Differential plant invasiveness is not always driven by host promiscuity with bacterial symbionts
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Differential plant invasiveness is not always driven by host promiscuity with bacterial symbionts
Metha M. Klock 1
Luke G. Barrett 0
Peter H. Thrall 0
Kyle E. Harms 1
Guest Editor: David Richardson
0 CSIRO Agriculture Flagship , Canberra, ACT 2601 , Australia
1 Department of Biological Sciences, Louisiana State University , Baton Rouge, LA 70803 , USA
Identification of mechanisms that allow some species to outcompete others is a fundamental goal in ecology and invasive species management. One useful approach is to examine congeners varying in invasiveness in a comparative framework across native and invaded ranges. Acacia species have been widely introduced outside their native range of Australia, and a subset of these species have become invasive in multiple parts of the world. Within specific regions, the invasive status of these species varies. Our study examined whether a key mechanism in the life history of Acacia species, the legume-rhizobia symbiosis, influences acacia invasiveness on a regional scale. To assess the extent to which species varying in invasiveness correspondingly differ with regard to the diversity of rhizobia they associate with, we grew seven Acacia species ranging in invasiveness in California in multiple soils from both their native (Australia) and introduced (California) ranges. In particular, the aim was to determine whether more invasive species formed symbioses with a wider diversity of rhizobial strains (i.e. are more promiscuous hosts). We measured and compared plant performance, including aboveground biomass, survival, and nodulation response, as well as rhizobial community composition and richness. Host promiscuity did not differ among invasiveness categories. Acacia species that varied in invasiveness differed in aboveground biomass for only one soil and did not differ in survival or nodulation within individual soils. In addition, acacias did not differ in rhizobial richness among invasiveness categories. However, nodulation differed between regions and was generally higher in the native than introduced range. Our results suggest that all Acacia species introduced to California are promiscuous hosts and that host promiscuity per se does not explain the observed differences in invasiveness within this region. Our study also highlights the utility of assessing potential mechanisms of invasion in species' native and introduced ranges.
Acacia; biological invasions; interactions; invasive; legume; mutualisms; rhizobia
Non-native species are a threat to native ecosystems,
particularly when they colonize new areas and rapidly
expand in abundance. Collectively, invasive species have
negative impacts at both local and global scales,
threatening biodiversity, accelerating global change and
VC 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
Term Definition Reference
Invasive Non-native species that (1) have Richardson et al.
self-sustaining populations (2011)
which, for a minimum of 10
years have reproduced by seed
or ramets without (or despite)
human intervention, and (2)
have spread and established
reproductive populations at large
distances from parent plants
Naturalized Non-native species that have
escaped cultivation and
populations but have not
spread to the extent of invasive
Richardson et al.
causing economic losses
(D’Antonio and Vitousek 1992;
Vitousek et al. 1996; Mack et al. 2000; Pimentel et al.
. Although not all introduced species become
invasive, those that do variously alter food sources for native
wildlife, change fire regimes, outcompete native species,
and impact soil communities, for example, by altering
microbial structure and soil nitrogen levels
D’Antonio 1998; Mack et al. 2000; Brooks et al. 2004)
better understand how species become invasive in new
environments, in-depth investigations of mechanisms
driving species invasions are needed.
Diverse mechanisms and hypotheses have been
proposed for why introduced species become invasive.
Many of the better-investigated drivers of invasiveness
are based on antagonistic or competitive interactions
(Blossey and Notzold 1995; Callaway and Aschehoug
2000; Keane and Crawley 2002; Levine et al. 2003)
work to date has investigated the role of enemy-release
in facilitating species invasions (i.e. invaders that prosper
in new environments because they leave their parasites,
pests, and predators behind [Keane and Crawley 2002]).
The Evolution of Increased Competitive Ability
Hypothesis predicts that adaptive evolution of invaders
provides a competitive advantage in novel ranges
(Blossey and Notzold 1995)
. Although overcoming
adversity imposed by antagonists and competitors may be
the driver of invasiveness for some species, mutualistic
interactions may also play a key alternate or synergistic
role in some invasions
(Richardson et al. 2000)
A growing body of work has examined the role of
mutualisms in the invasion of non-native species
et al. 2000; Birnbaum et al. 2012; Wandrag 2012)
The Enhanced Mutualism Hypothesis proposes that
species encounter novel beneficial symbionts in their native
range, which enhance their ability to survive and spread
(Richardson et al. 2000)
. The Accompanying
Mutualist Hypothesis suggests that invasive species are
introduced concurrently with their native mutualistic
partners, thereby enhancing their ability to survive in
novel habitats (
such as those between legumes and their symbiotic
nitrogen-fixing soil bacteria (i.e. rhizobia) may be
particularly important in explaining the ability of this group of
species to establish and expand abroad. Elucidating the
potential role that mutualistic interactions play in species
establishment and colonization may point towards
mechanisms driving differential levels of species invasion.
Australian Acacia species (Family: Fabaceae) are a
diverse group of legumes that form symbiotic relationships
with rhizobia. They have been introduced throughout the
world for a variety of purposes, including ornamental
use, fuel wood, erosion control, and forestry
Rangan 2008; Carruthers et al. 2011)
. Many Acacia
Non-native species that do not es- Richardson et al.
tablish populations without the (2000); Jepson
aid of humans (also ‘waifs’) Flora Project
species that have been introduced outside their native
range have become invasive abroad
(Richardson et al.
. Of the more than 1000 Acacia species occurring in
(Miller et al. 2011)
, 400 species have been
introduced outside their native range, with 6 % becoming
invasive, 12 % becoming naturalized and 82 %
remaining as casuals
(Richardson et al. 2011; Rejmanek
and Richardson 2013)
(see Table 1 for definition of
Globally, acacias vary in the number of regions they
have invaded [regions defined by
and Rejmanek and Richardson (2013)
include North America, Europe, Middle East, Asia,
Indonesia, Pacific Islands, New Zealand, Australia,
Indian Ocean Islands, Africa (southern), Africa (rest),
Atlantic Islands, South America, Caribbean Islands, and
Central America]. Differences in acacia invasiveness
among regions may be due to variation in invasive
capacity of these species, lower propagule pressure in
particular regions or differences in incidence reports among
(Richardson and Rejmanek 2011)
Within geographic regions, there is also evidence that
acacias vary in invasiveness. For example, sixteen
Australian Acacia species have been introduced to
California and differ in their invasive status in this region
(Jepson Flora Project 2015)
(Table 2). Whereas all these
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species except for two (A. cultriformis and A. redolens) are
invasive in at least one part of the world, they vary
markedly in their ability to invade and expand population sizes
in California. Acacia species were first introduced to
California for ornamental purposes and sold through the
nursery trade beginning in the mid-1800s
. Two species, A. dealbata and A. melanoxylon, are
currently designated as invasive in California
, five species as naturalized and nine species as
(Jepson Flora Project 2015)
(Table 2). Definitions of
invasiveness categories used for the purpose of this study
can be found in Table 1. Understanding the mechanisms
that enable multiple closely related species to
differentially establish and colonize natural areas in one particular
region is important for understanding what controls and
promotes species establishment in general (Klock et al.
One mechanism that may be an important determinant
of invasion success for acacias is their symbiotic
relationship with rhizobia. The legume–rhizobia interaction
has been long recognized as critical for the growth and
establishment of many legumes
(Thrall et al. 2005)
. Rhizobia are Gram-negative
bacteria that convert atmospheric nitrogen to a form
usable by the plant
(Bauer 1981; Sprent and Sprent 1990)
Within nodules, the plant provides rhizobia access to
carbon substrates and micronutrients, and also protects
them from desiccation
. When legumes form
an association with compatible symbiotic bacteria they
obtain a direct source of nitrogen unavailable to other
plants. Soil nitrogen availability for plants is often low
(Masclaux-Daubresse et al. 2010)
, so species that are more
readily able to form such associations may have a
competitive advantage over other plant species, particularly in
(Funk and Vitousek 2007)
The selectivity of different plant hosts for particular
rhizobial symbionts (hereafter referred to as “host
promiscuity”) may contribute to the differential ability of Acacia
species to establish and expand abroad. Hosts that are
more promiscuous (i.e. are able to effectively associate
with a wider range of rhizobial strains) may have a
competitive advantage when introduced to novel areas, where
they are likely to encounter unfamiliar nitrogen-fixing
(Richardson et al. 2000; Rodrıguez-Echeverrıa 2010;
Birnbaum et al. 2012)
. Previous research suggests that
widely distributed acacias in their native range are more
promiscuous rhizobial hosts, whereas those with more
limited distribution are more specific hosts
(Thrall et al.
. In addition, acacias that have become invasive in
multiple regions of the globe appear to be more
promiscuous hosts than naturalized or casual acacias (Klock et al.
2015). Variation in host promiscuity among Acacia species
introduced to California may help explain why certain
species have differentially invaded this region.
The goal of this study was to characterize the
nodulation ability of a suite of Acacia species that have become
differentially invasive within California. To examine this,
we used multiple Acacia species representing different
invasiveness categories and performed whole soil
inoculation experiments with a range of soils from different
environments and two different continents. Examining
species in their native and introduced ranges can provide
essential information for understanding the
contextdependent mechanisms influencing the invasion of
(Shea et al. 2005)
. By better understanding
the biological attributes of species in their home range,
we can predict and compare their responses abroad,
thereby gaining insight into which mechanisms are
influencing species survival and expansion in different
(Hierro et al. 2005)
. Using species of Acacia and
their rhizobial mutualists, we aimed to assess whether
the mechanisms promoting establishment and survival
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at home are the same that facilitate invasion abroad.
The purpose of conducting this experiment in the native
range was to challenge acacias with unfamiliar rhizobial
communities in areas where they naturally occur. This
mimics the conditions legume hosts face when
introduced abroad (although potential rhizobial mutualists
are likely to be more closely related to those they
typically associate with). Our approach also allowed us to
determine if observed patterns are maintained in the
invasive range, where rhizobial mutualists may be more
In particular, we evaluated aboveground plant growth
(biomass), survival, and nodulation responses. We
examined whether treatment of acacias with different soil
inoculants influenced plant performance. We also used
terminal restriction length polymorphism (T-RFLP), to
examine the composition and richness of rhizobial strains
associating with acacias in different invasiveness
categories. We hypothesized that invasiveness of non-native
acacias in California would be influenced by host
promiscuity with rhizobial strains, with the following
predictions: (1) invasive acacias would have higher biomass,
survival and nodulation responses (i.e. plant
performance) in both native and introduced ranges across a
greater number of soils than naturalized or casual
acacias; and (2) invasive acacias would associate with a
greater number of rhizobial strains (as measured by
number of ribotypes, or unique terminal restriction
fragment lengths) in both native and introduced ranges than
naturalized or casual species.
The genus Acacia (Fabaceae: Mimosoideae) is native to
Australia, with over 1000 species occurring variously
across the continent
(Miller et al. 2011)
(Fig. 1). We
focused on seven species that have been introduced to
California and have become invasive (A. dealbata and A.
melanoxylon), naturalized (A. baileyana and A. longifolia)
or remained casual aliens (A. cultriformis, A. pycnantha
and A. verticillata) in this region
(Cal-IPC 2006; Jepson
Flora Project 2015)
(see Fig. 1 for Acacia range
distributions in Australia and California). Five of these species
have been previously characterized for levels of host
promiscuity (A. dealbata, A. cultriformis, A. longifolia, A.
melanoxylon and A. pycnantha) using pure rhizobial cultures
(Thrall et al. 2000; Bever et al. 2013; Klock et al. 2015)
whereas two species have not (A. baileyana and A.
verticillata). All species examined here are native to
southeastern Australia and range from broadly distributed to
narrowly restricted within their native region (AVH 2015)
(Fig. 1). Previous research has provided at least some
evidence that more widely distributed acacias in their
native range are more promiscuous rhizobial hosts than
those that are narrowly distributed
(Thrall et al. 2000)
and that globally invasive acacias are more promiscuous
hosts than those that are naturalized or casual aliens
(Klock et al. 2015). Given the analogous variation in the
occurrence of our selected species within their novel
range, we used these species to examine whether
invasiveness in California might also be linked to variation in
Soil inoculant collection and preparation
Soil samples were collected from multiple sites in Acacia
species’ native (Australia) and introduced (California)
ranges to obtain a diverse suite of rhizobial communities
for use in glasshouse inoculation studies (Fig. 2 [see
Supporting Information—Table S1]). Whole soil inocu
lations were used rather than individual rhizobial
cultures to challenge acacias with rhizobial communities
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they have not previously been exposed to, thereby
reflecting more accurately the conditions acacias may face
when introduced abroad. Soils likely contained
organisms other than just rhizobia; however, all soils were
bulked and mixed within soil collection site, and all
Acacia species inoculated with soils from each site to
achieve homogenous treatment conditions.
In Australia, we collected soils from ten sites within a
150 km radius of Canberra, ACT, during July 2011. Sites
varied in disturbance regimes, from a highly disturbed
agricultural field, to an abandoned paddock, to an
undisturbed diverse native legume site. In California, we
collected soils from ten sites within a 50 km radius of San
Francisco, CA, during December 2011 (Fig. 2; [see
Supporting Information—Table S1]). Weather conditions
in Australia and California were very similar during the
sampling periods (high temp 11.2 C vs. 14 C; low temp
1.4 C vs. 1.7 C; precipitation 0.04 cm vs. nil) (www.
ncdc.noaa.gov, last accessed 03 August 2016).
In both ranges, we chose sites that did not contain
any of the Acacia species used in this study to challenge
all of the study species with unfamiliar rhizobial
communities. This was done to mimic conditions that hosts
might encounter when introduced to a new area. Soils
were collected over the course of one week. Soil samples
were excavated using a clean shovel and stored in paper
bags until processing. We collected multiple samples
from within each site and then bulked them within
replicates, with site as the level of replication, to make a
single composite for each of the 10 sites. Following
VC The Authors 2016
collection, soils were dried for up to 6 days. Once dry,
they were sieved through 3-mm mesh to remove rocks
and other debris and stored in paper bags until use. Soils
collected in California were shipped to Louisiana State
University (LSU) for use in the introduced range
glasshouse experiment. Temperatures at which soils were
stored fluctuated due to transport and handling
requirements but otherwise were held constant at 4 C. Previous
research has shown that the abundance of rhizobial
strains can decline over time in dry soil storage; however,
rhizobial strains are still abundant in soils after 6 months
(Martyniuk and Oron 2008; Thrall and Barrett pers. obs.)
In addition, as each Acacia species was subject to each
soil treatment, exposure to available rhizobial strains
was the same among species.
We conducted two glasshouse experiments to examine
the promiscuity of Acacia species in different
invasiveness categories. For the first experiment (hereafter called
the “native experiment”), glasshouse facilities were
located at CSIRO’s Black Mountain site in Canberra, ACT,
Australia. For the second experiment (hereafter called
the “introduced experiment”), glasshouse facilities were
located at LSU in Baton Rouge, Louisiana. The native
experiment was conducted from July to November 2011.
Seeds of all Acacia species used in this experiment were
collected within Australia and obtained from the
Australian Seed Company. The introduced experiment
was conducted from March to July 2012. For this
component, seeds of all Acacia species were collected directly
from plants in California in September 2011, shipped to
LSU, and stored in paper bags until use.
For both experiments, seeds were subjected to a
boiling water treatment to induce germination (boiling water
was poured over seeds and they were left to imbibe
water for 24 h). No further seed sterilization methods were
undertaken; however, seedlings were observed for
nodule presence at time of planting and none were
nodulated. In addition, while the native experiment control
did experience a moderate level of contamination at
final harvest, samples in the introduced range control
treatment showed no contamination, suggesting that
the source of contamination for the native experiment
was not the vertical transmission of rhizobia. Seeds were
transferred to trays of steam-sterilized vermiculite and
watered daily with sterile water for 14–20 days, or until
germination occurred. Seedlings were grown in the
glasshouse under local natural light conditions.
Once germinated, seedlings were transferred to
individual pots inoculated with soils collected from each of
the 10 sites. In the native experiment, for each of the
bulked soils, 10 replicates of each Acacia species were
planted in 8 15 cm pots filled 3=4 with sterilized sand
and vermiculite (1:1 volume), 50 g of an individual soil
treatment as a live inoculant and topped with additional
sterilized sand and vermiculite (1:1 volume) to avoid
cross contamination. For the introduced experiment,
seedlings were similarly planted and inoculated,
however replication varied due to availability of seed for
individual species (10 replicates of A. baileyana, A. longifolia,
A. melanoxylon and A. verticillata; 5 replicates of A.
dealbata and A. pycnantha; 4 replicates of A. cultriformis).
A rhizobia-free (N–) control was also included in both
experiments in which plants were not inoculated. For both
experiments, Acacia species soil combinations were
spatially randomized by glasshouse bench such that
each bench contained one replicate of each species
soil combination. Pot placement on the bench was
randomized. All plants were watered twice weekly with
sterile N-free McKnight’s solution
sterile water as needed. Plants were spaced well apart
on glasshouse benches to minimize cross-contamination
Plants were grown for 16 weeks in a
temperaturecontrolled glasshouse ( 20 C) and harvested in
November 2011 (native experiment) and July 2012
(introduced experiment), respectively. At harvest, seedlings
were clipped at the soil surface and aboveground
material was stored in paper bags. For the native experiment,
aboveground material was oven dried at 70 C for 48 h
and weighed. A malfunction with the drying oven
destroyed aboveground material for the introduced
experiment, therefore biomass data were lost. Belowground
material for both experiments (roots and attached
nodules) of each plant was stored individually in plastic bags
and frozen at –20 C until processing for molecular
analysis. Roots were scored at harvest for nodulation
quantity (0, <10, 10–50, >50) and quality (none, ineffective
[black or very small white nodules], intermediate
[mixture of small to medium white/pink nodules] and good
(Thrall et al. 2011)
Isolation of DNA and T-RFLP
We used terminal restriction length polymorphism
(T-RFLP) to identify community composition and genotypic
richness of rhizobia nodulating with Acacia species in the
glasshouse experiments. This technique is frequently used
for examining taxon richness of bacterial communities
et al. 1997)
. To extract DNA from root nodules collected
during harvest, 2–10 intact nodules per plant (depending
on availability) were first snipped from roots stored at –20
C. Nodules were surface sterilized by immersion in 90 %
ethanol for five to ten seconds, transferred to 3 % sodium
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hypochlorite and soaked for 2–4 minutes, and rinsed in
five changes of sterile water. Nodules were crushed using
liquid nitrogen, and DNA was extracted using Mo Bio
PowerPlant VR DNA Isolation kits following the protocol of
the manufacturer (Mo Bio Laboratories, Inc., Carlsbad, CA,
USA). Nodule processing and DNA extractions for the
native experiment were conducted at CSIRO laboratories in
Canberra, Australia, and for the introduced experiment at
LSU in Baton Rouge, LA. DNA extractions from the
introduced experiment were shipped to CSIRO laboratories
where all additional molecular analyses were conducted.
For all samples, we amplified the 16S rRNA gene using the
primers GM3 (5’-AGA GTT TGA TCM TGG C-3’) and GM4
(5’TAC CTT GTT ACG ACT T-3’) and the following PCR program:
initial denaturation at 95 C for 2 min, followed by 35
cycles of 95 C for 30 s, 50 C for 30 s and 72 C for 90 s,
followed by a final extension step at 72 C for 10 min and a
final holding temperature of 4 C. We digested the PCR
product using the restriction enzyme MspI (New England
BioLabs) in 30 ll reaction mixtures, and analysed the
fragment sizes using a 3130 l genetic analyzer (Applied
Biosystems, Warrington, United Kingdom). We used
GeneMapper version 5 (Life Technologies, Grand Island,
NY, USA) to examine T-RFLP profiles and included peaks
over 50 bp for further analysis. We quantified resulting
peaks using the local southern method
Peaks were binned using Ramette’s interactive binner
in the R statistical programming
language version 3.2.0
(R Core Team 2015)
DNA extracted from nodules contained both acacia
plastid and rhizobial DNA. While the GM3/GM4 primers
can also amplify mitochondrial and chloroplast DNA,
insilico analyses of restriction-fragment polymorphisms
for all Acacia species plastid sequences obtained from
Genbank indicated that polymorphisms in plastid DNA
were unlikely to contribute any variation to our T-RFLP
dataset. Specifically, to identify which
restrictionfragments corresponded to acacia plastid DNA, we
conducted an in-silico T-RFLP analysis by searching for the
primer sequences and restriction enzyme cut sites in
acacia plastid DNA sequences downloaded from
GenBank. We found that the restriction enzyme MspI cut
sites for acacia plastid sequences generated DNA
fragments greater in size than the cut-off for fragments used
in our analysis (i.e. the largest restriction-fragment in our
analysis was 545.3 bp, whereas the smallest restriction
fragment for acacia plastid DNA was 553 bp). Because
our cut-off was lower than the largest acacia plastid
restriction-fragment, any peaks corresponding to acacia
plastid DNA were excluded from our analysis. In
addition, review of polymorphisms attributable to individual
host species showed there were no polymorphisms
unique to all replicates of a host species (or group of host
species), further indicating that acacia plastid DNA did
not explain variation in the dataset.
Plant growth, survival and nodulation response
We examined the responses of acacias representing
three invasiveness categories to inoculation with 20
different soils (10 soils each in the native and introduced
ranges) collected from habitats in which the acacias
used in this experiment do not occur. We measured
differences among the invasiveness categories by
assessing aboveground biomass (native range only), survival,
nodulation presence/absence and nodulation index of
effectiveness. The nodulation index of effectiveness
categorizes the number of nodules found on the roots of
plant specimens, and is divided into levels of none, low,
medium, and high, delineated as follows: 0 nodules ¼
score of 0; 1–10 nodules ¼ score of 1; 11–50 nodules ¼
score of 2; >50 nodules ¼ score of 3.
We examined these four variables for the entire data
set using generalized linear mixed models (GLMM) and
used AIC to select the best models
. Acacia species was included in the
models as a random effect to include individual variation
of species in each invasiveness category. Aboveground
biomass and nodulation index were modelled using a
Gaussian distribution, and nodule presence/absence and
survival were modelled using a binomial distribution with
a logit link function. Negative control samples were not
included in models for the native experiment, as almost
all control specimens did not survive; however, they were
included in models for the introduced experiment.
We used the R statistical package “lme4” version 1.1-9
(Bates et al. 2012)
to determine whether main effects
(soil, invasiveness category and Acacia species)
contributed significantly to the models of interest, and whether
there were interactions among main effects. Acacia
species was maintained in all models as a random effect.
Models with the lowest AIC score were selected for
further analysis; models with a difference in AIC values
of <2 were considered equally likely
Anderson 2002; Bolker et al. 2009)
. Further analysis
consisted of conducting multiple comparisons of means
(MCMs) with Tukey contrasts using the R statistical
package “multcomp” version 1.4-1 (Hothorn et al. 2008),
which allowed us to determine whether there were
significant differences among invasiveness categories for
the response variables of interest (i.e. biomass,
nodulation presence and nodulation index) for individual soils,
while maintaining Acacia species in the model as a
We also examined biomass (native experiment only),
nodulation presence/absence, and survival for individual
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Acacia species to assess species-specific responses to
individual soil inoculants. We used ANOVA to compare
biomass among species x soil combinations and logistic
regression to analyse survival and nodulation presence.
We used a post-hoc Tukey’s HSD test to compare
biomass of different species to each soil inoculant using the
R statistical package “agricolae” version 1.1-2
. Analyses were conducted using the R
statistical programming language version 3.2.0
Rhizobial community composition and richness
We analysed binary data obtained from T-RFLP analysis
using non-metric multidimensional scaling (NMDS)
based on a Jaccard similarity matrix. We used the R
statistical package “vegan” version 2.3-0
(Oksanen et al.
to conduct ordination and Permutational ANOVA
(PerManova; function “ADONIS”) to test for differences in
rhizobial community composition among invasiveness
categories and soil types. If differences were detected
we ran pairwise comparisons between groups using
“ADONIS” with a Holm correction. We used ANOVA to
examine whether there were differences in ribotype
richness among invasiveness categories. Analyses were
conducted using the R statistical programming language
(R Core Team 2015)
We detected a significant interaction between soil and
invasiveness category for aboveground biomass (DAIC ¼
19.1, wi ¼ 1.00) (i.e. the best fitting model had an AIC
value >2 than all other models), indicating that the
growth response of species in different invasiveness
categories was influenced by the soil in which they were
grown [see Supporting Information—Table S2]. We,
therefore, examined each soil individually using MCMs
with Tukey contrasts and found that plants in different
invasiveness categories differed significantly in average
biomass response for only one soil (Fig. 3 [see
Supporting Information—Table S3])
ANOVA results indicated that biomass varied for
individual Acacia species across soil treatments (F9,605 ¼
470.21, P < 0.001); we also found a significant
difference in biomass across Acacia species (F6,608 ¼
346.80, P < 0.001), and an interaction between species
and soil treatment (F54,545 ¼ 135.01, P < 0.001) (Table
3). From here on, individual Acacia species are
indicated in the text as I (invasive), N (naturalized) and C
(casual). Using as a comparison the soil where biomass
was lowest for each species, post-hoc Tukey’s HSD test
showed that A. longifolia (N) and A. melanoxylon (I)
had significantly greater biomass for three soils, A.
baileyana (N), A. cultriformis (C), A. dealbata (I) and A.
verticillata (C) for two soils, and A. pycnantha (C) for
one soil [see Supporting Information—Fig. S1 and
The model with the best support for plant survival
included soil inoculation as a main effect with species as
a random variable (DAIC ¼ 3.87, wi ¼ 0.87) [see
Supporting Information—Table S5A], indicating that
variation in survival was driven by individual soils
rather than invasiveness category. Survival across soils
was generally high for all invasiveness categories
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(>50 % for all soils for the naturalized and casual
categories and nine out of ten soils for the invasive
category) (Fig. 4A [see Supporting Information—Table
Source df SS F P
Host species 69 346.80 61.93 <0.001
Survival for individual species was also generally high
across soils. We observed over 50 % survival for each
species in a minimum of seven soils (A. pycnantha [C])
and a maximum of all ten soils (A. longifolia [N] and A.
verticillata [I]) [see Supporting Information—Fig. S2
and Table S7A].
There was a moderate level of contamination in the
negative controls (nodules were found on 33 % of
samples), and very few samples that were not contaminated
survived, therefore, they were excluded from all native
The model with best support for nodulation presence
included soil inoculation as a main effect with species
as a random variable (Native experiment: DAIC ¼ 3.92,
wi ¼ 0.88) [see Supporting Information—Table S8A],
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indicating that differences in nodulation presence were
driven by individual soils rather than invasiveness
category. The presence of nodules across soils was generally
high for all invasive categories (>50 % in ten soils for the
casual category and nine soils for the naturalized and
invasive categories) (Fig. 5A [see Supporting Information—
Nodulation presence for individual species was also
generally high across soils, with over 50 % nodulation
presence for each species in a minimum of seven soils
(A. baileyana [N]) and a maximum of all ten soils (A.
cultriformis [C], A. longifolia [N], A. melanoxylon [I] and A.
verticillata [C]) (see Supporting Information—Fig. S4
and Table S10A]).
We found a significant interaction between soil and
invasiveness category for nodulation index of effectiveness
(DAIC ¼ 9.7, wi ¼ 0.97) [see Supporting Information—
Table S11A]. This indicates that there was an effect of
individual soils on nodulation index, such that the number
of nodules on plants belonging to different invasiveness
categories depended on the soil in which they were
grown. We, therefore, could not generalize nodulation
index response for invasiveness categories across all soils,
and examined nodulation index for each soil individually
using MCMs with Tukey contrasts. When soils were
examined individually, we found no significant difference in
nodulation index among invasiveness categories
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Visual assessment of ordination diagrams did not
indicate a clear difference in rhizobial community
composition among Acacia species invasiveness categories
(Fig. 7A). However, PerManova results from T-RFLP
analyses indicated a small but significant difference in
rhizobial community composition between the invasive
and casual categories (ADONIS, R ¼ 0.08, adjusted
P ¼ 0.024). Despite this slight difference in community
composition, there was no significant difference in
rhizobial richness among invasiveness categories (F ¼ 1.287,
P ¼ 0.284).
No contamination occurred in the introduced range
experiment so all control samples were retained for all
analyses. The best-supported model for survival in
the introduced range experiment included soil
inoculation as a main effect with species as a random
variable (DAIC ¼ 2.73, wi ¼ 0.79) [see Supporting
Information—Table S5B]. Similar to the native
experiment, survival across soils was generally high for all
invasiveness categories (>50 % for all soils for the
invasive and casual categories, and 8 out of 10 soils for
VC The Authors 2016
the naturalized category) (Fig. 4B [see Supporting
Survival for individual species was also high across
soils. We observed over 50 % survival in a minimum of
six soils (A. baileyana [N]) and a maximum of all 10 soils
(A. longifolia [N], A. melanoxylon [I] and A. verticillata [C])
[see Supporting Information—Fig. S3 and Table S7B].
The model with the best support for nodulation
presence included only soil inoculation as a main effect with
species as a random variable (DAIC ¼ 3.99, wi ¼ 0.88)
[see Supporting Information—Table S8B]. In contrast
to the native experiment, nodule presence across soils
was low for the introduced experiment (>50 % in six soils
for the casual category, three soils for the invasive
category and one soil for the naturalized category) (Fig. 5B
[see Supporting Information—Table S9B]).
Nodulation presence was generally low across soils for
individual species as well, with over 50 % nodulation
presence for each species in a maximum of six soils
(A. longifolia [N], A. pycnantha [C] and A. verticillata [C])
and a minimum of zero soils (A. baileyana [N] and A.
cultriformis [C]) [see Supporting Information—Fig. S5 and
We found a significant interaction between soil and
invasiveness category for nodulation index of
effectiveness (DAIC ¼ 22.32, wi ¼ 1.00) [see Supporting
Information—Table S11B]. We, therefore, examined
soils individually using MCMs with Tukey contrasts and
found no significant difference in nodulation index
among invasiveness categories (Fig. 6B).
Visual assessment of ordination diagrams indicated no
significant difference in rhizobial community composition
among acacia invasiveness categories (Fig. 7B).
PerManova results lent further support to this conclusion,
with no significant difference in rhizobial community
composition found among invasiveness categories (ADONIS, R
¼ 0.08, P ¼ 0.21). In addition, we found no significant
difference in rhizobial richness among categories of
invasiveness (F ¼ 1.224, P ¼ 0.31).
Summary of results: native and introduced experiments
Biomass results from the native experiment showed no
significant difference in aboveground biomass among
acacia invasiveness categories except in one soil.
Survival did not differ among categories in both the
native and introduced experiments. Nodule presence and
index of effectiveness was generally high across all
invasiveness categories for the native experiment, but low
for the introduced experiment. We found no
circumstances in which multiple models were equally likely for
individual response variables (i.e. differed by <2, see
above) for either the native or introduced experiments.
Rhizobial composition differed slightly among
invasiveness categories in the native experiment only; richness
did not vary among categories for either the native or
The goal of this study was to examine whether variation
in host promiscuity with rhizobial symbionts plays a role
in the differential invasion of Acacia species in California.
VC The Authors 2016
We found that host promiscuity as measured by plant
growth in the native experiment, and survival and
nodulation response in both native and introduced
experiments did not differ among acacia invasiveness
categories. However, acacias in the native experiment
(regardless of invasive status) were able to develop
nodules in a greater number of soils than in the
introduced range experiment. We found limited variation in
rhizobial associations among acacias that vary in
invasiveness in California. While rhizobial community
composition differed slightly among acacia invasiveness
categories in the native experiment, rhizobial richness or
the number of strains with which host species in these
groups formed an association was not significantly
different. Results from the introduced experiment showed
no difference in community composition or richness of
rhizobia associating with Acacia species in different
invasiveness categories. Plant growth response, paired with
belowground rhizobial richness results, suggests that
variation in host promiscuity may not be a major
determinant of invasiveness of Australian Acacia species in
Results from T-RFLP analyses indicated a slight
difference in the rhizobial communities acacias in different
invasiveness categories associated with, when paired with
Australian soils. However, no such differences were
evident with Californian soils, perhaps reflecting a greater
diversity of compatible rhizobial strains in Australian
soils. We found no difference in rhizobial richness among
invasiveness categories for either set of experiments.
Together, these results suggest that partner choice as
opposed to partner breadth may be more important in
explaining how interactions with rhizobia influence
potential for invasiveness in this set of Acacia species.
However, more work is required to generalize these
Birnbaum et al. (2012)
found similar results
when examining acacias that have become invasive
within their native continent; species examined
associated with the same abundance of rhizobial strains in
both native and novel ranges, and for two species tested
(A. longifolia and A. melanoxylon), they associated with
similar rhizobial communities between ranges.
In the native range experiment, rates of nodulation
and survival were similarly high across almost all soils.
Although we paired acacias with soils in which they did
not occur in their native range, effective rhizobial strains
may be broadly distributed, as has been previously found
(Barrett et al. 2012)
and mycorrhizal fungal
symbionts of acacias in Australia
(Birnbaum et al. 2014)
In their introduced range, acacias may be more likely to
encounter rhizobial strains that are more distantly
related to those with which they have co-evolved, or
appropriate strains may be completely absent, such that it
is more difficult to find suitable partners (perhaps
partially explaining the generally lower nodulation rates we
observed in the introduced experiment).
In addition to rhizobia, other organisms in soil
communities may have influenced plant performance. We
used whole-soil inoculation treatments, which may host
multiple rhizobial symbionts as well as pathogens and
other mutualistic microorganisms, and plant response
may be influenced by the presence of such organisms
(Thrall et al. 2007)
. The presence of non-rhizobial
mutualists may have had a greater effect on plant
performance in the experiment utilizing Australian soils,
because acacias native to Australia likely have higher
compatibility with the resident microorganisms. The
presence of pathogenic organisms such as fungi or
nematodes as well as interactions among co-occurring soil
biota may also affect Acacia species growth response,
influencing the potential positive benefit of being a
promiscuous rhizobial host. However, whether pathogenic
interactions are more likely to have influenced plant
growth in Australian or Californian soils is difficult to
assess. Other, more complex synergistic or antagonistic
interactions may also occur when using whole soil
inoculations. For example, rhizobial competition arising
from the presence of multiple rhizobial genotypes within
soils may have influenced mutualistic outcomes.
et al. (2014)
found evidence that acacias paired with
multiple rhizobial strains suffered diminished plant
growth response, likely due to altered patterns of
rhizobial association. Hence, an important caveat is that we
are unable to tease apart the complex species
interactions that may occur among the myriad organisms
occurring in natural soil communities, and which may have
influenced plant performance in this study.
A previous study has shown that more invasive Acacia
species are more promiscuous rhizobial hosts. Klock
et al. (2015) paired 12 rhizobial strains ranging in
effectiveness with 12 Acacia species differing in global
invasiveness (four invasive, four naturalized and four casual
species). In regard to plant growth, invasive acacias were
generally more promiscuous hosts, able to associate and
have a positive growth response with more rhizobial
strains than naturalized and casual acacias. However, in
this previous study, acacias were paired with single
rhizobial genotypes rather than whole soil inoculations and
acacia invasiveness was categorized on a global, rather
than regional scale (Klock et al. 2015). Acacia species
tested in this study vary in invasiveness in California;
however, all except for one are invasive in at least one
region of the world
(Richardson and Rejmanek 2011;
Rejmanek and Richardson 2013)
. We were interested in
what drives differences in invasiveness on a regional
scale; however, since all Acacia species tested here are
VC The Authors 2016
invasive at least somewhere in the world, they may very
well all be promiscuous rhizobial hosts, constrained by
mechanisms other than host selectivity for rhizobia from
becoming invasive in California. Host promiscuity with
rhizobia may indeed influence the ability of acacias to
invade novel regions, but other biotic and abiotic factors
likely contribute to the establishment and colonization of
these species, limiting some species from invading
particular regions, and promoting the invasiveness of
The lack of aboveground biomass data in the
experiment using soils collected in California reduced our
ability to determine whether patterns of plant performance
are consistent between native and introduced
experiments. While we were able to assess the ability of
acacias in different invasiveness categories to nodulate with
rhizobia and their subsequent survival, we do not know
whether this resulted in a beneficial growth response in
the introduced range experiment. This limits our ability
to assess whether acacias in the introduced range
responded in a beneficial manner as a result of being
paired with unfamiliar rhizobial symbionts. Future
studies would benefit from assessing aboveground biomass
of acacias paired with soils from their introduced range.
Still, results from our nodulation, survival and molecular
analyses provide strong evidence that acacias in
different invasiveness categories tested here do not vary in
host promiscuity with rhizobial symbionts.
Rhizobia-related mechanisms other than host
promiscuity may influence the invasiveness of acacias
introduced to novel regions. There is increasing evidence that
some legumes have been introduced abroad with their
native rhizobial symbionts
Crisostomo et al. 2013; Ndlovu et al. 2013)
similarity in associated rhizobial strains across native and
novel ranges (Birnbaum et al. 2016). The introduction of
both invasive species and their co-evolved beneficial
symbionts may circumvent any need for introduced
species to develop novel mutualistic rhizobial associations.
Acacia pycnantha, a native Australian species that has
become invasive in South Africa
(Ndlovu et al. 2013)
been found to associate with rhizobial strains more
closely related to those of Australian origin
(Ndlovu et al.
. Both A. longifolia and A. saligna associate with
rhizobia of Australian origin in Portugal
(RodrıguezEcheverrıa 2010; Crisostomo et al. 2013)
. Legumes native
to Portugal were also found to form associations with
rhizobial strains of Australian origin in areas where A.
longifolia occurred (
Birnbaum et al. (2016)
found evidence for three Acacia
species associating with the same rhizobial strains
between native and novel ranges within their native
continent. Dual invasion of symbiotic plant and microbial
species may thus be occurring in regions where acacias
have been introduced, or certain rhizobial strains may be
particularly widespread, potentially contributing to both
above and belowground structural changes in native
Acacias that become invasive in California may benefit
from mutualistic interactions other than the
legume–rhizobia symbiosis that aid in their establishment and
colonization. As indicated here, host promiscuity with
rhizobia alone does not appear to delineate invasiveness
of acacias in California. However, as a general trait
promoting invasiveness, host interactions with other taxa
may be important to the establishment, colonization and
survival of these species. Ant mutualists may aid in seed
dispersal and seed bank accumulation as well as
protection from herbivores for Acacia species that become
invasive in their novel range
(Holmes 1990; Montesinos
and Castro 2012)
. Acacia species that have become
invasive in California may also develop successful
mutualisms with avian seed dispersers
(Glyphis et al. 1981;
Underhill and Hofmeyr 2007; Aslan and Rejmanek 2010)
Being hosts for a variety of mutualistic organisms may
increase the opportunity for Acacia species to develop
self-sustaining, spreading populations that invade novel
Species that have become invasive in multiple areas of
the world may be constrained from establishing and
colonizing all regions where they are introduced. Identifying
as well as ruling out potential mechanisms influencing
expansion of species that have become invasive globally
but are constrained regionally can inform management
of species introduced abroad. We found that acacias
varying in invasiveness in California do not differ in their
ability to form symbioses with nitrogen-fixing bacteria,
as evidenced by a lack of difference in plant performance
and rhizobial richness when paired with diverse soil
inoculants. Invasive status of introduced acacias in
California, therefore, does not appear to be determined
solely by the ability to associate with larger numbers of
Due to the demonstrated capacity of almost all Acacia
species introduced to California to invade at least one
other region of the world, and previous research showing
that globally invasive acacias are promiscuous hosts, all
Acacia species, whether currently invasive or not in
California should be monitored closely for further
colonization and expansion in their introduced range. Just as
species differentially establish in their native ranges, the
levels of invasiveness that species accomplish when
VC The Authors 2016
introduced abroad may also vary. Our results suggest
that taking scale into account when examining the
factors that drive invasion of species is important; those
species that are deemed invasive on a global scale may
not be so on a regional scale, and different mechanisms
may be influencing their capacity to invade novel
regions. By identifying the mechanisms that both promote
and constrain acacia invasion in particular regions, we
can better inform management and future introduction
of these species abroad, thereby mitigating their
potential to cause negative impacts on native communities.
Sources of Funding
Our work was funded by a Louisiana Board of Regents
EPSCoR grant, funding from the National Science
Foundation (DEB-1311290) (United States), a Sigma Xi
Grant-in-Aid of Research (United States) and a Louisiana
Environmental Education Commission Research grant.
Contributions by the Authors
M.M.K., L.G.B., P.H.T. and K.E.H. conceived of and
developed the idea. M.M.K. and L.G.B. analysed the data; and
M.M.K. led the writing with input from all authors.
Conflicts of Interest Statement
No conflicts of interest.
We thank Caritta Eliasson, Mohammad S. Hoque, Kristy
Lam and Alexandre de Menenzes, for help and guidance
in the laboratory, glasshouse, and with data analysis. We
thank Meredith Blackwell, James T. Cronin, Hallie Dozier,
Bret Elderd and Richard Stevens for support and insight
throughout the development of the project. We thank
Sandra P. Galeano, Katherine Hovanes and Hector Urbina
for discussion in manuscript preparation. We thank
Matthew Ritter for discussion and invaluable insight. We
thank the guest editor and two anonymous referees for
helpful comments that improved the manuscript.
The following additional information is available in the
online version of this article —
Table S1. Soil collection sites for Acacia species
inoculants in Australia and California.
Table S2. GLMM Models predicting difference in
aboveground biomass (g) among California invasive rankings.
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VC The Authors 2016
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