The sensitivity of pollen as an indicator of elevation in neotropical lowland and Andean forests was measured using modern pollen samples collected from moss-polsters along a transect between 340 m and 3530 m elevation and from surface sediments in lowland swamps (240 m) of Madre de Dios, Peru. A blind study, using samples collected from the same transect in the following year, provided a test of reproducibility. The results show (1) clear elevational distribution patterns and (2) the ability of calibration data to predict the altitude of the blind samples. Characteristic associations of pollen taxa are found under differing hydrologies and elevations. The floodplain pollen assemblages are characterized by abundant Mauritia, Sloanea, Ficus, Iriartea and Arecaceae pollen types. At higher elevations, these lowland types decrease or are absent. Alchornea, Urticaceae/Moraceae, Bignoniaceae and Cecropia are dominant components of the pollen rain of the low-elevation zone (<1000 m). Acalypha, Alchornea, Cecropia, Rubiaceae and Urticaceae/Moraceae are important between 1000 m and 1600 m elevation. Pollen of Hedyosmum, Alnus, Poaceae and Combretaceae/Melastomataceae are abundant between 1600 m and 2000 m. Cecropia pollen dominates samples from low- to mid-elevation disturbed forests. Alnus pollen is most abundant, and Poaceae becomes rare, between 2000 and 2700 m. At high elevations above 2700 m, Asteraceae, Poaceae, Polylepis, Muehlenbeckia-type and Myrsine pollen are dominant. Statistical analysis of the data set using Detrended Correspondence Analysis (DCA) shows a precise correlation between community composition and elevation. The DCA axis 1 values are strongly correlated with sample elevation, exhibiting a linear relationship (r2=0.904). The results provide an estimate of the sensitivity of pollen analysis in the Neotropics as a proxy for measuring elevation and, by inference, temperature.

In temperate trees, the extension growth history of stems and branches is easily inferred from the positions of the rings of over-wintering bud-scale scars, allowing one to assess environmental responses from surveys of saplings (Beaudet & Messier 1998, Canham 1988, King 2001). But tropical trees rarely show such distinct demarcations of growth flushes. Furthermore, the frequency of flushing may vary with tree size and environment, and a number of tropical species, often light demanding, show continuous leaf production and extension growth throughout the year (Coley 1983).

This study describes a pollination system in a species of Araceae that involves three species of beetle, one of which is also a fruit predator. In a tropical cloud forest in Colombia, inflorescences of Xanthosoma daguense opened at dusk, releasing a sweet scent and raising their temperature 1–3 °C. Soon after, two species of Scarabaeidae (Dynastinae; Cyclocephala gregaria and C. amblyopsis) and one species of Nitidulidae (Macrostola costulata) arrived with pollen. Cyclocephala beetles remained inside the inflorescence for 24 h. The next night, Cyclocephala beetles left the inflorescence after picking up the freshly shed pollen, almost always moving to the nearest inflorescence available. The probability of inflorescence abortion and number of fruits set after the visit of one individual was equivalent for both Cyclocephala species. However, C. gregaria was much more abundant than C. amblyopsis, so it was the most important pollinator. There was a positive relationship between the number of dynastine visits and the number of fruits produced. Besides carrying pollen to the inflorescences, nitidulid beetles had a negative effect on female reproductive success through fruit predation. Nitidulid larvae developed inside the infructescence and preyed on up to 64% of the fruits. However, 8% of inflorescences not visited by dynastines were probably pollinated by nitidulids, because hand-pollination experiments showed that self-pollination was unlikely. Inflorescences potentially pollinated by nitidulids comprised 25% of the fruit crop in the year of our study. This interaction with a fruit predator that is also a potential pollinator resembles brood-site pollination systems in which pollinators prey on part of the fruit set (e.g. Ficus, senita cacti, Yucca), making this system substantially more complex than previously described dynastine-pollinated systems in aroids.

In neotropical forests many species of plant depend on animals for pollination and seed dispersal and it has been well documented that bats play an essential role in dispersal of many flowering plants (Gardner 1977). Bats are responsible for colonization of plants into forest gaps because they often disperse the seeds of plants adapted for growth in disturbed areas. Species of Cecropia, Piper, Solanum and Vismia are especially important pioneer plants, and bats play a critical role in the dispersal of these secondary woody species into both secondary and primary forests, and into the seed bank (de Foresta et al. 1984, Lobova et al. 2003).

Variation in the ant species associated with myrmecophytes (plants that provide their associated ants with nesting space, and sometimes with a complete diet) has been noted at both the regional and local levels, with plant distribution generally wider than that of the ants (Fonseca & Ganade 1996). This is the case for Cordia nodosa Lamark (Boraginaceae, subfamily Ehretioideae) whose most frequent associate ant species in Peru are Allomerus demararae (Wheeler) and Azteca spp. (Yu & Pierce 1998), while Azteca sp.1 and Allomerus octoarticulatus are the most frequent in Amazonian Brazil and French Guiana, respectively (Fowler 1993, Solano et al. 2003). Cordia nodosa plants are understorey treelets mostly less than 2 m tall, but taller individuals can be found. Their domatia are swollen, hollow stem nodes that form as the growing shoot tip invaginates through a subapical pore that then closes over when the domatium is mature (Yu 2001). So, plant-ants must reopen the chamber where the pore used to be, thus forming a kind of prostoma.

Habitat fragmentation causes a sharp increase in the forested area affected by edge effects (Murcia 1995). Compared with the core of the forest, edges present higher litterfall rates, air and soil humidity reduction, higher temperature and increased wind incidence (Bierregaard et al. 1992, Didham & Lawton 1999, Laurance et al. 1998, Matlack 1993). These features of edges may increase the probability of fire occurrence, especially if the surrounding vegetation is composed of grasses (D'Antonio & Vitousek 1992, Freifelder et al. 1998). After a fire, the damaged edge will grow substantial amounts of herbaceous vegetation, extending the inflammable area into the forest, and thereby creating a positive feedback system of fire susceptibility and intensity (Cochrane & Schulze 1999, Cochrane et al. 1999). Fires may affect the structure and composition of the vegetation (Cochrane & Schulze 1999, Didham & Lawton 1999, Sanaiotti & Magnusson 1995), favouring, in the long-term, the dominance of the community by species characteristic of disturbed habitats and making the habitat unsuitable to primary forest species (Malcolm 1994, Possingham et al. 1994). Unfortunately, there are few studies on the effects of fires on neotropical small-mammal populations (Borchert & Hansen 1983, Ojeda 1989, Vieira & Marinho-Filho 1998). The objective of this study was to analyse the influence of a fire on populations of two rodent species, Akodon cursor (Winge) and Oecomys concolor (Wagner) in two fragments of Atlantic Forest in Brazil.

The Old-World tropics encompass one of the floristically richest zones of the world and some of the hot spots of ant diversity. This results in a large variety of ecological interactions between both groups. One of them is the phenomenon of myrmecochory, seed dispersal by ants, which is also well known from temperate forests (Gorb & Gorb 2003, Ulbrich 1919), and which is most prominent in sclerophyll shrublands of Australia and southern Africa (Andersen 1988). Beattie (1983), who reviewed the distribution of ant-dispersed plants (at least 80 plant families worldwide) proposed that species richness and abundance of myrmecochores and diaspore-dispersing ants increases with decreasing latitude and thus predicted a greater variety of ant-dispersal systems in the tropics. However, up to now, few tropical myrmecochores have been described (Horvitz 1981, Horvitz & Schemske 1986), especially in the palaeotropics (Kaufmann et al. 2001). Here we report myrmecochory in two species of rain-forest herb of the Zingiberaceae, give the first evidence for seed dispersal by ants in this plant family and present a list of seed-dispersing ant species. An important benefit of myrmecochory is the dispersal distance of the ant-transported seeds (Andersen 1988), that has been found to be positively correlated with ant size (Gomez & Espadaler 1998a, Pudlo et al. 1980). In this study, we checked whether this correlation is also true for the conditions of the tropical rain forest, where Globba plants occur.

The production of domatia is a crucial step in the evolution of myrmecophytism (Fiala – Maschwitz 1992). Current models for the evolution of domatia predict that the ancestral state is one in which domatia production occurs later in a plant's ontogeny, while more derived ant-plants produce precocial domatia (Brouat – McKey 2000). I report that the neotropical tree Cordia gerascanthus L. represents a case of the former, in which domatia appear in mature trees. Through field and herbarium investigations, I have discovered that some C. gerascanthus individuals produce swollen, hollow, stem cavities that are ant-inhabited, while other C. gerascanthus individuals produce only solid stems. I suggest that C. gerascanthus, a species that produces domatia later in ontogeny, is an ideal system in which to investigate the basis of domatia evolution.

Many rain-forest trees in South-East Asia, including the dominant canopy family Dipterocarpaceae, reproduce in gregarious mast-fruiting events once every 2–11 y (Ashton et al. 1988). The dominant hypothesis for the evolution of masting is that predators are incapable of consuming mast seed crops, so that natural selection has favoured parent trees that fruit in synchrony (Janzen 1974, 1976). Mast flowering and fruiting are visually spectacular and quantified in harvest records for dipterocarp species producing large, oil-rich tengkawang seeds (Curran et al. 1999). Seedling recruitment following a mast is less obvious and has no immediate commercial value. However, a number of pulses of recruitment have been documented (Ashton et al. 1988, Chan 1980, Fox 1972, Liew & Wong 1973). These, together with general acceptance of the satiation hypothesis, have led to the widespread assumption that masts reliably increase seedling density and generate distinct seedling cohorts (Whitmore 1998). Indeed, foresters in Malaysia and Indonesia often recommend harvesting only after a mast, to ensure high densities of seedling regeneration.

Over recent decades, studies on post-dispersal seed predation have revealed some general patterns despite dealing with a highly variable phenomenon (reviews in Crawley 1993, Hulme 1998). Most of these studies, however, were carried out with plant species whose population density range from low to moderately high (Hulme 1998). Only rarely has seed predation been studied in very dense stands of a given plant species (Schupp 1988a, b, 1992). In such a situation, seed predators might respond to the local availability of seeds in a manner that differs from their response under conditions of low seed density. For instance, a high density of seeds beneath a fruiting plant may attract seed predators, causing a positive correlation between predation and seed density (Janzen 1970). Schupp (1992), however, noted that at a population scale, a high density of seeds could satiate predators, resulting in a negative correlation between predation and seed density (see also Burkey 1994).

The ecology and forest management of tropical montane cloud forests of the Neotropics have attracted little scientific attention so far, and understanding of the ecosystem is still fragmentary (Churchill et al. 1995). The montane forests of southern Ecuador are an outstanding biodiversity hotspot for vascular plants (Barthlott et al. 1996), including a multitude of tree species (Madsen & Øllgard 1994). However, identification of species, studies of phenology, productivity, seed production and growth of seedlings have only just begun (Homeier & Breckle 2002). None of the trees has been investigated with respect to their mycorrhizal status, although it is well established that mycorrhizal symbioses are not only important for survival of trees in nutrient-poor habitats (Read 1991) and play a key role for nutrient cycling and nutrient retention in the humus layers (Medina & Cuevas 1993, Rilling et al. 2001), but also have an important impact on the composition of plant communities (Allen et al. 1995, Kottke 2002, van der Heijden et al. 1998).

Tropical cyclones, which are frequent along the north-eastern Australian coast, can result in severe disturbances to rain forests in the region (Grove et al. 2000, Webb 1958). Branch breakages and tree falls result in high levels of light penetration to the forest floor, which is normally heavily shaded (Turton 1992). This change in microclimate stimulates the growth of normally suppressed seedlings, the germination of seeds that are triggered by sunlight (Chazdon 1988), and often, invasion by weeds. Fragmented rain forests, that are common in the region, are particularly vulnerable to impacts of cyclones because of their large edge to forest area ratio. Appropriate management of such rain forests, following catastrophic disturbance, requires a thorough understanding of recovery processes at a number of temporal and spatial scales (Grove et al. 2000).

In forested ecosystems leaf litter is generally the primary pathway through which nutrients are cycled from the canopy to the forest floor (other pathways include throughfall, stemflow and animal faeces; Jordan 1985). Consequently, any disturbance that alters the quantity or quality of litter can have dramatic impacts on nutrient cycling and the availability of essential nutrients to plants (Vitousek 1984). Fragmentation of tropical forests has been demonstrated to cause several changes in both the biotic (Cosson et al. 1999, Laurance et al. 1998, Saunders et al. 1991) and abiotic environments (Camargo & Kapos 1995, Debinski & Holt 2000, Laurance 2002, Laurance et al. 2002) and thus may influence litter accumulation in the remnant patches (Carvalho & Vasconcelos 1999, Didham 1998, Laurance et al. 2002).

Prey depletion is a major (but often neglected) factor driving the current decline of tigers (Panthera tigris L.) (Karanth & Stith 1999). Thus conservation planning for the tiger in a fragmented, human-dominated landscape requires reliable information on prey densities from various parts of its range. We report numerical and biomass density of six herbivores from the tropical dry-deciduous forests of Ranthambhore Tiger Reserve (RTR, 26°N and 76°E), Rajasthan, India, where these are important prey for tigers (Bagchi et al. 2003).

Figs (Ficus species) are an important resource for a diverse array of organisms in most tropical forests (Janzen 1979 but see Gautier-Hion & Michaloud 1989). The inflorescence of Ficus, hereafter referred to as the ‘fig’, is an enclosed receptacle lined with unisexual flowers. The flowers of Ficus species are pollinated by wasps that feed on galled fig ovules as larvae and that lay eggs in fig flowers as adults (Weiblen 2002). Ripe figs are consumed by vertebrate frugivores, which are the primary dispersers of fig seeds (Shanahan et al. 2001). The interaction between figs, pollinators and frugivores introduces the potential for conflict between the roles of raising fig wasps and dispersing seeds. Specifically, the pollination mutualism could be compromised if frugivores consumed figs containing pollinator larvae. This conflict is resolved in very different ways according to the breeding system of the fig.

In a now classic study, Zinke (1962) showed that a single Pinus contorta tree growing on a sand dune along the coast of California modified the chemistry of the soil underneath its crown. He found distinct patterns of pH, exchangeable cations and nitrogen (N) content moving from the bole outward to the crown drip zone, because the acidic bark and stemflow were concentrated around the bole (Zinke 1962). Subsequent studies in temperate forests have also found tree species to affect soil chemical properties such as pH, organic carbon (C) and rates of N mineralization (Boerner & Koslowsky 1989, Boettcher & Kalisz 1990, Finzi et al. 1998). Presumably, these species-specific effects are caused by inter-specific differences in organic acid exudation, nutrient uptake, litter quality or quantity, decomposition rates or nutrient outputs (Binkley & Giardina 1998, Knops et al. 2002, Rhoades 1997). Regardless of the causes, species-generated soil heterogeneity has implications for stand-level estimates of biogeochemical processes such as soil C storage and N-cycling as well as implications for plant diversity and regeneration (Finzi et al. 1998). Although a number of studies have demonstrated that tree species modify soil environments in temperate forests or monospecific tree plantations in the tropics (Fisher 1995, Rhoades 1997), few studies have investigated these processes in species-rich tropical forests (but see Rhoades et al. 1994).

Relatively little is known of the relationships between many lepidopteran species and their larval food plants in tropical rain forests, compared with temperate ecosystems. Species are often known only from the adult form, and the larvae and aspects of ecology, host plant, etc. are unknown (Williams 1971). Many species in the moth genus Achaea sometimes occur at high enough population densities to be defoliators, and several species are associated especially with forest trees: some are agricultural pests on groundnuts and castor oil plant (Ricinus communis) (Pinhey 1975, M. Shaw, pers. comm.). Achaea catocaloides occurs from West Africa through Uganda to East Africa (Seitz 1925) and has been recorded in interactions with ants in Cameroon (Dejean et al. 1991).

Most savanna areas are affected by fire (San José & Fariñas 1983, Scholes & Archer 1997), and presumably all of the fauna that normally occurs in them have characteristics that allow them to maintain populations despite regular fires. Fires are frequent in Amazonian savannas and the vegetation is fire adapted. The Amazonian savanna at Alter do Chão suffers regular fires that affect vegetation structure (Sanaiotti & Magnusson 1995). Fires pass quickly in this area, the vegetation recovers most of the pre-fire cover within 6 mo, and there appears to be no long-term effect on soil structure. Trees are generally little affected by fires and they retain most of their foliage after the fire passes.

Rainfall fluctuation influences the physiology and leaf production of the trees in tropical rain forests (Burslem et al. 1996, Condit et al. 1995, Reekie & Wayne 1992, Reich 1995), and consequently it may have an impact on the community of herbivores that depend on forest trees for their food resources (Aide 1988, 1992; Basset 1991, Coley 1998, Lowman 1985, van Schaik et al. 1993).

Tropical forests can recover after anthropogenic disturbances of light to moderate intensity (Aide et al. 1995, 1996, 2000; Uhl 1987, Uhl et al. 1988); however, severe disturbances (e.g. compaction or loss of soil) often result in conditions that prevent forest recovery. These degraded sites are often dominated by grasses (Aide & Cavelier 1994, Cabin et al. 2002, Cavelier et al. 1998, Uhl et al. 1988) and ferns (Cohen et al. 1995, García et al. 1994, Slocum et al. 2000, Walker & Boneta 1995) that can impose barriers for tree regeneration and arrest the succession process. Important barriers for tree regeneration include: (1) competition with grasses and ferns for soil moisture, nutrients and light (Aide & Cavelier 1994, Guimarães-Vieira et al. 1994, Holl et al. 2000, Nepstad et al. 1996, Russell et al. 1998, Walker 1994, Zimmerman et al. 2000), and (2) dispersal limitation given that grasses and fernlands offer few resources that attract seed dispersers (Guevara & Laborde 1993, Holl et al. 2000, McDonnell & Stiles 1983, Slocum & Horvitz 2000, Zimmerman et al. 2000).