Vulnerability Of Biodiversity In Tropical Estuaries

13.7.1 Land-use change

A 1991 workshop on the status of mangroves of Southeast Asian coastlines (Sasekumar 1993) reported that the region has lost large areas of mangroves in the Philippines (80%), Thailand (50%), Indonesia (50%) and Malaysia (32%). This pattern is likely to continue as greater demands are placcd on forest and fishery resources, along with land-use changes along coastlines and in upland watersheds; the result will necessarily cause a change in the ecological characteristics of tropical estuaries. Many of the species guilds and biodiversity components described above are sensitive to changes in physical conditions (salinity, turbidity), chemical balances (eutrophication) and biological changes (exotic species).

Indirect loss of mangrove biodiversity components has resulted from human alterations of upland watersheds causing rediversion of freshwater (dams and canals), and deterioration of water quality from input of toxic materials (heavy metals, oil spills, pesticides) and nutrients to rivers and coastal waters. Regional-scale changes in freshwater surface inflow into mangrove areas are associated with reduction in secondary productivity of tropical estuarine ecosystems due to degradation of habitat and water-quality of those ecosystems. Changes in species composition of mangrove communities alters the quality of leaf litter, resulting in different rates of decomposition and an altered quality of organic matter export (POC vs. DOC) to the adjacent estuary (Boto and Bunt 1981; Twilley 1985, 1988; Snedaker 1989). Species substitution along zones of edaphic conditions is limited in mangroves due to narrow species-specific tolerances (Rabinowitz

1978; Lugo 1980; Snedaker !982); therefore, eliminating a given species may alter specific types of refugia available to consumers (e.g. species with prop roots vs. those with pneumatophores).

River (and surface runoff) diversions that deprive tropical coastal deltas of freshwater and silt result in losses of mangrove species diversity and organic production, and alter the terrestrial and aquatic food webs that mangrove ecosystems support. Freshwater diversion of the Indus River to agriculture in Sind Province over the last several hundred years has reduced the once species-rich Indus River delta to a sparse community dominated by Avicennia marina; it is also responsible for causing significant erosion of the sea front due to sediment starvation and the silting-in of the abandoned spilt rivers (Snedaker 1984a,b). A similar phenomenon has been observed in southwestern Bangladesh following natural changes in distributary rivers of the Ganges and the construction of the Farakka barrage that reduced the dry season flow of freshwater into the mangrove-dominated wester Sundar-bans. Freshwater starvation, both natural and man-caused, has had negative impacts on the rich vertebrate fauna (e.g. arboreal primates, deer, gavial, large cats) of the Ganges River delta (Hendricks 1975; Das and Siddiqi 1985). In the delta region of the Magdalena river, rediversion of freshwater has resulted in the loss of about 50% (21 778 ha) of mangroves in the Santa Marta lagoon (Ciénaga Grande) (Botero 1990). The loss of mangroves and decline in water quality are associated with the loss of fishery resources in this region. These case studies demonstrate the sensitive nature of mangrove ecosystems to changes in freshwater diversion, particularly in dry coastal climates. They represent examples where reductions in the various biodiversity components change the function of tropical estuarine ecosystems.

The coupling of mangroves to coastal waters is considered to be the most important link in sustaining commercial and recreational fisheries that are associated with estuaries and related nearshore marine habitats. Utilization of mangrove as forest plantations promotes sustainable use of this valuable resource for forest products such as timber, fuelwood, tannins, pulpwood and charcoal (see Watson 1928; Saenger et al. 1983), albeit only, until recently, in the Old World tropics (Snedaker 1986). Recent forms of direct exploitation include the destruction of biodiversity components of mangrove forests by land uses such as aquaculture (shrimp ponds), agriculture (rice and salt ponds), urban development and forest clear-felling for economic gain and other purposes (vide Pannier 1979). The so-called "soil reclamation" projects in Africa, as well as in parts of Asia (cf. Ponnam-peruma 1984), for agriculture (and aquaculture) have reduced regional levels of coastal productivity owing to loss of mangrove habitats. In many instances the conversin of organic-rich, pyritic mangrove soils leads to the formation of acid sulfate soils that are extremely difficult to further reclaim or to make support the original diversity of the landscape (cf. Dost 1973; Moorman and Pons 1975).

13.7.2 Global climate change

Tropical estuarine ecosystems are also vulnerable to changes in coastal environments due to the global perturbations resulting from increased greenhouse gases in the atmosphere. Mangroves occur at the interface between land and sea, and therefore are very sensitive to changes in sea-level. COt and other greenhouse gases may double by the year 2050 as compared with the amounts present at the start of the industrial revolution, warming the earth's surface between 2 and 4°C. If the average temperature increases by 3°C by 2050 and remains constant thereafter, the sea level will probably rise approximately 1 m by 2100 (50-91 cm per 100 years); a global warming of 6CC by 2100 could result in a sea level rise of 2.3 m (>100 cm per 100 years; Intergovernmental Panel on Climate Change 1990). These figures represent an increase over present rates of sea level rise, and are important relative to the rise in sea-level rates observed during the late-Holocene phase (Scholl et al. 1969; Parkinson 1989; Wanless et al. 1994).

There is much controversy over the threshold level of sea-level rise that mangroves can tolerate. Scholl and Stuiver (1967) and Parkinson (1989) have demonstrated that mangrove peat production and accumulation rates were unable to keep pace with a rising sea-level of 27 cm per 100 years, and mangrove colonization was maximum during periods when sea-level rates decreased to 4 cm per 100 years (Scholl 1964a,b). Ellison and Stoddart (1991) reviewed Holocene stratigraphic records and sea-level change for a number of sites worldwide and emphasized that mangroves in low islands can only keep up with a sea-level rise of up to 8-9 cm per 100 years, and rates of 12 cm per 100 years will collapse these systems. However, Woodroife (1990) showed that mangroves in Belize and Jamaica, characterized by autochthonous sediment, have persisted for 6000 years. He suggests that mangroves may be able to keep pace with rates of sea-level rise of 50-80 cm per 100 years for short periods. In addition, mangroves in Key West, Florida, have expanded both seaward and landward in the last 56 years in spite of a rise in sea level equivalent to about 23 cm per 100 years (Maul and Martin 1993; Snedaker et al. 1994).

Changes in the species richness of mangroves during horizontal migration inland in response to changing sealevel will depend on the species-specific responses of mangroves to increased inundation and erosion (Clarke and Hannon 1970; McMillan 1971; Ellison and Farnsvvorth 1993; McKee 1993), and effects of propagule size to tidal sorting along the intertidal zone (Rabinowitz 1978; Jiménez and Sauter 1991). Both of these factors indicate that the depth of tidal inundation will be a primary factor in regulating the species zonation with rise in sea level. Most studies summarize that Rhizophora is more tolerant of low oxygen availability caused by tidal inundation and waterlogging than Avicennia. If no inland barriers are behind mangroves, mangroves would migrate inland facing a rising sea le%'el. Assuming other ecological factors keep relatively constant, Rhizophora, with larger propagule size and higher tolerance to inundation, will invade and dominate the higher zone previously occupied by Avicennia and Laguncu-laria. Avicennia and Laguncularia will retreat to newly formed saline, shallow intertidal areas, and the fringe mangroves, basically consisting of Rhizophora, will eventually disappear (Snedakcr 1993). Bacon (1994) argues that most predictions of how wetlands in the Caribbean will respond to rise in sea level are too simplistic because they do not account for the site-specific responses of wetlands to changes in hydrology.

Temperature is the basic climatic factor governing the northern and southern limits of mangrove distribution. The responses of mangrove forest to decreasing temperature are reductions in species richness (Tomlinson 1986), forest structure (Lugo and Patterson-Zucca 1977), forest height (Cintron el al. 1985) and biomass (Twilley el al, 1992; Saenger and Snedakcr 1993). Although mean air or water temperature show some correlation with mangrove distribution in the world (Chapman 1977; Tomlinson 1986; Clough 1992), extreme temperature may be the principal controlling factor. In this regard, it has been suggested that the frequency, duration and/or severity of freezing temperature is a prime factor governing the distribution and abundance of mangroves in the northern Gulf of Mexico (Sherrod et al. 1986). Avicennia germinans and Laguncularia racemosa appear to be more tolerant to freezing temperature in the neotropics than Rhizophora (McMillan 1975; Lugo and Patterson-Zucca 1977; Sherrod and McMillan 1981; McMillan and Sherrod 1986; Sherrod et al. 1986; Olmsted et al. 1993). The greater resprouting ability of Avicennia and Laguncularia results in greater recovery from freeze damage (Sherrod and McMillan 1985; Snedaker et al. 1992; Olmsted et al. 1993). The different tolerance to low temperature among individual mangrove species is usually inferred from their natural distribution and morphological adaptation. However, genetic diversity has been demonstrated to influence the tolerance of mangroves to chilling (McMillan 1975; Markley et al. 1982; Sherrod and McMillan 1985; McMillan and Sherrod 1986; Sherrod et al. 1986). For example, analysis of isozyme patterning in Avicennia germinans revealed a divergence of phosphoglucose mutasc and phosphoglucose isomerase among the Gulf of Mexico-Caribbean populations (McMillan 1986).

Several studies indicate that the frequency and intensity of tropical storms and hurricanes are likely to increase under warm climate conditions (dcSylva 1986; Emanuel 1987; Hobgood and Cerveny 1988). Since mangroves are distributed in latitudes where the frequency of hurricanes is high, it is important to understand how tropical storms and hurricanes affect forest development, (i.e. forest structure, species composition, etc.) and community dynamics, including biodiversity, in mangrove ecosystems. Yet this type of information is very limited, and restricted to a few areas in tropical and subtropical latitudes. For example, studies in Florida (Davis 1940; Eglcr 1952; Craighead and Gilbert 1962; Alexander 1967), Puerto Rico (Wadsworth 1959; Glynn et al. 1964), Mauritius (Sauer 1962) and British Honduras (Vermeer 1963; Stoddart 1969) describe the effects of hurricanes on defoliation and tree mortality, but most of the information lacks quantitative assessments of the damage. Recent studies in Nicaragua (Roth 1992) and Florida (Smith et al. 1994) have provided a more quantitative evaluation of the effects of hurricanes on mangrove forests. High density of seedlings and fast recovery of mangroves in Isla del Venado, Nicaragua, suggest that they are not threatened as a community from hurricane damage. Ogden (1992) pointed out that mangrove forests in Florida will be able to recover following Hurricane Andrew, but Smith et al, (1994) were more cautious based on the dynamic role that soil status (e.g. redox and sulfide concentrations) can have in controlling tree growth and development.

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