A central question in ecology and conservation is the degree to which populations are limited by bottom-up vs. top-down forces (Pace et al. 1999; Polis 1999; Terborgh et al. 1999; Hunter 2001). Recent studies have shown that top-down processes in terrestrial systems can have fundamental effects on population demography and community composition (Pace et al. 1999; Terborgh et al. 1999; Maron & Simms 2001; Ripple et al. 2001; Roemer, Donlan & Courchamp 2002). However, abiotic forces, such as El Niño Southern Oscillation (ENSO) events, are important community drivers (Meserve et al. 1999; Stapp, Polis & Sanchez Pinero 1999; Wright et al. 1999; Jaksic 2001). There is still only limited understanding of the degree to which interactions among species and the environment affect community processes (Oksanen & Oksanen 2000; Hunter 2001). A factor that has contributed to this limited understanding is the paucity of trophic studies that are both large in spatial scale and experimental in nature (Estes, Crooks & Holt 2001). Despite the challenges of large-scale experiments, they are both invaluable and irreplaceable in ecology (Likens 1985; Carpenter 1996).
Ecosystems are ‘the basic units of nature’ (Tansley 1935), and consequently the units of many applied ecological problems (Ormerod & Watkinson 2000). The importance of knowledge gained by large-scale and whole ecosystem experiments will increase as we adopt food web and ecosystem approaches to management and conservation (sensuPower 2001; Zavaleta, Hobbs & Mooney 2001). A collaboration between ecological research and management offers a particularly promising opportunity to test experimentally ecological hypotheses at relevant spatial and temporal scales (Schmitz & Sinclair 1997). However, the integration of large-scale experiments with conservation application, in the spirit of adaptive management, has been limited largely to aquatic ecosystems (Walters 1986; Carpenter & Kitchell 1993). This lack of large-scale terrestrial experiments is partly due to intrinsic logistical and moral challenges coupled with the recent dominating influence by experimentalists on the methodological philosophy of ecology (Estes 1995; Oksanen 2001). Oksanen (2001) provides a philosophical framework that attempts to resolve some issues put forth by experimentalists (Hurlbert 1984). To overcome logistical and moral challenges, creative approaches are needed (Estes 1995). Examples include exploring (i) habitat fragments with and without predators; (ii) the recovery and/or decline of an over-exploited species; and (iii) reintroductions (Breitenmoser & Haller 1993; Estes & Duggins 1995; Crooks & Soulé 1999; Stahl et al. 2001). Here, using islands as a model system, we report on a novel approach to large-scale experiments with direct application to conservation and management.
Many island biotas now include exotic mammals. Due to the lack of both top predators and a history of vertebrate herbivory in the evolutionary background of island plants (Carlquist 1974; Bowen & Van Vuren 1997), exotic herbivores are known to play key roles in island ecosystems (Hunter 1992). The study of these herbivores offers an opportunity to explore potential top-down effects of species additions, and their removal provides a powerful approach to understanding these effects. While the eradication of non-native mammals from islands is important in conservation (Towns, Atkinson & Daughtery 1990; Donlan et al. 2000; Simberloff 2001), it also provides opportunities for large-scale applied experiments on ecosystems. These opportunities, which are morally justified due to conservation gains (Soule 1990), have been largely under-utilized by researchers (notable exceptions include Dilks & Wilson 1979; North, Bullock & Dulloo 1994).
In this study, we used the removal of exotic herbivores from the San Benito Islands, Mexico, to investigate the impact and recovery of an island plant community from herbivory. We did this by contrasting one island, from where exotic herbivores were removed, with another adjacent and similar island, where they remained. We began this programme with the intent of evaluating the following hypotheses: (i) with herbivore removal, plant community structure changes due to the release of top-down regulation; and (ii) the response by the plant community is predictable from the hierarchy of herbivore food preference. Specifically, we expected the largest relative increases by the most preferred plants.
An understanding of the community-level responses to exotic herbivory has important applied implications, aiding in the management and restoration of grazing systems and degraded habitats (Bullock et al. 2001; Nugent, Fraser & Sweetapple 2001). However, the majority of evidence on community-wide impacts of exotic species remains anecdotal (Blossey 1999; Parker et al. 1999; although see Norbury 2001; Roemer, Donlan & Courchamp 2002). We therefore augmented our large-scale manipulation with smaller-scale replicated experiments. This approach allowed insight on both exotic herbivore and abiotic (i.e. precipitation) effects on an island plant community at the ecosystem scale.
BACKGROUND AND STUDY LOCATION
The San Benito Islands (Fig. 1; 28°18′30″ N and 115°34′00″ W) are located c. 65 km west of Point Eugenia, Baja California, Mexico, one of the driest areas in North America (mean annual rainfall < 100 mm; Crosswhite & Crosswhite 1982). These Sonoran desert islands are low in diversity but high in endemism, including at least eight endemic plants (Junak & Philbrick 2000). They have no native terrestrial mammals. The San Benito Islands are an important seabird nesting site, with 12 species nesting in large numbers (Boswell 1978). Seabirds probably play an important role in terms of nutrient input to the islands (C.J. Donlan, unpublished data; Anderson & Polis 1999). The study system consisted of an experimental island (San Benito West, SBW) and a control island (San Benito East, SBE). SBW (3·5 km2) and SBE (1·1 km2) are separated by c. 2 km and are similar both in vegetation and fauna (Junak & Philbrick 2000). The dominant plant community on both islands is maritime desert scrub, consisting of shrubs and suffrutescent perennials [Agave sebastiana (Greene), Atriplex barclayana (Benth.), Euphorbia misera (Benth.), Frankenia palmeri (I.M. Johnst.), Lycium brevipes (Benth.), Lycium californicum (Nutt.), Malva pacifica (M.F. Ray), Suaeda moquinii (Greene) and two species of cacti, as well as winter annuals, Cryptantha spp., Eschscholzia ramosa (Greene), Hemizonia streetsii (A. Gray) and Perityle emoryi (Torr.)] (Junak & Philbrick 2000).
European rabbits Oryctolagus cuniculus (L.) were introduced to the San Benito Islands during the early 1990s and have since caused significant damage to the vegetation (Donlan et al. 2000). On SBW, rabbits were introduced in 1991; goats Capra hircus (L.) and donkeys Equus asinus (L.) have had a discontinuous presence since 1948 (Junak & Philbrick 2000). At the beginning of the study, rabbits were abundant and a few goats and donkeys (11 in total) were present on SBW. On SBE, rabbits were introduced between 1995 and 1996; rabbits were abundant by March 1996 (Donlan et al. 2000). There is no history of goats or donkeys on SBE. Given the small numbers of donkeys (one to four) and goats present on SBW, both historically (Moran & Lindsay 1951) and just prior to the study, rabbits were considered the primary vertebrate herbivore on both islands.