+ Author Affiliations
- ↵*Corresponding authors’ e-mail address: guillaume.besnard@univ-tlse3.fr, peter.cuneo@rbgsyd.nsw.gov.au
- Received February 23, 2016.
- Accepted August 1, 2016.
Abstract
Invasive trees are generally seen as
ecosystem-transforming plants that can have significant impacts on
native vegetation,
and often require management and control.
Understanding their history and biology is essential to guide actions of
land managers.
Here, we present a summary of recent research into
the ecology, phylogeography and management of invasive olives, which are
now established outside of their native range as
high ecological impact invasive trees. The parallel invasion of European
and African olive in different climatic zones of
Australia provides an interesting case study of invasion, characterized
by
early genetic admixture between domesticated and
wild taxa. Today, the impact of the invasive olives on native vegetation
and ecosystem function is of conservation concern,
with European olive a declared weed in areas of South Australia, and
African
olive a declared weed in New South Wales and
Pacific islands. Population genetics was used to trace the origins and
invasion
of both subspecies in Australia, indicating that
both olive subspecies have hybridized early after introduction. Research
also indicates that African olive populations can
establish from a low number of founder individuals even after successive
bottlenecks. Modelling based on distributional data
from the native and invasive range identified a shift of the realized
ecological niche in the Australian invasive range
for both olive subspecies, which was particularly marked for African
olive.
As highly successful and long-lived invaders,
olives offer further opportunities to understand the genetic basis of
invasion,
and we propose that future research examines the
history of introduction and admixture, the genetic basis of adaptability
and the role of biotic interactions during
invasion. Advances on these questions will ultimately improve
predictions on the
future olive expansion and provide a solid basis
for better management of invasive populations.
Key words
Introduction
Exotic plant invasions are a major factor of global change and a significant threat to biodiversity (Mooney 2005).
Increasing rates of plant introductions, linked to the expansion of
global trade, suggest they will continue to pose conservation
challenges in the future (Gallagher et al. 2010). In the recent past, the rate and risk associated with alien species introductions has increased enormously due to the rapid
escalation in human alteration of the environment (Pimentel et al. 2001). Plant invaders can significantly alter the fire regime, nutrient cycling, hydrology and energy budget of native ecosystems
(Mack et al. 2000). Through competitive interactions, they can directly reduce native plant diversity and abundance (Lake and Leishman 2004; Pyšek and Richardson 2006). To limit the impact of invasive species, it is essential that management actions are guided by knowledge of ecological
requirements and evolutionary drivers favouring invasiveness.
Woody invaders are generally seen as ecosystem-transforming plants (Richardson and Rejmánek 2011; Nuňez and Dickie 2014). First, shading from invasive trees and shrubs has been repeatedly identified as detrimental to native understory diversity
(e.g. Hobbs and Mooney 1986; McKinney and Goodell 2010). The conversion of vegetation from an open stand to a closed canopy will generally be accompanied by microclimatic changes
such as higher humidity and lower temperatures (Gordon 1998). Second, woody plants have a high dependence on mutualists, both aboveground (for seed dispersal and/or pollination) and
belowground organisms (e.g. mycorrhizal fungi; Nuňez and Dickie 2014). Consequently, tree invasions are generally associated with disruption of mutualistic interactions by species exclusion
or recruitment of particular local species by facilitation (e.g. Mitchell et al. 2006). Tree invasions have also been linked to the co-introduction of soil microbes (e.g. Dickie et al. 2010; Ndlovu et al. 2013; Le Roux et al. 2016), potentially leading to major shifts in soil nutrient cycling that could in turn result in invasional meltdown (Dickie et al. 2014). Such ecosystem changes at both the macro- and micro-bial levels often make the restoration of invaded habitats a challenge
for land managers (Traveset and Richardson 2011).
The resilience of ecosystems is, however, variable and remains
difficult to predict, and the role of mutualism and antagonistic
interactions needs to be better documented (e.g. Palmer et al. 2008; Kaiser-Bunbury et al. 2010; Moeller et al. 2015).
Understanding the processes and factors leading to successful tree invasion has become a major topic in invasion biology (e.g.
Richardson et al. 2007; Richardson and Rejmánek 2011; Zenni et al. accepted). Many trees and shrubs that have become invasive were introduced for specific purposes, for example forestry trees
[e.g. Acacia spp. (Richardson et al. 2011; Thompson et al. 2015), Eucalyptus spp. (Doughty 2000), Pinus spp. (Richardson et al. 2007), Prunus serotina (Pairon et al. 2010)], ornamentals [Ligustrum spp. (e.g. Hoyos et al. 2010), Miconia calvescens (Le Roux et al. 2008)] or crops [e.g. Olive (Cuneo and Leishman 2006), Psidium cattleianum (Ellstrand et al. 2010)], while others have been accidentally introduced (e.g. Le Roux et al. 2008).
Population genetic studies of invasive trees has revealed different
histories of introduction, with invasion also strongly
linked to the level of propagule pressure [e.g.
usually high pressures in forest trees or crops, with multiple
introductions
from distinct provenances (Pairon et al. 2010; Le Roux et al. 2011; Mandák et al. 2013; Zenni et al. 2014) vs. low pressures in some ornamentals (Le Roux et al. 2008)]. Native populations of tree species often cover large areas and show considerable genetic variation in adaptive traits
to fit local environmental conditions (e.g. seed size, seed dormancy, cold hardiness, bud phenology; Mátyás 1996; Savolainen et al. 2007). In the invasive range, genetic admixture between distinct provenances of trees has been documented (e.g. Le Roux et al. 2011; Zenni et al. 2014). This admixture may offer the possibility to produce new genotypic combinations (Ellstrand and Schierenbeck 2000), but natural selection remains the main driver in adaptive switches to a new range (Zenni et al. 2014). Importantly, phenotypic plasticity has been also reported to play a central role in the rapid evolution of invasiveness
(e.g. Richards et al. 2012; Matesanz et al. 2012).
In this review, we examine the European olive (Olea europaea L. subsp. europaea), which is a major crop species and iconic tree of the Mediterranean region, and the related African olive (O. europaea subsp. cuspidata) which is a tropical wild olive primarily from southern and eastern Africa. Both these subspecies of O. europaea
have become vertebrate dispersed invasive trees following horticultural
introduction outside of their native range, particularly
in Australia (Cuneo and Leishman 2006).
As European and African olives originate from different climatic zones
and are known to hybridize, they provide an interesting
case study of parallel invasion with potential
admixture in Australia that can be traced with molecular markers. We
also provide
an overview of the ecology of these invasive
olives, and identify further research needed to guide future management
and invasion
risk.
Geographic distribution and diversity of olives
Olives (O. europaea L., Oleaceae) are native to the Old World (Médail et al. 2001; Green 2002). Wild olives are naturally distributed over three continents in highly variable environments and thus considered to have
high genetic diversity and adaptive capacity for naturalization and invasion in a wide range of habitats (Médail et al. 2001; Green 2002). Six olive subspecies are currently recognized but only two taxa have a large native distribution: the European olive (subsp.
europaea) in the Mediterranean basin, and the African olive (subsp. cuspidata) from South-East Himalaya to Southern Africa. Molecular data have been used to investigate the diversification of the olive
complex (e.g. Besnard et al. 2009). Several lineages have been described for olives based on various genetic markers, and their common ancestor dates back
to the Late Miocene (Besnard et al. 2009).
It is believed that the formation of the Saharan desert created a major
geographic barrier to gene flow between North African-Mediterranean
and Tropical African olives. Olive taxa are easily
distinguished based on genetic data, but can hybridize when in contact
leading to genetic admixture (e.g. Besnard and El Bakkali 2014; Cacères et al. 2015).
The cultivated olive tree originated from the Mediterranean Basin (Green 2002). It was probably first domesticated during the Copper Age in the Near East and underwent secondary diversification in central
and western Mediterranean areas (Kaniewski et al. 2012; Zohary et al. 2012; Besnard et al. 2013). Since the beginning of historical times, the cultivated olive and wild relatives have been spread by humans for various
reasons (e.g. olive production, rootstocks, ornament or forage; Carrión et al. 2010; Margaritis 2013; Besnard and Rubio de Casas 2016). In the native range, population turnover of olives is considered to be slow since millennial wild or cultivated trees are
known in different places (e.g. Baali-Cherif and Besnard 2005; Arnan et al. 2012; Bernabei 2015). In addition, when the tree is cut or destroyed aboveground, for instance through fire or heavy frost, even ancient trees
are able to resprout (e.g. Baali-Cherif and Besnard 2005; Therios 2009). This strategy allows for long persistence of olive individuals, and may partly explain the symbolism associated to this
species (Kaniewski et al. 2012).
Since the 19th century, European and
African subspecies have been introduced and become invasive in South
East Australia and
New Zealand, but the African olive has also
established as invasive in distant tropical oceanic islands (e.g.
Hawaii, Norfolk,
Kermadec, Saint Helena; Cuneo and Leishman 2006; GISD 2010).
The European olive tree was one of the earliest plants introduced to
Australia by gardener George Suttor in 1800 and agricultural
pioneer John Macarthur in 1805, but multiple new
clones (varieties) have since followed (Sweeney and Davies 1998; El Kholy et al. 2012). The African olive was also brought into Australia by the Macarthur family, and is listed in the 1843 Nursery catalogue
of Camden Park (Campbelltown, NSW; Fig. 1A). The reasons for the initial introduction of this wild taxon are unclear, but its uses as rootstocks for the cultivated
olive and as a hedging plant are believed to be the main purposes.
View larger version:
Figure 1.
Illustrations of invasive olives in Australia. (A) Centennial olive tree (identified with genetic markers as a hybrid between
European and African olives; Besnard et al. 2014)
in the Camden Park Estate, NSW. This property (detail at the bottom
right) is well known for the horticultural introduction
of numerous plants into Australia during the
19th century; (B) European olive invasion in Adelaide Hills, SA.
Abandoned pastures
are rapidly colonized; (C) European olive
‘halo’ around the base of native eucalypt due to bird dispersal of seed;
(D) low-plant
diversity beneath dense African olive canopy
(Mount Annan, NSW); (E) ‘Cut and paint’ herbicide application technique.
Plants
are cut as close as possible to ground level
and undiluted glyphosate is quickly applied to the stump. Other methods
can be
used in areas where physical olive cutting is
not practical such as the ‘Drill and Inject’ technique or the
application of
a broadleaf selective herbicide that leaves
grasses intact to provide competition against weed re-establishment (Blood 2002);
(F) mechanical control of African olive using ‘forest mowing’ technique
(Mount Annan, NSW). This approach is used extensively
in the arboriculture and vegetation
management industry, and commonly utilizes earthmoving equipment. This
type of machinery
is able to mulch and process all above ground
vegetation in situ, which is retained on site as mulch to prevent erosion and provide a level of weed suppression. Olive stumps are retained
in situ to ensure soil stability,
with stumps > 30 mm spot treated with herbicide and diesel; (G) field
grass germination after direct
sowing on cleared olive site; (H) olive lace
bug (Frogattia olivinia), an Australian native pest of olives. This insect is a sap-sucking insect which feeds on the underside of the leaf, causing
a yellow mottling of the leaf surface.
A phylogeographic approach has been used to identify the origins of invasive olive populations in Australia and other locations
(Fig. 2; Besnard et al. 2007 ,2014).
Polymorphisms from maternally (plastid DNA) and biparentally inherited
(nuclear DNA) genomes revealed that European olive
populations from South Australia (SA) mostly
originated from the Mediterranean Basin, and derived from multiple
cultivar introductions
(Besnard et al. 2007).
In contrast, African olive populations from NSW, Norfolk Island,
northern New Zealand, Hawaii and Saint Helena mainly originated
from South Africa, harbouring two plastid
haplotypes detected in the Western Cape (Besnard et al. 2014). African olive seed or tree introduction from South Africa was certainly from an area around Cape Town controlled by Europeans
during the early 19th century (Wilmot 1869)
and accessible for collecting propagules. For example, African olive is
abundant in the Kirstenbosch National Botanical
Garden (which was established during the 17th
century), and is frequently seen in the surrounding countryside where it
can
colonize anthropogenic habitats (G. Besnard,
personal observation). Nuclear genes also indicated that hybridization
between
the two introduced olive subspecies has occurred in
Australia, both in SA and NSW (Fig. 2A).
This hybridization was probably early after the introduction of olives,
since a high density of hybrids (all the 26 analysed
individuals) was detected in the historic property
of Camden Park, well known for the horticultural introduction of plants
to Australia during the 19th century (Besnard et al. 2014).
View larger version:
Figure 2.
Scenario of an invasion story. (A) Early human-mediated spreads of European cultivated olives (Olea europaea subsp. europaea) and wild African olives (Olea europaea subsp. cuspidata)
before the end of the 19th century. The quasi-simultaneous introduction
of cultivars and African olives in Australia (between
the 1800s and 1840s) was followed by early
admixture in SA and NSW, as testified by genetic analyses (Besnard et al. 2007, 2014). The same phenomenon is likely to have occurred in St Helena, Northern New Zealand and Norfolk Island, where olives have
also been early introduced (Cuneo and Leishman 2006; Lambdon 2012). Additionally, admixture between European and African olives could have as well occurred in the native range where cultivars
have been introduced in the Cape area early after the European settlements (Wilmot 1869).
These hypotheses (indicated by question marks) still need to be
confirmed with genetic analyses. It is also not yet excluded
that independent introductions from the Cape
area to Australia, New Zealand and St Helena have occurred, and
additional genetic
analyses based on nuclear markers are
necessary to test these hypotheses; (B) after a lag time (i.e. selection
of new gene
combinations more adapted to local
conditions) olives have spread in the landscape and mainly colonized
anthropogenic habitats
(e.g. abandoned pastures). Such African olive
expansion was also observed within the native range, in Western Cape
(e.g. Stellenbosch
countryside, G. Besnard, personal
observation). In parallel, new European olive cultivars were still
introduced in the same
areas increasing the probability of admixture
with invasive olives. Further secondary introductions of African olive
from
the invasive range (NSW) to new ranges (e.g.
Hawaii) were demonstrated with genetic analyses (Besnard et al. 2014).
Additional recent translocations from South Africa to new areas are
poorly documented but likely to have occurred (e.g.
trade of ornamental trees at La Réunion). The
depauperate nuclear genetic diversity in African olives from Australia
and Hawaii
as well as the occurrence of the same two
African plastid haplotypes in the whole invasive range (i.e. East
Australia, Norfolk,
Kermadec, North New Zealand, Hawaii and St
Helena) suggest a spread from a limited number of founders from the Cape
area (Besnard et al. 2014).
Multiple variety introductions from different origins have maintained a relatively high overall diversity in the Australian
cultivated European olive germplasm (Sweeney and Davies 1998). As a consequence, relatively high gene diversity was detected in olive populations from South Australia (Besnard et al. 2007, 2014). In contrast, successive bottlenecks occurred during the invasive history of the African olive, first from South Africa
to Australia, and then from Australia to oceanic islands (Besnard et al. 2014; Fig. 2B). Testing population demography scenarios suggested that the initial effective population in Maui (Hawaii) was very small,
likely with less than ten individuals (median Ne = 9.8; Besnard et al. 2014). As a consequence, the African olive gene pool on Hawaii is particularly depauperate.
Biology and ecological impact of invasive olives: lessons from Australia
Seed and pollen dispersal:
Efficient gene dispersal mediated by pollen and fruits should favour
colonization of suitable habitats and maintain connectivity
between distant patches (e.g. Sork 2015).
In olives, while pollen is mainly wind-dispersed, fleshy fruits are
rich in oil and particularly attractive to frugivores
(e.g. birds, rodents) that ensure their natural
dispersal. In the invasive range, fruit size and avian dispersal are key
factors
driving the spread of olives. Fruit dispersal has
been reported over tens of km (Aslan and Rejmánek 2011). Birds are less able to manipulate and swallow fruits wider than 11.83-mm diameter (Altcántara and Rey 2003); however, the normal fruit size of ∼7 mm for African olive is optimal for dispersal by native and introduced birds (Cuneo and Leishman 2006). When cultivated olive groves of European olive are abandoned, the fruit of self-seeded olive trees are smaller than the
original cultivars, and avian fruit dispersal may thus increase (Spennemann and Allen 2000).
Growth and initial establishment: Although multiple factors are involved in the success of an invasive tree species, growth rate has been identified as a
good predictor of invasiveness (Lamarque et al. 2011). On one hand, there are no known studies comparing the growth rates of O. europaea
in native and invasive range, however, observations indicate that
growth rates are similar between native and invasive locations
where the climatic conditions are similar (P. Cuneo
and G. Besnard, personal observation). On the other hand, the capacity
to produce profuse dense seedling ‘mats’ beneath
the canopy of established trees is a remarkable feature of the African
olive
in its invasive range (Cuneo and Leishman 2006). Indeed, African olive seedling densities of 950 seedlings/m2 are commonly observed in Australia, a regenerative capacity never observed in the native range (Cuneo and Leishman 2006). Such a high establishment capacity has also been reported in other invasive trees, in particular in privets (Ligustrum spp.; Hoyos et al. 2010; Greene and Blossey 2012).
Impact of olives in Australia: European olives are now well established as invasive in the Adelaide hills, SA, where the climate is comparable with the
Mediterranean region (Fig. 1B and C).
The majority of invasive stands occur in areas of former woodland with
fertile, slightly acidic soils and 400–600 mm rainfall.
The invasive populations harbour diseases and pests
such as the olive fly that can cause crop losses in managed olive
groves,
but their key ecological impacts are the
displacement of native vegetation and increased fire risk (DWLBC 2005).
European olive reduces the abundance and diversity of native plant
species, altering the canopy structure of the woodland
and preventing native regeneration. Native canopy
cover may be reduced by up to 80 % and native species diversity up to 50
% (Crossman 2002).
The presence of the invasive African olive has similar ecological impacts in NSW. A mapping study showed clearly that African
olive is widespread and well established as invasive across a large region of western Sydney (Cuneo et al. 2009). In Cumberland Plain Woodland (South-West Sydney, NSW), where the native vegetation cover has been reduced to 13 % of its
original extent (DECCW 2010),
field surveys and a manipulative shading experiment showed that light
levels under African olive cover were substantially
reduced compared to native woodlands (canopy
openness of 4 % and 50 %, respectively), and there were 78 % fewer
native species
under African olive (Fig. 1D) compared with un-invaded woodland sites (Cuneo and Leishman 2013). African olive was able to maintain an 88 % survival rate under a dense olive canopy (Cuneo and Leishman 2013). A study of invertebrate species richness for successive stages of African olive invasion (Nguyen et al. 2016)
also revealed that diversity was significantly reduced under mature
African olive stands compared to early-stage olive and
mature native woodland. These studies confirmed the
adaptability of African olive and its ability to act as an ‘ecosystem
transformer’ by decreasing native plant and
invertebrate diversity and substantially modifying native communities.
The ecological impact of invasion by both olive subspecies on native plant diversity is relatively high compared with the
impact of other woody invasives. For example, Ligustrum sinense (Chinese privet) invasion resulted in a 41 % decrease in native plant diversity (Merriam and Feil 2002), and Acer platanoides (Norway maple) invading native Fagus grandifolia forest resulted in 36 % fewer native species beneath the shade of the Norway maple canopy (Wyckoff and Webb 1996). A study of Acacia saligna invasion of native fynbos vegetation in South Africa was also shown to reduce native plant diversity by 63 % at three long-term
invaded sites (Holmes and Cowling 1997).
Considering their negative impacts on native ecological communities in
Australia, olives are now recognized by state agencies
as a significant threat to the remnant vegetation,
in particular the African olive in the Cumberland Plain and the Hunter
Valley, north of Sydney. There are 12 ecological
communities of the Cumberland Plain listed as endangered under state and
federal legislation (DECCW 2010), justifying management actions to reduce the impact of the African olive.
Management and control of invasive olives in NSW
Both European and African olives are considered by land managers and bush regeneration practitioners as persistent woody weeds
difficult and expensive to control (West 2002; Crossman et al. 2003; DWLBC 2005).
In the case of African olives, maintenance of a ‘natural’ fire regime
(5–10 year recurrence interval) or the use of fire
as a management tool has not proven effective in
controlling the introduced in favour of native species, as established
African
olive (> 20 mm stem diameter) individuals are
able to resprout from a lignotuber after fire in a manner similar to
native
species such as Eucalypts and Banksias (von Richter et al. 2005).
Control of olives invading native
vegetation is best achieved at the incipient stage, particularly as
olive seedlings first
appear as ‘halos’ beneath large perch trees and
there are components of the native understory layer still present (Fig. 1C). In contrast, the control of dense monoculture African olive forests (Fig. 1D)
that are widespread throughout South-West Sydney (NSW) is a critical
issue. Control of this advanced ‘forest’ stage of African
olive invasion is expensive, and is currently
achieved through a combination of herbicide application and mechanical
land
clearing (Fig. 1F; Cuneo and Leishman 2015). Cleared olive sites are then monitored for re-sprouting stumps and germinating seeds.
Cuneo et al. (2010)
showed that the persistence of African olive seed in soil was ∼2.4
years. Its seed viability thus declines rapidly and provides
a narrow window of opportunity for germination and
regeneration. Persistence in the soil seed bank is indeed short compared
with other invasive species, particularly
hard-seeded legumes such as Broom (Cytisus scoparius) which forms persistent soil seed banks (>5 years; Thompson et al. 1993).
This also means that once mature African olive trees are removed,
control of seedlings germinating from the seed bank should
be required only for 2–3 years, along with
monitoring of seedlings derived from dispersal by birds into the managed
site.
In contrast, European olive seeds differ from those
of the African olive, as it has a thicker woody endocarp and
physiological
dormancy of the embryo that leads to slow
germination under horticultural and field conditions. This dormancy of
the embryo
is present even when the endocarp is removed (Rinaldi 2000), and seeds can retain high germinability after storage for three years (Fabbri et al. 2004). Although there are no studies of European olive persistence in the soil seedbank available, the combination of resistant
endocarp and dormancy suggests longer seed persistence than for the African olive.
Removal of invasive olives using the
methods described above will not be sufficient to restore the original
native eucalypt
woodland vegetation, particularly for degraded
sites where dense olive stands have developed over several decades.
Active
restoration and promotion of native regeneration
will be required as part of a strategy to promote native plant
diversity,
control weeds and achieve sustainable woodland
landscapes (Prober and Thiele 2005). The combination of short soil seed bank persistence in the African olive combined with relatively unaltered soil chemistry
after long-term African olive invasion (Cuneo and Leishman 2015) provides an opportunity to restore these degraded sites through direct seeding of native species and stimulation of the
native soil seed bank. Experimental work by Cuneo and Leishman (2015) showed that native grasses were absent from the soil seed bank in highly degraded African olive sites but direct seeding
was able to re-establish a native perennial grass cover (Fig. 1G),
which was resistant to subsequent weed invasion. This grass cover could
be managed as an important first stage in woodland
restoration, with exotic broadleaf species
controlled by fire and/or selective herbicide. The resilience of native
species
was evident in the fire-stimulated germination of
several hard seeded native species from the soil seed bank after 15
years
of African olive invasion (Cuneo and Leishman 2015).
The results of this restoration experiment were used to develop a
‘bottom-up’ model of ecological restoration, where restoration
efforts focus initially on the establishment of a
dense perennial grass cover as an early successional stage. Fire can be
used in subsequent years to provide interstitial
gaps for further direct seeding, and additional stimulation of soil seed
bank germination.
Ecological modelling and predictions of future invasive dynamics
There are a number of bioclimatic modelling studies (based on climatic, soil and land cover variables) which are relevant
to the potential olive distribution under current and future climates in Australia (Crossman and Bass 2008; O’Donnell et al. 2012; Cornuault et al. 2015; Roger et al. 2015). The parallel invasion of European and African olives was recently investigated in south-eastern Australia (Cornuault et al. 2015; Fig. 3).
By comparing the ecological requirements of native and invasive olives,
it was shown that the spatial segregation of the
two subspecies in their non-native range was partly
determined by differences in their native niches (i.e. niche filtering;
Cornuault et al. 2015). However, a realized niche shift occurred through a contraction of the native niche in both subspecies. Although niche shifts
are considered to be rare in invasive plants (Petitpierre et al. 2012), such changes were already highlighted by Gallagher et al. (2010) on 20 species exotic to Australia and also in the invasive range of Pinus taeda by Zenni et al. (2014).
The reason for these rapid shifts is not yet identified. The adaptation
and expansion of olives into these new habitats
could be due to the selection of new gene
combinations and/or the high level of observed phenotyic plasticity,
combined with
an absence of some stresses in a new range (see
below).
View larger version:
Figure 3.
Distribution of both olive subspecies (subspp. europaea and cuspidata) in the Australian range as predicted by Species Distribution Models calibrated with data from the invaded range (based on
observed occurrences indicated by small green points). The GARP software (Stockwell and Peters 1999)
was used for the niche modelling analyses based on 37 environmental
variables. These consisted of 19 climatic variables
(including annual trends, seasonality and
extreme climatic parameters measured over the last 30 years), 10
variables of soil
properties, 7 land cover variables and
altitude [see Cornuault et al. (2015) for detailed analyses]. Although the infestation presently occurs around Adelaide (subsp. europaea) and Sydney (subsp. cuspidata),
models predict a potential spread over large areas in south-eastern
Australia. Note that the predicted distribution is
not continuous with a large gap between NSW
and SA/Victoria states, but potentially suitable habitats for both
subspecies
largely overlap in NSW.
It was also shown that olives have not yet colonized their full potential distribution area in Australia based on current
conditions (e.g. O’Donnell et al. 2012; Cornuault et al. 2015; Roger et al. 2015). Cornuault et al. (2015) predicted that suitable habitat for the European olive covers a large region from Adelaide to Melbourne, expanding further
north into the plains west of Sydney and Brisbane (Fig. 3). In contrast, the invasion of African olive invasion should be concentrated in eastern NSW and south-eastern Queensland.
According to Cornuault et al. (2015), African and European olives share suitable habitats in NSW, Queensland and north of Melbourne (Fig. 3). Modelling presented in O’Donnell et al. (2012) and other studies on invasive species under future climate (Scott et al. 2008; Kriticos et al. 2010; Roger et al. 2015) predict a pattern of coastal and southerly retreat for temperate exotic plant species as a result of a warmer and drier
climate. Future predictions of range contraction in invasive O. europaea have to be taken with caution, however, since the assumption of niche stasis may be unreasonable (Cornuault et al. 2015), especially with admixture between the two subspecies. In addition, it is increasingly evident that the response of invasive
species to future climate change is likely to be strongly species- and context-dependent (Leishman and Gallagher 2015).
Future research directions
As presented in this review, olives are
becoming increasingly naturalized and invasive, and are now considered
to be ‘next
generation’ invasive trees. The combination of
abundant seed crops/propagule pressure and vertebrate seed dispersal are
key
factors in their establishment and spread, but
other biotic and abiotic factors could also be involved. Invasive olives
provide
an excellent case study of parallel invasion of two
closely related taxa with considerable research opportunities,
particularly
based on the genetic data now available. The key
future research questions that target the success of invasive olives are
presented in Table 1,
and include; history of introduction with possible admixture between
distant provenances, the genetic basis of their adaptability
during invasion, the role of biotic interactions
(e.g. with belowground native or co-introduced mutualists) and options
for
bio-control. Advances on these questions outlined
below will ultimately improve our predictions on future expansion, and
provide
a solid basis for better management of invasive
olive populations.
View this table:
Table 1.
Summary of main topics on olive invaders and proposed methodologies to address specific questions.
History of the invaders: While the great lines of the invasive olives’ history have been studied with phylogeographic and population genetic approaches
(Besnard et al. 2007, 2014; Besnard and El Bakkali 2014), some important questions remain to be addressed on this matter (Fig. 2).
First, the origins of some invasive
populations of African olive have not yet been properly investigated. In
particular, we
know that invasive populations from Saint Helena
share a common maternal origin (from Western Cape) with those of
Australia,
New Zealand and Hawaii (Besnard et al. 2014), but it is unclear whether these introductions are independent or sequential. This could be tested with population genetics
(e.g. Estoup and Guillemaud 2010). Identifying introductions as independent or sequential could be essential to interpret patterns of invasion (e.g. multiple
evolution of invasiveness vs. expansion from a common cradle).
Second, the earliest introductions of cultivated European olive in the Western Cape are potentially ancient (i.e. following
early stable European settlements in the Cape area at the 17th century; Wilmot 1869) and occurred before the introduction of the African olive to Australia. A first contact between African and European olives
could have thus happened in the native range (as reported in Acacia pycnantha; le Roux et al. 2013).
It seems likely that hybridization between these two subspecies has
taken place in South Africa, because the African olive
is often present in anthropogenically disturbed
habitats in close contact with cultivated olive groves (as in NSW). The
importance
and scale of this phenomenon is, however, unclear.
It could thus be relevant to compare patterns of genetic admixture and
recent population dynamics of olives in the
invasive range (Australia, New Zealand) and in natural and
anthropogenically disturbed
habitats of the Cape area. Describing the genome
structure of both native and invasive trees will allow testing the
hypothesis
of introgression (i.e. incorporation of a gene or
small genomic blocks from one entity into the gene pool of a second,
divergent
entity) from one subspecies to another during the
early steps of invasion. Comparative genomics would also allow for
testing
the role and relative importance of inter-taxa
recombination, and whether this has the potential to increase
evolutionary
changes and produce phenotypes that are better
suited to colonize novel environments (e.g. Ellstrand and Schierenbeck 2000; Facon et al. 2006; Lavergne and Molofsky 2007; Zenni et al. 2014).
Prediction of areas at risk for future invasion and causes for the rapid adaptation to new habitats:
Considering the need to limit the expansion of the olives in Australia,
it is now essential that existing populations are
mapped accurately, and any new incursions in areas
identified as being bioclimatically suitable are closely monitored or
controlled
at an early stage. Refined models to confidently
predict olive distribution should be very useful to better identify
areas
at risk for future invasion. In particular,
potential niche shifts have to be considered carefully in such models
(see above).
Feedbacks with native and co-introduced biota are
poorly known and not integrated in most predictive models (Guisan et al. 2014). It is also essential to identify the drivers behind the realized niche shifts reported by Cornuault et al. (2015).
Non-genetic or genetic factors could be involved. Indeed, the olive is
able to modulate expression of phenotypes (i.e. plasticity)
according to surrounding conditions (e.g. García-Verdugo et al. 2009; Rubio de Casas et al. 2011), while early admixture between the two olive subspecies, as reported in Australia, may offer the possibility of new gene
combinations (Besnard et al. 2014).
Common garden experiments with both
invasive and native trees may be done to compare the phenotypes of
individuals in native
and invasive ranges, in order to determine the
importance of genetic factors and plasticity in the expression of traits
in
these different environments (e.g. Kueffer et al. 2013; Heberling et al. 2016).
We suggest investigating fitness and growth performance of a set of
genotypes (i.e. wild, cultivated, naturalized and invasive)
in different environments (i.e. native and
introduced ranges). Such an experiment, however, would not be easy to
carry out,
and it may be difficult to obtain relevant data due
to the olive's longevity. We expect that invasive olives are more
likely
than non-invasive olives to have traits that favour
them in a changing environment; these traits include broad
environmental
tolerance, short juvenile periods (with rapid and
profuse seedling emergence) and ability for long-distance dispersal (Hellman et al. 2008). Traits promoting a better adaptation to anthropogenically disturbed habitats could have also been essential to colonize
abandoned pastures, and domesticated olives may have brought positively selected genetic factors (e.g. Ellstrand et al. 2010; Hufbauer et al. 2012; Thompson et al. 2012).
Association mapping and population genetics may potentially help the
identification of genomic blocks with such genes promoting
adaptation to new habitats in the invasive range.
For instance, selective sweeps (i.e. reduction of DNA variation in a
genomic
block with a mutation under recent and strong
positive selection) could be observed in populations from the invasion
front
compared with native populations (e.g. Zenni and Hoban 2015).
Invasive olives could also provide further insight into how trees like the African olive, have successfully adapted and invaded
large areas despite relatively narrow genetic variation within populations (Besnard et al. 2014).
Inbreeding depression is expected to limit the success of introduced
species, but this ‘invasion paradox’ of strong bottleneck(s)
combined with invasion success has been repeatedly
reported (e.g. Sakai et al. 2001), even in trees (Le Roux et al. 2008). Some authors have argued that the invasion success of genetically impoverished populations is dependent on environmental
factors such as temporary or permanent release from environmental stresses in the new range (Schrieber and Lachmuth 2016). Other important factors are the initial quality of propagules and population demography (e.g. Hufbauer et al. 2013).
In particular, further research could examine the role of domestication
and successive bottlenecks in reducing the mutation
load before or during the establishment of invasive
olives. High genetic load is expected in large, natural populations of
self-incompatible perennials, such as wild olives (Byers and Waller 1999). The European olive domestication is a complex story of inbreeding, and of admixture between distinct genetic pools (Díez et al. 2015). It is still unclear whether the domestication process has contributed to, or reduced the mutation load in the cultivated
olive pool. For the African olive, inbreeding phases observed in the introduced range (Besnard et al. 2014) may have also allowed either purging or fixation of deleterious alleles. Declining heterozygosity could reduce fitness due
to fixation load (e.g. Mattila et al. 2012), but in contrast, an efficient purging of deleterious mutations could avoid this phenomenon (e.g. Glémin 2003; Facon et al. 2011; Marchini et al. 2016).
The genetic load on some traits could be compared between native
(genuinely wild or cultivated) and non-native olives at
different stages of the invasion process, in order
to assess the impact of successive inbreeding phases (bottlenecks) in
purging
deleterious mutations. For each subspecies,
early-growth stage performance (i.e. germination, growth) of progeny
resulting
from controlled crosses of native trees, of
invasive trees and between native and invasive trees could be compared
(e.g. Keller and Waller 2002).
Such a study might help to disentangle the relative importance of the
genetic load (i.e. low fixation load) and a release
from stress (i.e. phenotypic plasticity) in the
‘tolerance’ of the African olive to sequential, strong bottlenecks as
revealed
in NSW and Hawaii (Besnard et al. 2014).
Biotic interactions and monitoring of invasive populations:
The role of biotic interactions in the success of invasive olives also
deserves to be investigated. The enemy release hypothesis
posits that the success of some invasive species is
related to the scarcity of natural enemies (e.g. parasites) in the
introduced
range compared with the native range (Keane and Crawley 2002).
For more a decade, this hypothesis has received much attention but is
probably too simplistic, because both antagonistic
and mutualistic interactions can be involved with
any organism which either limit or favour the spread of an exotic
species.
First, interactions with microbes must be
documented. Indeed, mutualistic interactions with the soil biota may
facilitate
plant invasions, and some invasives are known to
alter soil-borne mutualists in ways that affect recipient plant
communities
(Richardson et al. 2000). Two of the strongest soil mutualisms involve mycorrhizal fungi and nitrogen fixing bacteria, both of which improve the
nutrient status of their host plants (Reinhart and Callaway 2006).
The interaction between these mutualists and invasives also has the
potential to alter soil chemistry. An initial comparison
of soil properties between native woodland areas
and African olive invasion sites indicated no major differences for soil
pH or key soil elements (Cuneo and Leishman 2015),
however, the interaction between invasives and soil biota/chemistry
deserves further investigation. The diversity of organisms
associated to invasive olives and native vegetation
thus needs to be studied, in order to better understand the role of
mutualistic
and/or antagonistic interactions in the olive
invasion, which includes bacteria, mycorrhiza and microfauna such as
nematodes
or arthropods (Aranda et al. 2011; Montes-Borrego et al. 2014; Abdelfattah et al. 2015; Palomares-Rius et al. 2015). Co-invasion between trees and associated ectomycorrhizal fungi has been reported (e.g. Dickie et al. 2010); olives, however, have arbuscular mycorrhiza (Glomeromycota; Montes-Borrego et al. 2014) and should tend to associate with generalist, cosmopolitan fungal species or to form novel associations with native soil
fungi (e.g. Nuňez and Dickie 2014). In the future, the use of techniques such as metabarcoding or metatranscriptomics (e.g. Montes-Borrego et al. 2014; Abdelfattah et al. 2015)
may greatly facilitate the taxonomic and functional characterization of
the microbiome and microfauna associated to the
olive, both in non-invaded and in invaded woodland
habitats. It may help unravel changes in local communities during olive
invasion. Such changes could impact not only soil
biogeochemical cycles but also affect the whole ecosystem, for example
by
altering competitive interactions between native
and invasive plants (Callaway et al. 2004).
Alteration of local ecosystems through
olive invasion is evident at the level of the macrofauna and
invertebrates. For example,
the formation of an African Olive canopy causes
changes in woodland bird assemblages through changes in vegetation
structure
and fruit availability (NSW Scientific Committee 2010). These patterns of seed dispersal and utilization by both native and non-native animals should be quantified (e.g. Perea and Gutiérrez-Galán 2016). Furthermore, the extent of European olive cultivation throughout south-eastern Australia has highlighted the presence of
several insect pests of olives. As an example, the olive lace bug (Frogattia olivinia; Fig. 1H), which is native to NSW and southern Queensland, is known as a pest insect of olives both in SA and NSW (Spooner-Hart et al. 2002; Bean 2006; P. Cuneo, personal observation). Bean (2006)
found that this insect was able to impact olives through leaf damage
and reduced branch growth. Unfortunately, this study
was only conducted over one season and did not
assess the impact of olive lace bug infestation on fruit production and
long-term
health of trees.
Lastly, efforts have been already made to restore habitats heavily invaded by the African olive in NSW (Cuneo and Leishman 2015). Especially, direct seeding techniques have been developed to re-establish ground layer (Fig. 1G).
The impact of these practices on biotic interactions could be assessed
in field experiments. The communities of microbes
and microfauna (nematodes, insects) could be
compared between various habitats to test the resilience of biotic
interactions
in restored stands. Altogether, these
considerations emphasize the need to consider total ecosystem function
if we want to
better assess the impact of olive invasion and
anticipate and control its spread.
Concluding remarks: reclaim the past or accept a novel landscape?
Studies of woody plant invasions have shed light on many crucial aspects of invasion ecology (Richardson and Rejmánek 2011). With invasive species now well-established worldwide, the ecological role of such species in their ‘new’ habitats and the
determination of appropriate ecological restoration targets (Hobbs et al. 2009)
is now the focus of considerable debate. On a global scale, human
modification of ecosystems is the major cause of biodiversity
loss, a process that is being accelerated through
the spread of invasive species by human activity and trade (Pimentel et al. 2001).
Against this dominant backdrop of human ecosystem modification and
invasive species impacts, some ecologists are beginning
to question the feasibility of restoring ecosystems
to their original or ‘historic’ condition (e.g. pre-European settlement
of Australia) and whether the new combinations of
species (novel landscapes) might not offer valuable ‘ecosystem services’
in a changing world (Ewel and Putz 2004).
In this review, we have described how
invasive olives function as ‘ecosystem transformers’, and particularly,
in the Australian
context, are able to transform temperate eucalypt
woodlands with a diverse grassy understory to a closed canopy system
with
a depauperate understory. The long life span of
olive trees (100 years +) and resultant low forest floor light levels
result
in eventual displacement of the eucalypt woodland
(including canopy trees), rather than co-existence. Olive canopy is
still
able to provide ecosystem services such as soil
stability and fauna habitat, but this fundamental shift in vegetation
structure
causes changes in woodland bird and invertebrate
assemblages (NSW Scientific Committee 2010; Nguyen et al. 2016).
Decisions about how much conservation and
restoration investment is appropriate will depend on shifting cultural
values about
historic fidelity and ecological integrity,
sentimentality about ecosystems of the past, local species diversity,
priorities
for livelihood and sustainability (i.e.
historically faithful restorations versus ecosystem services-oriented
projects), and
designs for resilience (Hobbs et al. 2009).
In the Australian context, there is an intrinsic cultural value placed
on landscape identity, which is largely dominated
by the eucalypt in its myriad of forms, either as a
tall forest, lone paddock tree or distinctive silhouettes on a
ridgeline.
Understanding the biology and achieving effective
control of woody invasive species such as European and African olives is
about retaining ecosystem function, but also about
retaining eucalypt woodlands—a core element of the Australian landscape
identity.
Sources of Funding
G.B. is supported by TULIP (ANR-10-LABX-0041) and PESTOLIVE (ARIMNet action KBBE 219262).
Contributions by the Authors
Both authors equally contributed to this review.
Acknowledgments
We thank J. Hackel, L. Zinger, J. Bruxaux and J. Cornuault for fruitful discussions and help for writing the paper.
Footnotes
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Guest Editor: Heidi Hirsch
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Citation: Besnard G, Cuneo P. 2016. An ecological and evolutionary perspective on the parallel invasion of two cross-compatible trees. AoB PLANTS 8: plw056; doi:10.1093/aobpla/plw056
- © The Authors 2016. Published by Oxford University Press on behalf of the Annals of Botany Company.
This is an Open Access article distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/4.0/), which permits unrestricted reuse, distribution, and reproduction in any medium, provided the original work is properly
cited.
