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Biological invasions, as a result of increased international trade and transportation, are one of the single most important threats to the health of native ecosystems. Among these invasive species, exotic pathogens easily pass undetected during an initial lag phase. Although in the majority of cases exotic pathogens do not cause significant damage (Jones & Baker, 2007), there are many examples in which they have caused enormous damage to plant populations (Desprez-Loustau et al., 2007). Initial survival of the invasive pathogen depends on a susceptible host(s) and a disease-conducive environment. Subsequent spread will be mediated by ecological and biological interactions between the hosts and the pathogen (Burdon & Jarosz, 1988; Gilbert, 2002). Unlike native pathogens that may modify structure and diversity of plant communities slowly over time, invasive pathogens can result in rapid and severe transformation of plant communities as they are brought into contact with naïve populations of a previously unexposed host (Burdon et al., 2006). When the new host is a large keystone species, structural change in the plant community can be far-reaching because of the cascading effects on associated animal and plant life which may lead to increased or decreased diversity (Burdon et al., 2006). Dramatic examples of changes in plant community structure following disease introductions include loss of diversity in native jarrah (Eucalyptus marginata) communities in south-western Australia after the introduction of Phytophthora cinnamomi (Hardham, 2005), increased diversity with the emergence of new plant communities composed of co-dominant shrubs and trees to replace chestnut (Castanea dentata) in eastern North American forests after the introduction of Asian Cryphonectria parasitica (Stephenson 1986) and the replacement of elm (Ulmus spp.) by shade-tolerant species in eastern North American forests following invasion by Ophiostoma ulmi (Parker & Leopold, 1983). Although tree mortality may provide increased nesting sites for birds and animals (Franklin et al., 1987), loss of heavy seed-producing species may have adverse effects on diversity of wildlife (Castello et al., 1995; Monahan & Koenig, 2006).
Symptoms of disease in natural forests and woodlands commonly are highly variable through time and space. This variability results from interactions among biotic and environmental layers that are inherently heterogeneous. Such interactions present formidable challenges and opportunities to understand and manage the dynamics of disease spread. Landscape pathology attempts to integrate host and pathogen systems and to develop spatially explicit models that help in the prediction of disease spread and dynamics (Holdenrieder et al., 2004; Ostfeld et al., 2005; Hamelin, 2006). In addition to spatial interactions, the physiological dependence between host and pathogen that results from phenological changes in the host also determines the development of disease (Biere & Honders, 1996, Kennelly et al., 2005; Blachinsky et al., 2006). Native pathogens evolve synchronicity with host phenological cycles, particularly for diseases of ephemeral organs such as flowers and fruits (Ngugi & Scherm, 2006). However, incidence of disease caused by introduced pathogens will be variable depending on chance matching between host and pathogen phenology. Such variation could explain in part the patchy distribution of disease that is often observed in natural forests and woodlands. For pathogens causing canker diseases, the availability of newly differentiating vascular tissues in the host may be a prerequisite for disease progression. If the timing of sporulation events is seasonal and it coincides with the period of resumption of cambial activity in the host, infection is likely. However, hosts with late cambial activity would escape infection during the main sporulation event. This is likely to be particularly true of splash-dispersed pathogens in Mediterranean environments, where peak sporulation events are rainfall and temperature limited. Variations in host phenology could therefore partly explain patchiness in host mortality; when the pathogen peaks, the host must be available.
Sudden oak death (SOD), an emerging disease caused by an introduced pathogen, has provided the opportunity to evaluate the importance of synchronicity between host and pathogen in terms of the incidence and spatial distribution of disease. Mortality of oaks (Quercus spp. Sect. Lobatae) and tanoak (Lithocarpus densiflorus), resulting from infection by Phytophthora ramorum (Rizzo et al., 2002), is threatening to transform the coastal woodlands of central and northern California. The spatial distribution of disease on coast live oak (Quercus agrifolia) is uneven (Kelly & Meentemeyer, 2002), but the causes are still poorly understood. Sporulation of the pathogen is at a maximum between the months of December and May when climate is cool and humid (Davidson et al., 2005). Coast live oak is a keystone species of California's coastal woodlands, providing habitat for a wide range of insects, birds and mammals. It forms a vegetation type in central and southern California that is primarily single species (78% of basal area) and is relatively dense compared with other oak species in California (Waddell & Barrett, 2005).
Our earlier work indicated significant heritable variation in lesion size on branch cuttings of coast live oak in response to inoculation with P. ramorum (Dodd et al., 2005). Here, we extend this earlier work to ask whether variations in host cambial phenology could contribute to the patchiness of mortality of coast live oak and whether genetic resistance can be detected through shifts in susceptibility when a stand has been exposed to an epidemic of the disease. We used three approaches. First, we followed the response of branch cuttings to inoculation throughout a full annual cycle in 2003–2004 to test for seasonal variation in responsiveness. We estimated clonal repeatability to determine whether variations in response had a genetic base. Secondly, we sampled bud and cambial tissue of the same trees in spring 2006 to determine variations in phenology of bud burst and cambial activity. We recognize that it would have been preferable to have phenological data for the same year as inoculations. However, variations among trees are likely to remain constant across seasons because of high heritability of phenological traits in trees (Morgenstern, 1996; Rehfeldt et al., 1999). Over the relatively small spatial scale of our sampled trees, local climatic variations from one year to the next are unlikely to be important. We then investigated whether individuals with later resumption of shoot and cambial activity also produced maximum lesions at a later date. Finally, we anticipated that stands having suffered high mortality would have been purged of the more susceptible genotypes. We hypothesized that this could be detected as smaller average lesion sizes and lower variability in lesion size. For this, we compared two neighbouring stands that contrasted in levels of prior mortality as a result of SOD disease.