Ecosystem recovery is a type of ecological restoration in which the expectation is that the system will regain desirable attributes through ‘natural’ processes (Meffe, Carroll & Pimm 1997). For example, once disturbances imposed by invasive non-indigenous organisms on native species are reduced by control measures, it may be possible for an ecosystem to recover to a pre-disturbance state (Paine, Tegner & Johnson 1998). However, where disturbance impacts exceed the capacity of species to recover, ecosystems can change permanently, even when disturbances are reduced (Hobbs & Norton 1996). Barriers to vegetation community restoration in many instances result from reduced dispersal and seedbank recruitment of pre-disturbance species, competition, and biogeochemical and physical changes associated with disturbances (Suding, Gross & Houseman 2004). Feedbacks between species and ecosystem properties may further prevent recovery to the pre-disturbance state, and in some situations, positive feedbacks themselves may initiate trophic cascades that permanently shift ecosystems to alternate states (Bazely & Jefferies 1997; Wolf, Cooper & Hobbs 2007).
Grazing and browsing by large mammalian herbivores is one form of disturbance that can have a major impact on ecosystems. For example, high densities of deer in North America, Europe, Japan and New Zealand have been responsible for altering the structure and composition of forests (Gill 1992; Coomes et al. 2003; Côtéet al. 2004) and open-hill habitats (Rose & Platt 1987; Takahashi & Kaji 2001; Albon et al. 2007; Wolf et al. 2007). These changes can reduce the recruitment of deer-preferred species and lead to modifications in nutrient cycling and successional trajectories that may be difficult to reverse (Côtéet al. 2004). As such, deer are a conservation problem in many countries where they have been introduced, and also in their native ranges, when populations are allowed to increase. However, the long-term responses of ecosystems to deer population control are not well described, mainly because there are few places where deer control has been imposed for a length of time that is meaningful in terms of the generation times of long-lived plant species. Although restoration of plant communities is possible following release from high levels of deer browsing, the time scale for monitoring recovery often requires decades rather than years (Anderson & Katz 1993; Rose, Suisted & Frampton 2004; Webster, Jenkins & Rock 2005; Long, Pendergast & Carson 2007). Studies of long-term (>30 years) patterns of vegetation change rarely consider the influence of mammalian herbivory, and those that do are in the context of increasing herbivore populations (Rooney & Dress 1997; Schütz et al. 2003; Rooney et al. 2004) or focus on the recovery dynamics of individual species (Anderson & Katz 1993; Lee et al. 2000; Webster et al. 2005) or habitat types, e.g. forests (Stewart, Wardle & Burrows 1987; de la Cretaz & Kelty 2002) or grasslands (Rose & Platt 1987; Rose et al. 2004). Hence, much is still unknown about the outcomes of different restored communities within multi-habitat landscapes.
In New Zealand, several species of deer introduced for recreational hunting have become abundant in natural habitats. By the 1960s, red deer Cervus elaphus scoticus ranged over most of the country (King 2005), measurably impacting forest composition and structure by reducing populations of the most palatable species (Nugent et al. 2001b; Wardle et al. 2001; Husheer, Coomes & Robertson 2003; Husheer 2007). Similar impacts have been observed in grasslands, where the recovery times for native plants following reductions in deer densities may be in the order of decades (Rose & Platt 1987; Lee et al. 2000; Rose et al. 2004). Introduced red deer colonized our study site, the Murchison Mountains, Fiordland National Park, New Zealand, in the 1930s and reached relatively high densities (c. 11 deer km−2) in the late 1950s (Parkes, Tustin & Stanley 1978), measurably impacting vegetation (Wilmshurst 2003). Browsing by deer reduced the abundance of palatable species in forests and subalpine shrub (Wardle, Hayward & Herbert 1971) and in grasslands, overgrazing reduced the stature and density of palatable tussocks (Mills, Lee & Lavers 1989). In their native range, red deer similarly alter vegetation composition, structure and dynamics when present at densities >10 deer km−2 (Côtéet al. 2004). However, the impacts are likely to be greater in New Zealand because of the slow growth rate of native plants (Bee, Kunstler & Coomes 2007) and lack of specific adaptations of plants to mammalian herbivores (Bond, Lee & Craine 2004).
The objectives of our study were to test whether control of introduced red deer since 1962 has led to: (i) vegetation changes across a 518 km2 landscape in New Zealand over a period of 39 years, and (ii) the recovery of earlier depleted deer-preferred species. Our study site is a unique example of long-term deer control, because the mountains have been prioritized for deer culling by successive governments to protect the habitat of the only remaining wild population of takahe Porphyrio hochstetteri, an endangered flightless rail (Mills & Mark 1977). We predicted that deer-palatable species would be initially uncommon, but that regeneration of these species would occur by the end of our study period, given that culls reduced deer to the levels at which regeneration of palatable species is thought to be possible (<2 deer km−2; Husheer 2007) by 1986 (Nugent & Sweetapple 1989).