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The Argentine ant, Linepithema humile: natural history,
ecology and impact of a successful invader
Elena Angulo1,**, Benoit Guénard2,**, Paride Balzani3, Alok Bang4, Filippo Frizzi5,
Alberto Masoni5, Sílvia Abril6, Andrew V. Suarez7, Benjamin Hoffmann8,
Giovanni Benelli9, Hitoshi Aonuma10, Lori Lach11, Palesa Natasha Mothapo12,
Theresa Wossler12, Giacomo Santini5,*
1 Estación Biológica de Doñana (CSIC), Av. Américo Vespucio 26, 41092 Sevilla, Spain
2 School of Biological Sciences, The University of Hong Kong, Kadoorie Biological Sciences Building, Pokfulam Road,
Hong Kong, PRC
3 Faculty of Fisheries and Protection of Waters, South Bohemian 38925, Research Center of Aquaculture and Biodiversity
of Hydrocenoses, University of South Bohemia in České Budějovice, Vodňany, Czech Republic
4 Biology Group, School of Arts and Sciences, Azim Premji University, Bhopal 462022, Madhya Pradesh, India
5 Department of Biology, University of Florence, Via Madonna del Piano 6, 50019 Sesto Fiorentino, Italy
6 Department of Environmental Sciences, University of Girona, Maria Aurèlia Campmany, 69, 17003 Girona, Spain
7 Department of Evolution, Ecology and Behavior and Department of Entomology, University of Illinois, 320 Morrill Hall,
505 South Goodwin Avenue, IL61801 Urbana, USA
8 CSIRO Health and Biosecurity, Tropical Ecosystems Research Centre, PMB 44, Winnellie, NT 0822, Australia
9 Department of Agriculture, Food and Environment, University of Pisa, Via del Borghetto 80, 56124 Pisa, Italy
10 Graduate School of Science, Kobe University, 1-1 Rokkodai, Nada-ku, Kobe, 657-8501 Hyogo, Japan
11 College of Science and Engineering, James Cook University, PO Box 6811, Cairns, QLD 4870, Australia
12 Department of Botany and Zoology, Stellenbosch University, Private Bag X1, Matieland 7602 Stellenbosch, South Africa
* Corresponding author: giacomo.santini@uni.it
** These authors contributed equally to this work
With 1 gure
Abstract: The Argentine ant, Linepithema humile, is one of the world’s worst invasive species, with established popula-
tions in at least 40 countries on six continents. In this review, we synthesise the vast literature on this species in four areas,
concentrating on its introduction to natural systems. The rst section reviews its distribution, habitat preferences, and the
factors promoting its invasion success. Second, we review current knowledge of its ecological impacts on invertebrates,
vertebrates and ecosystem functions. The third section deals with behaviour and genetics, particularly traits promoting
invasiveness. Finally, we address applied issues, emphasising the quantication of the economic costs and eradication strat-
egies associated with L. humile invasion. Despite tremendous research eorts, especially over the past 40 years, numerous
knowledge gaps remain in the understanding of the distribution, ecology, impacts, management, and economic costs of this
species. We conclude by highlighting the most critical gaps and propose a research agenda to tackle the future challenges
in the study of L. humile biology.
Keywords: biological invasions; climate change; Dolichoderinae; invasive species; pest control
1 Introduction
Invasive alien species are transported through human trade
to new regions outside of their native range and ultimately
cause adverse impacts to introduced ecosystems and their
associated biota. The Argentine ant, Linepithema humile
(Mayr 1868, syn. Iridomyrmex humilis; Dolichoderinae), is
listed among the 100 of the world’s worst invasive alien spe-
cies (Global Invasive Species Database 2023). Originating
from a small area in South America (Wild 2004), it was intro-
duced into new regions since at least the mid-19th century
and is now established on all continents, except Antarctica.
Entomologia Generalis, Vol. 44 (2024), Issue 1, 41–61 Open Access
Published online February 25, 2024
© 2024 The authors
DOI: 10.1127/entomologia/2023/2187 E. Schweizerbart’sche Verlagsbuchhandlung, 70176 Stuttgart, Germany, www.schweizerbart.de
Review Article
The Argentine ant is one of the most widely studied ant
species, being the subject of more than 1100 scientic papers
published since 1945. Research on Argentine ants has been
expansive since the late 1980s, peaked around the late 2010s,
and inspired studies on other invasive ants (Suarez et al.
1999) (Supplementary Fig. S1). The topics covered in the lit-
erature encompass all the themes of invasion biology; how-
ever, the interest in this species extends far beyond this eld,
as understanding the behavioural and physiological adapta-
tions and the genetic mechanisms that facilitate its success
pose interesting questions in a broader evolutionary ecology
context. Additionally, due to the ease of maintenance in the
laboratory, the Argentine ant has been used as a model organ-
ism for investigating collective behaviour, decision-making
and self-organisation (e.g. Goss et al. 1989).
This review summarises the current knowledge on this
species to provide a synthesis of hundreds of papers on a spe-
cies with widespread environmental and economic impacts
and identify priorities for future research. Our review cov-
ers distribution and spread, impacts, behaviour, population
biology, as well as eradication strategies. We conclude with
a research agenda, highlighting specic challenges to be
addressed in future studies.
2 Spread and updated distribution
2.1 Global spread (Fig. 1)
Determining the native range of introduced species can be
particularly challenging, especially for widespread species
whose actual distribution is the result of repeated introduc-
tions occurring over several centuries (Suarez et al. 2001;
Wetterer et al. 2009). Nonetheless, with advancements in
taxonomic (Wild 2004), molecular (Tsutsui et al. 2001),
and species distribution modelling studies (Roura-Pascual
et al. 2004), together with thorough samplings in geographi-
cally distant regions (Suarez et al. 2001), the native range
of L. humile is now one of the best understood among
globally-distributed invasive alien ants, even if some gaps
may remain. Linepithema humile is native to the Río Paraná
drainage basin and its tributaries (Río Paraguay and Río
Uruguay), which span across eastern Bolivia, southern
Brazil, Paraguay, Uruguay and northern Argentina. Within
this area, most known records are, however, restricted to a
few kilometres along the river drainage (Wild 2007), while
some populations may have been introduced early on, espe-
cially within more urbanised environments (Tsutsui et al.
2001; Wild 2004).
Outside its putative native range, the Argentine ant has
been recorded in 59 countries, including Argentina and
Brazil, where it is considered non-native in some regions
(Wild 2004; Guénard et al. 2017). It is thus among one of
the most widespread invasive alien ants (Wong et al. 2023),
being recorded from all 12 biogeographic realms colonised
by ants.
It is paramount to discriminate among dierent introduc-
tion stages to discern the ant’s spreading pattern, potential
impacts, and prevent over-predictions in species distribution
modelling. Indeed, invasion stages (introduction, establish-
ment, spread, and impact) reect various ecological lters,
which, once overcome, may lead to the establishment and/
or spread of introduced populations (Wong et al. 2023). A
list of the terms and denitions used throughout the manu-
script is provided in Supplementary Table S1. To date, non-
native outdoor populations of L. humile have been recorded
in 40 countries, including several overseas territories, while
records for 19 other countries have been limited to quarantine
interceptions or indoor populations (Guénard et al. 2017).
Outdoor populations can become invasive when the popula-
tion spreads and causes ecological harm. In most instances
studied, the Argentine ant has been ecologically dominant
and displaced most native ant species (Castro-Cobo et al.
2021); however, sometimes the Argentine ant does not
achieve ecological dominance, coexisting with native spe-
cies, persisting in this status for > 10y (Castro-Cobo et al.
2020b). Investigation of such establishments, without domi-
nance, may yield important insights into the mechanisms for
achieving successful invasion elsewhere.
Linepithema humile most readily invades Mediterranean
and subtropical ecosystems. The invasion history of L.
humile includes at least seven invasion events from the
native to non-native regions, as well as several long-distance,
human-mediated dispersal between invaded ranges (Suarez
et al. 2001; Tsutsui et al. 2001; Sunamura et al. 2009a; Vogel
et al. 2010).
Historical records place the Saharo-Arabian biogeo-
graphic realm as the earliest region invaded by L. humile,
as early as 1858. The island of Madeira was invaded rst,
followed by other Macaronesian Islands, except for the Cape
Verde archipelago (Wetterer et al. 2009). The ant established
populations in the Maghreb in the early 1920s, mainly lim-
ited to coastline areas even if it now occasionally expands
inland (Slimani et al. 2020). Towards the end of the 20th
century, additional records were reported East within the
Arabian Peninsula (Oman, Yemen) and later in Iran (Ghahari
et al. 2009), while records from Turkey require conrmation
(Kiran & Karaman 2020).
Following its introduction in the Western Palearctic realm
as early as 1890, the ant appeared for several decades limited
to a thin fringe of the southern Atlantic and Mediterranean
Sea coastlines to about 25 km inland (Bernard 1983, but see
Espadaler & Gómez 2003), including several islands (e.g.
Balearic Islands, Corsica, Crete; Gómez & Espadaler 2006;
Blight et al. 2009; Masoni et al. 2020; Salata et al. 2020).
Newly established populations have, however, been recently
reported farther north along the western coastal regions of
France (Blatrix et al. 2018; Charrier et al. 2020), inland in
Italy (Frizzi et al. 2023), and temporary outdoor populations
observed in England (Fox & Wang 2016), the Netherlands
(Boer et al. 2018) and Germany (Seifert 2018), apparently
42 Elena Angulo et al.
associated with buildings. For these regions, long-term
survival outdoors remains to be conrmed. Biogeographic
dispersion from at least three European geographic sources
appears likely (Blight et al. 2010a, 2012).
In the Nearctic realm, the Argentine ant was rst recorded
in 1891 in New Orleans, Louisiana (Suarez et al. 2001), and
is now found from Mexico to northern California on the West
coast, and from Florida to North Carolina on the East Coast
of the USA nearly continuously. Non-coastal populations
remain patchier, likely due to multiple local introductions,
and anthropogenic habitat modication creating pockets of
suitable conditions in otherwise unsuitable areas (Menke
et al. 2007; Brightwell et al. 2010; see section 2.2). These, in
turn, provide several bridgeheads expanding the introduced
range further (Espadaler & Gómez 2003; Menke et al. 2007;
Brightwell & Silverman 2011).
Within the Afrotropical realm, most records are from
South Africa, while other records are from Cameroon,
Namibia and Zimbabwe, with population survival in these
last three countries requiring conrmation (Wetterer et al.
2009). In South Africa, where L. humile has been well stud-
ied, it was rst recorded in 1893, and again in 1901 and cur-
rently forms two distinct supercolonies (Mothapo & Wossler
2011) thought to originate from direct introductions from the
native range (Tsutsui et al. 2001; Vogel et al. 2010).
In the Australian and Oceanian realms, Argentine ants
were detected in 1939 and 1990 in Australia and New
Zealand, respectively, with established populations in
New Zealand originating from southern Australia (Corin
et al. 2007b). In Australia, L. humile is mainly found along
coastal areas in the southern half of the country, including
Tasmania and Norfolk Island, and has expanded north as far
as Brisbane (Homann et al. 2011), from potentially two dis-
tinct introduction events (Suhr et al. 2009). In the Pacic,
the species has been recorded from a few islands, including
Hawaii (1916), Easter Island (1987) and the Mariana Islands,
with the latter to be conrmed (Wetterer et al. 2009).
Within the Sino-Japanese realm, outdoor populations
of L. humile have been known from Japan since 1993
(Sugiyama 2000) and have since spread within the south-
ern part of Honshu along a 900 km long region (ranging
from Yamaguchi to Tokyo prefectures), and more recently
to Shikoku island (Ohara & Yamada 2012), with multiple
introductions suspected within the country (Sunamura et al.
2009a). In South Korea, populations of L. humile were
detected in 2019 in the southern part of the country (Busan
Fig. 1. (A) Global distribution of the Argentine ant (Linepithema humile) showing its native (blue) and introduced (orange and yellow)
ranges. Outdoors (orange) and indoors/quarantine (yellow) records are shown separately as they illustrate dispersal and environmen-
tal limitations for the species. Individual records with available coordinates are shown as dots, while polygons, where the species is
recorded, are coloured accordingly, with region denitions following Guénard et al. (2017). In regions where records of dierent types
are reported (e.g. both established and non-established populations), priority is given to native and then established populations. Grey
polygons show countries where no valid records are known. (B) Native range showing intermixing of potential introduced populations
with native populations. Abbreviations: Pr. Paraguay, Ur.: Uruguay.
Ecology and management of the Argentine ant 43
and Kwangyang regions; Lee et al. 2020), where the species
now appears to have become locally dominant. The lack of
records from China is notable considering the presence of
both suitable environments (Jung et al. 2022) and sympatric
species from its native range now established in the country
(e.g. Solenopsis invicta, Wasmannia auropunctata).
Within the Neotropical realm, the distribution of L.
humile appears extensive, but patchy outside its native range,
as the species has been recorded from southern Argentina and
Chile up to the Guyana shield, Panama or Chiapas (Mexico),
although some historical records (e.g. Costa Rica) are now
considered dubious (Wild 2004; Wetterer et al. 2009).
In 2023, L. humile was recorded from Reunion Island at
1200 m elevation, representing the rst population estab-
lished outdoors within the Malagasy realm (Colindre 2023).
Additionally, models based on environmental conditions
predict further suitable areas for its spread in the region
(Jung et al. 2022, but see Roura-Pascual et al. 2011).
Overall, the Argentine ant, despite being established out-
side its native range for over 165 years, keeps expanding
globally through human-mediated long-distance dispersal
(e.g., Iran, South Korea, Mascarene Islands) fuelled by both
primary and secondary introductions (Vogel et al. 2010),
but also through more regional spread with further expan-
sion inland as observed in several countries such as Algeria,
France, Italy or the USA.
In several regions, records of Argentine ants are restricted
to indoor populations or intercepted specimens (e.g., quar-
antine), as observed in the northern parts of the Western
Palearctic and Nearctic realms, especially beyond the north-
ern half of the USA (from Virginia to Oregon). These records
are insightful as they provide information about the human-
mediated capacity of this species to reach new regions while
facing environmental conditions that potentially prevent its
survival and establishment outdoors. However, the situation
within the Oriental realm is likely to be dierent. Records
from Peninsular Malaysia, the Philippines, Taiwan, Thailand,
and Vietnam lack evidence for outdoor establishment,
despite potential climatic suitability within some of these
regions (e.g. Taiwan, Vietnam; Roura-Pascual et al. 2011;
Jung et al. 2022) and further studies conrming propagules
arrival in those regions would be welcome. Therefore, data
from biosecurity interceptions are particularly relevant for
understanding propagule pressure, even in countries where
the Argentine ant is already established (Australia: Suhr
et al. 2019; New Zealand: Corin et al. 2007b), and to dene
the main introduction pathways worldwide. In Australia, sur-
prisingly, most propagules originated from regions unknown
to host outdoor or any Argentine ant populations (Suhr et al.
2019), a pattern also observed in New Zealand (Corin et al.
2007b), suggesting that viable populations may exist near
trade hubs (e.g. Singapore, Fiji, Thailand), but remain unde-
tected; or that other factors regarding the production of trade
records should be evaluated critically.
2.2 Habitat preferences and factors promoting
local spread
The spread of Argentine ants involves two separate pro-
cesses: short-range diusion by budding and long-range
dispersal by human transport (Suarez et al. 2001). The suc-
cessful establishment of an invasive species is inuenced
by a combination of the abiotic and biotic conditions of
the recipient environment and the biological and functional
traits of the invader (Blackburn et al. 2011).
Abiotic factors limit the spread of this temperate-ant into
Mediterranean-type and some sub-tropical areas because
the species cannot tolerate high temperatures (thermal limit
CTmax: 38–40 ºC; Jumbam et al. 2008) and requires high
humidity and moisture levels (Holway et al. 2002b; Menke
& Holway 2006; Menke et al. 2007). It exhibits seasonal
polydomy, retracting into overwintering nests during cold
months and spreading out during warmer months (Heller
& Gordon 2006; Diaz et al. 2014). Couper et al. (2021)
reported a population retraction after an extreme 4-year
drought in northern California. In South Africa, Luruli
(2007) showed that L. humile nests close to waterways, riv-
ers, gardens and regularly watered agricultural sites. Arnan
et al. (2021) showed its ability to occupy empty European
climatic niches. Indeed, L. humile is most abundant closer
to anthropised areas (Holway & Suarez 2006), including
disturbed environments like agricultural elds and clearings
with loose soil providing proper nesting sites (Way et al.
1997; Vonshak & Gordon 2015).
Regarding biotic factors, interspecic competition with
native ants can limit Argentine ant spread. However, evidence
varies, from no evidence (Holway 1998b; Castro-Cobo et al.
2019), to specic native species slowing the invasion spread
(Thomas & Holway 2005; Walters & Mackay 2005; Menke
et al. 2007; Blight et al. 2010). Recently, specic traits and
competitive abilities of native ants have been investigated,
because the species providing resistance seem to be ecologi-
cally dominant, and with strong competitive abilities such as
mass recruitment (Castro-Cobo et al. 2020a). Argentine ant
spread is also facilitated by the mutualistic relationship with
aphids (Grover et al. 2007; Tillberg et al. 2007; Mothapo &
Wossler 2017). Local spread is limited because the species
does not have nuptial ights, and spreads by budding, but it
could be fostered by scavenging vertebrates (Castro-Cobo
et al. 2019; Castro-Cobo et al. 2021).
Finally, dependence on abiotic factors suggests that
global climate change will aect the current distribution pat-
terns of L. humile, retracting in tropical areas but expanding
in higher latitudes (Roura-Pascual et al. 2004; Cooling et al.
2012; Bertelsmeier et al. 2016). However, some populations
have persisted for a long time (Castro-Cobo et al. 2021),
while others that have declined could recover with climate
change (Cooling et al. 2012): in any case, the distribution of
invaded areas will change according mainly to local-scale
environmental conditions (Menke & Holway 2020).
44 Elena Angulo et al.
3 Ecological and environmental impacts
3.1 Impacts on ant communities
Once successfully established in a new area, the Argentine
ant usually displaces much of the local ant fauna resulting
in a change in the structure of communities (e.g. Holway
et al. 2002a; Sanders et al. 2003; Lessard et al. 2009), with
evidence for long-term impacts accumulating (Menke et al.
2018; Achury et al. 2021). Loss of local species richness and
diversity have been reported both in disturbed and undis-
turbed areas, from urban parks (Touyama et al. 2003) and
agroecosystems (Zina et al. 2020) to uninhabited islands
(Naughton et al. 2020). Nevertheless, species displacement
is not the rule, possibly owing to the superior competitive
abilities of certain ant species (Heller 2004). For example,
species of the Tapinoma nigerrimum complex and Lasius
niger have comparable competitive abilities to the Argentine
ant (Blight et al. 2010b; Cordonnier et al. 2020). Small and
hypogeic species (e.g. Stenamma spp., Solenopsis sp. and
Heteroponera imbellis) also appear less aected in some
areas (Ward 1987; Holway 1998a; Rowles & O’Dowd
2009a). Similarly, the small Mediterranean ant Plagiolepis
pygmaea can probably co-occurrt with L. humile due to
its submissive behaviour during interspecic encounters
(Abril & Gómez 2009; Zina et al. 2020). Other native ants
can co-occur because of dierent thermal requirements
along the day, e.g., the thermophilic Cataglyphis oricola
and C. tartessica (Angulo et al. 2011) or dierences in ther-
mal breadth that reduce coexistence along the seasons, e.g.,
Prenolepis imparis (Nelson et al. 2023). In South Africa,
some native ant species (Tetramorium spp. and Meranoplus
spp.) show increased abundance in invaded areas than in
uninvaded ones, although the facilitating role of L. humile
remains unclear (Mothapo & Wossler 2017; Devenish et al.
2021).
The sympatry of L. humile with other invasive ant spe-
cies has been observed and studied in several parts of the
world, leading to various ecological outcomes ranging
from local exclusion to co-occurrence. In the USA, the
Asian needle ant Brachyponera chinensis may displace L.
humile because of its broader seasonal period of foraging
activity and colony expansion (Rice & Silverman 2013).
Contrarily, L. humile can prevail on Solenopsis invicta by
attacking its queens (Brinkman 2006). Instead, in Bermuda,
the Argentine ant co-occurred for at least 25 years with
Pheidole megacephala, which had invaded the area
approximately fty years earlier, without either of them
displacing the other (Haskins & Haskins 1988). In Spain,
where Lasius neglectus and L. humile overlap, coexis-
tence is achieved through spatial segregation (Trigos-Peral
et al. 2021). Future investigations should assess whether
L. humile directly favours or impairs the invasion of other
invasive ants, and under which circumstances (O’Loughlin
& Green 2017).
3.2 Impacts on non-ant arthropods, plants and
ecosystem functions
The substantial alteration of ant communities might aect
native myrmecophilous arthropods, which may not thrive
when their ant host is absent. Populations of the myr-
mecophilous cricket Myrmecophilus kubotai declined in
invaded areas due to its inability to use Argentine ant colo-
nies after native ants have been displaced (Takahashi et al.
2018). On the contrary, the brood production of the lycae-
nid buttery Narathura bazalus was similar in invaded and
uninvaded areas, thus suggesting that the buttery larvae can
also be attended by the Argentine ant (Ikenaga et al. 2020).
It has been estimated that 10,000–100,000 myrmecophiles
may exist; hence understanding the impact of the Argentine
ant on their survival is of the utmost importance (Parker &
Kronauer 2021).
Argentine ants have a wide range of impacts on other
arthropods in the soil. These eects have arguably been det-
rimental in Hawai’i, where the native fauna has no evolution-
ary history with ants (Reimer et al. 2019); their introduction
to the high-elevation shrublands led to a drastic reduction in
diversity and biomass of many arthropod groups (Cole et al.
1992; Krushelnycky et al. 2008). Where arthropods are not
naive to ants, eects vary, likely largely due to some com-
bination of their behaviour, defences, resource overlap with
Argentine ants and indirect eects. In northern California,
several taxa, such as ies and collembolans, were not
detected in invaded areas, whereas other taxa, such as ground
beetles and isopods, increased (Human & Gordon 1997). In
New Zealand, taxa in invaded sites were dierently aected;
for example, collembolans increased, and isopods, amphi-
pods and fungus-feeding beetles declined (Stanley & Ward
2012). In contrast, studies in Californian riparian woodlands
(Holway 1998a), Santa Cruz Island (Hanna et al. 2015), and
coastal scrub in southeastern Australia (Rowles & O’Dowd
2009a) report no eects of Argentine ants on the richness
and abundance of non-ant arthropods. This complex picture
reveals that local factors are probably crucial in determining
the eects of L. humile on these taxa, and their identication
is not always straightforward. The behavioural plasticity of
the Argentine ant to local conditions is probably one of the
keys to understanding the incongruity of its eects (Sagata &
Lester 2009), with future studies focusing on the dierential
sensitivity of non-ant arthropods to invasion needed.
As with many other invasive ant species, Argentine ants
aect plant-associated arthropods, with cascading eects on
plants. Linepithema humile often increases the abundance
of honeydew-producing hemipterans by protecting them
against predators and parasitoids (e.g. Powell & Silverman
2010). However, not all natural enemies are equally aected,
and considerable variability in the eects has been shown.
For example, Daane et al. (2007) showed that in California
vineyards invaded by L. humile, the density of parasit-
oids is lowered and that of predators increased. Contrarily,
Ecology and management of the Argentine ant 45
Calabuig et al. (2015) found that in citrus orchards in Spain,
the abundance of generalist predators decreased while that of
parasitoids increased in the presence of Argentine ants. The
Argentine ant may, however, reduce herbivore populations,
with positive eects on plants (Stanley et al. 2013, but see
Henin & Pavia 2004).
In invaded communities, the Argentine ant disrupts the
benets of myrmecochory either through the displacement
of native ant species and without contributing signicantly
to seed dispersal (Gómez & Oliveras 2003; Frasconi Wendt
et al. 2022) or by reducing seedling emergence rates (Gómez
et al. 2003). In addition, L. humile may modify plant com-
position by favouring direct seed dispersal of invasive over
native plant species, through a selection process relying on
seed size and dispersal distance (Rowles & O’Dowd 2009b).
Ultimately, these mechanisms may lead to important changes
in plant community composition (Christian 2001; Devenish
et al. 2019)
Linepithema humile also disrupts plant-pollinator mutu-
alisms (Blancafort & Gomez 2005) through pollinators’ dis-
placement, competition, predation or deterrence (Lach 2007,
2008; LeVan et al. 2014; Liang et al. 2022). Consequently, a
reduction in plant reproductive success has been associated
with their invasion (Blancafort & Gómez 2005). Therefore,
L. humile can threaten the conservation of pollinators and the
plants relying on them (Lach 2013).
All the relationships between L. humile and other organ-
isms discussed in this section might have cascading eects
on several ecological processes and functions. The alteration
of the community of soil invertebrates, accompanied by the
reduction of the soil microbial biomass, slows the decom-
position in the invaded areas, resulting in a higher C:N ratio
and nutrient content in the soil of invaded sites (Stanley &
Ward 2012).
3.3 Impacts on vertebrates
Recorded eects on vertebrates can be either direct or indi-
rect, and a comprehensive list of known impacts is shown in
Supplementary Table S2.
The displacement of native ants has cascading ecosystem
impacts by disrupting the interactions among native ants and
native predators (Pintor & Bayers 2015). For example, the
displacement of native ants by Argentine ant was interpreted
as one of the causes of the decline of habitat suitability for
the coastal horned lizard, Phrynosoma coronatum, a highly
specialised ant predator (Suarez & Case 2002). A similar
result was found recently for amphibians, which consumed
relatively fewer ants in invaded than uninvaded areas, espe-
cially for the most ant-specialist species, the natterjack toad,
Epidalea calamita (Alvarez-Blanco et al. 2017). In these
cases, the Argentine ant was more dicult to detect, cap-
ture and consume by the predators (but see Ito et al. 2009).
Moreover, if the predator cannot move beyond the invaded
area, the growth and survival of individuals can be compro-
mised. A decrease in juvenile growth has been shown in P.
coronatum under laboratory conditions, in E. calamita and
in the spadefoot toad (Pelobates cultripes) when the main
prey oered was the Argentine ant instead of native ants;
and E. calamita juveniles also suered from lower survival
(Suarez & Case 2002; Alvarez-Blanco 2019). Lack of suit-
able prey decreases territory quality and alters behavioural
patterns. For example, adults of the natterjack toad moved
to uninvaded areas, being less abundant in invaded areas
(Alvarez-Blanco et al. 2017). Therefore, the cascading food
web consequences of Argentine ant invasions could lead to
population extirpations of the most myrmecophilous verte-
brates (as suggested for the horned lizard, Fisher et al. 2002).
Cascading eects of the invasion can alter the prey of
other vertebrates that do not usually feed on ants. In cork
oak forests in northeast Spain, insectivorous bird community
composition diered in invaded and uninvaded areas (Pons
et al. 2010), and there was a reduction of caterpillar biomass,
an essential prey in the hatchling diet (Estany-Tigerström
et al. 2010). Invaded areas represented lower breeding qual-
ity areas for the insectivorous blue tit, Cyanistes caeruleus
with reduced clutches and growth, lighter edglings with
slightly yellower and duller plumage, but with nest and
hatching success higher in these areas, possibly due to the
blue tit predation shift towards alternative preys still avail-
able (Estany-Tigerström et al. 2013). In areas where envi-
ronmental factors could be limiting, the negative impacts of
the Argentine ant could be more pronounced. For example,
a population of the great tit (Parus major) inhabiting a sub-
optimal environment at Doñana National Park (Spain), in the
southern limit of the species distribution, reared poorer qual-
ity ospring in invaded areas (Alvarez-Blanco et al. 2020).
Vertebrate avoidance of invaded areas could also be
caused by direct harassment by the Argentine ant, which
is dicult to determine without specic experiments or
observational data. For example, in the case of Argentine
ant feeding on dead chicks, it is challenging to determine
whether the ants caused the death or if they just recruited
to the carrion (Hooper-Bui et al. 2004; Flores et al. 2017;
Varela et al. 2018). Quantication of nest failure showed
that the Argentine ant has limited impacts on the breeding
of birds: > 2% of failed nests of the dark-eyed junco Junco
hyemalis (Suarez et al. 2005), and even less for Bulwer’s
petrel Bulweria bulwerii (Boieiro et al. 2018). In some of
these cases, ants were observed feeding on the egg contents
while hatching (as also seen in California in the Least bell’s
vireo, Vireo bellii pusillus; Peterson et al. (2004)), swarmed
over hatchling chicks or were seen attacking nestlings that
were still living and later died.
Impacts on bird reproductive success have been linked to
Argentine ant disturbance or harassment. In Spain, the qual-
ity of nestlings of the great tit was negatively aected by
the Argentine ant (Alvarez-Blanco et al. 2020): chicks reared
in invaded areas were smaller, lighter, had lower nutritional
condition and altered oxidative stress balance compared with
chicks reared in uninvaded areas. Because invaded and unin-
46 Elena Angulo et al.
vaded territories were interspersed and shared overlapping
foraging areas, direct disturbance was suggested. Similarly,
in California, Nell et al. (2023) showed that the breeding suc-
cess of the coastal cactus wren (Campylorhynchus brunnei-
capillus sandiegensis) was negatively related to Argentine
ant abundance, with the causal mechanism being harassment.
Lower nest box occupancy in invaded areas was observed in
the great tit (Alvarez-Blanco et al. 2020), in contrast to the
blue tit (Estany-Tigerström et al. 2013).
Direct attacks of the Argentine ant have been observed
toward newly metamorphosed amphibians of three Iberian
species (Alvarez-Blanco et al. 2021), using numerical domi-
nance and spraying defensive compound from the pygidial
gland. The venom’s main compounds are iridomyrmecin,
dolichodial and iridodial, which were considered to be
mainly used as trail and alarm pheromones (Choe et al.
2012). Iridomyrmecin was demonstrated to be the venom
causing amphibian death, as it penetrates the toad skin, caus-
ing paralysis and is ultimately lethal at high doses (Alvarez-
Blanco et al. 2021). The eect of the defensive compound
has also been tested on three amphibians in the Argentine
ant’s native range, for which the chemical is toxic but sug-
gested to be ineective in the eld (Llopart et al. 2023).
4 Behaviour, physiology, and genetics
4.1 Foraging behaviour and trophic niche
Argentine ant colonies in invaded areas can contain mil-
lions of workers with interconnected nests spread over
thousands of square-metres (Tsutsui & Case 2001; Pedersen
et al. 2006; Heller et al. 2008). Large colonies facilitate
resource discovery and retrieval through chemical commu-
nication and recruitment (Beckers et al. 1989). Observations
in California estimated that over 250,000 workers foraged
daily on each tree of a citrus orchard, and nearly 300,000
workers visited urban bait stations each night (Vega & Rust
2003). Extreme polydomy allows for dispersed central place
foraging, facilitating rapid discovery and resource recruit-
ment (Holway & Case 2000; Robinson 2014). Whenever
a resource appears, they are typically the rst species to
nd and monopolise it, with 24-h foraging and persistent
trails observed under suitable conditions (Abril et al. 2007;
Flanagan et al. 2013).
In invaded areas, Argentine ants consistently locate
resources more quickly than native ants do (Human &
Gordon 1996; Holway 1999; Gomez & Oliveras 2003;
Angulo et al. 2011).
Another advantage of polydomy is that once workers from
one nest locate resources, they can be rapidly distributed to
colony mates inhabiting nearby nests. Markin (1968) used a
radioactive marker (32P labelled sugar water) and found the
tracer “had spread almost entirely” through an 81-tree citrus
grove covering 400 m2 in just three days. Using dye in sugar
solution over two weeks, Heller et al. (2008) found the food
was shared up to an area of 647 m2, and Vega & Rust (2003)
found the marker in over 50% of ants up to 61 m away from
the bait station (the maximum distance examined).
Argentine ants are omnivorous, and are predators of
other insects and small vertebrates, scavenge on carrion,
gather plant material, including small seeds, and consume
liquid carbohydrate resources, including insect honeydew
and plant nectar (Holway et al. 2002a; Rowles & O’Dowd
2009b). However, directly quantifying the relative contribu-
tion of dierent food sources to a colony can be challeng-
ing, and stable isotope analysis has allowed more insights
into this topic. Using laboratory colonies of Argentine ants,
Menke et al. (2010) found that workers fed an articial ani-
mal-based diet had d15N values 5.5% greater than those fed
a plant-based diet. Similarly, colonies with access to honey-
dew-producing aphids had d15N values 6% lower than colo-
nies without access to aphids.
Colonies from native populations tend to be more carniv-
orous than colonies from introduced populations (Tillberg
et al. 2007). However, there is also considerable variation
in relative trophic position among Argentine ant colonies
within sites. For example, in their native range, separate
Argentine ant colonies can vary by an entire trophic level
(d15N variation up to 3 ‰) at a single site (Tillberg et al.
2006) or up to 2.6% over one year in the introduced range
(Menke et al. 2010). This variation may be related to the
location of colonies relative to resources (Hanna et al. 2017,
Mothapo & Wossler 2017), or the seasonal production of
brood (Menke et al. 2010). Tillberg et al. (2007) tracked the
leading edge of an invasion of Argentine ants over eight years
and found that the relative trophic position of workers was
lower behind the invasion front relative to those at the inva-
sion front, probably because they become more dependent
on plant-based resources over time (Tillberg et al. 2007).
However, after another eight years, d15N of Argentine ants
at this site increased, suggesting that the decline in relative
trophic position was temporary or relative trophic position
could uctuate over larger time scales (Baratelli et al. 2023).
These results together point to the potential for substantial
spatial and temporal variation in resource assimilation and
the need for more research identifying mechanisms respon-
sible for variation in the Argentine ant diet.
Notably, the type of food source used can inuence the
behaviour of the Argentine ant and, in turn, have cascad-
ing eects on the interactions with other ant species and
overall activity. Sucrose deprivation reduces Argentine ant
worker aggression and overall activity (Grover et al. 2007).
Similarly, the availability of oral nectars increases Argentine
ant activity (Mothapo & Wossler 2017). The monopolisation
of plant-based resources has also been linked to the invasion
success. For example, access to aphid honeydew increases
propagule survival, colony growth and worker activity rate
(Shik & Silverman 2013), and even sucrose can increase
local Argentine ant abundance and spread into forested habi-
tats (Rowles & Silverman 2009).
Ecology and management of the Argentine ant 47
4.2 Reproductive behaviour
Knowledge about the reproductive behaviour of L. humile
comes from experimental and eld studies in the invaded
range. Like many invasive ants, L. humile displays second-
ary polygyny, resulting from gyne acceptance or colony
fusion (Passera et al. 1988). Queen number is inversely pro-
portional to queen fecundity (Keller 1988; Abril et al. 2008;
Abril & Gómez 2020). A few queens lay almost all of the
colony’s eggs, while others contribute few to none (Abril
& Gómez 2014; 2020). However, this inequality disappears
when queens articially experience monogynous conditions,
suggesting cohabiting queens exert some form of reciprocal
reproductive inhibition (Abril & Gomez 2020). The under-
lying mechanism could be queen pheromones since greater
quantities of certain cuticular hydrocarbons (CHCs) are
associated with higher rates of queen productivity and sur-
vival (e.g. Abril & Gómez 2020). Less is known about the
physiological or behavioural eects of these compounds on
nestmate queens, including how they aect fecundity among
mature queens. Additional research must clarify whether
these CHCs are queen pheromones or whether they serve to
signal queen fertility.
In L. humile, the queen number is regulated by adopt-
ing or executing queens. Colonies seem more likely to adopt
non-nestmate queens whose CHC proles are similar to those
of the host colony’s queens (Vasquez et al. 2008). Execution
in this species occurs in late spring; workers execute up to
90% of their colony’s mature queens (Keller et al. 1989).
CHCs also play an important role in this process, since the
survivors display higher levels of certain compounds cor-
related with queen productivity (Abril et al. 2018; Abril &
Gómez 2019), suggesting that workers execute less produc-
tive queens to increase colony productivity. Moreover, the
number of queens executed positively correlates with the
queen number (Abril & Gómez 2019). According to Vargo
& Passera (1992), L. humile workers regulate queen num-
bers to control levels of queen inhibitory pheromones in the
colony. Indeed, mature queens use queen pheromones to
inhibit gyne development in three ways: they cause workers
to behave aggressively towards queen larvae, including can-
nibalism (Bach et al. 1993; Passera et al. 1995); they prevent
dealation and egg-laying in virgin queens (Passera & Aron
1993a); and they prompt workers to attack and kill virgin
alates (Passera & Aron 1993b). The massive execution of
queens in the spring coincides with the period of larval sexu-
alisation. Therefore, by eliminating mature queens, workers
cause a drastic drop in the colony’s levels of queen phero-
mones, allowing new gynes to be produced (Vargo & Passera
1992). However, further research is needed to identify the
causal link between queen pheromones and caste develop-
ment in this species.
A key facet of its invasiveness is that little time elapses
between sexuals emerging and the new queens laying eggs.
Gynes emerge later than males; they reach sexual matu-
rity and mate just a few hours after emergence. As in many
other invasive ants, mating occurs within the natal nest, and
young, mated queens start to lay eggs within a few days
(Passera & Keller 1992). Work remains scarce on the spe-
cies’ reproductive biology in its native range, especially on
queen fecundity, the regulation of queen number, and the
seasonal execution of queens. This knowledge could help us
better understand whether the species’ reproductive biology
has changed throughout its invasion history.
4.3 Aggressive interactions
Intraspecic aggression among non-nestmates in invasive
populations of Argentine ants has been investigated in detail
due to its impact on the formation of supercolonies via
unicoloniality. Reduced aggression or its complete absence
between non-nestmates can yield a numerical advantage to
L. humile, allowing it to exert aggression on other supercolo-
nies and towards other species.
4.3.1 Intraspecific aggression among non-nestmates
Although unicoloniality (see Supplementary Table S1) is
observed within the native range of Argentine ants, super-
colonies expand over short distances (0.05–6 km) compared
to introduced populations (Vogel et al. 2010). Beyond these
distances, ghting was commonly observed at all spatial
scales, suggesting sub-structuring of populations (Suarez
et al. 1999; Tsutsui et al. 2003; Heller 2004; Blight et al.
2017). In the invaded range, individuals from nests separated
by hundreds or thousands of kilometres do not show aggres-
sion, forming supercolonies that spread over vast distances.
Indeed, unicoloniality is common in the invaded range, as
seen in the USA (Holway et al. 1998; Tsutsui et al. 2003;
Thomas et al. 2006), Europe (Giraud et al. 2002; Blight
et al. 2012; Castro-Cobo et al. 2021), Japan (Sunamura
et al. 2009a; Inoue et al. 2013), South Africa (Mothapo &
Wossler 2011), Australia (Björkman-Chiswell et al. 2008;
Suhr et al. 2011), and New Zealand (Corin et al. 2007a). In
these populations, aggression between non-nestmates can
be completely absent (Tsutsui et al. 2003), mildly present
(Giraud et al. 2002) or replaced by increased rates of allog-
rooming and antennation (Björkman-Chiswell et al. 2008).
The aggression in these introduced populations remains sig-
nicantly lower than levels observed among non-nestmates
in the native populations of L. humile (Blight et al. 2017).
In experiments involving aggression bioassays between
supercolonies from dierent continents, it was found that the
major supercolonies of Europe, California, Japan, Australia,
New Zealand and Hawai’i showed no reciprocal aggres-
sion, but showed aggression to secondary supercolonies or
smaller colonies from South Africa, California and Hawai’i
(Sunamura et al. 2009b; van Wilgenburg 2010b), indicating
that there is one supercolony (the “main” or the “large”) that
has a trans-continental spread. There has been evidence for
the existence of high levels of aggression between supercol-
onies from the same country, sometimes of the same region,
as seen in the USA (Chen & Nonacs 2000; Buczkowski et al.
48 Elena Angulo et al.
2004; Thomas et al. 2006; van Wilgenburg et al. 2022), Japan
(Sunamura et al. 2009a), South Africa (Mothapo & Wossler
2011) and Mediterranean Europe (Blight et al. 2009; Abril &
Gómez 2011; Blight et al. 2012).
There are endogenous and exogenous sources of nest-
mate recognition in Argentine ants, modulating aggression
in the invaded range. Endogenous sources are related to
low genetic diversity (see section 4.4). This causes similar-
ity in CHC proles and ultimately results in unicoloniality
over large areas. Genetically more homogenous colonies
also attack genetically diverse colonies (Tsutsui et al. 2003).
Wherever genetic diversity is higher, as it occurs in native
populations, or where multiple introduction events raise the
genetic diversity in certain spatial pockets such as airports
or ports, non-nestmates display higher levels of aggression
(Suarez et al. 1999) or allogrooming and antennation (Giraud
et al. 2002; Björkman-Chiswell et al. 2008) as compared to
populations with low genetic diversity.
The role of non-heritable exogenous environmental com-
ponents, such as diet, has been debated in lowering aggres-
sion between non-nestmates. In general, similar diets could
lead to similar CHC proles, reduced aggression among
non-nestmates and may lead to colony fusion (Buczkowski
et al. 2005, but see Suarez et al. 2002). Aggression has also
been reduced with time under laboratory conditions (Chen
& Nonacs 2000), or by the absence of certain macronu-
trients such as carbohydrates in the diet (see section 4.2).
Inversely, processes that interfere with an individual’s CHC
prole, such as a dierent diet, or even engaging with the
prey leading to the transference of the prey’s CHC onto the
ants, can lead to disruption of colony integrity (Liang &
Silverman 2000; Liang et al. 2001; Silverman & Liang 2001;
Buczkowski & Silverman 2006). Colony identity can also
play a role, as dierent colonies respond dierently to diet
changes (Buczkowski & Silverman 2006). Moreover, highly
aggressive colonies maintain their aggressiveness despite
dietary changes, whereas medium- or low-aggression colo-
nies reduce their aggression when diet becomes similar
(Buczkowski et al. 2005). However, diet is proposed to be
ineective in creating conditions suitable for unicoloniality
(Thomas et al. 2005; van Wilgenburg et al. 2022).
Other modulators of aggression are memory and prior
experience of an aggressive interaction with a non-nestmate.
Individuals escalate aggression in future encounters based
on their prior experience of facing a non-nestmate, and, in
response, showing or receiving aggression. This experi-
ence can modulate future interactions for up to a week (van
Wilgenburg et al. 2010a). At the colony level, less aggressive
colonies raise their aggression levels in future interactions
after encounters with a hostile and highly aggressive colony,
transitioning from asymmetrical to symmetrical colony
interactions concerning aggression (Thomas et al. 2005).
Finally, intraspecic aggression among non-nestmates in
Argentine ants is also a context-dependent process based on
social and ecological contexts of discrimination of a non-
nestmate. For example, increased aggression is observed
when the context related to nest proximity exists in the form
of numerous nestmates or familiar territory, but the aggres-
sion is reduced when such contexts are removed (Buczowski
& Silverman 2005). Such context-dependent cues often give
dierent results in dyadic interactions versus colony inter-
actions. The role of such exogenous and context-dependent
cues in modulating aggression also highlights the impor-
tance of using appropriate social and ecological context-
based behavioural assays for testing aggression in L. humile
in the future.
4.3.2 Interspecific aggression
Argentine ants are usually highly aggressive towards other
species, and their ability to break the discovery-dominance
trade-o (Human & Gordon 1996; Holway 1999) is part
of their success as invasive species (section 3.1). There
are inherent dierences between dierent supercolonies of
Argentine ants, and they are dierentially aggressive towards
other ant species. The dierential invasion success of dier-
ent supercolonies could also be a function of their inherent
aggressiveness, with aggressive supercolonies attaining a
wider range and milder colonies receiving more aggression,
resulting in smaller territories and fewer resources, and a
narrower range (Abril & Gómez 2011).
Argentine ants display a “bourgeois strategy” in ght-
ing, behaving either as “hawks” (behaviourally aggres-
sive) or “doves” (behaviourally submissive) depending on
group size (Carpintero & Reyes-López 2008). Owing to
their small size, Argentine ants usually lose one-to-one con-
tests or numerically-matched interspecic contests between
smaller colonies or groups of workers (Frizzi et al. 2023).
However, as the group size grows, Argentine ants show a
change in behavioural strategy, with rapid recruitment of
individuals towards aggressive interactions (Buczkowski
& Bennett 2008), active cooperation among individuals
where multiple individuals of L. humile ght a single indi-
vidual of the native ant (Sagata & Lester 2009; Blight et al.
2010b; Bang et al. 2017; Leonetti et al. 2019), and chemical
defences combined with physical aggression (Buczkowski
& Bennett 2008; Welzel et al. 2018). Native ant species who
are behaviourally aggressive can withstand Argentine ant
invasions, as happens in Tapinoma c.f. nigerrimum (Blight
et al. 2010b), and the Australian native ant, Iridomyrmex
rufoniger (Walters & Mackay 2005). Interestingly, unlike
many native ant species, L. humile individuals show immu-
nity to their nestmates’ chemical defences, and these serve
a dual purpose, incapacitating the opponents and simultane-
ously acting as an alarm pheromone to recruit more nest-
mates to aggressive interactions (Buczkowski & Bennett
2008; Welzel et al. 2018). Such social facilitation around
aggressive interactions allows L. humile to advance from a
local numerical advantage of ghting in groups to a global
numerical advantage of higher absolute numbers, and assists
in invading new areas and outcompeting larger and ecologi-
Ecology and management of the Argentine ant 49
cally dominant native ants (Human & Gordon 1996; 1999;
Holway 1999).
4.4 Genetic variability/population genetics
As in other invasive ants, genetic diversity is higher in native
than in introduced populations, due to founder eects (e.g.
genetic bottlenecks and single introductory events, Suarez
et al. 2008; Vogel et al. 2010). Low genetic diversity is linked
to its unicolonial social organisation, which is an important
attribute of the Argentine ant’s invasive potential. Sib mat-
ing, queen executions and “genetic cleansing” (i.e. the death
of individuals with rare alleles) may further contribute to the
reduction of genetic diversity, the decrease in the number of
haplotypes, and the increase of nestmate relatedness, which
nonetheless remains very low (Giraud et al. 2002; Keller &
Fournier 2002; Inoue et al. 2015). Such a condition contrib-
utes to reducing intraspecic competition and aggressive-
ness (Inoue et al. 2015; but see Sanmartín-Villar et al. 2022)
and confers these population’s further local ecological domi-
nance and an unrestricted growing potential (Holway et al.
1998), at least during the rst phase of their invasive process
(Lester & Gruber 2016).
The analysis of the genetic structure of L. humile super-
colonies using both nuclear and mitochondrial DNA showed
that these are characterised by i) the presence of many
queens in each nest, ii) nestmates relatedness not dierent
from zero (Pedersen et al. 2006), iii) the presence of a single
mitochondrial haplotype per supercolony (Sunamura et al.
2009a; Vogel et al. 2009), and iv) a strong genetic dieren-
tiation between supercolonies, primarily in their native range
(Thomas et al. 2006). All these features further support that
supercolonies are closed breeding units, with no signicant
inbreeding and limited gene ow. The competition occurring
where dierent supercolonies come into contact contrib-
utes to regulating the stability and evolution of unicolonial
structures and their gene pool in native and invaded ranges
(Sanmartín-Villar et al. 2022).
The invasion history at the global scale (section 2.1) is
still debated, and genetic analysis, together with behavioural
assays, can help elucidate the relationship between dier-
ent populations (Corin et al. 2007b). Analysing maternally
inherited mitochondrial DNA is a key tool for investigat-
ing invasion histories based on founding queen dispersal
(Tsutsui et al. 2001; but see Vogel et al. 2010). The analysis
of mitochondrial marker genes, such as cytochrome c oxi-
dase (CoI, CoII) and cytochrome b (Cytb), allowed identi-
fying more than 19 dierent haplotypes in both native and
introduced ranges (Sunamura et al. 2009a; Vogel et al. 2010;
Park et al. 2021). Several have an intercontinental distribu-
tion due to long-distance human-mediated dispersal events
(Suarez et al. 2001). For example, the LH1 haplotype is
widespread across Europe, North America, Australasia and
Japan, while LH3 was found in South (Chile and Ecuador)
and North America, but also Asia (Vogel et al. 2010; Inoue
et al. 2013; Seko et al. 2021a).
5 Applied issues
5.1 Economic impacts
Although the evidence for the ecological impacts of inva-
sive ants, including the Argentine ant, has accrued rapidly
in the past few decades (section 3), the synthesis of their
economic impacts have only recently been collected and
analysed (Diagne et al. 2020; Angulo et al. 2022). Between
1980–2020, L. humile caused economic impacts worth
US$ 19.2 million to the global economy (standardised to
US$ 2017 values), thus causing global economic impacts
of US$ 480,000 annually. However, even if no costs were
recorded, the species started to become a severe urban and
agricultural pest one century before, around the end of the
19th century (Newell & Barber 1913). Compared to other
invasive ants, the Argentine ant is the fourth costliest spe-
cies after Solenopsis spp. (US$ 32 billion), W. auropunc-
tata (US$ 19 billion) and A. gracilipes (US$ 66 million).
The decadal costs of L. humile have remained stable in the
last four decades, with a spending of about US$ 3–5 million
per decade (Angulo et al. 2022). However, these gures are
highly underestimated, as they consider only costs reported
in reliable publications, mainly in English, creating a lin-
guistic and geographic skew in cost representation (Angulo
et al. 2022).
Although L. humile has invaded more than 30 countries
worldwide (section 2.1), the economic impact information
has been reported only from 7 of these, from which 6 are
high-income economies (Australia, Ecuador, Japan, New
Zealand, Portugal, Spain, and the USA; Supplementary Fig.
S2). Among them, the number of locations reporting costs
is two to three orders of magnitude lower than the number
of locations from where Argentine ants are reported in these
countries (Angulo et al. 2022). Furthermore, the costs among
these seven countries are highly disparate, with the highest
costs coming from Australia (84% of total costs), followed
by New Zealand (5%), USA (4–5%) and Japan (3–4%). The
costs in Europe represent <1%, and Africa and mainland
Asia do not have any reliable cost reported (Angulo et al.
2022), despite the spread of L. humile on these continents.
The under-representation of costs from the low- and middle-
income economies are likely language-related and/or related
to the emerging nature of the discipline in these regions
(Angulo et al. 2021; Bang et al. 2022).
Most economic costs generated by L. humile invasions
were in human-modied environments (87% of total costs),
open forests (3%) and scrub forests (2%) (Angulo et al.
2022). Over 99% of costs are due to management, and more
than 90% are related to post-invasion management, including
control programs, eradication campaigns and containment
operations (Supplementary Fig. S2; Angulo et al. 2022). This
gure contrasts with those obtained when all invasive ants
or all invasive species are considered, where management
accounts only for 10% or less (Diagne et al. 2021; Angulo
et al. 2022). This dierence could be due to a bias toward
50 Elena Angulo et al.
high-income countries whose governmental services or o-
cial organisations can aord management. A more uniform
geographical representation of economic impact entries
could mitigate this trend.
5.2 Eradications and control
Due to the high environmental impacts in its invaded range,
Linepithema humile has been the target of several control and
eradication attempts. The use of chemicals against L. humile
dates back to the end of the 19th century (Homann et al.
2009). The control of invasive ant populations, including L.
humile, has been the subject of previous reviews, and we refer
to these papers for a comprehensive description of the meth-
ods employed (Silverman & Brightwell 2008; Homann
et al. 2009; 2016). Here we focus on eradication attempts,
trying to elucidate why some of them failed and others were
successful. A summary of the literature about management
published since 2010 is reported in Supplementary Table S3.
Reports of eradications for eleven species with estab-
lished populations outside their native ranges exist, among
which, L. humile has been eradicated from the greatest area
(~16,000 ha) and the greatest number of times (~3,000 dis-
crete populations) (Homann et al. 2011; 2016). Most of
these population-level eradications were from a single pro-
gram conducted in Western Australia (Van Shagen et al.
1994), with an average population size of about 10 ha, and
the largest of approximately 300 ha (Homann et al. 2011).
Most of those eradications were achieved using toxic sprays,
primarily of organochlorines prior to their deregistration.
However, sprays are now rarely used for broadscale (>tens
of ha) eradications because of signicant non-target issues.
Eradication attempts of L. humile using products other
than sprays (granular baits and gels/pastes) had mixed results
(Silverman & Brightwell 2008), with only two reported to
have achieved eradication. The rst, on Tiritiri Matangi Island
in New Zealand, eradicated two populations covering 10 and
1 ha, respectively (Green 2019). The eort took 16 years,
involving paste baits containing 0.01% pronil dispersed
manually at densities no less than one per 3 m2 over the entire
areas multiple times and only wherever residual populations
were found. The second, in Japan, also involved two popula-
tions covering 8.5 and 16 ha, respectively (Sakamoto et al.
2017). That program primarily used paste baits containing
0.005% pronil placed every 5–10 m along buildings, also
using a spray containing 0.005% pronil whenever brood
was found. Other eradications that remain unreported in the
scientic literature relate to three populations on Norfolk
Island, Australia, covering 2 ha and less, that were treated
with paste baits containing 0.6 g/kg pronil placed every
few m2 over the entirety of the areas an unknown number of
times. Inoue et al. (2015) used paste baits with pronil over
a period of 11 months at two doses and showed 99.8% reduc-
tion of L. humile populations while having limited eects on
other arthropods. Buczkowski & Wossler (2019) highlighted
that pronil is eective against L. humile even at very low
doses (ng). Interestingly, they showed that high secondary
mortality can be achieved through horizontal transfer within
the population; in eld plots, releasing pronil-sprayed
workers led to a > 90% reduction of the L. humile population
within 24 h (see also Hooper-Bui et al. 2015).
So why are there so few eradications of L. humile using
products other than sprays, and what can be done to bolster
change? Aside from numerous administrative and technical
reasons why eradications fail (Myers et al. 2000; Simberlo
2009), the greatest issue appears to be the relatively low
ecacy of individual treatments. Eective baiting can be
dened as one or both scenarios leading to the extirpation of
populations: the reproductive queens being incapacitated, or
the workers being killed. The greater the number of queens
and/or workers aected by individual treatments, the fewer
repeat treatments needed to achieve eradication, but the rela-
tive merits of the two scenarios to achieve eradication remain
to be determined (Silverman & Brightwell 2008).
For unknown reasons, the eciency of individual eld
treatments on L. humile populations is lower than for some
other invasive ant species often targeted for eradication.
This disparity is most striking with using hydrogels as bait
(Buczkowski et al. 2014). Even when using extremely low
concentrations (<0.006–0.0007%) of pronil or thiameth-
oxam, which allows workers to return to the nest and share
the bait with nestmates before they die, L. humile abundance
decline after each treatment is more of a noticeable drop (e.g.
~20%) than a crash (e.g. > 80%) (Boser et al. 2014; Rust
et al. 2015). Notably, no eradication of an entire L. humile
population has been achieved yet, even after 12 or more
applications spaced approximately weekly to monthly apart
(Boser et al. 2017; Homann et al. 2023). In comparison,
a single treatment using the same hydrogel bait contain-
ing pronil will kill over 99% of yellow crazy ant, A. gra-
cilipes, workers, and three treatments spaced three months
apart give almost certainty of eradication (Homann et al.
2023). Whether this same disparity also occurs with granular
products is unclear because no eradication programs have
reportedly used granular products. Ultimately, this dierence
between outcomes for dierent species suggests that it is not
the concentration of the active compound that is the pre-
dominant issue, but something about biology that requires a
change in baiting techniques.
One possibility is a simple dierence in food dispersal
within colonies between queens, brood, and workers. Markin
(1970), however, showed that the number of ants to receive
food, the speed and the distribution patterns among castes
were similar to that of non-invasive species. Another possi-
bility is that baits are non-preferential to environmental food
sources, and therefore not enough of the active ingredient is
being taken to the nests. Notably, ants often prefer complex
nectars over simple sugars (Blüthgen & Fiedler 2004), and
hydrogel baits used to date against L. humile were composed
solely of sucrose. Also, when there is an abundant alterna-
tive food source, workers that have fed on a bait would have
Ecology and management of the Argentine ant 51
reduced opportunity to share the bait with other individuals
because fewer individuals would be hungry (Markin 1970).
Another possibility could be that, per standard body weight,
workers require less toxicant to be killed than queens, at
least for hydramethylnon (Hooper-Bui & Rust 2000); poten-
tially with workers dying quickly before providing the lethal
dose to a queen. There also seems to be a disparity between
bait constituents and queen feeding preferences. Liquid and
hydrogel baits are composed solely of sugar, but queens are
preferentially fed protein (Markin 1970). Ultimately, this
may be inconsequential if only workers need to be extirpated
to achieve eradication.
6 Future research
Although Argentine ants are among the most studied ant spe-
cies, this review identied several important knowledge gaps
that could be important future priorities for research.
One poorly-understood phenomenon with huge manage-
ment implications is why some introductions result in suc-
cessful invasions, whereas others result in the establishment
of the ant without spreading. The latter are underreported;
however, their frequency and the factors that cause them are
essential to improve our ability to predict local spread and to
attempt its control. Comparative or meta-analysis studies at
the global scale, including multiple invaded sites and native
areas for comparison, will provide insights into the fre-
quency of establishments that do not result in invasions and
their drivers. In particular, more records could be searched in
areas with long-term monitoring, such as California (Menke
& Holway 2020) or Japan (Inoue et al. 2013), or obtained by
extensively resampling regions invaded in the past.
Despite the vast literature on the ecological eects of
L. humile on other species and ecosystems, there are still
knowledge gaps in the comprehension of why some non-ant
arthropod communities are more sensitive to the invasion
than others. Regarding vertebrates, future studies should
pay special attention to areas and seasons where vertebrates
coexisting with the Argentine ant are at their most vulner-
able stage, and to the eects on specialised ant predators.
Particularly relevant are the sublethal impacts of the venom
associated with reduced body condition, survival and devel-
opment of juvenile amphibian and avian ospring (Alvarez-
Blanco 2019; Llopart et al. 2023), which may show delayed,
long-term consequences on populations that can be easily
overlooked. The impact of the venom should be extensively
investigated to understand its toxicity for other vertebrates,
such as mammals, altricial birds or reptiles. Finally, irido-
myrmecin can deter pollinators and hence have hidden cas-
cading eects on plants that are still to be fully quantied
(e.g. Wilson et al. 2020).
The link between the monopolised food sources (e.g.
carbohydrates and plant-based resources vs. proteins) with
ants’ overall activity and aggressiveness and its eects on
invasion success deserve further studies. Also, the adaptive
change in dietary breadth as key to success is another mecha-
nism worth exploring in depth (Seko et al. 2021b).
Quantifying the economic impacts of invasive species
such as the Argentine ant is a necessary step in highlighting
that biological invasions pose not only ecological but also
socioeconomic challenges. Despite its worldwide distribu-
tion, estimating the costs of Argentine ant invasions is still
in its infancy. The current gures merely represent the lower
limit of the actual economic impact due to widespread gaps
in reporting from several geographic areas, economies, habi-
tats, and sectors.
Further trials and research are needed to elucidate
the causes inhibiting eradication success. One important
advancement could be linking genetic analysis to eradica-
tion techniques to use the most eective products for each
specic population. For example, Hayasaka et al. (2015)
showed that supercolonies characterised by dierent domi-
nant haplotypes have dierential susceptibility to pronil
baits, and this nding opens new perspectives towards tar-
geted eradication/control strategies. Control based on RNA
interference (RNAi) is also promising. This technology is
based on a highly specic, post-transcriptional gene inacti-
vation process, triggered by double-stranded RNA (dsRNA)
homologous to the gene sequence to be suppressed. Ideally,
it is possible to design a dsRNA that interrupts one or more
vital genes, such that only the target species would be killed.
RNAi is now functional as a topical spray for some agri-
cultural pests (Hoang et al. 2022), but remains to be fully
functional as an ingested toxicant just like typical ant baits.
The Argentine ant could be the ideal subject for testing and
developing such a technique, which is gaining attention to
control invasive insects, including ants (e.g. Allen 2021).
To conclude, L. humile, one of the 100 worst globally
invasive species, has been extensively researched, espe-
cially in the last 100 years, coinciding with its global range
expansion. This review has synthesised historical and recent
advances regarding the global invasion of the species, with
a focus on its natural history, behaviour, genetics, ecological
and socioeconomic impacts and future research directions.
Such reviews have the potential to inform researchers, prac-
titioners and policymakers on the existing knowledge and
knowledge gaps on species of global concern.
Acknowledgements: EA acknowledges support by the Junta de
Andalucía, Consejería de Universidad, Investigación e Innovación,
PROYEXCEL_00688 within the PAIDI 2020. AB acknowledges
Azim Premji University’s Grants Programme for support (UNIV
RC00326). GS and AM were supported by the National Biodiversity
Future Center. BG was supported by a Visiting Professor grant by
Università degli Studi di Firenze. We are grateful to Laurent Colindre
for kindly sharing with us his unpublished materials on new ndings
of L. humile in the Reunion Island. Comments from two anonymous
reviewers improved a previous version of the manuscript. We thank
K. P. Akilan for Figure S2 in the Supplementary Materials.
52 Elena Angulo et al.
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Manuscript received: July 4, 2023
Revisions requested: August 23, 2023
Revised version received: October 4, 2023
Manuscript accepted: January 11, 2024
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Ecology and management of the Argentine ant 61
