Imported crazy ant displaces imported fire ant, reduces and homogenizes grassland ant and arthropod assemblages
A recently introduced, ecologically dominant, exotic ant species, Nylanderia fulva, is invading the Southeastern United States and Texas. We evaluate how this invader impacts diversity and abundance of co-occurring ants and other arthropods in two grasslands. N. fulva rapidly attains densities up to 2 orders of magnitude greater than the combined abundance of all other ants. Overall ant biomass increases in invaded habitat, indicating that N. fulva exploits resources not fully utilized by the local ant assemblage. At high density, as N. fulva spreads, it eliminates the current ecologically dominant invasive ant, red imported fire ants (Solenopsis invicta). Compared to imported fire ant dominated habitat, N. fulva invasion zones have lower non-ant arthropod species richness and abundance with impacts differing by trophic category. Further, N. fulva reduces abundance and species richness of the remainder of the ant assemblage and does so in a non-random manner: impacting species with small sized workers much less than species with larger workers. In these and other ant assemblages with a large exotic component, the exotics tend to be small bodied species. As a result, N. fulva almost completely eliminates regionally distributed species, but leaves globally distributed species largely unaffected, thereby systematically favoring introduced over native diversity. S. invicta impacts wildlife and arthropod assemblage structure and is nearly ubiquitous in non-forested habitats of the Southeastern United States and Texas. Its displacement by N. fulva has critical implications for the natural systems of this region.
KeywordsBiotic homogenization Tramp ants Nylanderia sp. nr. pubens Invasive ants Tawny crazy ant Hairy crazy ant Rasberry crazy ant Paratrechina fulva
Biological invasions drive diversity loss in the current extinction crisis (Sala et al. 2000), primarily by causing local, although typically not global, extinction of native species (Powell et al. 2011). Secondarily, they homogenize the biota by replacing regionally distinct species assemblages with assemblages comprised of more cosmopolitan taxa (McKinney and Lockwood 1999), thereby making biogeographically distant assemblages more similar.
The Southeastern United States of America (USA) has suffered a series of invasions by ecologically dominant invasive ant species native to central South America. Argentine ants (Linepithema humile Mayr), colonized the USA through New Orleans prior to 1891, and, within 50 years became widespread throughout the Southeastern USA. Argentine ant invasions dramatically reduce diversity of native ant assemblages (Holway et al. 2002). Around 1918, black imported fire ants (Solenopsis richteri Forel), colonized the USA through Mobile, Alabama (Wilson 1951). Red imported fire ants (Solenopsis invicta Buren), also introduced into Mobile (Wilson 1958), followed S. richteri in the 1930s. S. invicta spread and rapidly displaced S. richteri throughout most of the Southeast USA (Tschinkel 2006). S. invicta also apparently displaced Argentine ants (Buren et al. 1974) currently restricted to small scattered populations (Buczkowski et al. 2004). Ecological impacts of the spread of S. invicta were documented in central Texas where it impoverished ant, and non-ant arthropod faunas (Porter and Savignano 1990; Morris and Steigman 1993; Camilo and Phillips 1990), as well as negatively impacting many types of ground nesting birds and reptiles (Allen et al. 2004). In a low disturbance, woodland environment, ant and non-ant arthropod assemblage diversity largely recovered within 12 years of invasion by S. invicta (Morrison 2002). In contrast, in a xeric grassland, evidence of S. invicta impacting ant assemblage structure has persisted in some but not all habitats (LeBrun et al. 2012).
The region is now entering the next chapter in this history of serial invasion by ecologically dominant ants. An exotic ant species, Nylanderia (formerly Paratrechina) fulva (Mayr), colonized the Houston, Texas area around 2002 and, at the same time, dense infestations began to be reported in Florida. Initial attempts to identify this ant did not succeed (Meyers and Gold 2008), which led to it being referred to variously as N. sp. nr. pubens, N. pubens, and Rasberry crazy ant (Cook et al. 2012; Meyers and Gold 2008; Valles et al. 2012). A recent combined genetic and morphological analysis demonstrates that all populations of this ant in the USA are N. fulva, with the potential exception of South Florida that was not tested (Gotzek et al. 2012). By 2012, populations of these ants had established at sites in 21 counties in Texas (Center for Urban and Structural Entomology 2012). Dense, scattered populations of this ant are also present throughout Florida (Valles et al. 2012), in southern Mississippi (MacGown and Layton 2010), and in southern Louisiana (Hooper-Bui et al. 2010).
Nylanderia fulva evolved in central South America where it is a conspicuous, ecologically dominant member of the ant fauna (LeBrun et al. 2007; Wilder et al. 2011; Feener et al. 2008). Introduced N. fulva populations in Colombia have been responsible for widespread ecological and agricultural damage (Zenner de Polania 1990). Similar to most other globally widespread, ecologically destructive invasive ant species (Holway et al. 2002), populations of N. fulva in Texas are supercolonial with no aggression evident between workers from distant nests (Horn 2009). Nests, commonly containing 0–5 queens, males, workers, and brood (LeBrun pers. obs., Horn 2009), occur in soil cavities or in objects on the surface, such as pots, predisposing them to human transport. Its diet appears to be the same as documented in introduced populations in Colombia, consuming insects and the sugary exudates of plants and homopteran insects (Zenner de Polania and Bolaños 1985). No flights of winged reproductives have been reported for this species. Thus, new colonies are likely founded exclusively through colony fission. As a result, isolated populations of N. fulva spread slowly averaging about 20–30 m/month of outward spread (Meyers 2008). The outcome of resource competition with S. invicta appears to depend upon relative density with N. fulva monopolizing the majority of resources in laboratory conditions of equal abundance (Horn 2009). Population abundances of this ant inside invaded areas are tremendous (see “Results”). However, no published information exists on ecological impacts of these dense populations.
In Texas, N. fulva has invaded both urban and natural environments. Natural environments invaded to date include both highly modified and native pastureland habitats as well as highly disturbed wood lots and intact floodplain forest (E. G. LeBrun pers. obs.). The demonstrated capacity of N. fulva to invade wild lands demands an examination of its ecological impacts in natural environments. As a first step in this process, we contrast impacts of N. fulva on co-occurring ant species at two grassland sites on the Texas Gulf Coast.
Employing a natural experiment design, we use pitfall traps to sample areas invaded by N. fulva and adjacent, uninvaded areas to evaluate whether N. fulva impacts diversity and abundance of co-occurring ants and non-ant arthropods. Further, we examine whether N. fulva alters the body size distribution of the resident ant assemblage (McGlynn 1999), and whether, as a consequence of body size biased impacts, it exerts a homogenizing influence by differentially impacting regionally versus globally distributed species.
We sampled two N. fulva invasion fronts at sites separated by 90 km within the Northern Gulf Coastal Prairie Ecoregion (Griffith et al. 2004): Iowa Colony: 29.44°N, 95.44°W, and El Campo: 29.23°N, 96.31°W. The Iowa Colony site supported a higher density N. fulva population than did El Campo (see “Results”). These sites will be referred to respectively as the “high density” (HD) and the “moderate density” (MD) sites. Both sites are disturbed coastal prairie containing a mixture of native and introduced plant species. The HD site has a history of intensive grazing and off road vehicle trails throughout. During periods of heavy rain, this site is prone to shallow flooding. Dominant plant species at the HD site are bushy bluestem (Andropogon glomeratus Walt.), dewberry (Rubus trivialis Michx.), McCartney rose (Rosa multiflora Thunb.), and Sesbania (Sesbania drummondii Rydb.). The MD site is an ungrazed fragment of disturbed prairie vegetation in a matrix of row crop agriculture. Vegetation at the MD site is dominated by little bluestem [Schizachyrium scoparium (Michx.) Nash], Johnson grass [Sorghum halepense (L.) Pers.], dewberry (R. trivialis Michx.), and goldenrod (Solidago cf. altissima). Red imported fire ants (S. invicta) invaded this region around 1967 (Hung and Vinson 1978). Given this 40 year interval, regionally these systems are either at equilibrium or in a process of recovering diversity lost during the initial S. invicta invasion.
Sampling was conducted during October 2009, and August 2010. Ant and arthropod assemblages were quantified using grids of pitfall traps. In this natural experiment design, accurate inference from comparisons of invaded and adjacent uninvaded habitat assumes these areas possessed similar arthropod assemblages prior to invasion. For this reason, within both sites, all areas sampled were matched by: tree cover (trees absent), herb and shrub species composition, and known disturbance history. To further protect against spurious conclusions, we carried this work out at two independent sites. Finally, dense imported fire ant populations are a universal characteristic of open, grassland habitats in the Northern Gulf Coastal Prairie ecoregion (E. G. LeBrun pers. obs.), making any random deviation from this norm very unlikely.
Pitfall traps, 250 ml plastic cups with snap cap lids, and 7 cm diameter openings, were installed, charged with a solution of water, odorless laundry soap, and 25 % propylene glycol, and then left closed for 3 days to allow soil disturbance effects to dissipate. Traps were opened for 5 days. Pitfall trapping stations (HDS: 71 stations, MDS: 54 stations) were separated by 30 m, and contained 2 traps each spaced 2 m apart. Contents of both traps were combined for data analysis.
All ants were identified to species and counted. Arthropods were sorted to morphospecies, excluding all morphospecies of Diptera, Lepidoptera, Hymenoptera, and Neuroptera that had wings. These highly mobile groups, likely immigrated to the area of the trap and their numbers do not reflect local conditions. All Collembola were counted but not sorted to morphospecies. Morphospecies, with the exception of spiders apart from Lycosidae and Salticidae, were identified to family level or lower, and, based upon this ID, assigned to a trophic category: herbivore, predator, or detritivore.
The head widths and body masses of ant species were measured. Head width, taken with a dissecting microscope and optical micrometer, was the average of 3 measures from the largest available workers. To assess average worker mass, 10–20 workers chosen randomly from pitfall traps were placed in a drying oven for 48 h and then weighed in batches of 5 on a Mettler-Toledo UMX2 Ultra-microbalance™.
The N. fulva population edge was defined as the line bounding all points where N. fulva was present at that station and at least one other adjacent station. For some analyses, habitat was dichotomized into areas at least 100 m behind the edge of the N. fulva population, here forward “Nf-abundant areas”, and uninvaded areas near and in front of the N. fulva population edge, here forward “Nf-scarce areas”. At both sites, some N. fulva workers were present in isolated stations within the Nf-scarce areas. The distance between a sampling station and the N. fulva population edge was measured on Google Earth (http://www.google.com/earth/index.html). These distances provide a proxy measure for the relative amount of time a station has been occupied by N. fulva.
At the MD site, area uninvaded by N. fulva with habitat matching the invaded area was limited. At this site, to provide sufficient replication for within site comparisons, Nf-scarce stations included all stations in front of the N. fulva population edge plus stations <10 m behind the edge. In this rapidly expanding population, stations 10 m behind the edge had very limited exposure to N. fulva prior to sampling, and N. fulva was present at very low abundance (6.7 ± 9.2 workers/trap) (mean ± SD). No between site statistical comparisons are made, so the lack of a precise between site match in the definition of Nf-scarce stations does not impact the results.
Unless specified otherwise, all analyses were performed using JMP® 9.0 (SAS Institute 2011). Non-parametric, Wilcoxon Rank Sums tests were preferred for analysis of ant abundance data which did not meet assumptions of normality. Contingency tables were used to analyze presence–absence data characterizing ant prevalence.
We compare species density using Ecosim (Gotelli and Entsminger 2004). Species density is the number of species encountered across the array of traps, and provides the best available measure of species richness for each habitat. Estimating species richness requires extrapolating from species encountered to total species present, including those not observed, and relies on prevalence of rare species to estimate the number of unobserved species (Colwell 2011). In comparisons of recently invaded habitat and adjacent uninvaded habitat, species abundances are in flux, making this extrapolation inherently unreliable. To calculate species density, the average and standard deviation in species observed per sample accumulated was calculated by repeatedly aggregating randomly selected stations sampled with replacement (Gotelli and Colwell 2001). To test whether accumulation curves from areas differ, a Z score was calculated using the mean richness for each curve at the maximum observed sample size for the less well sampled area, and the standard deviation from the more intensively sampled area for that level of replication (Zarr 1999). N. fulva were removed from the data set prior to analysis creating some empty traps. To preserve these traps in the data set, a place holder species designated as present in all traps was inserted (LeBrun et al. 2012). Prior to presentation, all means have been back transformed.
To quantify variation in abundance, we define “prevalence” as the fraction of stations where a species was present in the pitfall trap. The “incidence abundance” of a species refers to the number of workers per pitfall trap considering only stations where that species occurred. The product of prevalence and incidence abundance yields the “abundance” of a species: the average workers per pitfall trap across all stations.
Logistic regression was used to evaluate the association between worker body size, and impact of N. fulva on species abundances. Responses of species to N. fulva were dichotomized. If average abundance of a species within a site declined by 25 % or more comparing Nf-scarce to Nf-abundant areas, it was characterized as negatively impacted by N. fulva. Response of all other species were designated as neutral to positive. All species occurring at both sites exhibited the same categorical response to N. fulva at both sites. All species observed at multiple stations at either site were included. Data was taken from the site where a species was most abundant.
To examine whether ant body size dependent impacts alter the biogeographic makeup of the assemblage, we categorized species by their ranges. Species restricted to the Nearctic zoogeographic province (North America north of Central Mexico) were categorized as “regional”. “Global” species are species that occur in more than one zoogeographic province. Finally, “invasive” species are exotic species that penetrate natural environments, reach high densities, and have documented negative ecological impacts on ecosystems they colonize. Species distribution information was obtained by summarizing occurrence information from the website AntWeb (AntWeb 2011), Creighton (1950), and any taxonomic literature published after 1950 that is listed on the Hymenoptera Online database (Hymenoptera Online (HOL) 2011).
Regression analyses of distance from the edge of the expanding N. fulva population and arthropod densities were performed separately for stations ahead of the population edge and stations behind the edge. The processes at work in these two environments were distinct and no uniform functional relationship spanning this boundary was expected. For non-ant arthropod data, a few stations captured extreme numbers of individual morphospecies, distorting the overall analyses of trophic categories. To address this, we reduced the abundance of any morphospecies present in a trap at >5 times its median abundance across all traps in which it occurred to the maximum abundance recorded for this species below that threshold. This constraint limited the effect of extreme outliers while preserving the occurrence and relative abundances of morphospecies. It reduced counts for 2 morphospecies at a total of 3 stations. Due to extreme abundance, Collembola were not included in the detritivore trophic category but were analyzed separately from the other arthropods.
Impacts on ants: high density and moderate density N. fulva sites
Results of analyses contrasting ant abundance between stations in Nf-scarce and Nf-abundant areas at high density (HD) and moderate density (MD) N. fulva sites
HD N. fulva site
MD N. fulva site
Impacts on non-ant arthropods: high density N. fulva site
Results of linear regression analyses examining relationships between abundances of non-ant arthropods and the distance from the edge of the N. fulva population at the high density site
N. fulva +
N. fulva −
27.2 + 0.05 × x
11.4 + 0.02 × x
7.5 + 0.009 × x
8.3 + 0.02 x x
Within a year of colonizing, N. fulva attains numerical abundances up to 2 orders of magnitude greater than other ants in the system combined, including red imported fire ants (S. invicta). Because pitfall traps quantify relative abundance reliably, but tend to oversample fast-moving species, like N. fulva, relative to others (Andersen 1991), how accurately this measures the actual increase in biomass must be interpreted cautiously. Studies of the abundance of some other invasive ant species indicate that these species extreme abundance in the introduced range partly result from more effective exploitation of honeydew and extrafloral nectar relative to the native assemblage (Helms et al. 2011; Tillberg et al. 2007; Wilder et al. 2011). Identifying the factors that underlie the increase in N. fulva biomass requires further study. Understanding how this high biomass is maintained may help to understand the causes of the substantial between site variability in N. fulva abundance, and concomitant between site variability in ecological impacts, documented in this study.
As N. fulva invades, S. invicta suffers strongly. At the HD site, by 200 m behind the front S. invicta is entirely eliminated. Extrapolating from the expansion rate of 180 m/year observed at this site in 2010–2011, N. fulva displaces all fire ant colonies slightly more than a year after spreading into an area. However, S. invicta declines in abundance but persists throughout the MD site. Between fall of 2010 and 2011, the invasion front at the MD site advanced between 150 and 200 m indicating that stations most distant from the N. fulva population edge had been colonized by N. fulva 3–4 years previously. Thus some S. invicta colonies can persist in this environment. Whether S. invicta can establish new colonies and thus co-exist long-term is unknown.
Nylanderia fulva reduces the species richness and abundance of these ant assemblages (Figs. 2, 4). The original invasion by S. invicta reduced the richness and abundance of native ant assemblages regionally (Porter and Savignano 1990; Morris and Steigman 1993; Camilo and Phillips 1990). Recovery of these assemblages has been documented in some systems (Morrison 2002), while in others the evidence of its impact has persisted (LeBrun et al. 2012). Thus the impacts exerted by N. fulva occur against a historical backdrop of repeated assemblage disruption, impoverishment, and partial recovery. The patterns of species responses provide insight into the longer-term trajectory of these assemblages subject to serial invasion.
Interestingly, N. fulva impacts local ant assemblages in these disturbed grasslands in a non-random manner. Comparing HD N. fulva environments with nearby uninvaded habitat, smaller ant species show a mixed response but the majority of species either do not change or increase in abundance. In contrast, large species uniformly decline in abundance. In these ant assemblages, small-bodied ants are disproportionately represented by globally distributed tramp species. The net outcome is a reduction in both biological and functional diversity, accompanied by an intense biotic homogenization of the local ant assemblage (McKinney and Lockwood 1999). By limiting the size distribution of the ant assemblage, N. fulva reduces the functional diversity of ants, the most abundant class of consumers in the leaf litter (Bihn et al. 2010), potentially altering ecosystem function (Kaspari 2005). Further, the preferential removal of a suite of regional species while leaving globally distributed species intact, represents a remarkable process biogeographically homogenizing the local assemblage.
In recently invaded habitat, dense N.fulva populations saturate the environment, and their nests commonly occur in abandoned nest mounds of S. invicta (E. LeBrun pers. obs.). Further, the larvae and brood of other ants was reported as a food source of N. fulva in Colombia (Zenner de Polania and Bolaños 1985). Potentially, the decline of the ant assemblage generally may arise as a consequence of nest raiding by N. fulva either for food or space. Under this proposed mechanism, differential species loss arises because nests of small bodied species, by virtue of having small tunnels, are more resistant to raiding. Suggestively, average head width of the largest workers, a trait intimately related to tunnel diameter, explains more variation in the logistic relationship with impact than does body mass of an average worker. Also consistent with nest raiding, the size threshold above which all ant species suffer negative impacts from N. fulva invasion is slightly larger than the average head width of an N. fulva worker. Testing these and other hypotheses for what drives size biased displacement requires further study.
Is replacement of large-bodied, regional species with small-bodied, global species a general property of ant invasions? Non-native ant species are documented to be smaller than their regionally restricted congeners (McGlynn 1999; Passera 1994; Ness et al. 2004; but see King and Porter 2007). Further, in a study comparing arriving propagules with successfully established ant species, small body size was the most important variable explaining successful establishment (Lester 2005). Thus, in assemblages with many introduced species, a large fraction of small-bodied ants are likely to be global species. Several studies report small bodied ant species resisting displacement by opportunistically nesting, invasive ant species [Anoplolepis gracilipes (Hoffmann and Saul 2010), Pheidole megacephala (Hoffmann et al. 1999), L. humile (Ward 1987; Human and Gordon 1997; Holway 1998)]. These studies attribute resistance of small species to either their foraging niche [hypogeic foragers resist displacement by L. humile (Ward 1987)], or to their functional role [small opportunists and cryptic species resist displacement by P. megacephala (Hoffmann et al. 1999)]. Our results suggest an alternative hypothesis: resistance of small bodied ant species to displacement may be in part driven by resistance to nest raiding, and may represent a general property of invasions by opportunistically nesting ant species.
The suppression by N.fulva of herbivorous arthropod species richness as well as the abundances of all arthropod trophic categories has potential to impact ecosystem function. The functional consequences of the reduction in predator abundance are at least in part offset by the impact of N. fulva on lower trophic levels. The functional consequence of reducing arthropod detritivore abundance, the least impacted group, may be to slow or speed rates of decomposition and nutrient cycling (Gessner et al. 2010). Thus, it is unlikely that N. fulva will strongly alter decomposition rates. However, herbivores of the grass and forb matrix are the most strongly impacted group, declining both in abundance and species richness. In addition to likely direct predation of eggs and early instars, the strength of this reduction may arise because these species need to be stationary to feed, difficult for an insect within an ant population of this density. These changes should translate into reduced rates or patterns of herbivory, potentially altering relative abundances of plant species over time. In grassland systems, insect herbivores tend to enhance plant diversity by preventing dominant species from forming monocultures (Carson and Root 2000). Thus, reduction in herbivore abundance and diversity may reduce long-term co-existence among plant species. Similar impacts have been observed in tropical forests invaded by the ant Wasmannia auropunctata (Roger) which reduced rates of herbivory by leaf-chewing insects (Dunham and Mikheyev 2010).
In habitat uninvaded by N.fulva, we expected no relationship between distance from the edge and arthropod abundance. There was no significant relationship. However, arthropod abundance tended to be highest at the population edge. This trend may result from some unanticipated underlying gradient that N. fulva reverses, or it may arise from insects emigrating from habitat densely occupied by N. fulva and settling in low density areas.
Solenopsis invicta is nearly universal in non-forested habitats in the Southeastern USA west to Central Texas. If N.fulva replaces S. invicta as it spreads, differences between these ants will have critical consequences for natural and agricultural systems regionally. For example, S. invicta is scarce in closed canopy woodlands. However, in addition to thoroughly occupying grasslands, N. fulva reaches high densities in forested environments (E. LeBrun pers. obs.). Thus forest invertebrate and vertebrate fauna that have been less impacted by S. invicta may be under greater threat from this invader. Further, N. fulva exhibits a dramatic seasonal cycle in local abundance. Populations are extraordinarily dense in the summer and fall, but much less so in spring and early summer (E. LeBrun pers. obs.). As a consequence, in contrast with S. invicta (Allen et al. 2004), species that nest in spring may be less directly impacted by these ants. However, because N. fulva reduces general arthropod abundance, food for insectivorous wildlife will be reduced and these animals may suffer greater net impacts as a result.
The long-term environmental consequences of this invasion remain unclear. N.fulva has demonstrated that it can penetrate natural ecosystems. Although undisturbed systems are very rare regionally, N. fulva has invaded intact bottomland forest habitats (E. G. LeBrun pers. obs.) and its success in pasturelands indicates that it will invade intact prairie reserves once it encounters them. The role of disturbance in determining the densities this invader achieves or the impacts it has remains to be investigated. There has been no attempt to eradicate this invader or limit its expansion by regulating transport of goods likely to contain colonies. Thus, in all likelihood, N. fulva will become widespread in urban and non-urban environments in the Gulf Coast region. Lacking data on its aridity and freeze tolerance limits, we cannot predict its potential future range. Equally it is unclear for how long these extremely dense populations will persist, or whether the negative effects we document are transitory (Cooling et al. 2011). However, short-term consequences of this invasion for arthropod communities are clear: reduction or displacement of S. invicta, further degradation of native ant diversity, and a generalized reduction in arthropod abundance. Currently, there is a pressing need for more information about impacts of this invader on other aspects of the habitats it invades, an increased understanding of its physiological tolerances, and its basic ecology both in the USA and in its native range.
We thank Perry Herman and Land Tejas Companies, Ltd for allowing access to their properties. Residents of Iowa Colony provided information on the history of the area. Jerod Romine, Drew Townsend, and Mike Marischen assisted in conducting field work. Alejandro Calixto and Danny Macdonald provided information on the location of established populations. Funding was provided by the Helen C. Kleberg and Robert J. Kleberg Foundation, and the Lee and Ramona Bass Foundation.
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