Introduction
Invasive insects can profoundly affect ecological communities, threatening biodiversity (Pimentel et al. 2001; Englund 2008; Hill et al. 2013), disrupting important ecological processes (forest canopy structure, biogeochemical cycles, Gandhi & Herms 2010; suppressing native foundational species, McGeoch et al. 2015), and imposing large economic costs (Bradshaw et al. 2016; Painiet al. 2016). Invasive ants comprise >240 species, and can affect the behavior, functional role, and abundance of their native counterparts (Holway et al. 2002; Bertelsmeier et al. 2017). In so doing, ants can restructure pollination networks (Vanbergen et al. 2018), interrupt seed dispersal (e.g., Horvitz & Schemske 1986; Rodriguez-Cabal et al. 2012) and pollination (e.g., Fuster et al. 2020), and spread diseases in pollinator communities (Vanbergen et al. 2018).
Although their community-level impacts are well documented and diverse, the consequences of ant invasions for biogeochemical cycles are poorly understood. In particular, effects of invasive ants might be expected to reverberate throughout ecosystems via shifts in carbon dynamics for several reasons. First, native ants, which are often displaced by invasive ants (Ness & Bronstein 2004; Milligan et al. 2016), can increase the spatial variability of soil carbon (e.g., wood ants in subalpine forests, Risch et al. 2005; Finér et al. 2013) as a byproduct of foraging and ground-nesting. Second, invasive ants can feed on extrafloral nectar of host plants (Ness & Bronstein 2004; Lachet al. 2009) and collect honeydew from heterospecific insect partners (Beardsley et al. 1982; Zhou et al. 2017; Demian & Tarnita 2019; Anastasio 2020): both activities remove carbon from their host plant’s active carbon pool that would otherwise support aboveground growth and development (Pringle 2016). Ant interactions with nectaries or with phloem-feeding insects can affect the carbon source-sink ratio of host plants (Albani et al. 2010; Del-Claroet al. 2016; Prior & Palmer 2018) which can affect leaf carbon exchange rates (Goldschmidt & Huber 1992; Nebauer et al. 2011). Third, invasive ants can deter or facilitate herbivory on host plants with consequences for plant growth and overall canopy size (e.g., Savageet al. 2009; Lach & Hoffmann 2011; Kulikowski II 2020), which may combine with changes in leaf carbon exchange rates to affect whole-plant carbon fixation. Finally, invasive ants could influence ecosystem carbon cycling by displacing the ant defenders of ant-plants, some of which are dominant primary producers in communities (e.g., devil’s gardens, Frederickson et al. 2005; Acacia drepanolobium savannas, Goheen & Palmer 2010), with potentially large effects on local carbon cycles.
We investigated how invasion by Pheidole megacephala Fabricius (the “big-headed ant”) affects carbon cycling in a widespread and monodominant foundation species, the whistling thorn tree (Acacia drepanolobium ). Pheidole megacephala has invaded tropical and subtropical ecosystems around the world (Wetterer 2012), extirpating native ant mutualists (Ness & Bronstein 2004; Riginos et al.2015), forming facultative partnerships with phloem-feeding insects (e.g., Beardsley et al. 1982; Gaigher et al. 2013), and suppressing the abundance, distribution, and diversity of many native insects (Ness & Bronstein 2004; Hoffmann & Parr 2008; Riginos et al. 2015; Milligan et al. 2016). In savannas underlain by clay-rich vertisols (i.e., ‘black-cotton’ savannas) in Laikipia, Kenya,A. drepanolobium comprises >95% of woody cover (Young et al. 1996) and forms obligate mutualisms with four native ant species (Crematogaster mimosae Santchi, Crematogaster nigriceps Emery, Crematogaster sjostedtiMayr, and Tetraponera penzigi Mayr). Host plants produce extrafloral nectar and hollow spine domatia (e.g., Hocking 1970; Huntzinger et al. 2004) which are consumed and occupied by single colonies consisting of thousands of defensive ants (Palmer 2004). The most common mutualist, C. mimosae , consumes nectar and honeydew (Prior & Palmer 2018) and reduces herbivory by large mammals (Stanton & Palmer 2011) including elephants (Goheen & Palmer 2010). In invaded habitats, C. mimosae mutualists are completely extirpated byP. megacephala , which does not deter herbivores (Riginos et al. 2015).
Because native ant mutualists impose high continuous metabolic costs on their host plants but provide protection for their tree against destructive herbivory by elephants and other large herbivores, we expected that the replacement of defensive C. mimosae with non-defensive P. megacephala would cause distinct short- and longer-term consequences for A. drepanolobium trees. In the short-term, the removal of nectar- and honeydew-consuming C. mimosae by P. megacephala may free up carbohydrates, which would support leaf growth, photosynthetic upregulation, management of water (e.g., Inoue et al. 2017; Zhang et al. 2019), and other metabolic processes of the host tree (Wiley & Helliker 2012; Glanz-Idan & Wolf 2020). We also would not expect P. megacephala workers to act as a direct resource sink for the tree, because they fail to induce nectar production on host plants, resulting in a 55-98% decline in active nectaries on invaded trees (Riginos et al. 2015; Palmeret al. 2020): thus, host trees should experience energetic savings immediately after invasion. However, the loss of native ant mutualists increases the risk of rare but heavy damage by elephants and other large herbivores for A. drepanolobium trees (Goheen & Palmer 2010), such that over the longer-term, invaded communities experience more instances of mild to catastrophic herbivory (Riginoset al. 2015). King and Caylor (2010) demonstrated that the prevention of herbivory by native ants influences photosynthetic rate of the host tree, but direct ant-plant interactions and the role of this invasive ant were not investigated in their study. These prior studies informed our predictions that 1) invasive P. megacephala affects leaf-level photosynthetic rates by removing an energetic sink for the host tree immediately after invasion, but also 2) that P. megacephala reduces canopy-level photosynthesis over longer time scales by rendering trees vulnerable to canopy damage.
We conducted field experiments and observations over a 2 yr period to investigate how P. megacephala invasion affects carbon fixation in A. drepanolobium . Because the effects of invasion frequently lag behind the initial arrival of the invader (Simberloff 2011), we evaluated both how P. megacephala influences host plant carbon fixation after a recent invasion (<1 year ago) and in “longer-term” invasive sites (invaded ca . 5 years ago). We investigated these short- and longer-term impacts of invasion in wet and dry seasons during which host plant rates of photosynthesis can substantially differ (King & Caylor 2010). We addressed three research questions regarding A. drepanolobium : (1) Does leaf photosynthetic rate of A. drepanolobium change shortly after the extirpation of costly ant mutualists by P. megacephala ? (2) Does leaf photosynthetic rate of A. drepanolobium further change in longer term invasion sites, and how is that rate influenced by ant-plant and vertebrate herbivore-plant interactions? (3) How do vertebrate herbivores and invasive ants contribute to changes in canopy-level photosynthesis for longer-term invaded trees?