Ecology Letters, (2014) 17: 499–507
LETTER
John L. Maron,1* Harald Auge,2,3
Dean E. Pearson,1,4 Lotte Korell,2,5
Isabell Hensen,3,5 Katharine N.
Suding6 and Claudia Stein6†
doi: 10.1111/ele.12250
Staged invasions across disparate grasslands: effects of seed
provenance, consumers and disturbance on productivity and
species richness
Abstract
Exotic plant invasions are thought to alter productivity and species richness, yet these patterns
are typically correlative. Few studies have experimentally invaded sites and asked how addition of
novel species influences ecosystem function and community structure and examined the role of
competitors and/or consumers in mediating these patterns. We invaded disturbed and undisturbed
subplots in and out of rodent exclosures with seeds of native or exotic species in grasslands in
Montana, California and Germany. Seed addition enhanced aboveground biomass and species
richness compared with no-seeds-added controls, with exotics having disproportionate effects on
productivity compared with natives. Disturbance enhanced the effects of seed addition on productivity and species richness, whereas rodents reduced productivity, but only in Germany and
California. Our results demonstrate that experimental introduction of novel species can alter
ecosystem function and community structure, but that local filters such as competition and herbivory influence the magnitude of these impacts.
Keywords
Community assembly, exotic species, grasslands, invasion, local filters, plant competition, plant
productivity, small mammals, species richness.
Ecology Letters (2014) 17: 499–507
Community productivity and species richness are fundamental
attributes of plant community structure. As such, there has
been much work aimed at understanding the determinants of
these community attributes, as well as exploring how they
might influence each other (Grime 1973; Adler et al. 2011).
We know that standing biomass can be affected by abiotic
factors such as soil fertility and precipitation, as well as by
biotic factors such as variation in local plant diversity (Rosenzweig 1968; Tilman et al. 2001; Hooper et al. 2005). Local
species richness can be affected by stochastic processes such
as dispersal limitation (Hubbell 2001; Turnbull et al. 2000;
Zobel et al. 2000; Stein et al. 2008) as well as deterministic
factors including niche partitioning, competition and consumer impacts (MacArthur & Levins 1967; Maron et al. 2012;
Germain et al. 2013; Kempel et al. 2013).
Interestingly, invasion of exotic species into communities
can also affect community productivity and local species richness. Plant productivity is often higher in heavily invaded
than nearby uninvaded sites (Wilsey & Polley 2006; Liao et al.
2008; Vila et al. 2011). As well, invasion can also enhance spe-
cies richness, even at small spatial scales (Stohlgren et al.
1999; Stadler et al. 2000; Sax & Gaines 2003; Sax et al. 2005).
This typically occurs where many ‘weak invaders’ establish at
low density; invasion can clearly reduce species richness in
cases where a particular invader forms dense stands that
crowds out natives (Vil
a et al. 2011).
The association between invasion, increased productivity
and altered species richness raises several important questions. First, what is the cause–effect relationship between
invasion and increased productivity or altered species richness? Almost all studies that have documented an association
between invasion and increased productivity or species richness have been correlative (but see Zavaleta & Hulvey 2004;
Maron & Marler 2008). Thus, it is unclear to what extent
increased productivity at invaded sites is due to invasion
itself, or driven by underlying site characteristics (such as
nutrient status of soils) that facilitate both invasion and
increased productivity. It is also generally untested whether
the increase in production at invaded sites (or changes in
species richness) is due to the addition of a new species to a
system or whether it is caused by the addition of new exotic
species to a system. Some would argue that given similar
1
5
INTRODUCTION
Division of Biological Sciences, University of Montana, Missoula, MT, 59812,
USA
2
Department of Community Ecology, UFZ, Helmholtz Center for Environmen-
Institute of Biology, Martin Luther University Halle-Wittenberg, D-06108,
Halle, Germany
6
Environmental Science, Policy & Management, University of California at
tal Research, D-06120, Halle, Germany
Berkeley, 130 Mulford Hall, Berkeley, CA, 94720-3114, USA
3
†Present address: Biology Department, Washington University in St. Louis,
German Centre for Integrative Biodiversity Research (iDiv) Halle-Jena-Leipzig,
D-4103, Leipzig, Germany
Campus Box 1137, One Brookings Drive, St. Louis, MO 63130-4899, USA
4
*Correspondence: E-mail: john.maron@mso.umt.edu
Rocky Mountain Research Station, USDA Forest Service, Missoula, MT, 59801,
USA
© 2014 John Wiley & Sons Ltd/CNRS
500 J. L. Maron et al.
circumstances, invading systems with novel native species
should produce similar patterns as invading them with
exotics (Davis et al. 2011).
Second, what is the role of disturbance and consumer pressure in mediating the impacts of colonising species on community structure or ecosystem processes? Disturbance that
removes resident species can lead to greater recruitment of
added species (Turnbull et al. 2000; Zobel et al. 2000; Myers
& Harms 2009; Maron et al. 2012; Kempel et al. 2013).
Generalist rodent consumers, as both granivores and herbivores, can suppress the recruitment of individual species
(Howe et al. 2006; Bricker et al. 2010; Pearson et al. 2011,
2012; Maron et al. 2012) and also reduce resident plant
cover/productivity (Batzli & Pitelka 1970; Keesing 2000;
Heard & Sax 2013). While much interest has centred on the
role of generalist native consumers as sources of biotic
resistance (Elton 1958; Parker et al. 2006; Liu et al. 2007;
Pearson et al. 2012), standardised parallel experiments across
systems that quantify the magnitude of biotic resistance to
staged invasions are uncommon. In most cases we do not
know how strongly native consumers suppress the productivity or diversity caused by newly arriving species. Differences
in consumer pressure on natives versus exotics could, at least
in part, account for the widespread correlation between invasion and enhanced productivity.
Finally, how might disparate grassland systems differ in
their inherent invasibility? This has been a topic of longstanding interest in invasion biology (Lonsdale 1999), yet
large-scale comparisons of invasibility have been challenging
because simple contrasts in the number of exotic species that
different systems support, or the density of exotics in those
systems, are often confounded by history and differences in
propagule pressure. Addressing this question requires quantifying invasibility across systems by controlling propagule pressure and experimentally examining how in situ filters affect
invasion.
We staged experimental invasions across grasslands in Montana, California (USA) and Germany. These grassland systems vary in underlying productivity and soil nutrients, the
composition of the generalist rodent herbivore assemblages
(granivores dominant in Montana, herbivores dominant in
California and Germany), background levels of invasion as
well as underlying disturbance regimes (German grasslands
are maintained by mowing; see Appendix S1 for a detailed
comparison of these grasslands). At sites across each grassland we added seeds of 19–20 native and 19–20 exotic species
(species number depending on region) to subplots that were
either disturbed (to remove resident competitors) or not in
and out of larger rodent exclosures. By performing seed addition experiments across such divergent grasslands we could:
(1) determine the strength and consistency with which addition of novel species influences productivity and species richness, (2) assess whether these effects differ predictably
depending on seed provenance (i.e. whether added species are
native vs. exotic) and (3) compare the strength of local in situ
filters (i.e. impacts of generalist consumers and/or resident
competitors) in their ability to ‘resist’ the impacts of seed
addition.
© 2014 John Wiley & Sons Ltd/CNRS
Letter
METHODS
Experimental design
We established experiments in three distinct grassland systems:
(1) perennial caespitose grasslands of the Blackfoot Valley,
western Montana, (2) mixed exotic, annual and native
perennial-dominated, coastal-influenced grasslands on the
Pepperwood Preserve, northern California, and (3) perennialdominated semi-dry grasslands in central Germany (Appendix
S1).
Within each region, we established experiments at 10 sites
(nine sites in California), with sites separated by 1–54 km,
0.1–3.5 km and 1.6–31.6 km in Montana, California and
Germany respectively. At each site within each grassland type,
we randomly selected a location to establish a rodent exclosure and an adjacent paired rodent exclosure control plot of
the same size. Rodent exclosures were 10 9 10 m (or at some
sites 10 9 15 m) in Montana, and 5.5 9 7 m in California
and Germany. All were constructed from 0.625 9 0.625 cm
wire mesh fencing buried 40–60 cm deep and which extended
60 cm aboveground. Fencing was topped with metal flashing
to prevent rodents from climbing over the top. We maintained
snap traps within exclosures to ensure they were secure.
Rodent exclosures did not exclude birds, invertebrates or
ungulates, but observations (and experimental seed depots
placed in exclosures in Montana and California) revealed that
animals other than rodents removed few seeds (J.L. Maron,
D.E. Pearson & C. Stein unpublished data) and there was little
evidence of ungulate browsing in experimental plots at any
site.
Within each rodent exclosure and rodent exclosure control
plot, we established 12 permanently marked 0.5 9 0.5 m subplots. Subplots were randomly assigned to a unique factorial
combination of seed addition (native, exotic or no seeds
added) and disturbance, with each treatment replicated twice
(two
replicates 9 +/ disturbance 9 native/exotic/control
seed addition = 12 subplots). In Montana, vegetation within
subplots assigned to the disturbed treatment was killed in
mid-growing season in 2009 using the broad spectrum,
low-persistence herbicide Roundup (Monsanto Corporation,
St Louis, MO, USA). Several weeks after the herbicide application, we disturbed the top 10 cm of soil and removed dead
vegetation from each subplot using a hoe. In California and
Germany, no herbicide was used and instead the top 15 cm of
soil was mechanically turned over and all aboveground plant
parts and coarse roots were removed.
We added seeds of 19 or 20 native species and 19 or 20 exotic species (depending on the region, see Appendix S2) to subplots designated for seed addition (with seed provenance
randomly assigned to subplot). Added species were grassland
species that occurred in the regional species pool, but were
mostly uncommon locally within each region’s grassland;
some species were shared across regions (Appendix S2). As
added species differed in seed size, we added fewer seeds of
large vs. small seeded species to account for the seed size/seed
number trade-off (Haig & Westoby 1988; Appendix S2). Plant
species were chosen so that within each region, the distribution of seed size, life span and affiliation to functional group
Letter
were as similar as possible between the native and the exotic
species pool. Seeds were collected locally within each region.
In Germany and California, seeds were added to the two replicate subplots of each treatment combination at the same
time, in fall 2009. In Montana, we replicated in time in addition to space; seeds were added to one subplot of each treatment combination in fall 2009 and again to a separate
replicate subplot of each treatment combination in 2010. Separate control no-seeds-added subplots of each treatment combination were also established in 2009 and 2010.
We estimated the invasibility of grasslands, and how disturbance and rodents influenced invasibility, by examining how
addition of native vs. exotic seeds affected the aboveground
biomass, species richness and evenness in seed addition subplots. In summer 2012, 3 years after seeds were added to subplots, we scored the presence/absence of all species in each
seed addition (and control) subplot (in Montana we only
scored the presence/absence of added species) using a quadrat
with 25 squares of 10 9 10 cm (California and Germany) or
100 squares of 5 9 5 cm (Montana), and then we harvested
all aboveground biomass in each seed addition and control
subplot at all sites within the three regions when vegetation
was at peak biomass. In Germany, biomass and species richness data were averaged across the two replicate subplots of
each treatment combination. As grasslands in Germany are
managed by mowing, we simulated mowing by annual biomass harvests in and out of rodent exclosures in the years
prior to 2012 as well. In California, only one replicate was
harvested, and in Montana, only subplots established in 2009
(i.e. the same year the experiment was established in California and Germany) were scored and harvested. In Montana,
harvested plants were sorted to species, bagged, placed in a
drying over 60 °C until dry and weighed. In California and
Germany, plants were not sorted by species, but bagged, dried
and weighed as above. In Montana, in subplots to which we
added exotic seed we clipped all flowers/seed heads from
plants on which these occurred to prevent new seed input. At
the end of the experiment we killed all exotics using Roundup.
This was not done in Germany and California.
To examine how seed addition influenced the evenness of
local assemblages, in California and Germany, we visually
estimated the per cent cover of all species within each seed
addition and control subplot at peak biomass in summer
2012. Percentage cover of species was then used to calculate
Pielou’s species evenness of each subplot using the index:
J = ∑ (Pi 9 ln Pi)/ln S with Pi being the relative contribution of the ith species to total cover, and S the total number
of species on subplot.
Analysis
To examine how seed addition influenced aboveground plant
biomass, at each site within each grassland type, we calculated
a log response ratio, defined as ln(seeds added/control) where
control was the total aboveground biomass of resident vegetation in the no-seeds-added subplots (of the same disturbance
and rodent exclusion treatment as the seeds-added subplot
used in the numerator) and seeds added was the total aboveground biomass in a seeds-added subplot of the same site/
Experimental invasion, productivity and species richness 501
treatment combination as the denominator. By performing
analyses on the log response ratio (rather than simply the difference in productivity between seed addition and control subplots) we could standardise across our three grassland types
that varied in background productivity. We used a four-way
ANOVA to determine how region (Montana, California and
Germany), seed provenance (native vs. exotic seeds), rodent
exclusion and disturbance (all fixed factors) and their interactions influenced the log response ratio of aboveground biomass. Site (nested within region) was treated as a random
variable and rodent exclusion was treated as a whole-plot factor with disturbance and seed addition as split-plot factors.
For data from California, we had to eliminate three sites from
the analysis because voles breached rodent exclosures at these
sites and affected vegetation inside rodent exclosures. Data
were analysed in the GLIMMIX module in SAS (ver. 9.3)
(SAS Institute, Cary, NC, USA) using an underlying Gaussian distribution. We used orthogonal contrasts to test a priori
hypotheses and to decompose significant interactions. In particular, as the rodent community is dominated by voles in
California and Germany, whereas in Montana mice are more
common, we performed an a priori contrast to compare the
average of California and Germany to Montana. To investigate whether adding native or exotic seeds increased productivity over the control without seed addition, we used t-tests
to test whether there was a significant difference between least
square means of log response ratios with zero. Any value significantly larger than zero indicates that adding seeds
increases biomass.
To determine how seed addition influenced total species
richness, for data collected from California and Germany, we
calculated the difference between total number of species on
seed addition plots, and total number of species on control
plots (of the same disturbance/rodent exclusion treatment
combination) in 2012. This difference represents the number
of newly established added species on seed addition plots. In
Montana, as we only recorded the presence of added species
within seed addition subplots, we calculated the difference in
the total number of added species on seed addition subplots
and control subplots of the same treatment combination. We
then ran an identical four-way ANOVA as above, with the
response variable being the difference in species richness
between a seed addition subplot and control no-seeds-added
subplot. A priori contrasts were constructed as discussed
above. To determine how seed addition influenced the proportional change in overall species richness (i.e. resident + added
species), we performed the same ANOVA as above, but using
the log response ratio for species richness (defined as for biomass) as the response variable. We also ran three additional
four-way ANOVAs, structured as above, but for data from
Germany and California only. The response variables for
these analyses were as follows: (1) the difference in the total
number of non-added resident species in each seed addition
and no-seeds-added control subplot (paired within site and
rodent exclusion/disturbance treatment), (2) the difference in
species evenness values between seed addition and no-seedsadded control subplots, again paired within site and rodent
exclusion/disturbance treatments, and (3) the difference
between paired seed addition and no-seeds-added control sub© 2014 John Wiley & Sons Ltd/CNRS
502 J. L. Maron et al.
Native
Exotic
1.2
Montana
0.9
0.6
0.3
Log response ratio of above-ground biomass
plots in the summed total percent cover of all resident vegetation (i.e. non-added species) in each subplot. These analyses
allowed us to examine how seed addition affected evenness
and to what extent changes in species richness and cover
reflected added species vs. changes in resident community
structure. Finally, to explore how particular functional groups
(grasses, non–N-fixing forbs, N-fixing forbs; see Appendix S2)
responded to disturbance and/or rodent exclusion, and how
this differed based on seed provenance and across regions, we
calculated the average cover of all added species of each functional type in Montana, California and Germany using their
scores in the gridded 50 9 50 cm quadrats. In Montana, we
omitted two species, Collinsia parviflora and Veronica verna
because these are both early season ephemeral species that
had senesced by the time we estimated cover later in the season. As some added species showed high abundances in noseed addition control plots, we corrected their cover in seed
addition plots by their cover in the paired control plot. We
then ran a four-way ANOVA (as above) for each of the three
functional groups separately using average cover as the
response variable. If a particular species did not recruit in any
subplot at a particular site, that species was omitted from the
data at that site.
Letter
0.0
–0.3
California
1.2
0.9
0.6
0.3
0.0
–0.3
Germany
1.2
0.9
0.6
0.3
0.0
RESULTS
Invasibility-effects of seed addition on productivity
Across disturbance and rodent exclusion treatments, exotic
seed addition led to proportionately greater productivity than
did native seed addition (Fig. 1; F1,133 = 25.62, P < 0.001). As
well, exotic seed addition substantially increased productivity
over standing biomass in no-seeds-added plots (t = 6.42,
P < 0.001 across all sites; overall effect of adding native seeds
not significant: t = 1.55, P > 0.05). We calculated the difference in productivity between seeds added and no-seeds-added
subplots, and compared the magnitude of this difference
between exotic vs. native seed addition subplots. The difference in the effect of exotic vs. native seed addition on productivity was substantial and averaged 332.8 g/m2, 52.4 g/m2 and
45.2 g/m2 in Germany, California and Montana respectively.
The log response ratio for how seed addition influenced productivity was marginally non-significantly different among
regions (F2,23 = 2.7, P = 0.09).
In disturbed seed addition subplots there was proportionately more biomass compared with undisturbed seed addition
subplots (Fig. 1; F1,133 = 16.94, P < 0.0001). However, the
magnitude of this effect also differed among our three regions
(region 9 disturbance interaction, F2,133 = 4.12, P < 0.02),
with California differing from Germany (contrast, t = 2.61,
P < 0.01), whereas Montana did not differ from the average
of the other two regions (t = 1.57, P > 0.05). In absolute
terms, the difference in productivity between disturbed and
undisturbed seed addition subplots (calculated as described
earlier) was 217.2 g/m2, 12.4 g/m2 and 45.6 g/m2 in Germany,
California and Montana respectively. Interestingly, disturbance had similar proportionate effects on productivity across
both exotic and native seed addition subplots (non-significant
disturbance 9 seed
addition
interaction;
F1,133 = 2.11,
© 2014 John Wiley & Sons Ltd/CNRS
–0.3
Undisturbed
No rodents
Undisturbed
Rodents
Disturbed
No rodents
Disturbed
Rodents
Figure 1 Effects of rodent consumers and disturbance on aboveground
plant biomass in subplots to which we added native or exotic species as
seed (see Methods for details). Shown are least square mean ( SEM) log
response ratios for aboveground biomass, defined as ln(seeds added/
control) where control values are those from no-seeds-added control
subplots of the same disturbance and rodent exclusion treatment
combination as a particular seed addition subplot. Values above zero
indicate that seed addition increased aboveground plant biomass
compared with plots to which no seeds were added. Least square means
calculated from the region 9 rodent exclusion 9 disturbance 9 seed
addition interaction.
P > 0.05). Although the main effect of rodent exclusion on
productivity was not significant (Fig. 1; F1,23 = 0.36,
P > 0.05), the effects of these consumers varied significantly
among our regions (region 9 rodent interaction, F2,23 = 5.59,
P < 0.02). In Germany and California, voles were extremely
abundant and vole damage to vegetation was visually striking.
In contrast, this was not the case in Montana (because voles
are rare and granivorous mice are more important; Maron
et al. 2012). As such, the suppressive effects of rodent consumers on plant production in Germany and California were
greater than effects of rodent consumers in Montana (contrast, t = 2.89, P < 0.01). The difference in plant biomass
between seeds-added and no-seeds-added subplots in rodent
exclosures averaged 364 g/m2, 59.12 g/m2 and 18.8 g/m2 in
Germany, California and Montana respectively. In contrast,
the average difference in productivity between seeds-added
and no-seeds-added subplots outside of rodent exclosures was
156.4 g/m2, 4.6 g/m2 and 22.3 g/m2 in Germany, California
and Montana respectively. Rodents similarly suppressed plant
biomass in native vs. exotic seed addition subplots
Letter
Experimental invasion, productivity and species richness 503
(rodent 9 seed addition interaction, F1,132 = 0.21, P > 0.05).
All other two-, three- and four-way interactions were not statistically significant (Appendix S3).
Invasibility-effects of seed addition on species richness and diversity
Across our three regions, adding native seeds or exotic seeds
increased species richness on average by 4.7 species (t = 11.61,
P < 0.001) or 5.0 species (t = 12.35, P < 0.001) respectively.
There were significant differences among our regions in how
seed addition influenced species richness (Fig. 2; F2,23 = 5.57,
P < 0.02), with seed addition increasing species richness to a
greater extent in Montana compared with Germany and California (contrast, t = 2.91, P < 0.01), and in Germany compared with California (contrast, t = 2.08, P < 0.05). Averaged
across our regions, invading subplots with exotic species did
not lead to greater species richness than invading with natives
(Fig. 2; F1,138 = 0.67, P > 0.05), although this differed among
grasslands
(region 9 seed
addition;
F2,138 = 19.73,
P < 0.0001). Exotic seed addition led to relatively greater
species richness in comparison to native seed addition in
Montana compared with Germany and California (contrast,
t = 5.86, P < 0.001). Disturbing seed addition subplots
increased species richness (Fig. 2; F1,138 = 14.23, P < 0.001),
12
Native
Exotic
Montana
8
4
Change in species richness
0
12
California
8
4
0
12
Germany
8
4
0
Undisturbed
No rodents
Undisturbed
Rodents
Disturbed
No rodents
Disturbed
Rodents
Figure 2 Effects of rodent consumers and disturbance on species richness
in subplots to which we added native or exotic species as seed. Shown are
least square mean ( SEM) differences in species richness between seed
addition plots and plots with the same treatment that did not receive seed
addition. Least square means calculated from the region 9 rodent
exclusion 9 disturbance 9 seed addition interaction.
although this effect also varied among regions (disturbance 9 region; F2,138 = 7.69, P < 0.001). Disturbance had
substantially greater impacts on species richness in Germany
compared with California (contrast, t = 3.59, P < 0.001),
whereas Montana did not differ from California and
Germany combined (t = 1.57, P > 0.05). As well, there was
a significant disturbance 9 region 9 seed provenance interaction (F2,138 = 6.60, P < 0.002). In Montana, disturbance
greatly enhanced the species richness of exotics in comparison
to natives, whereas this was not the case in Germany and
California (contrast, t = 3.38, P < 0.001). Finally, excluding
rodents had no significant effect on how seed addition influenced species richness (Fig. 2; F1,23 = 1.22, P > 0.05), and no
other two-, three- or four-way interaction was statistically
significant (Appendix S3). When we used the log response
ratio for species richness as the dependent variable to examine
the proportional change in species richness as a result of adding seeds, we found significant differences among regions
(F2,23 = 48.84, P < 0.0001) but no other significant main
effects or interactions. The addition of seeds resulted in a
greater proportional increase in species richness in Montana
compared with Germany and California (contrast, t = 9.81,
P < 0.0001).
Seed addition did not change species richness of resident
non-added species in California (lsmean = 0.04, t = 1.0,
P > 0.05), but slightly decreased it in Germany
(lsmean = 1.15, t = 3.38, P < 0.01), with the difference
between regions being significant (F1,14 = 8.16, P < 0.05).
While seed provenance had no significant influence
(F1,84 = 2.04, P > 0.05), the number of resident species in
California and Germany was reduced in disturbed but not
undisturbed subplots (undisturbed lsmean = 0.24, t = 0.73,
P > 0.05, disturbed lsmean = 0.95, t = 2.93, P < 0.01;
F1,84 = 12.67, P < 0.001). There were no other significant
main effects or interactions in effects on resident species richness (Appendix S4).
Seed addition of native species as well as of exotic species
slightly
increased
evenness
(exotic
seed
addition:
lsmean = 0.051, t = 2.78, P < 0.01; native seed addition:
lsmean = 0.056, t = 3.06, P < 0.01), in both Germany and
California; however, neither the effect of seed provenance nor
any other main or interaction effect on evenness was significant (Appendix S4). The total cover of resident species was
not reduced by seed addition in California (lsmean = 3.4,
t = 0.65, P > 0.05), but was significantly decreased in Germany (lsmean = 25.4, t = 6.27, P < 0.0001), with the difference between regions being significant (F1,14 = 11.02,
P < 0.01). Disturbance decreased the total cover of resident
species (F1,84 = 33.54, P < 0.0001), but these effects also differed among regions (disturbance 9 region; F1,84 = 15.62,
P < 0.001), with disturbance having limited effects on total
resident cover in California (lsmeans; undisturbed = 0.7,
t = 0.12, P > 0.05, disturbed = 6.2, t = 1.07, P > 0.05),
but decreasing resident cover in Germany (lsmeans; undisturbed = 10.9, t = 2.46, P < 0.05, disturbed = 39.9,
t = 8.97, P < 0.0001). All other main effects and interactions
were non-significant (Appendix S4).
The average cover of added species within each of the three
functional types (grasses, non–N-fixing forbs, N-fixing forbs)
© 2014 John Wiley & Sons Ltd/CNRS
504 J. L. Maron et al.
Letter
turbance, F1, 134 = 10.45, P < 0.002). All other interactions
were non-significant.
The average cover of N-fixing forbs differed among regions
(F2,20 = 8,56, P < 0.003) and was higher in California and
Germany than Montana (contrast t = 2.3, P < 0.04; Fig. 3).
Exotic N-fixers tended to be more abundant than native
N-fixers (F1,108 = 8.16, P < 0.005), but the region 9 seed
provenance interaction was significant (F1,108 = 9.83,
P < 0.001) with the average cover of exotic N-fixers greater in
California and Germany than in Montana (contrast t = 3.44,
t < 0.0008). Finally, the effect of rodent exclusion differed
across regions (region 9 rodent exclusion, F2,20 = 4.38,
P < 0.03), with rodent exclusion tending to enhance the cover
of N-fixers to a greater extent in California and Germany vs.
Montana (contrast, t = 1.9, P = 0.07; Fig. 3).
had variable responses to disturbance and/or consumer
removal. Average cover of added grass species did not change
significantly across any treatment combination, and this was
consistent across regions (effect of all main effects and interactions P > 0.11). The average cover of added non–N-fixing
forb species was higher in California and Germany than in
Montana (effect of region, F2,22 = 19.94, P < 0.0001; contrast,
t = 3.57, P < 0002), higher for exotics than natives
(F1,134 = 11.09, P < 0.002), greater in disturbed than undisturbed subplots (F1,134 = 127.6, P < 0.0001) and marginally
non-significantly higher inside vs. outside of rodent exclosures
(F1,22 = 3.72, P = 0.067; Fig. 3). As well, disturbance
enhanced the average cover of added non–N-fixing forbs more
in California and Germany than in Montana (region 9 disturbance, F12,134 = 32.8, P < 0.001; contrast, t = 2.44,
P < 0.02), and more in Germany than in California (contrast,
t = 7.25, P < 0.0001; Fig. 3). Added exotic non–N-fixing forbs
had greater cover than added native non–N-fixing forbs in
California and Germany compared with Montana
(region 9 seed provenance, F2,134 = 4.75, P < 0.013; contrast,
t = 2.69, P < 0.008; Fig. 3). Finally, added exotic non–N-fixing forbs had greater average cover in disturbed subplots than
did added native non–N-fixing forbs (seed provenance 9 dis-
0.30
Average cover
Average cover
0.25
Native
Exotic
DISCUSSION
Our replicated seed addition experiment in grasslands in
Montana, California and Germany yielded several novel
results. Perhaps most noteworthy was that invading with exotics led to substantially increased standing biomass, both over
background productivity in no-seeds-added subplots as well as
Montana
0.30
Non N-fixing forbs
0.25
0.20
0.20
0.15
0.15
0.10
0.10
0.05
0.05
0.00
0.00
0.30
California
0.30
0.25
0.25
0.20
0.20
0.15
0.15
0.10
0.10
0.05
0.05
0.00
0.00
N-fixing forbs
Average cover
Germany
0.40
0.40
0.35
0.35
0.30
0.30
0.25
0.25
0.20
0.20
0.15
0.15
0.10
0.10
0.05
0.05
0.00
0.00
Undisturbed Undisturbed Disturbed Disturbed
No rodents Rodents No rodents Rodents
Undisturbed Undisturbed Disturbed Disturbed
No rodents Rodents No rodents Rodents
Figure 3 Effects of rodent consumers and disturbance on the average cover of native and exotic species of non-N-fixing forbs (left panels) and N-fixing
forbs (right panels). Shown are least square means ( SEM) calculated from the region 9 rodent exclusion 9 disturbance 9 seed addition interaction.
© 2014 John Wiley & Sons Ltd/CNRS
Letter
over subplots to which we added native species. Our results
therefore experimentally demonstrate that patterns found in
correlative studies where invasion is often associated with
higher plant productivity (Liao et al. 2008; Vil
a et al. 2011)
may arise because exotics themselves increase productivity. By
overcoming propagule limitation through seed addition, we
also increased local species richness. However, unlike for productivity, invading with natives and exotics generally had similar impacts on species richness. Competition from resident
vegetation, as well as herbivory/granivory by generalist
rodents, mediated how strongly added species affected community structure and productivity, with the strength of these
interactions varying across grassland types. Disturbance
allowed for greater final biomass and species richness in seed
addition subplots. In Germany and California, these effects
occurred irrespective of whether we invaded with native or
exotic species. In Montana, however, disturbance increased
exotic species richness more than native species richness.
Exclusion of small mammals, primarily generalist voles in
Germany and California, led to substantial gains in the productivity of seed addition subplots, regardless of whether the
species added were native or exotic. In Montana, rodent
exclusion had minimal effects on the productivity.
Averaged across disturbance and rodent exclosure treatments, exotic seed addition increased productivity over background conditions (i.e. productivity of no-seeds-added control
subplots) by an average of 40–163%, depending on grassland
type. The increased production in exotic seed addition subplots was generally due to just a few species obtaining high
abundance, but was not consistently related to affiliation to
any of the plant functional groups (grasses, non–N-fixing forbs or N-fixing forbs). In Germany, exotic N-fixing species
(Appendix S2) were disproportionately dominant, whereas in
California, Daucus carota and Plantago lanceolata were the
dominant added exotic species. Interestingly, local filters such
as competition and/or disturbance did not universally favour
exotic production over natives. For example, across our three
grasslands exotic added species did not out produce native
added species with disturbance. Thus, other factors likely contribute to the overall greater production of plots with added
exotics vs. added natives or controls.
Interestingly, we found no significant differences among
regions in the impact of seed provenance on productivity or species richness (averaged across all treatments), indicating that
the three disparate grasslands had consistently similar responses
to native vs. exotic seed addition. Moreover, when all local filters were in place (i.e. when we analysed only those subplots
that were undisturbed and open to rodent consumers), there
were no significant differences among regions in the impacts of
seed addition on the log response ratio of productivity
(F2,23 = 0.03, P > 0.05) or species richness (F2,23 = 0.57,
P > 0.05). Thus, no one grassland was inherently more open to
invasion, so long as they were undisturbed and generalist consumers were present. Furthermore, in a separate analysis we
found no consistent pattern across regions in how total per species seed weight influenced the ultimate abundance of those species (J.L. Maron, H. Auge, D.E. Pearson and C. Stein,
unpublished data). The fact that seed addition in the presence
of local filters increased local species richness, and did so simi-
Experimental invasion, productivity and species richness 505
larly for natives vs. exotics, indicates dispersal limitation is common, and relatively equal in magnitude for the native and
exotic species we added. That gains in species richness due to
seed addition were similar in magnitude across grasslands (in
the presence of filters) suggests congruence across systems in the
magnitude of dispersal limitation. The minor impacts of seed
addition on overall species evenness, the number of resident
species and on their cumulative cover in Germany and California suggest that gains in both productivity and species richness
due to seed addition did not come at great expense to resident
vegetation. While we did not collect data to analyse these effects
in Montana, we observed qualitatively similar results. Had the
experiment run for a longer duration, we would expect that exotics would ultimately start to decrease resident abundance and
perhaps even diversity, particularly in disturbed plots protected
from rodents.
Few studies have quantified how different grassland systems
vary in the strength with which resident competitors or consumers inhibit establishment of colonising species, or compared how the intensity of these interactions vary depending
on whether colonisers are native or exotic species. Metaanalyses that amalgamate results from different studies in
divergent systems have found conflicting results. Some reviews
have shown that generalist native consumers can provide biotic resistance to invaders (Parker et al. 2006; Kimbro et al.
2013). Others indicate that native plants incur greater enemy
damage than exotics (Liu & Stiling 2006; Hawkes 2007),
although effects of consumers on plant performance are generally similar between natives and exotics (Chun et al. 2010).
Finally, other meta-analyses have indicated that the strength
of biotic resistance imposed by resident competitors and consumers is roughly similar (Levine et al. 2004). Although useful, these results are often difficult to interpret because
‘invasion resistance’ is often defined differently among studies.
Some studies examine effects of native consumers/competitors
on the success of established invaders (i.e. Kimbro et al.
2013), whereas others examine how native species influence
invader establishment or impact (i.e. Levine et al. 2004).
Moreover, as meta-analyses summarise results from single
experiments done in different systems, where methodology
and other factors are not standardised, the generality of such
results are not always clear.
Using standardised parallel experiments in grasslands in
Montana, California and Germany, we found that adding seeds
of a diverse suite of species to disturbed subplots led to universally greater productivity and species richness compared with
seed addition into undisturbed subplots that supported resident
competitors. These results confirm that seed addition combined
with removal of resident species often leads to greater recruitment of added species compared with seed addition alone
(Gross & Werner 1982; Turnbull et al. 2000; Zobel et al. 2000;
Myers & Harms 2009). Although competition by resident vegetation reduced the impacts of added species (on productivity
and species richness) in all three grassland types, the magnitude
of these effects also differed among systems. In western Montana, large resident bunchgrasses provided strong competitive
resistance to invasion (Maron et al. 2012). Disturbing subplots
in Montana allowed more added species (particularly exotics;
Fig. 2) to establish than in Germany and California, partly
© 2014 John Wiley & Sons Ltd/CNRS
506 J. L. Maron et al.
because resident vegetation in Montana recovers very slowly
after disturbance. This slow recovery of vegetation also led to
low biomass in disturbed seeds added subplots through time
(although less so for subplots invaded by fast-growing exotics
vs. slower growing natives; Fig. 1), which accounts for the more
muted effects of disturbance on productivity in Montana vs. the
other two grasslands. In Germany and California, disturbance
not only facilitated the establishment of invading species, but
some of these attained high biomass very quickly. This lead to
enhanced production in disturbed vs. undisturbed subplots and
also in comparison to disturbed subplots in Montana. Interestingly, however, resident vegetation appeared to ‘resist’ invasion
similarly for exotics and natives (at least in Germany and in
California; in Montana, exotics were favoured more than
natives by disturbance).
Generalist rodent consumers also provided substantial resistance to ‘invasion’, but only in Germany and California, where
abundant voles had large but similar effects on both native and
exotic biomass. However, vole impacts, along with pocket
gophers in California, were not completely indiscriminant. In
California, species with belowground structures such as bulbs
(e.g. native Sisyrinchum bellum) or large taproots (exotic Daucus
carota) were particularly reduced, whereas in Germany, voles
decimated exotic N-fixers (e.g. exotic Medicago 9 varia, Onobrychis viciifolia and Vicia villosa). At Montana sites, voles are rare
(Maron et al. 2010), and although deer mouse seed predation
can be intense, it appears to have greater effects on abundance
of specific species, particularly large-seeded ones (Bricker et al.
2010; Pearson et al. 2011; Maron et al. 2012), rather than on
total aboveground plant biomass. Ground squirrel herbivory
on added species was negligible (J.L. Maron & D.E. Pearson
personal observations). These results, together with a lack of
rodent 9 seed provenance interaction for any of the plant functional groups, suggest that feeding preferences of generalist
native herbivores and granivores in the three regions are not
driven by plant provenance, but by other plant characteristics
such as defence and nutritional value.
Our experiment provides significant insights regarding how
in situ filters, theorised to control community composition
and ecosystem function (Weiher & Keddy 1999), actually
apply to both native and exotic plants across multiple systems. When all filters were in place, exotics had greater effects
than natives, but the differences in effects were limited. However, reducing competition or herbivory resulted in gains in
productivity and species richness, with the magnitude of these
effects varying depending on grassland. The fact that ‘invasion’ by exotics had larger impacts on productivity and (at
least in Montana) species richness than was produced through
‘invasion’ by natives suggests that seed provenance does matter, at least when local filters are disrupted. The weak effect
of seed provenance when filters were in place is consistent
with the notion that exotics are not inherently benefitting over
natives due to release from their specialist natural enemies.
However, when we experimentally disrupted local filters, it
exaggerated the effects of exotics on productivity, indicating
that release from generalist natural enemies (rodent consumers) and native competitors both enhance exotic productivity.
Yet there was weak evidence for additive or synergistic effects
of these two factors (sensu Blumenthal 2006). So why do the
© 2014 John Wiley & Sons Ltd/CNRS
Letter
impacts of exotics increase more than natives from disruption
of in situ filters? One possibility is that exotics are not a random sample of introduced species because immigration filters
select for species that have inherently greater productivity
(Van Kleunen et al. 2010), which can be expressed once local
filters are relaxed. Another possibility is that exotics alter belowground processes in ways that enhance their own production,
whereas this is less the case for natives. Examining these mechanisms in more detail, and exploring the relative influence of in
situ filters in affecting native vs. introduced species could yield
rich dividends for understanding invasion biology and community ecology.
ACKNOWLEDGEMENTS
We thank the many people who helped with field/lab assistance in Montana, Germany and California, who are too
numerous to mention individually here. We greatly appreciate
the Montana Department of Fish Wildlife and Parks, the
U.S. Fish and Wildlife Service, Pepperwood Preserve, the
University of Montana and numerous private land owners in
Montana and Germany for allowing us to work on their
lands. This research was supported with grants from the
USDA Cooperative State Research, Education and Extension
Service (2005-35101-16040 to JLM/DEP, 2006-01350 to KS),
NSF (DEB-0614406 to JLM/DEP, DEB-1001807 to KNS),
the German Academic Exchange Service (DAAD 50750649)
to HA, the Graduate School HIGRADE to LK and by generous travel support from the Global Invasions Network NSF
(RCN DEB-0541673; PIs R. Hufbauer and M. Torchin).
STATEMENT OF AUTHORSHIP
JM, DP, HA and CS designed and initiated the experiments,
and LK, IH and KS helped with collection of field data. JM
wrote the first draft of the manuscript and HA performed the
statistical analyses. JM, DP, HA, IH, LK, KS and CS edited
the manuscript.
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SUPPORTING INFORMATION
Additional Supporting Information may be downloaded via
the online version of this article at Wiley Online Library
(www.ecologyletters.com).
Editor, Jonathan Chase
Manuscript received 13 August 2013
First decision made 17 September 2013
Second decision made 5 December 2013
Manuscript accepted 24 December 2013
© 2014 John Wiley & Sons Ltd/CNRS