, R., Hodgson, J.G., Cornelissen, J.H.C., Rorison, I.H., Hendry, G.A.F., other 28 authors, 1997. Integrated screening validates primary axes of specialization in plants. Oikos, 79, 259-281. Hooper, E., Condit, R., Legendre, P., 2002. Responses of 20 native tree species to reforestation strategies for abandoned pastures in Panama. Ecological Applications, 12, 1626-1641. Ibarra-Manríquez, G., Sinaca, S., 1995. Lista florística comentada de la Estación de Biología Tropical “Los Tuxtlas”, Veracruz, México. Revista de Biologia Tropical, 43, 75-115. Ibarra-Manríquez, G., Sinaca, S.,1996. Lista florística comentada de la Estación de Biología Tropical “Los Tuxtlas”, Veracruz, México (Mimosaceae a Verbenaceae). Revista de Biologia Tropical, 44, 41-60. Ibarra-Manríquez, G., Sinaca, S.,1996a. Lista florística comentada de la Estación de Biología Tropical “Los Tuxtlas”, Veracruz, México (Violaceae a Zingiberaceae). Revista de Biologia Tropical, 44, 427-447. Martínez-Ramos, M. 1985. Claros, ciclos vitales de los arboles tropicales y regeneración natural de las selvas altas perennifolias. Investigaciones sobre la Regeneración de selvas altas en Veracruz, México. (eds. A. Gómez-Pompa and S. Del-Amo), 191240. Alhambra Mexicana S.A. de C.V., Mexico. Martínez-Ramos, M., 1994. Regeneracion natural y diversidad de especies arboreas en selvas humedas. Boletin de la Sociedad Botanica de Mexico, 54, 179-224. Parkhurst, D.F., Loucks, O.L., 1972. Optimal leaf size in relation to environment. Journal of Ecology, 60, 505-537. Pearcy, R.W., Sims, D.S., 1994. Photosynthetic acclimation to changing light environments: scaling from the leaf to the whole plant. In: Caldwell, M.M., Pearcy, R.W. (Eds.), Exploitation of environmental heterogeneity by plants, Academic Press, USA, pp. 145-173. Poorter, H., Remkes, C., 1990. Leaf-area ratio and net assimilation rate of 24 wild -species differing in relative growth-rate. Oecologia, 83, 553-559.
112 Ricker, M., Siebe, C., Sánchez, S., Shimada, K., Larson, B.C., Martínez-Ramos, M., Montaginini, F., 2000. Optimising seedling management: Pouteria sapota, Diospyros digyna, and Cedrela odorata in a Mexican rainforest. Forest Ecology and Management, 139, 63-77. Sokal, R.R., Rohlf, F.J., 2003. Biometry: The principles and practice of statistics in Biological Research. Third edition. W. H. Freeman and Company, New York. Soto-Esparza, M., 1976. Algunos aspectos climaticos de la región de Los Tuxtlas. In: Gómez-Pompa, A., C.Vázquez-Yanes, S., Del Amo, R., Butanda, C. A. (Eds.), Investigaciones sobre la Regeneración de Selvas altas en Veracruz, México, CECSA, Mexico, pp. 70-111. Thomas, S. C. and Bazzaz, F. A. 1999. Asymptotic height as a predictor of photosynthetic characteristics in Malaysian rain forest trees. Ecology, 80, 16071622. Walters, M.B., Kruger, E.L., Reich, P.B., 1993. Relative growth-rate in relation to physiological and morphological traits for northern hardwood tree seedlings species, light environment and ontogenic considerations. Oecologia, 96, 219-231. Welden, C.W., Hewett, S.W., Hubbell, S.P., Foster, R.B., 1991. Sapling survival, growth, and recruitment - relationship to canopy height in a Neotropical forest. Ecology, 72, 35-50. Whitmore, T.C., 1989. Canopy gaps and the two major groups of forest trees. Ecology, 70, 536-538.
6. SYNTHESIS AND IMPLICATIONS
Deforestation of tropical habitats is occurring at a rapid and in many places accelerating pace (Bawa and Dayanandan, 1997). Conservation efforts only may not be enough to preserve the forest, given the inevitable lost of species in fragmented forest (Turner, 1996). However, at the same time that deforestation is taking place, tens of thousands of hectares of denuded land are annually abandoned to succession back to forest (Moran et al., 1994). Restoration around forest remnants, abandoned land undergoing secondary succession, and open areas is urgently needed to decrease the impact of fragmentation in border areas and to protect the remaining core forest. Restoration ecology uses natural succession as basic tool; this managed succession may have a variety of objectives (Higgs, 1997, Parker 1997; also Howe 1994, 1999). Recovery of gross ecosystem functions, such as material retention and loss, energy inputs and outputs, and the outlines of food-web structure may require only a small subset of the 30 to 100 tree species ha-1 representative of forest structure in tropical Mexico (see Ehrenfeld and Toft, 1997, Palmer et al., 1997). To recover and maintain the high biodiversity of the tropical forest characterized by myriads of plant-animal interactions, many more species are needed. Which species should be chosen to start restoration projects depends on particular circumstances. Land is abandoned in different degrees of perturbation; a gradient of perturbation may go from little in small clearings of different size surrounded by mature
113
114 forest to highly degraded abandoned land covered by exotic grasses. In most of the cases, small seeded, highly vagile pioneer species are in the soil bank or arrive by dispersal events; they establish first after a perturbations occur, growing fast under high light levels. Abandoned lands far from the sources of forest seeds may remain under “arrested succession” as grasslands for long time (Zanne and Chapman, 2001). In these places succession does not start because of the lack of seeds of pioneer species (Nepstad et al., 1990, Holl, 1999). If and when pioneer trees do arrive, they soon provide a canopy that shades and kills grasses, thereby immediately improving the micronenvironmental conditions required by other tree species for germination and establishment. In places under arrested succession, the planting of pioneer species is mandatory, they will restore ecosystem function in a short time, and if trees bearing fleshy fruits are planted, birds and bats will be attracted, and they will further increase biodiversity by dropping seeds of other species. However, the planting of only pioneer species in restoration projects is not totally desirable. During the normal process of secondary succession in neotropical habitats, a mixture of fast growing pioneer species and a species-poor cohort of long-lived tree species establish for up to several decades (e.g., Denslow, 1985, Uhl et al., 1988, Guariguata et al., 1997). Pioneer species show low survival per capita and low diversity as a group (González-Montagut, 1996; Martínez-Ramos, 1985), therefore, few pioneer species dominate perturbed areas. This pioneer forest will not resemble the original forest for a long time, if ever.
115 A pioneer forest may provide ecological services (i.e. forest cover, soil retention), however, some biological function will not be there. Once a canopy of pioneer species exists, species diversity may keep increasing due to input of seeds brought by birds and bats attracted to these places. However, the intrinsic low dispersal ability of most deepforest, non-pioneer species may preclude their presence in extensive second-growth forests far from large remnants for many decades to centuries. A general consequence of natural or human-induced secondary succession with dispersal limitation of deep forest species is that isolated forest remnants may be surrounded for decades by species-poor assemblages of a subset of local early successional trees which overall comprise ≤ 20% of tropical tree species (MartínezRamos, 1994, Welden et al., 1991). Here, “letting nature takes its course” encourages continuing random species loss from primary forest remnants during what could be recovery of floristic and faunistic diversity in and near forest fragments (see Janzen, 1988). A second general consequence is that this usual process of secondary succession precludes recovery of biotic diversity representative of either forests prior to deforestation, or of remaining fragments, for decades. Restoration efforts in these places should ecourage the increase of biodiversity of deep forest species with low dispersal ability that otherwise may reach the secondary forest very slowly, if at all. Most tree species in the highly diverse mature forest are deep-forest, non-pioneer species (80 %, Martínez-Ramos, 1985). They germinate and establish inside the forest in shade (Swaine and Whitmore, 1988) and are usually long-lived species, although there are many exceptions in more species-rich tropical forests (Foster et al., 1986). Even when these species are those that sustain the highly diverse myriad of biological interactions,
116 they are not used in restoration projects for a number of reasons: non-pioneer species show lower growth rates than pioneer species (Swaine and Whitmore, 1988), therefore, planting only these species will delay the recovery of highly desired ecosystem functions (i.e., soil retention). More important, the little knowledge about non-pioneers survival and growth outside the forest preclude their use in restoration porjects. Given that nonpioneer tropical tree species provide much of the structural diversity as well as food resources for fruit- and seed-eating animals, they should be used to increase biodiversity in early successional ecosystems and planted as mixed stands with pioneer species in places under arrested succession. These plantings should sharply accelerate the process of re-vegetation of complex communities. Two to 10 years may be lost in slower growth of some trees, but decades saved in total community recovery. Dozens of non-pioneer species have been evaluated for survival and growth rates in abandoned pastures (Haggar et al., 1998, Leopold et al., 2001), while no information exist on their performance in places undergoing secondary succession. Given the high biodiversity in tropical ecosystems, it is impractical to hope to test more than a small fraction of individual species under certain conditions. One approach is to determine vegetative characteristics that predict success in survival and growth of trees under the stressful conditions of open pasture or abandoned land undergoing secondary succession, and then use those characteristics as criteria for selection of species to enrich or dominate forest restorations. Determination of which trait or traits best predict(s) species survival and growth in different scenarios of restoration (i.e. early successional ecosystems, pastures) will obviate the need to test all tropical tree species individually to decide which of them to
117 use. Maximizing survival per planted seedling reduce the cost of nursery maintenance and replanting in restoration projects. By resolving which trait is a better predictor of species growth rates in in different restoration scenarios we will minimize the time needed to regenerate a forest of high structural and species diversity. How can one predict survival and growth rates of tropical tree species under different scenarios of restoration? To survive, plants need to display a minimum leaf area (Kohyama 1987); to grow, photosynthetic products should be sufficient to maintain and increase shoot and root biomass. Morphological and physiological adjustments in plant traits to different microenvironmental conditions that occur within the forest are in part responsible of the higher survival and growth of individuals in variable environments (King, 1994). This phenotypic plasticity allow plants to survive and grow in a wide range of environmental conditions The phenotypic plasticity in leaf traits shows predictable behavior for different levels of light and water. Individuals exposed to sun conditions develop leaves with higher leaf mass per unit area (specific leaf mass; SLM), higher leaf water content (WC: fresh leaf weight-dry leaf weight/ leaf area, g m-2) and thicker leaves, and lower stomatal density in comparison with shade leaves (Popma et al., 1992, Oberbauer and Strain, 1986, Pearcy and Sims, 1994). These traits increase whole-plant productivity by maximizing photosynthesis and minimizing loss of water (Björkman, 1981, Bongers and Popma, 1988, Popma et al., 1992). Reduction in leaf area and increase in thickness in sun conditions alleviates heat load and decreases loss of water (Chiariello, 1984). On the other hand, plants in shade have larger leaves for light capture and maximization of sunfleck use.
118 The amount of phenotypic plasticity that a species displays may depend on the environmental heterogeneity that it has to endure through its ontogenetic development and to the most prevalent environmental conditions that it experiences. At Los Tuxtlas, Veracruz, Mexico, I measured leaf traits in eight late-successional tree species at three ontogenetic stages (seedlings, juveniles and adults) and at contrasting light levels (sun vs. shade). I found that seedlings, juveniles and adults showed similar phenotypic plasticity in leaf traits (leaf flexibility; the ratio of sun and shade measurement in leaf traits) for leaf size, SLM, leaf density, water content and toughness, though higher leaf flexibility in thickness and physiological attributes were found for adults compared to juveniles in the ivy, Hedera helix for contrasting light levels (Hoflacher and Bauer, 1982). Individuals of tropical tree species are exposed to horizontal and vertical light gradient inside the forest and they are able to adjust their leaf traits at all ontogenetic stages perhaps better than species in other ecosystems with less environmental heterogeneity. The most prevalent environmental conditions that a tree species experiences for most of its life may be represented by its height (Thomas, 1996, Westoby, 1998). For 36 Malaysian tropical tree species, it was found that leaf size, measured in adults fallen leaves and juveniles freshly leaves, decrease from juveniles to adult stage for tall species while short species show the opposite trend (Thomas and Ickes, 1995). For 28 Malaysian tropical tree species, it was found for SLM, that taller species showed higher SLM at adult stage than conspecific juveniles and the direction of this change was similar for short species but the magnitude of this change was higher for taller species (Thomas and Bazzaz, 1999). For eight late-successional species which maximal tree height that ranges from 8 to 40 m, I found that taller species show higher change in leaf size and SLM in
119 different ontogenetic transitions (ontogenetic leaf variability) than short species, which agrees with the previous studies even when a different measure of tree height was used. The best measure of tree height is the average final height of a population of a given species; this is called asymptotic tree height (Thomas, 1996). For my study, longterm growth records were not available, only range of maximal tree height. Maximal tree height refers to “champion trees”, for which final size is affected by environmental and genetic variation. For Los Tuxtlas, this data was taken from Ibarra-Manriquez and Sinaca (1995, 1996, 1996a). Therefore, even when caution in using “champion tree” data is mandatory (Thomas, 1996), existing data of maximal tree height is a useful first approach. Non-pioneer species that reach the canopy (tall species) are exposed to a higher heterogeneity in light levels and water availability than short species during their expansion to attain the canopy, but once there, they experience the highest light levels in the upper part of their crown. Those species that inhabit the understory all of their lives show lower ability of changing leaf traits through ontogeny given that they may remain in the dark understory from seedlings to adult stages. How will the variability in leaf traits, related to prevalent habitat that adults experience aid to predict performance of species in different scenarios of restoration? In extensive abandoned pastures, where no natural regeneration is taking place, high light conditions and low water availability are the prevalent environmental conditions. Therefore, leaf traits related to high light conditions might provide the means to better survive and grow there, and taller species are expected to show those leaf traits. From all leaf traits previously mentioned, SLM have been found to better correlate with
120 growth rates for many species in many ecosystems (e.g. Grime et al., 1997; Reich et al., 1998; Westoby, 1998). For twelve late-successional tree species growing in pastures where incoming vegetation was cut once a month to mimic arrested succession, I found that, mean SLM measured in individuals growing in pastures (high light levels), or mean SLM and mean Leaf density measured in the closest secondary forest (lower light levels), to some extent predict growth rates in diameter for individuals growing in pasture environments. Those species that show lower mean SLM show higher growth rates than those species with higher mean SLM, with regressions coefficient (R2) that range from 0.27 to 0.45. This agrees with previous studies done under controlled conditions, for example Grime and 28 colleagues (1997) found a Spearman rank correlation coefficient of - 0.50 of growth rates with SLM, while Reich and colleagues (1998) found a Pearson correlation coefficient of - 0.90 for growth of nine boreal species with their SLM. The lower regression coefficient that I found are possible due to the error components associated to measurements in the field: unknown environmental conditions during the development of leaves and the genetic variation in leaf traits independent of environmental conditions. A greenhouse experiment with two late-successional tree species further supports my results from the field. Brosimum alicastrum and Pouteria campechiana were grown in pots under three light levels (30, 50 and 100 % availability) and two water availabilities (50 and 70 % of field capacity) during 6 months. Pouteria showed higher increments in height than Brosimum (3.76 ± 0.1 and 1.06 ± 0.01 cm / month respectively). The significant interaction between species and light levels (Table XXVI, Appendix L) revealed that Brosimum showed significantly higher increments in height at
121 30 % of light than at 100 % while Pouteria showed similar increments in height at all light levels (Figure 17, Appendix M). Species identity and light levels significantly affected leaf traits (Table XXVII, Appendix N and Table XXVIII, Appendix O). The regression of Log of Increments in height on mean SLM per each combination of light and water availability (Figure 18, Appendix P) showed the same pattern observed in field conditions: those individuals and species with low mean SLM show higher growth rates. Secondary forest has become a large portion of tropical forests (Moran et al., 1994). The increase of plant biodiversity with species that show very high dispersal limitation is mandatory if lost animal diversity is to eventually recover. If mean leaf traits (specifically leaf density and SLM) predicted increments in height for plants growing under one set of environmental conditions, the variation of leaf traits might explain the survival and growth rates of late-successional species growing in the different microhabitats of an early successional environment (i.e., secondary forest, edge and open pastures). For 23 late-successional tropical tree species I found that the intraspecific variation in SLM (measured as the coefficient of variation of SLM) explained survival and growth of individuals in early successional environments. Those species with higher variation in SLM showed higher mean increments in height and in diameter across the different microhabitats of an early successional environment. Maximal tree height was not correlated with survival and growth in early successional ecosystems in this study, as was expected. Thomas (1996), found that asymptotic maximal tree height predicted growth rates for adults and saplings of 28 Malaysian tree species. I did not find these correlations perhaps because I used maximal
122 tree height that refers to “champion trees” (trees favored by genetic and particular environments; Thomas 1996), instead of asymptotic tree height. Mean SLM and its variation offer easily-measured variables that may provide two of several criteria for selection of species for planting mixed species stands. For example, if a site far from sources of forest seeds is to be restored, a mix of pioneer and nonpioneer species should be planted there. Samples of leaves of as many as possible nonpioneer species from the region should be taken and evaluated for leaf traits. Three to five leaves from at least three individuals per species in only one microhabitat (sun or shade conditions) and one ontogenetic stage (saplings or adults) should be taken. Data about tree height and fruit type should be gathered when possible. Those non-pioneer species with lower mean SLM or leaf density will show higher survival and growth rates in longlasting high light conditions. A mix of those species and all known pioneer species should be planted. A high number of seeds of pioneer species will be needed because of their high initial mortality. Further selection of non-pioneer species may be done by fruit type, seed size, attractiveness to particular animal dispersal agents, and other desired criteria. For places alredy undergoing secondary succession, pioneer species might already be there. In this case, samples of leaves of as many as possible non-pioneer species, preferably those with the lowest dispersal ability should be evaluated for leaf traits. Three to five leaves per individual per species in different microhabitats should be evaluated: understory, gaps or borders and open areas when possible. Only one ontogenetic stage should be used, if adults are sampled, then, sun leaves at the top of the crown and shade leaves from the bottom of the crown can be used to evaluate variation in leaf traits. Also,
123 since variation in leaf traits is related to maximal tree height, tall trees may be choosen for further evaluation of leaf traits. A mix of as many non-pioneer species as possible should be planted, if necessary pioneer individuals should be cut or they canopy reduced. Further selection of species by local use or fruit type can be done to generate the desired restored forest.
6.1
Literature Cited
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APPENDICES
127
128
APPENDIX A TABLE XVI LEAF SIZE MEASURED IN EIGHT TROPICAL TREE SPECIES AT THREE ONTOGENETIC STAGES IN LOS TUXTLAS, VERACRUZ, MEXICO 1 ONTOGENETIC STAGE
ADULT
JUVENILE
SEEDLING
HABITAT SUN
SHADE
SUN
SHADE
SUN
SHADE
Calophyllum brasiliense
30.9 ± 1.7
36.3 ± 1.3
34.1 ± 2.7
33.3 ± 2.8
39.3 ± 2.3
35.1 ± 3.1
Brosimum alicastrum
51.8 ± 2.4
51.8 ± 2.6
77.9 ± 4.8
62.4 ± 3.6
29.0 ± 2.8
28.1 ± 3.7
Eugenia inirebensis
23.5 ± 1.0
21.6 ± 0.7
25.2 ± 1.5
22.7 ± 2.1
37.3 ± 3.6
27.9 ± 1.7
Pouteria rhynchocarpa
24.9 ± 0.8
33.7 ± 0.9
28.6 ± 1.3
27.6 ± 1.4
26.5 ± 2.7
47.2 ± 3.0
Licaria velutina
152.1 ± 4.3
149.7 ± 6.2
175.7 ± 6.5
127.4 ± 4.8
Pimenta dioica
39.1 ± 1.8
60.1 ± 2.5
53.1 ± 3.0
62.2 ± 6.7
28.9 ± 1.3
37.4 ± 4.5
Amphitecna tuxtlensis
43.7 ± 2.1
41.6 ± 1.7
40.0 ± 2.5
41.3 ± 2.1
37.1 ± 2.4
31.6 ± 2.8
Nectandra ambigens
26.8 ± 1.4
39.9 ± 2.2
58.9 ± 3.1
83.3 ± 6.7
42.7 ± 3.9
38.4 ± 2.8
SPECIES
1 Means and standard errors of Leaf size (cm2)
168.9 ± 14
87.0 ± 6.4
129
APPENDIX B TABLE XVII SLM MEASURED IN EIGHT TROPICAL TREE SPECIES AT THREE ONTOGENETIC STAGES IN LOS TUXTLAS, VERACRUZ, MEXICO 1 ONTOGENETIC STAGE
ADULT
JUVENILE
SEEDLING
HABITAT SUN
SHADE
SUN
SHADE
SUN
SHADE
Calophyllum brasiliense
85.6 ± 2.9
70.5 ± 1.6
64.2 ± 1.4
54.9 ± 1.4
57.2 ± 5.5
55.7 ± 3.6
Brosimum alicastrum
80.0 ± 2.5
61.2 ± 1.4
47.2 ± 1.0
58.3 ± 1.3
61.1 ± 1.9
40.3 ± 1.1
112.1 ± 4.8
82.8 ± 2.6
69.6 ± 2.2
68.5 ± 1.5
66.3 ± 3.1
72.8 ± 4.3
92.5 ± 1.1
71.4 ± 1.5
81.3 ± 2.9
55.1 ± 1.1
76.5 ± 3.1
60.7 ± 1.9
Licaria velutina
111.8 ± 4.2
103. 9 ± 3.8
97.6 ± 2.2
77.8 ± 2.0
99.5 ± 3.8
64.9 ± 1.2
Pimenta dioica
125.8 ± 3.2
100.0 ± 3.9
138.8 ± 13.3
76.5 ± 7.4
96.3 ± 2.7
67.5 ± 1.9
Amphitecna tuxtlensis
67.7 ± 1.6
56.5 ± 0.8
51.2 ± 2.2
51.4 ± 1.6
43.3 ± 3.0
59.7 ± 2.9
Nectandra ambigens
110.7 ± 2.9
91.5 ± 2.0
92.2 ± 1.6
71.0 ± 1.3
64.5 ± 3.2
67.2 ± 1.5
SPECIES
Eugenia inirebensis Pouteria rhynchocarpa
1
Mean ± standard error of specific leaf mass (g m-2)
130
APPENDIX C TABLE XVIII LEAF DENSITY MEASURED IN EIGHT TROPICAL TREE SPECIES AT THREE ONTOGENETIC STAGES IN LOS TUXTLAS, VERACRUZ, MEXICO 1 ONTOGENETIC STAGE
ADULT
JUVENILE
SEEDLING
HABITAT SUN
SHADE
SUN
SHADE
SUN
SHADE
Calophyllum brasiliense
0.36 ± 0.01
0.33 ± 0.01
0.35 ± 0.01
0.40 ± 0.02
0.31 ± 0.03
0.36 ± 0.02
Brosimum alicastrum
0.45 ± 0.01
0.45 ± 0.01
0.43 ± 0.01
0.43 ± 0.01
0.40 ± 0.01
0.48 ± 0.02
Eugenia inirebensis
0.52 ± 0.02
0.41 ± 0.01
0.41 ± 0.02
0.48 ± 0.03
0.33 ± 0.01
0.34 ± 0.02
Pouteria rhynchocarpa
0.64 ± 0.01
0.54 ± 0.01
0.58 ± 0.02
0.53 ± 0.02
0.73 ± 0.04
0.53 ± 0.02
Licaria velutina
0.44 ± 0.01
0.43 ± 0.01
0.44 ± 0.01
0.41 ± 0.01
0.53 ± 0.02
0.50 ± 0.02
Pimenta dioica
0.47 ± 0.01
0.41 ± 0.01
0.42 ± 0.01
0.39 ± 0.03
0.5 ± 0.02
0.41 ± 0.01
Amphitecna tuxtlensis
0.39 ± 0.01
0.36 ± 0.01
0.37 ± 0.01
0.45 ± 0.02
0.41 ± 0.04
0.83 ± 0.05
Nectandra ambigens
0.55 ± 0.01
0.47 ± 0.01
0.52 ± 0.02
0.41 ± 0.01
0.65 ± 0.04
0.60 ± 0.03
SPECIES
1
Mean ± standard error of Leaf Density
131
APPENDIX D TABLE XIX LEAF TOUGHNESS MEASURED IN EIGHT TROPICAL TREE SPECIES AT THREE ONTOGENETIC STAGES IN LOS TUXTLAS, VERACRUZ, MEXICO 1 ONTOGENETIC STAGE
ADULT
JUVENILE
SEEDLING
HABITAT SUN
SHADE
SUN
SHADE
SUN
SHADE
Calophyllum brasiliense
127.1 ± 2.4
122.1 ± 2.2
101.8 ± 2.3
101.2 ± 4.1
Brosimum alicastrum
115.2 ± 3.6
97.7 ± 4.0
93.6 ± 5.4
100.4 ± 4.8
Eugenia inirebensis
115.6 ± 2.7
121.0 ± 3.9
98.99 ± 5.4
97.7 ± 7.1
126.1 ± 8.7 118.8 ± 7.5
Pouteria rhynchocarpa
115.6 ± 3.9
106.8 ± 2.8
120.7 ± 4.4
102.1 ± 4.4
130.6 ± 6.3 113.3 ± 4.6
Licaria velutina
132.4 ± 4.3
130.1 ± 3.4
123.9 ± 6.2
117.9 ± 4.3
107.5 ± 6.3
78.3 ± 3.8
Pimenta dioica
147.2 ± 3.6
142.6 ± 2.5
151.2 ± 2.2
90.2 ± 3.5
151.8 ± 5.7
139.5 ± 4.1
Amphitecna tuxtlensis
158.0 ± 2.1
152.4 ± 2.2
141.6 ± 5.2
160.9 ± 2.6
114.7 ± 12.8 115.1 ± 8.2
Nectandra ambigens
144.5 ± 2.8
144.8 ± 2.3
140.3 ± 3.5
135.5 ± 3.8
122.6 ± 3.1 123.0 ± 4.2
SPECIES
1
Mean ± Standard Errors of Toughness (g)
98.5 ± 8.2 106.1 ± 6.6 117.5 ± 9.5
68.2 ± 5.3
132
APPENDIX E TABLE XX LEAF WATER CONTENT MEASURED IN EIGHT TROPICAL TREE SPECIES AT THREE ONTOGENETIC STAGES IN LOS TUXTLAS, VERACRUZ, MEXICO 1 ONTOGENETIC STAGE
ADULT
JUVENILE
SEEDLING
HABITAT SUN
SHADE
SUN
SHADE
SUN
SHADE
Calophyllum brasiliense
149.2 ± 2.9
141.5 ± 2.7
123.7 ± 5.4
90.6 ± 5.1
131.0 ± 7.1
97.3 ± 4.7
Brosimum alicastrum
103.0 ± 4.5
77.6 ± 2.9
62.7 ± 2.1
79.7 ± 3.7
91.6 ± 3.7
46.2 ± 3.7
Eugenia inirebensis
102.6 ± 4.8
118.1 ± 3.2
109.1 ± 7.3
85.3 ± 6.9
131.9 ± 4.9
140.5 ± 6.7
55.2 ± 2.7
64.7 ± 2.5
60.4 ± 3.6
53.3 ± 4.0
31.9 ± 6.1
55.4 ± 4.1
Licaria velutina
140.4 ± 5.6
135.1 ± 2.1
127.1 ± 3.5
110.8 ± 3.2
88.0 ± 5.1
66.5 ± 4.4
Pimenta dioica
145.1 ± 3.8
139.2 ± 2.5
195.5 ± 19.8
128.1 ± 16.3
96.5 ± 5.1
99.4 ± 5.4
Amphitecna tuxtlensis
107.9 ± 2.6
102.1 ± 2.3
89.3 ± 3.9
68.9 ± 4.0
66.5 ± 5.6
30.1 ± 5.6
Nectandra ambigens
93.5 ± 3.8
106.2 ± 3.1
94.3 ± 5.9
104.6 ± 2.8
41.2 ± 5.5
47.8 ± 5.3
SPECIES
Pouteria rhynchocarpa
1
Mean ± standard errors of Water Content (g m-2)
133
APPENDIX F TABLE XXI UNIVARIATE ANALYSIS OF VARIANCE OF FOLIAR FLEXIBILITY OF EIGHT TROPICAL NON-PIONEER TREE SPECIES IN LOS TUXTLAS, VERACRUZ, MEXICO 1 SPECIES
FOLIAR FLEXIBILITY LEAF SIZE
SLM
LEAF DENSITY
TOUGHNESS
WATER CONTENT
F (7,56) = 2.7*
F (7,56) = 4.4***
F (7,56) = 9.3****
F (7,56) = 3.23*
F (7,56) = 8.19***
Nectandra ambigens
0.99 ± 0.02
1.18 ± 0.10
1.00 ± 0.05
1.01 ± 0.04
1.23 ± 0.07
Brosinum alicastrum
1.17 ± 0.10
1.21 ± 0.10
0.94 ± 0.02
1.30 ± 0.15
1.43 ± 0.15
Calophyllum brasiliense
1.14 ± 0.10
1.10 ± 0.10
1.03 ± 0.06
1.01 ± 0.04
1.01 ± 0.07
Pimenta dioica
0.80 ± 0.10
1.35 ± 0.10
1.31 ± 0.10
1.14 ± 0.04
0.80 ± 0.12
Eugenia inirebensis
1.53 ± 0.20
1.31 ± 0.10
1.05 ± 0.02
1.16 ± 0.07
1.17 ± 0.07
Licaria velutina
0.81 ± 0.10
1.55 ± 0.20
1.14 ± 0.05
1.29 ± 0.12
1.28 ± 0.24
Amphitecna tuxtlensis
1.10 ± 0.10
0.97 ± 0.10
0.79 ± 0.10
0.98 ± 0.10
1.71 ± 0.36
Pouteria rhynchocarpa
0.86 ± 0.10
1.16 ± 0.10
1.25 ± 0.01
1.01 ± 0.02
0.90 ± 0.06
1
Flexibility is the ratio of sun over shade value in leaf characteristics for pairs of individuals growing under contrasting light levels. Means ± standard errores are reported. * P < 0.05, *** P < 0.0005, **** P < 0.0001
134
APPENDIX G TABLE XXII FAMILY, COEFFICIENT OF VARIATION OF LEAF TRAITS AND SAMPLE SIZE OF 23 LATE-SUCCESSIONAL TROPICAL TREE SPECIES Species
Family
CV Leaf size
CV SLM
CV Leaf density
CV Water Content
N
Ampelocera hottlei
Ulmaceae
29.08
6.82
11.76
22.14
89
Amphitecna tuxtlensis
Bignoniaceae
14.65
21.02
20.4
18.79
83
Calophyllum brasiliense
Clusiaceae
22.64
27.60
6.97
15.37
66
Ceiba petandra
Bombacaceae
34.88
26.66
8.3
14.27
9
venezuelanense
Sapotaceae
30.07
7.30
14.17
25.85
47
Cojoba arborea
Mimosaceae
43.13
27.30
19.61
44.67
62
Cordia megalantha
Boraginaceae
49.04
24.09
17.27
26.27
50
Cordia stellifera
Boraginaceae
17.41
29.47
8.35
16.00
149
Couepia polyandra
Chrysobalanaceae
33.71
16.62
15.63
23.06
47
Eugenia inirebensis
Myrtaceae
26.51
6.57
3.22
6.94
38
Guarea grandifolia
Meliaceae
34.01
15.48
11.64
11.79
32
Hirtella triandra
Chrysobalanaceae
56.38
14.49
17.11
28.00
13
Inga sinacae
Mimosaceae
52.40
25.11
13.25
19.05
64
Chrysophyllum
135
TABLE XXII (continued) Species
Family
CV Leaf size
CV SLM
CV Leaf density
CV Water Content
N
Licaria velutina
Lauraceae
29.88
15.30
7.2
9.15
115
Lonchocarpus cruentus
Fabaceae
57.59
51.49
30.56
41.84
112
Nectandra ambigens
Lauraceae
31.64
20.0
14.93
11.37
125
Persea schiedeana
Lauraceae
35.43
27.74
13.1
14.93
9
Pimenta dioica
Myrtaceae
33.33
21.08
15.4
25.41
55
Poulsenia armata
Moraceae
20.49
18.40
6.63
22.36
12
Pouteria rhynchocarpa
Sapotaceae
21.17
10.63
13.15
22.79
105
Rheedia edulis [Garcinia intermedia] Clusiaceae
24.41
8.49
2.12
9.02
31
Sapindus saponaria
Sapindaceae
40.68
12.28
30.93
63.27
90
Trichilia havanensis
Meliaceae
43.69
14.23
11.58
13.16
38
Virola guatemalensis
Myristicaceae
38.74
19.63
6.81
12.74
235
136
APPENDIX H
Log Increment Height (understory) = -0.22 + 0.04 CV SLM 2 R = 0.37, P< 0.05 N=10
Log of Increment in Height
1.0
Trhe
0.8 0.6
Insi
Cepe Coar Pesc
Live Sasa
0.4
Vigu Poar
Chve
0.2
Cost Copo Neam
Pidi Come Cabr Porh
Gugr Hitr
Amtu
0.0 Euin Amho
-0.2 -0.4
2
6
Log Increment Height (canopy) = -0.195 + 0.03 CV SLM 2 R = 0.35, P< 0.05, N=13
Rhed
10
14
18
22
26
30
34
Coefficient of Variation of SLM
Figure 16. Regressions of the Log of monthly Increment in height on the coefficient of variation of SLM for canopy and understory species growing in an experimental planting in the Cooperative of Lazaro Cárdenas, Los Tuxtlas, Mexico. Incremental growth refers to one mean of three periods of growth from 1998 to 2001. Open circles and thick line correspond to canopy and emergent species (> 25 m). Filled circles and thin line correspond to understory species. The key for the species follows Table X.
137 APPENDIX I TABLE XXIII SURVIVAL OF 12 LATE-SUCCESSIONAL TROPICAL TREE SPECIES IN PASTURES AND SECONDARY FORESTS Species
Family
N
Survival %
Survival %
Secondary
Pasture
Forest Amphitecna tuxtlensis
Bignoniaceae
106
94
65
Calophyllum brasiliense
Clusiaceae
21
83
73
Chrysophyllum mexicanum
Sapotaceae
87
90
100
Coccoloba hondurensis
Polygonaceae
83
96
80
Cymbopetalum baillonii
Annonaceae
63
77
39
Diospyros digyna
Ebenaceae
49
85
72
Guarea grandifolia
Meliaceae
37
84
44
Nectandra ambigens
Lauraceae
65
94
81
Omphalea oleifera
Euphorbiaceae
94
91
78
Pouteria campechiana
Sapotaceae
59
93
76
Rheedia edulis [Garcinia intermedia] Clusiaceae
72
93
63
Sapotaceae
68
83
56
Sideroxylon portoricense
138 APPENDIX J TABLE XXIV LEAF TRAITS AND TOTAL NUMBER OF INDIVIDUALS SAMPLED OF 12 LATESUCCESSIONAL TROPICAL TREE SPECIES Species
N
Leaf size1
SLM (g /m2) WaterContent2 Leaf density
Amphitecna tuxtlensis
54
21.9 ± 1.5
71.8 ± 2.8
81.5 ± 5.3
0.50 ± 0.02
Calophyllum brasiliense
13
35.1 ± 3.3
109.8 ± 4.1
152.6 ± 11.3
0.43 ± 0.03
Chrysophyllum mexicanum
50
15.2 ± 0.9
74.7 ± 2.1
99.0 ± 6.4
0.47 ± 0.02
Coccoloba hondurensis
50
64.7 ± 4.6
71.4 ± 2.2
131.2 ± 5.7
0.36 ± 0.01
Cymbopetalum baillonii
22
19.8 ± 2.6
50.0 ± 2.2
116.6 ± 6.9
0.31 ± 0.02
Diospyros digyna
34
42.1 ± 3.6
85.6 ± 1.9
114.1 ± 3.3
0.43 ± 0.01
Guarea grandifolia
11
32.8 ± 4.9
59.1 ± 3.2
149.9 ± 16.2
0.32 ± 0.05
Nectandra ambigens
37
61.1 ± 4.2
87.7 ± 2.5
123.32 ± 6.2
0.43 ± 0.02
Omphalea oleifera
42
176.9 ± 13.9
36.4 ± 2.6
146.3 ± 7.9
0.20 ± 0.01
Pouteria campechiana
33
32.2 ± 2.5
68.3 ± 2.9
82.7 ± 5.1
0.46 ± 0.02
[Garcinia intermedia]
35
14.5 ± 0.8
115.7 ± 4.3
116.4 ± 9.9
0.54 ± 0.03
Sideroxylon portoricense
34
24.2 ± 2.7
70.9 ± 3.5
89.0 ± 5.5
0.45 ± 0.01
Rheedia edulis
1
cm2
2
g /m2
139
APPENDIX K TABLE XXV PEARSON CORRELATION AMONG LEAF TRAITS MEASURED IN PASTURES AND SECONDARY FOREST FOR 12 TREE SPECIES AT LOS TUXTLAS, VERACRUZ, MEXICO1,2 Leaf size S F 3 Leaf size Sec Forest
1.00
Leaf size Pasture
0.98****
Leaf size P
SLM Sec For
SLM P4
WC Sec Forest
-0.40
-0.48
1.00
SLM Pasture
-0.43
-0.50
0.96****
1.00
Water Content S F
0.45
0.44
0.27
0.38 *
1.00
Water Content P
0.42
0.52
-0.27
0.38 *
0.57 ***
Leaf Density S F
0.68*
1
-0.53
1.00
-0.75**
0.79**
0.82**
-0.35
-0.65*
1.00
-0.64*
0.74**
0.81**
-0.25
-0.83**
0.91***
Pearson correlation and Bonferroni probabilities * P < 0.05, *** P < 0.001, ****P < 0.0001 3 Secondary Forest 4 Pastures 2
L Density S F
1.00
SLM Sec Forest
Leaf Density P
Water C P
140
APPENDIX L TABLE XXVI SUMMARY OF UNIVARIATE ANALYSIS OF VARIANCE OF INCREMENTS IN HEGHT FOR BROSIMUM ALICASTRUM AND POUTERIA CAMPECHIANA FROM A GREENHOUSE EXPERIMENT MAIN FACTOR
DF
F
P
