Background

It is estimated that of the total global extent of tropical forest, tropical or subtropical dry forest accounts for 42 % of total global forest cover (Murphy & Lugo 1986). They have been exposed to severe, large-scale changes, through the cutting of valuable trees, creation of pastures, accidental or intentional fires (Gerhardt & Hytteborn 1992) and as a source of fuel wood (Murphy & Lugo 1986) and virtually all of the tropical dry forests that remain are currently exposed to a variety of different threats, largely resulting from human activity (Miles et al., 2006). However, despite their over-exploitation there have been relatively few studies of tropical dry forest and even fewer studies done on their regeneration pathways, which can provide knowledge crucial to the restoration of these forests (Vieira & Scariot, 2006).

Tropical dry forests have particular natural regeneration attributes which are not currently well understood (McLaren & McDonald, 2003b; Vieiria & Scariot, 2006). Unlike tropical moist forests, regeneration in gaps is not the primary mechanism; and the availability of moisture plays a major role in successful establishment of seedlings. The timing and duration of rainfall is crucial as there is a high probability that a given seedling will be left stranded in a drying soil and there is therefore a low probability of successful regeneration from seeds (Ewel 1980). The pronounced seasonality affects patterns of seed production, germination, survival and seedling development (Khurana & Singh 2000). Seeds of a majority of dry tropical species mature in the dry season and they are dispersed at the beginning of the rainy season when sufficient moisture is available for germination and seedling growth (Singh & Singh 1992; McLaren & McDonald, 2005). The favourable growing period is restricted to short rainy seasons when seeds are expected to germinate and seedlings establish. The deciduous state of some or most tree species allows for an increase in irradiance that in the absence of moisture exacerbates desiccation in seedlings and hence higher rates of mortality (e.g. Gerhardt 1996a).

Vegetative regeneration may well be the primary regeneration mechanism in disturbed dry forest sites, where stem and roots remain in place (Ewel 1977; Murphy & Lugo 1986; Murphy et al. 1995; McLaren & McDonald, 2003b). In the wet tropics, where large-scale disturbance occurs as a result of clearing, burning and extensive storm damage, regeneration from stem coppice is also important (e.g. Byer & Weaver 1977; Ewel 1977; Stocker 1981; Uhl et al. 1981; Putz & Brokaw 1989; Kauffman 1991; Bellingham et al. 1994). It may however, be more important in dry forest sites where resprouting offers considerable resilience to disturbance in dry forests where successful regeneration by seed is highly susceptible to rainfall seasonality (Ky-Dembele et al., 2007; McLaren & McDonald, 2003c; Vieira et al., 2006; Vieira & Scariot, 2006). Also, because trunk bases are less subject to rapid decay in the dry tropics, trees in seasonally dry forests are much more prone to reproduce vegetatively through coppicing (Ewel 1980) in response to disturbance. However, the different abilities of species to produce shoots will affect long-term species diversity in disturbed forests.

The sensitivity to moisture availability renders the regeneration of the dry forests highly vulnerable to predicted climate change effects. Most predictions of the response of tropical forests to rising CO2 concentrations and temperatures and changing precipitation patterns have concluded that changes in precipitation will have the most impact, in that drying trends will remove drought-sensitive species from the forest (e.g. Condit, 1998). However, a recent assessment of neotropical rain forests over a 20 year period showed an increase in faster- growing species to the detriment of slower-growing sub-canopy species, apparently irrespective of their light or moisture requirements (Laurance et al., 2004). This may be due to the fertilising effect of rising CO2 concentrations. However, this is uncertain, and strong droughts have been shown to cause shifts in tree-community composition in Panama (Condit et al., 1996) and niche differentiation with respect to soil water availability has been shown to determine distributions of tropical trees at both local and regional scales (Engelbrecht et al., 2007). Körner (1998) postulated that changing plant water relations could become the most important of all elevated CO2 effects on tropical forests, as water is always a selective driver of plant growth. The length of dry periods is among the key determinants of the species structure of communities (Condit, 1998; Körner, 1998). 97% of the remaining global dry forest is estimated to be ‘at risk’, and in the Americas this is mostly from climate change (Miles et al., 2006). Most studies on the effects of climate change in the tropics have been centred on tropical rainforest with a noticeable absence of work in tropical dry forest (Körner, 1998). Given that seasonally dry forests are so heavily impacted in the neotropics, and largely surrounded by agricultural landscapes, it is unlikely that they will be able to spread their range in the face of hotter and drier climates because of the lack of habitat in anthropogenic landscapes (Pennington et al., 2004). Mayle et al., (2004) demonstrated that dry forest species merely shift their ranges rather than expand them in the face of drier climates in the Bolivian Chiqitano region. If this is generally true, then the future of the seasonally dry forests is very bleak as they will “die where they stand” (Pennington et al., 2004).

Drying trends in a dry forest environment may create conditions similar to those more usually found in xeric tropical environments where plants are more drought-tolerant and seedling survival is enhanced by both the direct effects of habitat amelioration by shade, and the indirect effects of shaded plants attaining greater size before the onset of extreme conditions (Hastwell & Facelli, 2003). This is regarded as facilitation, the positive effect of plants on the establishment or growth of other plants (Holmgren et al., 1997). In the last decade, plant ecologists have focused more on the occurrence of positive plant-plant interactions than ever before, especially in severe environments. These studies have shown that facilitative effects are stronger in these environments, leaving little doubt of their generality and importance and raising questions about the assumed ubiquity of competition as the dominant interaction between neighbouring plants. Postulated relationships between facilitation and stress (Bertness & Callaway, 1994; Callaway & Walker, 1997) propose that, as conditions for plant growth becomes increasingly adverse, facilitation becomes ‘usually common’ (Bertness & Callaway, 1994) or that the ‘importance’ and intensity of facilitation increases. Dormann and Brooker (2002) suggest a model of plant interactions whereby the importance of facilitation increases along a gradient of increasing environmental harshness, while the importance of competition decreases (Figure 1). They apply this model to harsh environments, such as the Arctic, where species removal experiments tend to find facilitative rather than competitive effects. This phenomenon has also been recorded in more mesic environments which encounter periods of environmental severity, particularly drought. Hastwell and Facelli (2003) found that the relationship between facilitation and environment severity is more complex than previously thought, as neither the intensity nor the importance of facilitation necessarily increases as conditions become severe. Maestre and Cortina (2004) found that competitive interactions dominated at both extremes of an environmental gradient and suggested that a shift from facilitation to competition under high abiotic stress conditions is likely to occur when the levels of the most limiting resource are so low that the benefits provided by the facilitator cannot overcome its own resource uptake.