Diversity and ecosystem function near steady state

There is substantial evidence from experiments and from agriculture, the latter derived from studies of monocultures versus mixed cropping systems, that two or more spccies that differ enough in their resource requirements can produce more biomass or retain more nutrients when grown in combination than can either species grown alone (Trenbath 1976; Ewe) et al. 1991; Swift and Anderson 1993; Fownes 1995). To our knowledge, such effects have not been demonstrated in any intact natural system, and it would be difficult to do so, because the spatial and temporal heterogeneity in ccosystcm properties and processes is large, and the precision of measurements necessary to demonstrate subtle effects is difficult to achieve.

Nevertheless, the striking pattern of decreasing species diversity with increasing distance from source areas that Mueller-Dombois (1990) described for Pacific high islands represents the best test-case we are likely to find for examining the direct effects of diversity on ecosystcm function in natural ecosystems. Mangroves in particular are a well-defined group of plants that can play an important role in ecosystem function. Woodroffe (1987) demonstrated that their diversity decreases markedly from west to east across the Pacific, from ~30 taxa in Papua New Guinea to no native taxa in the Society Islands (Figure 10.1). Detailed work on how this decline in species diversity might affect ecosystem properties and processes has not been reported, although Woodroffe (1987) did explore the possibility of a decline in annual litterfall in mangrove ecosystems in less diverse sites. The sparse data available did not suggest a decline.

On a broader scale, comparisons of annual above-ground litterfall (the most easily and widely measured component of net primary production) on islands versus more diverse continental systems at similar elevations suggest that the island systems are generally less productive (Tanner 1977; Veneklaas 1991). However, there is no evidence that diversity itself is the controlling factor. No effort has been made to examine production along a gradient in diversity across islands, and moreover, alternative explanations for the differences are tenable. For example, the prevalence of maritime cloud layers, and perhaps relatedly the well-known (but not fully explained) altitudinal compression of elevational zones on isolated mountains (the

Massenerhebung effect) (Grubb 1977; Bruijnzeel et al, 1993), could cause the observed pattern independent of any effects of diversity. Overall, we know of no clear evidence for direct effects of diversity on the functioning of island ecosystems near equilibrium, and we are not aware of any systematic attempts to gather such evidence.

10.2.2 Disturbance, diversity and ecosystem function

Can biological diversity reduce variability in ecosystem structure and function that might otherwise occur in response to environmental fluctuations, directional change or disturbance? Logically, we would expect that some of the species or populations in a community might respond positively to any given change, while others would respond negatively, and that a highly diverse community would be more likely to contain specics and populations capable of responding positively to a wide range of changes and perturbations. Therefore, it would be reasonable to expect that the rates of ecosystem processes might vary less following disturbance in more, as compared with less, diverse ecosystems, and/or that they might recover to prcdisturbance levels more rapidly (Lawton and Brown 1993; Vitousek and Hooper 1993).

There is some empirical support for this interaction between diversity and ecosystem function. MacNaughton (1977, 1993). and more recently Tilman and Downing (1994), demonstrated that productivity decreased less in response to drought or other environmental perturbations in diverse than in simple systems. The initial variation in diversity in these studies came about as a result of environmentally driven site-to-site variation (MacNaughton 197?, 1993), or as a response to experimental manipulation of another factor (Tilman and Downing 1994); it would be more satisfying to observe the same result in sites where diversity itself has been manipulated as an experimental factor. As discussed above, islands offer the closest approximation of a natural experiment that varies species-level diversity directly - is there relevant evidence from islands?

The phenomenon of stand-level dieback in island forests may provide useful evidence. Montane areas on oceanic islands often support forests dominated by a single canopy species (e.g. Metrosideros polymorpha in the Hawaiian Islands, Scalesia pedunculata in the Galapagos), while continents and continental islands generally support more diverse forests in comparable environments. Montane forests in both Hawaii and the Galapagos have gone through striking episodes of stand-level dieback, in which most individuals of the dominant canopy species have died more or less synchronously over wide areas (Muellcr-Dombois 1987; Itow and Mueller-Dombois 1988). Mueller-Dombois (1995) concludes that these monodominant stands are made up of a singie cohort of trees that reach a susceptible lifestage together, and that some trigger (not always identified) can then cause synchronous dieback. Similar diebacks may occur on the specics level in more diverse island ecosystems, or on continents, but their ecosystem-level consequences are small because most of the species in diverse sites survive dieback of a component specics with little change in overall ecosystem function.

The ecosystem-level consequences of stand-level diebacks have not been determined directly on islands - often the phenomenon has not been recognized until it is past - but more predictable canopy diebacks in high-latitude continental areas have been shown to alter production, nutrient loss and soil nutrient availability at least briefly (Sprugel and Bormann 1981; Matson and Boone 1984). If stand-level diebacks in low-diversity island systems could be anticipated, it would be useful to determine their ecosystem-level effects directly.

It could also be rewarding to evaluate the dynamics of diverse and simple ecosystems following exogenous destructive disturbances, asking: Are more diverse ecosystcms less affected by such disturbance? Do they recover more rapidly following disturbance? Hurricanes frequently strike a wide range of islands that differ substantially in their native biological diversity (Mueller-Dombois 1995. Bowden 1995); studies of the resistance and resilience of various island systems following catastrophic disturbance would allow us to test the interactions between diversity, catastrophic disturbance and ecosystem function.

(0.2.3 Invasion, Extinction and Ecosystem Function

Are island ecosystems more easily invaded by alien species than continental systems? Are island species more susceptible to extinction in the face of anthropogenic change? To what extent do invasion and extinction alter ecosystem structure and function on islands, and why?

There can be no doubt that on average island species have proven to be more susceptible to anthropogenic extinction than have continental species. For example, birds are a well-known group in which most species have been identified and described, and most extinctions are documented. The overwhelming number of recent extinctions have taken place on oceanic islands (Wilson 1992), and island species contribute disproportionately to all lists of endangered birds, particularly those of the most profoundly endangered taxa. Other groups of organisms are similar. There is even evidence that species that are endemic to a single island or archipelago may be more vulnerable to local extinction than indigenous species that also occur elsewhere, as Adserson (1989) demonstrated in the Galapagos.

Not all island endemics are equally susceptible to extinction; some become aggressive weeds in human-disturbed areas, and others (like the red crab of Christmas Island) may even contribute strongly to biotic resistance to invasion by aiicn species (Lake and O'Dowd 1991; Cushman 1995). The reasons why so many species are susceptible to extinction are not all clear. Small population sizes and other demographic factors may be important, but also the disharmony of island biotas, their consequent lack of major functional groups of species, and their susceptibility to disruption as a result of biological invasion must also contribute.

Conclusions concerning how easily island ecosystems can be invaded are not equally solid, largely because it is difficult to know how often potential invaders reach either island or continental areas and then fail to establish populations. Without knowing the fraction of potential invasions that are unsuccessful, it is difficult to generalize about the ease of invasion (Simbcrloff 1986; D'Antonio and Dudley 1995). Generally, there is good evidence that a significantly larger proportion of the flora and fauna on oceanic islands (especially in protected areas) is made up of alien species than is the case on continents or continental islands, but the absolute number of alien species in the flora and fauna may be no greater on islands (Loope 1992; D'Antonio and Dudley 1995; MacDonald and Cooper 1995).

In contrast, there is good evidence that ecosystem-level effects of invasion of islands can be dramatic, while it is more difficult to demonstrate ecosystem-level consequcnces of extinctions. A number of biological invasions have been shown unequivocally to alter ecosystem properties and/ or processes on islands.

1. The invasion of the Macaronesian tree Myrica faya into Hawaii. Myrica faya, an actinorrhizal nitrogen fixer, colonizes nitrogen-deficient young volcanic sites in Hawaii. In the process, it increases total inputs of nitrogen into demonstrably AMimited ecosystems by more than four-fold, and alters soil fertility in the areas it invades in a way that favors colonization by additional alien species (Vitousek and Walker 1989; Aplet 1990; Walker and Vitousek 1991). While the Hawaiian flora contains a number of symbiotic nitrogen fixers, none appears capable of colonizing young volcanic sites.

2. The effect of the introduced house mouse (Mux musculus) on the populations and activity of a flightless, litter-feeding moth (Pringleuphaga marioni) on Marion Island in the Indian Ocean. P. marioni is the major contributor to the physical breakdown of plant litter on Marion Island, and its reduction as a result of predation by mice decreases rates of litter decomposition and could affect nutrient availability to plants (Crafford 1990; Smith and Steenkamp 1990).

3. Invasion by fire-enhancing grasses on many islands. A number of species of grasses, many of them African, can colonize open-canopied woodlands and shrublands, and in the process cause the accumulation of enough fuel to permit the spread of fire in areas that did not normally support it. Following fire, these introduced grasses recover more quickiy than native species, converting diverse woodland/shrubland systems into grassland and increasing the chances of subsequent fires (Hughes et al. 1991; D'Antonio and Vitousek 1992). Grass invasion/fire cycles are important in many continental as well as island areas, but the simpler bio tie background on islands makes the dynamics and consequences of such invasions more amenable to analysis.

In all of these cases (except perhaps the last), the invader represents a novel type of organism added to an island ecosystem, and hence potentially a new functional group added to an existing community.

Examples of the ecosystem-lcvel consequences of extinction have not been worked out to a similar level of certainty. This difference in part reflects the fact that biological invasion is often a dramatic, rapid phenomenon that can be evaluated as it occurs. In contrast, by the time we consider a species to be in danger of extinction, its effect on ecosystem-level processes may already be much reduced, and it may no longer be possible to obtain meaningful background data. Nevertheless, it is likely that anthropogenic extinctions have altered island ecosystems substantially. The following examples are particularly strong candidates for such effects.

1. The loss of nesting sea birds to introduced predators and human hunting. These birds must have been significant vectors for the movement of nutrients (especially phosphorus) from marine to oceanic island ecosystems (Bowden 1995, Given 1995).

2. The extinction of herbivorous and frugivorous birds, many of them flightless, that once represented the dominant herbivores of many oceanic islands. These birds must have had multiple effects on island plant communities; their loss may have inhibited the dispersal of some species while releasing others from grazing (James 1995).

3. The loss of symbiotic nitrogen-fixing plants, which are important in maintaining soil fertility, but whose high protein contents may have made them particularly palatable once humans brought generalist grazers to islands.

Finally, in many cases extinctions have occurred simultaneous with - and probably as a result of - invasions, making it difficult to determine the effects of extinction alone on ecosystem function.

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