Many shades of green: the dynamic tropical forest–savannah transition zones

(a) Evolution

The emergence of tropical savannah ecosystems as a significant global biome at the expense of tropical forests is a striking and relatively recent feature of Earth history. It seems broadly associated with the decline in atmospheric CO₂ and the resulting cooling and drying of the atmosphere (as a result of increasing amounts of water being locked up in ice caps) that has occurred since the Eocene (around 50 Ma) [22,23], but key transitions have not been synchronous, suggesting that multiple factors are involved [24]. Decreasing atmospheric CO₂ and increasing daytime water stress favoured the C₄ photosynthetic pathway, which is physiologically competitive but metabolically expensive so cannot be sustained in shady conditions [25]. The opening up of forests into predominantly C₃ grasses occurred in the Early–Middle Miocene (11–24 Ma) [24,25]. As C₃ grasses do not have an intrinsic photosynthetic advantage over C₃ trees, this transition is unlikely to have been triggered by atmospheric CO₂ falling below a critical threshold value. Rather, it seems likely that this transition was driven by an increasing fire regime in an increasingly arid tropical biosphere, and further reinforced by the coevolution of ungulate grazers and mixed grazers/browsers (e.g. horses, rhinos, antelope and elephants) adapted to exploit the abundant and easily available ground vegetation and, through browsing or trampling, tree-sapling survival and growth [24].

C₄ grasses were a feature of these early tropical grassland landscapes soon after their origination, especially in drier parts of floodplains, but did not rise to global dominance until much later (6–8 Ma), and notably only a few taxa (in particular the Andropogonae) were able to spread to global dominance [23]. This suggests that, although low CO₂ may have created the opportunity for C₄ grasses to be competitive, it was other traits than C₄ photosynthetic pathway that led to the global spread and dominance of selected C₄ grass species, including high wet season growth rates and slow decay rates (both generating high fuel load and greatly facilitating fire), protected buds, storage organs and an ability for rapid resprouting [9]. These C₄ grasslands constructed and promoted a highly flammability niche, largely displaced the previous C₃ grasslands, and also spread further into forests, pushing back the forest–savannah transition [23].

Throughout the Quaternary (2.6 Ma to Present), the forest–savannah transition has moved back-and-forth in broad synchrony with the ice sheets. During cold periods (globally drier and with CO₂ ∼ 200 ppm), savannahs have advanced into forests, and during warmer interstadials and interglacials including the Holocene (globally moister and with CO₂ ∼ 280 ppm) forests and woodlands have encroached into savannahs [26]. Local differences in hydrology, soil fertility and herbivory modified the patterns of advance. In West and Central Africa, there is evidence of extensive retreat of forests into wet refugia during glacial maxima [27]; in Amazonia and southeast Asia, the retreats were widespread but more probably more muted in extent [28]. Even in the relatively stable climate of the Holocene, there have been substantive shifts with tropical rainfall regimes associated with solar precession; in a peak dry period approximately 3000 years ago savannahs were extensive in the Congo basin [27], and both Africa and Amazonia have witnessed an advance of forest into savannah over the past few thousand years [28]. Since the Late Pleistocene humans have been an increasingly important modifier of this system, primarily through changing fire regimes (sometimes intensifying fire regimes, but at other times modifying fire regimes through management) and through declines in megaherbivores, most notably in South America and Australia but to some extent in Asia and Africa [21].

Hence the broad history of the forest–grassy vegetation transitions shows a highly dynamic interaction between trees and C₄ grasses, mediated at largest spatio-temporal scales by rainfall patterns and atmospheric CO₂ concentrations, but reinforced by feedbacks with fire and herbivory and plant adaptations to all these factors [29].

(b) Current sharp boundaries

One remarkable feature of transitions between tropical forests and savannahs is that they are frequently quite abrupt, as opposed to being a gradual gradation in tree cover. In mesic environments, some studies argue that forest and savannah stands may represent alternative stable states modulated by fire-mediated feedbacks [30–32]. The alternative stable states theory postulates that two ecosystem states can occur under the same set of environmental conditions, and the presence of one or the other is determined by disturbance and historical pathway at the community and larger scale [33].

In an alternative stable states interpretation of the forest–savannah boundary (figure 2), the broad-scale vegetation cover is set by precipitation regime and seasonal water supply, with low rainfall unable to sustain a closed canopy of woody vegetation. However, at intermediate rainfall amounts, both open-canopy and closed-canopy stands can exist [31], but the shift in canopy cover is often abrupt. One proposed mechanism for this abrupt threshold is the presence of two positive feedback loops associated with fire and/or herbivory (figure 2) [34,35]. Within the savannah ecosystem, the relatively open-canopy cover allows a dry microclimate with plentiful light that supports establishment of a light-demanding and stress-tolerant grass layer, which in turn favours consumers (recurrent fires and/or large herbivores) [36], that in turn prevent forest tree seedlings breaking through a fire trap or herbivore trap, favouring the continued presence of the grass layer. Within the forest ecosystem, the closed canopy inhibits the establishment of the light-demanding grasses in the understory and also maintains a humid microclimate, which in turn suppresses fire occurrence and intensity, allowing tree seedling establishment and the maintenance of a closed canopy. These two feedback loops tend to drive the spatial gradients in vegetation cover away from a smooth and gradual reduction in tree cover in favour of relatively sharp transitions to alternative states (figure 2).

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One of the main empirical pieces of evidence supporting the alternative stable states theory is that the frequency distribution of remotely sensed tree cover in the tropics is not a smooth unimodal function but has three distinct modes, which seem to correspond to forest, savannah and treeless states [31,37,38]. A recent study of field based-data on tree basal area for African and Neotropical forests and savannahs also found evidence supporting the tri-modality (grasslands, savannahs and forests) [37].

However, the remote-sensing based evidence is still controversial. Reported discontinuities on tree cover may be an artefact of the statistical and classification method used [39,40]. Furthermore, the satellite-derived tree cover product cannot detect variation in tree cover less than 10% and is therefore of limited applicability below 30% tree cover [41]. Other studies suggest that coexistence may be confined to a well-defined edaphic/climate envelope, with both soil and climate playing a role as the key determinants in the relative location of forest and savannah [37], and that other mechanisms related to strong climate control may also be operating [42].

There are still many question marks on the scale of operability of these factors [43–45] and on the hysteresis (historical pathway dependence) operating in these systems [44,46]. Whether a forest or savannah dominates at any particular point depends on microsite factors (e.g. soils, drainage) but crucially may depend on historical pathway. For example, a long past temporary release of herbivory or fire pressure may have enabled tree saplings to grow to a size where they could survive fire or herbivory [34,36]. Similarly, if the historical climate on a given place has become gradually drier over time, ecological inertia and feedbacks may maintain a forest even when current climate conditions favour a savannah [47]. The scale of operability of hysteresis and current drivers and processes varies regionally. For example, Lehmann et al. [48], in a global study that encompassed field and remote-sensing data for Australia, Africa and South America, found that the magnitude of main drivers in determining savannah woody structure varied substantially across continents. They concluded that evolutionary history and environmental differences are likely to drive the regional variation in the functional relationships between woody vegetation, fire and climate.

At local scales, the potential of coexisting alternative stable states would tend to increase with spatial heterogeneity. Small differences in contemporary abiotic factors (e.g. soils, drainage), biotic factors (e.g. herbivore behaviour) or in disturbance pathway can result in radically different canopy cover. When considering changes over time, the two stabilizing feedback loops described above and in figure 2 tend to inhibit continuous and gradual shifts back-and-forth in the boundary in response to climatic variability. However, this may increase the likelihood of abrupt changes in response to extreme events, such as successful fire penetration of the forest during an extreme dry spell. Indeed, the migration of these sharp boundaries over longer time scales is another strand of evidence in favour of the alternative stable states hypothesis, suggesting that the location of a particular vegetation type is not fixed by local soil conditions.

(a) Water availability and species' hydraulic strategies

Despite precipitation being one of the primary drivers of the distribution of tropical forest and savannahs [43,48], there are no simple threshold precipitation values that determine the transition. While closed-canopy humid forests tend to predominate at annual rainfall values above 1800 mm [31,37], both forests and savannahs can be found at intermediate rainfall regimes (800–2500 mm) [30] and even at lower precipitation regimes areas exist with dry forest patches mosaicking with dry savannahs. Rather than total precipitation, rainfall seasonality [45] and the depth of the seasonal precipitation deficit can be a better predictor of biome boundaries at a coarse scale [52]. These factors interact with local soil depth, texture, hydrology and topography to determine actual seasonal local water availability, leading to mosaic landscapes of forest and savannah.

How grasses, savannah and forest tree species deal with water availability is reflected through their hydraulic strategies. In the herbaceous layer, two different growth forms associated with drought adaptation are found in savannahs [51]: (i) the drought escaping ephemeral or deciduous perennial grasses, which are active only in the wet season, and survive the period of drought stress as seeds or by going dormant; (ii) the deep-rooted phreatophytes, that escape the drought by gaining access to the water table.

For shrub and savannah trees and forest tree species, hydraulic strategies are varied and complex. Savannah tree species tend to be drought-resisters, exhibiting isohydric behaviour, a conservative strategy limiting transpiration rates and stomatal conductance during periods of water stress to maintain water potential [53], and buffering water supply through tissue water storage [54,55]. Studies on dry forest species suggest that most species are either drought-avoiders (deciduous species with sophisticated root morphology to maximize resource capture during a limited growth season [56]) or evergreen drought-resisters that minimize transpiration rates and cavitation risk to persist under dry conditions [57,58]. A third group consists of light-demanding pioneer species with anisohydric behaviour (i.e. they do not regulate stomatal conductance in order to maintain high productivity, but at the risk of hydraulic cavitation), but these species are usually precluded from very dry environments [58]. Nevertheless, in mesic environments they may play a crucial role in colonizing savannah environments especially if the conditions during the dry season still allow them to operate below their safety margins (e.g. if they establish during periods without extreme drought events).

At the forest–savannah transition, congeneric forest and savannah trees tend to employ different suites of functional traits to deal with limited water availability [59,60]. Brazilian savannah species tend to have greater leaf area-specific hydraulic conductance and greater transport capacity on a leaf area basis than congeneric forest trees [59,60] and higher leaf water potential and stomatal conductance in the late dry season relative to forest trees [60], but there is little difference in wood density and minimum water potential [60]. The proportion of stem-stored water used appears to be larger in savannah tree species (16–31%) [55] than in tropical tree species (9–15%) [61], but there appear to be no reported values on stem water use from forest species that establish near savannah boundaries. Another important feature of species and the zone of transition is their water uptake strategy, which varies among species and vegetation types, and is mostly reflected in differences in root morphology and root water uptake. For example, in Brazilian cerrado, woody savannah vegetation has deep roots that can uptake up to 82% of water from below 1 m soil depth, while more grassy environments (campo sujo) take up about 67% of the water from below the same depth [62]. When grasses and woody vegetation share the same space, grasses can take up topsoil water sufficiently rapidly to reduce drainage into the subsoil, as well as take up subsoil water directly, thus decreasing water supply to woody plants and limiting seedling and sapling establishment [9,63].

(b) Fire

At the global scale, fire regimes (the spatial and temporal distribution of fires, and their intensity) are closely associated with different biomes, with tropical forests usually being fire-sensitive ecosystems and grassy biomes usually fire-dependant or fire-adapted ecosystems [64,65]. However, fire is not unilaterally responsive to climate or to vegetation [66]. Although the fire regimes characteristic of the modern world can be explained with reference to current vegetation and climate, there are complex interactions among fire, climate, vegetation and human activities. As such, a biome can contain different fire regimes, and contrasting biomes may effectively converge on a common fire regime [66].

C₄ grassy biomes support the highest fire frequencies on Earth [67], and fire stimulates flowering of many grassy flora (e.g. [68,69]). Within these grassy vegetation types, fire frequency and intensity is usually higher in humid savannahs than semi-arid savannahs [70] because of higher productivity resulting in higher fuel availability. The occurrence of fire at a given location will be determined by the availability and continuity of flammable dry biomass, which is largely determined by dry grass biomass (horizontal fuel continuity) but also by other variables such as canopy height and structure (vertical fuel continuity). Fire return intervals are a key aspect of these transitions, as they determine the operability of a two threshold phenomena in tree–grass interactions [34]. At the community scale, the ‘fire-suppression threshold’ exists. When tree canopy cover closes sufficiently (typically more than 45–50% tree cover [33]), it reduces the community flammability by increasing air humidity and inhibiting light-demanding grasses and resulting fuel loads, thereby reducing and disconnecting the distribution, intensity and propagation of fires. The level of this threshold depends on the fire return interval, as well as on several attributes of the adult tree species and herbaceous understorey species, such as tree crown size and density, and the shade tolerance of grasses or other herbaceous understory species [34]. At the level of individual saplings, a second threshold exists, the ‘fire-resistance threshold’. Saplings need to grow above the level of the grass to avoid topkill by fire—the destruction of above-ground biomass [34,71,72]—or complete mortality. The suppression of juveniles by repeated topkill results in a demographic bottleneck often referred to as a ‘fire trap’, which results in many individuals being trapped in the grass flame zone for many years (a phenomenon also known as the ‘Gulliver effect’ [73]). Recruitment into an adult size class is possible only if the sapling experiences a fire-free interval of sufficient duration to allow it to reach a critical size at which it is no longer susceptible to topkill [34]. Other factors such as nutrient availability, water supply, browsing and grass competition can interact with this fire trap if they increase the growth rate of saplings and thereby enhance their chances of escaping the fire trap in a particular fire interval.

The ability of herbaceous and woody species to resist or to recover from fire depends on four main mechanisms, which are not mutually exclusive: fast growth rates, structural protection, flammability and regeneration. As explained above, those individuals with faster growth rates are more likely to escape the fire trap and avoid topkill [34,72]. Many species have developed several structural adaptations to fire such as woody geoxylic suffrutices with enlarged underground xylopodia or lignotubers, thick corky barks, or thick shoots and leaves concentrated at shoot tips [12,74,75].

Flammability can be defined through three characteristics that define how well the fuel ignites (ignitibility), how well it burns (combustibility) and how long it burns (sustainability) [76,77]. Studies on flammability along forest–savannah transitions are still scarce, but some evidence suggest that fires in the savannah are faster and more intense than in the nearby forests, primarily driven by the dominance of grassy fuels [78]. However, grasses are not homogeneous fuels to fire. A recent study on the flammability of grasses from South African fire-prone environments showed different degrees of flammability driven by different functional traits and ultimately by evolutionary history [79]. Finally, regeneration will depend on the ability of the already established vegetation to persist after fire by resprouting, or by favouring new individuals to germinate (e.g. serotinous species, that require heat to germinate) [80–82].

There are strong feedback mechanisms between dry atmospheric conditions and fire. A drier environment (defined as an extended period without rain with high temperatures and low air humidity) increases ignition probability by rapidly drying vegetation and making a larger portion of the vegetation available to be consumed by flames. The high amount of fuel increases the chance of fires that consume a larger fraction of the vegetation (e.g. tree canopies) than a fire occurring under less extreme dry conditions. Under exceptionally dry conditions, fires may exhibit extreme behaviour, and in exceptional cases produce firestorms that create and sustain their own wind conditions though strong convective effects [83]. In these conditions, high-intensity fires may penetrate into closed-canopy environments, creating light gaps that favour grass invasion and tree mortality. For example, in an experimental treatment in southeast Amazonia, the combined effect of drought and fires has led to massive tree mortality, reducing original tree cover by half [84].

In these vegetation transitions, the interaction between water availability and fire has played a major role in the evolution of vegetation, such as the geoxyle life forms in African savannahs [12,13]. Water–fire–vegetation interactions are especially complex in forest–savannah transitions located in seasonally dry regions, and the frequency and intensity of both fire and dry periods often determine the current state of those transitions in terms of abruptness and species composition. For example, in regions where fire has been excluded, one usually finds a suit of relatively drought-tolerant forest species that are able to establish into the open savannah environments [85]. Similarly, a fire event during an intense drought period may stimulate cavitation of woody species [86], thereby killing more individuals than a fire in less extreme dry conditions.

(c) Soil

Over large regions, different soil types deriving from different substrate support different vegetation types. Soil in grasslands and savannahs are often poor compared to forests [43,87]. For example, South African savannahs have lower cation exchange capacity than Central African rainforests, but also lower than South American savannahs [88]. Similarly, it is well known that the soil fertility of seasonally dry forests is higher than savannahs under comparable climates in Brazil [89,90]. Soil cation status appears to be a key determinant of vegetation formation-type distributions across tropical South America [91].

At the community scale, soils can also vary spatially along the forest–savannah transition in fertility (e.g. organic matter, cation exchange capacity, macronutrients), soil physical properties (e.g. percentage of sand), and soil depth, the latter two factors combining to determine soil water availability. This variation can limit tree seedling establishment [92] and determine different below-ground biomass allocation strategies [93]. However, despite the fact that soil fertility has long been recognized to be lower in savannah soils, some studies suggest that savannah soils are indeed low in nutrients on the topsoil, but that deeper soil layers would have adequate nutrient stocks to sustain forest [92]. This could perhaps explain the ability of some woody species to establish in savannah soil, but given the diverse species-specific nutrient-uptake strategies [94] this deserves further investigation.

The interaction of soil fertility with water availability and soil water holding capacity can strongly influence forest–savannah boundaries [43]. A key feature of many seasonal wet–dry climates is the formation of impermeable hardpans at 1–2 m depth, which determine and limit soil depth. Forest patches are usually found located in areas with higher soil moisture than savannah patches, but in seasonally flooded parts of the landscape the waterlogging can create anoxic conditions for roots, favouring short-lived grasses that grow in the non-flooded season. Conversely, a more regular water supply and shallower water table enables gallery forests to persist in savannah landscapes.

At the local scale, higher levels of soil fertility are found beneath tree crowns in savannahs of Belize [95] and Brazil [96]. This small-scale soil differentiation influences the nutritional status and productivity of grasses in these environments [97] and may play a pivotal role in tree seedling establishment. Fire recurrence also has an effect on soil fertility. The effect of ashes deposited in the soil surface after fires tends to increase pH in places with high fire recurrences and provides a higher availability of cations in the soil [98].

(d) Carbon dioxide

CO₂ is a global driver that can potentially increase productivity, tree growth rates and water use efficiency [99–101]. Atmospheric CO₂ concentration has increased from around 280 ppm in the pre-industrial era to around 400 ppm by 2015, an almost 50% increase, and a number of laboratory and field studies suggest that this may be stimulating the photosynthesis and water use efficiency of C₃ plants [102].

At the local scale, higher CO₂ concentrations favour the C₃ photosynthesis pathway over the metabolically more costly C₄ pathway. Water use efficiency is increased because the stomata need to open less to allow a certain amount of carbon uptake, reducing water loss from the leaves per unit of carbon uptake [101]. The observed long-term increase in biomass observed in old-growth forest plots across the tropics may be caused by increasing atmospheric CO₂ [103], thus favouring trees over most abundant lowland savannah grasses [104]. For example, there is evidence that elevated CO₂ levels stimulate resprouting cerrado species [80], and stimulate growth in South African acacia species but do not stimulate growth of C₄ grasses [100,105]. At the regional scale, therefore, increasing CO₂ can be expected to favour both woody thickening of savannahs and forest encroachment into savannahs, by altering the fine competitive balance at the transition between trees and C₄ grasses (see ‘Contemporary changes in forest–savannah transitions’ section). The net impact may be greater in mesic than in arid savannah environments [104,106]. Such a broad-scale shift in relative competitiveness would interact with other processes, for example, increasing the likelihood of tree saplings escaping fire and herbivory traps.

(e) Herbivory

Large herbivores (mass 45–1000 kg) and megaherbivores (more than 1000 kg) can influence vegetation structure through destruction and opening-up of woody vegetation, through browsing and trampling of tree saplings, and through grazing of ground layers that can promote the success of grasses over tree saplings [21,107]. On the other hand, they can favour woody vegetation by facilitating large-scale seed dispersal through their dung, and grazers interact with fire regimes by reducing grass biomass fuel loads and fire return frequency [108]. Herbivory accelerates nutrient cycling by the break-up and consumption of tough vegetation and excretion of resulting products in more labile forms [21]. Megafauna also strongly facilitate the lateral transfer of nutrients across landscapes, creating high fertility latrine hotspots in the short term, but on longer timescales facilitating the diffusion of nutrients along concentration gradients [21].

Proboscideans (elephants) seem particularly important as keystone species that shape the savannah environment and the forest–savannah threshold. Exclosure studies in African savannahs show that African elephants (Loxodonta africana) reduce tree cover by 15–95% and are the primary cause of tree death, far exceeding fire and drought in importance as agents of adult tree mortality [109]. In Kruger National Park in South Africa, they uproot up to 1500 trees per elephant per year [110], However, a recent study in the Serengeti suggests that repeated, low-intensity damage from elephants was more important to mortality than acute but less frequent damage [109]. The effect of bush elephants on the actual forest–savannah transition is less clear. Forest dwelling species can also influence the vegetation transition. For example, the African forest elephant Loxodonta cyclotis favours trails along the forest–savannah boundary, behaviour which may facilitate dispersion but suppress seedling recruitment, and also act as a moderate fire break limiting fire penetration of forests (Y Malhi 2016, personal communication).

At near-equilibrium abundances, an abundance of large herbivores may stabilize the forest–savannah boundary, limiting forest seedling establishment in the savannah and decreasing risk of fire penetration of the forest by decreasing grassy fuel loads. It follows that a decline in large herbivores may increase the instability of the transition, facilitating woody encroachment of the savannah but also increasing fire intensity and frequency, with the net result dependent on local climate trends and fire regimes [111]. A high abundance of megaherbivores, as seen in Kruger National Park, may also destabilize the transition and cause reductions and retreat in tree cover [110].

High densities of herbivores are today mainly associated with African savannahs, but until 10–50 000 years ago similarly high animal biomass was likely found in all major tropical savannahs [112]. The decline and loss of these megafauna is generally coincident with the arrival of humans, with the most severe extinctions (Australia, Americas) occurring where human arrival was abrupt and without a long history of earlier Homo presence. Even in Africa, with a long history of gradual human evolution, early Homo may have been responsible for the decline in many megafauna over the Pleistocene, including the large grazing elephant Elephas iolensis that once dominated the grasslands, and perhaps confined the browsing bush elephant Loxodonta africana to woodlands. Globally, the Late Pleistocene hosted 50 species of megaherbivore (more than 1000 kg), compared with only nine remaining today (the African forest and bush elephants and Asian elephant, five rhinoceros and the hippopotamus, all of which are generally severely depleted in range and abundance [21]). Proboscideans are particularly important shapers of vegetation and were a dominant feature of savannahs in every continent except Australia. For example, South American savannah woodlands hosted Stegomastodon platensis (a browser, mass 5 tonnes) and Stegomastodon waringi (a grazer, mass 7.5 tonnes) [112]. South Asian savannah woodlands hosted various species of Stegodon (mass up to 13 tonnes). By comparison, the African bush elephant, the largest extant land mammal, has a typical mass of 4 tonnes. Australia, Darwin's ‘separate creation’, was distinct in being the only tropical continent lacking probiscideans but hosted large browsing diprotodons (mass 2.8 tonnes). Prehistoric biomass and abundance is challenging to estimate, but it seems likely that the abundances of these ‘lost elephants' were similar to those of pre-nineteenth century African bush elephants, with consequent impacts on vegetation structure, fire regimes and seed dispersal. The lack of megaherbivores that once were abundant may partially explain broad differences in the forest–savannah boundary (e.g. the greater extent of woody savannahs in South America versus open grassland woodlands in African and the different rainfall thresholds at which the boundary exists in different continents).

(f) Biological invasions

Another key feature of contemporary global change is the rapid global mixing of species, either deliberately or by accident through global transport. Because they are sites of frequent fire disturbance, savannahs are among the ecosystems most vulnerable to biotic invasions [113]. Invasive species can significantly alter ecosystem functioning, fuel loads and fire regimes, and create new pressures on forest–savannah transitions [113]. In fact, invasions seem to be facilitated in ecotonal environments [114], where invasive species can include both grasses and woody vegetation. Most invasive grass species are used in pastures as forage species, and their invasion to the forest–savannah areas is closely associated with fire (e.g. [115–118]). A large body of literature exists on the grass invasion of savannahs in the Neotropics and northern Australia, where African grasses in particular have had major impacts. For example, many African grasses have successfully invaded Neotropical regions and nowadays species such as Melinis minutiflora and several Urochloa species have altered many forest–savannah boundaries across the Neotropics [119,120]. These grass species are often being reported to be the main drivers of the ‘savannization’ process [19]. In Africa, most invasive plants are woody species, with Lantana camara and Chromolaena odorata being the most prominent species [111]. Andropogon gayanus, which is a native African grass, is one of the most widespread exotic grasses in Australian tropical savannahs, but its presence is widespread also in Brazil and other Neotropical countries [111].

(g) Other relevant ecological processes