Introduction
Non-native 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 often lead to shifts in the behavior, functional role, and abundance of their native counterparts (Holway et al. 2002; Bertelsmeieret al. 2017). In so doing, non-native 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 directly modulate the spatial distribution of carbon (e.g., wood ants in subalpine forests, Risch et al. 2005; Finér et al. 2013), or engage in ant-plant mutualisms that influence host plant carbon dynamics (Pringle 2016). Second, invasive ants can feed on extrafloral nectar of host plants (Ness & Bronstein 2004; Lach et al. 2009) and collect honeydew from heterospecific insect partners (Beardsleyet al. 1982; Zhou et al. 2017; Demian & Tarnita 2019; Anastasio 2020). Ant interactions with nectaries or with phloem-feeding insects can affect the carbon source-sink ratio of host plants (Albaniet al. 2010; Del-Claro et al. 2016; Prior & Palmer 2018) which can increase or decrease 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., Savage et al. 2009; Lach & Hoffmann 2011; Kulikowski II 2020), which may combine with changes to leaf carbon exchange rates to affect whole-plant carbon fixation. Finally, invasive ants could especially influence ecosystem carbon cycling by invading ant-plants that are dominant primary producers in some communities (e.g., devil’s gardens, Frederickson et al.2005; Acacia drepanolobium savannas, Goheen & Palmer 2010), such that invasion would have disproportionate effects on local carbon cycles.
We investigated how invasion by Pheidole megacephala Fabricius (the “big-headed ant”) affects carbon cycling in a widespread and mono-dominant 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), but otherwise suppressing abundance, distribution, and diversity of 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 of Laikipia,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 exclusively house one native ant species at a time, producing extrafloral nectar and hollow spine domatia (e.g., Huntzinger et al. 2004) to feed and house thousands of 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). However, and because P. megacephala does not consume extra-floral nectar, host trees may experience energetic savings immediately after invasion, even as longer-term costs (through risk of intense herbivory) increase. 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. Thus, both ant-plant and vertebrate-plant interactions are potential modes by which invasive ants may impact leaf photosynthetic rate (via source-sink dynamics) and canopy carbon fixation (via canopy damage by herbivores).
We conducted field experiments and observations to investigate howP. 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 howP. megacephala invasion influences host plant carbon fixation over both the short- (<1 year) and long-term (ca . 5 years). We investigated these short- and long-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 the leaf photosynthetic rate of A. drepanolobium change shortly after the extirpation of costly ant mutualists by P. megacephala ? (2) Does the leaf photosynthetic rate of A. drepanolobium further change in long term invasion sites, and how is that rate influenced by ant-plant and vertebrate plant interactions? (3) How do vertebrate herbivores and invasive ants contribute to changes in canopy photosynthesis for invaded trees?