Pollinator conservation paradox: exotic wildflowers support native pollinators under global changes

Pollinator conservation paradox: exotic wildflowers support native pollinators under global changes | Research Square window.SnipcartSettings = { analytics: { enabled: false } }; (function() { var accessVector = localStorage.getItem('access_vector') || ''; window.dataLayer = window.dataLayer || []; if (accessVector) { window.dataLayer.push({ user: { profile: { profileInfo: { snid: accessVector } } } }); } })(); (function(w,d,s,l,i){w[l]=w[l]||[];w[l].push({'gtm.start':new Date().getTime(),event:'gtm.js'});var f=d.getElementsByTagName(s)[0],j=d.createElement(s),dl=l!='dataLayer'?'&l='+l:'';j.async=true;j.src='https://www.googletagmanager.com/gtm.js?id='+i+dl;f.parentNode.insertBefore(j,f);})(window,document,'script','dataLayer','GTM-K279D39R'); Browse Preprints In Review Journals COVID-19 Preprints AJE Video Bytes Research Tools Research Promotion AJE Professional Editing AJE Rubriq About Preprint Platform In Review Editorial Policies Our Team Advisory Board Help Center Sign In Submit a Preprint Cite Share Download PDF Research Article Pollinator conservation paradox: exotic wildflowers support native pollinators under global changes Rebecca Nelson, Eric Seabloom, Elizabeth Borer This is a preprint; it has not been peer reviewed by a journal. https://doi.org/ 10.21203/rs.3.rs-7188448/v1 This work is licensed under a CC BY 4.0 License Status: Posted Version 1 posted You are reading this latest preprint version Abstract Invasive wildflowers pose a conservation paradox: while they often reduce the diversity and abundance of native wildflowers, they can provide resources for native pollinators, including imperiled species. Previous work has framed wildflower invasions as outcomes of global change, but less is known about how interacting anthropogenic drivers influence both invasion and pollination. In particular, it remains unclear whether exotic wildflowers compensate for native floral losses under ongoing environmental change. To address this, we tested whether exotic wildflowers support native pollinators under two drivers of wildflower decline: eutrophication and defaunation. Using a factorial global change experiment at three sites in the highly invaded California floristic region, we tested whether increases in exotic wildflowers (1) sustain pollinator visitation and richness, (2) maintain pollinator composition and function, and (3) facilitate co-invasion by exotic pollinators. We found that eutrophication promoted exotic asters, which served as visually prominent, attractive hubs in plant-pollinator networks. These asters supported both generalist and specialist native pollinators but also increased visitation by exotic pollinators, raising the risk of invasional meltdown. Our results suggest that exotic wildflowers can buffer pollinator communities against global change, but may do so while shifting pollinator composition toward non-native species. Eutrophication Invasive Species Defaunation Plant-Pollinator Mutualisms Global Change Invasional Meltdowns Figures Figure 1 Figure 2 Figure 3 Introduction Invasive wildflowers pose a paradox for conservation. While invasive wildflowers can decrease the diversity and abundance of native wildflower species (Brown et al. 2002 , Powell et al. 2011 , 2013 , Vilà et al. 2011 , Goodell and Parker 2017 ), they can benefit native pollinators, including imperiled insect species (Harmon-Threatt and Kremen 2015 , Gibson et al. 2019 , Russo et al. 2019 ). Evidence for this paradox remains mixed (Stout and Tiedeken 2017 , Kaiser-Bunbury et al. 2017 , Szigeti et al. 2020 , Gaiarsa and Bascompte 2022 ). Prior efforts to tackle this paradox have begun with wildflower invasion as the focal global change (Carvalheiro et al. 2008 , Molina-Montenegro et al. 2008 , Bartomeus et al. 2008 , Parra-Tabla and Arceo-Gómez 2021 , Valdovinos et al. 2023 ). Yet, pollinators are declining under many global changes (Grixti et al. 2009 , Potts et al. 2010 , Dirzo et al. 2014 , Carvalheiro et al. 2020 , Janousek et al. 2023 , Edwards et al. 2025 ), and plant invasions that alter the quality of the environment for pollinators can be driven by interacting factors (Bradley et al. 2010 , Yang et al. 2022 ). This reality raises the question: do individual or interacting anthropogenic changes that favor exotic wildflowers compensate pollinators for losses of native wildflowers? Two such anthropogenic changes driving plant invasions and global declines in native wildflowers are eutrophication (Song et al. 2011 , You et al. 2017 , Li et al. 2022b , MacDougall et al. 2024 , Nelson et al. 2025 ) and defaunation (Borer et al. 2014b, Bråthen et al. 2021 , Price et al. 2022 , Borer and Risch 2024 , Zhu et al. in press), the loss of large wild mammalian herbivores (Dirzo et al. 2014 ). Under both eutrophication and defaunation, losses in wildflower abundance and diversity resulting from increased grass competition (DeMalach et al. 2017 , You et al. 2017 , Eskelinen et al. 2022 ) can decrease pollinator visitation (van Klink et al. 2015 , Filazzola et al. 2020 , Veen et al. 2024 ) and change pollinator community composition (van Klink et al. 2015 , David et al. 2019 , Carvalheiro et al. 2020 , Filazzola et al. 2020 , Wang et al. 2022 ). Unlike loss of large herbivores, eutrophication can also alter floral reward chemistry, with mixed effects on nectar and pollen attractiveness and quality (Ceulemans et al. 2017 , Majetic et al. 2017 , Russo et al. 2023 ). Moreover, large mammalian herbivores can rescue native wildflowers and their pollinators from declines under eutrophication (Veen et al. 2024 ). These past findings have largely focused on uninvaded ecosystems (van Klink et al. 2015 , Carvalheiro et al. 2020 , Wang et al. 2022 , Veen et al. 2024 ). Under eutrophication and defaunation, invasive wildflowers can increase or maintain their abundances (Hayes and Holl 2003 , Liu and Stiling 2006 , Seabloom et al. 2015 , Lannes et al. 2016 ). Herbivory also can counteract increases in invasive plants under eutrophication (Gianoli and Salgado-Luarte 2017 , Li et al. 2022a ). Thus, eutrophication and defaunation can lead to invasion, but their interaction can cause differing outcomes for pollinators, underscoring the need to consider the processes that lead to plant invasion. Paradoxically, invasive wildflowers may support pollinators in these rapidly changing landscapes. Pollinators adjust their foraging to track changes in floral abundance (Sih and Baltus 1987 ) and floral rewards (Ito et al. 2021 ), which may be altered in different ways by nutrient supply (Burkle and Irwin 2010 , Vaudo et al. 2022 ) or herbivores (Moreira et al. 2019 ). If exotic wildflowers become more abundant than native wildflowers under eutrophication or defaunation, native pollinators may visit a greater proportion of exotic wildflowers relative to native ones. Increases in exotic wildflowers under eutrophication and herbivore exclusion may maintain pollinator composition. Indeed, some exotic wildflower species can functionally replace native species as core network hubs (Russo et al. 2019 , Parra-Tabla and Arceo-Gómez 2021 ) that stabilize the network’s structure by supporting both generalist and specialist pollinators (Valdovinos et al. 2009 ). However, exotic wildflowers can decrease specialist pollinators, decreasing pollinator diversity (Kaiser-Bunbury et al. 2017 ). Alternatively, eutrophication and defaunation could interactively determine exotic vs native floral abundance (Burkepile et al. 2017 , Li et al. 2022a ) and consequently pollinator visitation. Increases in exotic wildflowers under eutrophication and defaunation could promote invasional meltdowns where exotic wildflowers co-invade with their exotic pollinators (Simberloff and Von Holle 1999 , Barthell et al. 2001 ), or alternatively have contrasting effects on exotic pollinators. Exotic and native plants can indirectly affect each other’s fitness both positively and negatively through competition for shared pollinators (Molina-Montenegro et al. 2008 , Etter et al. 2022 , Dritz et al. 2023 ), a well-documented form of apparent competition (Holt 1977 , Goodell and Parker 2017 ). Moreover, exotic pollinators, particularly the western honeybee ( Apis mellifera ) in its introduced range, can displace native pollinators on shared floral resources, decreasing native wildflower seed set (Goodell and Parker 2017 , Page and Williams 2023 ). The California Floristic Province provides an opportunity to examine this paradox. Although it is a hotspot for plant-pollinator biodiversity and specialization (Thorp and Leong 1998 , Frankie et al. 2009 , Harrison 2013 , Harrison et al. 2018 ), the California Floristic Province experiences high rates of plant invasions especially invasive asters (Levine and D’Antonio 1999 , Harrison 1999 , Seabloom et al. 2006 ), defaunation (Hayes and Holl 2003 ), eutrophication (Valliere et al. 2020 ), and introduced honeybees (Barthell et al. 2001 , Page and Williams 2023 , Travis et al. in press). To test the paradox that exotic wildflowers may support native pollinators under high nutrient conditions and altered herbivory, we replicated herbivore exclusion and nutrient addition experiments across three central California grasslands. We hypothesized that: increases in exotic wildflowers, particularly asters, and decreases in native wildflowers under eutrophication and herbivore exclusion will (1) increase native pollinator visitation and richness on exotic wildflowers relative to native wildfowers; (2) maintain native pollinator community composition via exotic asters acting as core network hubs for both specialist and generalist pollinators, but (3) increase exotic pollinator visitation via invasional meltdowns. Materials and Methods Study Sites : This research occurred at three grassland sites in California that spanned a 150 km longitudinal gradient with cool wet winters and hot, dry summers: the Sierra Foothill Research & Extension Center (39.24 N, -121.28 W, Dec 2023 precipitation: 121.77 cm), the University of California McLaughlin Reserve (38.86 N, -122.41 W, Dec 2023 precipitation: 164.07 cm), and the Hopland Research & Extension Center (39.01 N, -123.06 W, Dec 2023 precipitation: 196.27 cm). The exotic species pool contained primarily annual grasses and wildflowers from the Mediterranean region of Europe (Heady 1977 , Seabloom et al. 2006 ). Nutrient Network Experiments : All three sites contained replicates of a randomized block, factorial experiment that use identical methods to manipulate fertilization and herbivore presence as part of the globally replicated Nutrient Network (Borer et al. 2014a , Seabloom et al. 2015 ). Each site contained at least 3 randomized blocks of 5 m x 5m plots assigned to a factorial combination of the following treatments: Fencing (Control and Fenced) and Fertilization (Control and Fertilized with macro and micronutrients, NPKµ). This design created four distinct fencing and fertilization treatments: Control (unfenced and unfertilized), Fence (fenced and unfertilized), NPKµ, (unfenced and fertilized with NPKµ), and NPKµ, + Fence (fenced and fertilized with NPKµ). The fencing treatment was designed to exclude large, nonclimbing mammals. The fences were 230 cm tall with the lower 90 cm composed of 1 cm wire mesh, which included a 30 cm wide, outward-facing flange stapled to the ground to prevent digging. The upper portion of the fence is composed of strands of barbless wire. The fertilization treatment is composed of a combination of macro- and micro-nutrients (10 g N m − 2 yr − 1 as timed-release urea; 10 g P m − 2 yr − 1 as triple-super phosphate 10 g K m − 2 yr − 1 as potassium sulfate; and 100 g m − 2 yr − 1 of a micronutrient mix 6% Ca, 3% Mg, 12% S, 0.1% B, 1% Cu, 17% Fe, 2.5% Mn, 0.05% Mo, and 1% Zn). The macronutrients were applied annually, and the micronutrients were applied a single time at the start of the experiment. The primary large herbivores excluded at the sites include cattle and deer. Treatments started in 2008. Hopland and McLaughlin had three experimental blocks with these treatments, while Sierra Foothill had five experimental blocks. Our experimental design captured changes in the behavioral visitation of pollinators within a given site rather than population-level changes in pollinators. Plant-Pollinator Observations : We surveyed each plot for floral visitors (20-minute observations/plot) and wildflower abundance (# floral units/wildflower species/plot of non-gramminoid, herbaceous angiosperms) (Veen et al. 2024 ) at least three times over the course of the growing season (April-July) for two years, a minimum of 120 observer hours per plot per year in 2023 and 2024 across the full flowering season (see Supplemental Methods). Floral visitation surveys occurred only on calm sunny or partly sunny days and if flowering wildflowers were present. A floral visit occurred if an animal contacted the reproductive parts of the flower. When multiple floral visitors were present at the same time, we were able to simultaneously record visitation. We recorded the morphospecies identity of the floral visitor and the plant species visited. Floral abundance was estimated for each plot within ten days of the date a given pollinator survey occurred. We estimated floral abundance by counting the total number of floral units, defined as a 1 cm x 1 cm part of the flower, per flowering plant species in a given plot (Veen et al. 2024 ). Voucher specimens of each insect morphospecies were collected, pinned, and identified to species with the help of the Bohart Museum of Entomology (Davis, California, USA). We used these species data to classify pollinators into distinct morphospecies (see Supplemental Materials). When collecting specimens during our surveys, we paused the timer and resumed timing once collection was complete. For sensitive taxa (Papilionidae and Bombus spp.), we took photographs which were used to identify species. We ceased recording visits and collected vouchers once an individual left a plot to minimize the impact of collection. Data Analysis All data analyses were performed in R (v. 4.4.1). Models were checked for normality of residuals and overdispersion where appropriate.We performed rarefaction analyses for pollinator visitation data with the ‘rarecurve’ function in the ‘vegan’ package (Figure S4) (Dixon 2003 ). To test our first hypothesis, we used a generalized linear mixed effects models to test whether fencing and fertilization treatments affected exotic wildflower floral abundance, native wildflower floral abundance, native wildflower plant species richness, pollinator morphospecies richness on native vs exotic wildflowers, raw native pollinator visitation to native wildflowers (total pollinator visits), native pollinator visitation to native wildflowers per floral unit (per capita) (total pollinator visits/# floral units), native pollinator visitation to exotic wildflowers, and native pollinator visitation to exotic wildflowers per floral unit (per capita). To calculate per capita visitation values, we divided pollinator visitation by plot-level floral abundance of the plant species visited. For abundance data, we used negative binomial models due to overdispersion, while for richness data we used Poisson distributions. For per capita response variables, we used Gaussian distributions. We included a main effect of experimental treatments and a random effect of experimental block nested within site, and dropped the nested effect of block in cases where we had convergence issues (see Supplement). We used the emmeans package to compute estimated marginal least-square means for pairwise combinations of treatments (Lenth et al. 2024 ). We used generalized linear models to examine whether native pollinator visitation (# visits by native pollinator morphospecies) and native pollinator morphospecies richness were explained by native plant species richness, native floral abundance, and exotic floral abundance using negative binomial mixed effects models with a random effect tof site (Figure S2). To test our second hypothesis regarding pollinator composition, we compared how native pollinator community composition differed among treatment, plant species, plant family and local provenance using the vegan package in R to run PERMANOVAs with Bray-Curtis dissimilarity indices (Dixon 2003 ). We grouped native pollinator data by pollinator functional group (long-tongued bees, short-tongued bees, beetles, butterflies and moths, kleptoparasitic bees, and flies) using literature-based taxonomic and morphological information (LeBuhn 2013 ), and then compared how treatment affected native pollinator functional group abundance using differential multivariate generalized linear models with the mvabund package (Wang et al. 2012 ). To further test our the effects of plant provenance on network structure, for each site, we calculated bipartite networks of plant-pollinator interactions and measured the individual contribution of each wildflower species to network nestedness using the bipartite package in R (Dormann et al. 2009 ). Nestedness is the degree to which generalist plants (and pollinators) interact with both generalist and specialist pollinators (and plants) (Bascompte and Jordano 2007 , Bastolla et al. 2009 ). Nestedness is a core property of mutualistic networks associated with structural stability (Bastolla et al. 2009 , Valdovinos et al. 2016 ). By measuring the individual contribution of each plant species to network nestedness, we can determine whether or not invasive vs native plant species act as core hubs in the network attract diverse pollinators. To test our third hypothesis regarding invasional meltdowns, we examined how visitation of the two exotic pollinator species we observed responded to eutrophication and herbivore exclusion. We fit a negative binomial generalized linear effects model to test for the effects of treatments on visitation by the exotic western honeybee Apis mellifera wildflowers. We also counted visits by the exotic leafcutter bee Megachile apicalis by treatment and plant species. Becase we only recorded 14 total visits from M. apicalis , we did not have a sufficient sample size to fit a similar model for M. apicalis. We used multivariate abundance analyses with the mvabund package to examine whether exotic pollinator visitation differed by wildflower family. Results In total, we recorded 517 individual floral visitors to 31 wildflower species and 62 pollinator morphospecies across all sites and years and made 820 observations of floral abundances to 72 wildflower species across all sites and years (Tables S1-S2). At McLaughlin, we observed 31 pollinator morphospecies visiting 11 wildflower species, out of 29 total flowering wildflower species. At Sierra Foothills, we observed 32 pollinator morphospecies visiting 17 out of 42 total flowering wildflower species. At Hopland, we observed 28 pollinator morphospecies visiting 10 wildflower species out of 39 total flowering wildflower species. Common floral visitors included bee, syrphid fly, bombyliid fly and beetle species. We had 4 species of exotic asters that were visited by pollinators in our data: Carduus pycnocephalus , Senecio vulgaris , Sonchus asper , and Centaurea solstitialis . We observed two exotic pollinator species: western honeybee Apis mellifera and the leafcutter bee Megachile apicalis . We observed one imperiled pollinator species Crotch’s bumble bee ( Bombus crotchii ), visiting both native wildflowers and exotic asters. Wildflower Communities Fertilization decreased native wildflower floral abundance and diversity but increased exotic aster floral abundance, as measured by the number of floral units (Fig. 1 , Figure S1). Native wildflower species richness significantly declined under the combined effects of fertilization and herbivore exclusion relative to the unfenced controls (z = 3.18, p = 0.008, Table S3). When herbivores were excluded, native wildflower richness and native wildflower abundance significantly declined under fertilization (richness: z = 2.99, p = 0.015; abundance: z = 3.33, p = 0.005, Table S3). When herbivores were present, however, fertilization did not affect native wildflower floral abundance (Table S3) nor native wildflower species richness (Table S3). While the floral abundance of exotic wildflowers for all floral families combined together did not differ by treatment (Table S3), floral abundance of asters (Asteraceae), the most common family of exotic wildflower, increased under fertilization (z=-3.31, p = 0.005, Table S3) as well as its combined effects with herbivore exclusion (z=-4.23, p = 0.0001, Table S3) (Fig. 1 B). The increase in exotic asters under fertilized conditions occurred both when herbivores were excluded (z=-3.66, p = 0.001, Table S3) and present (z=-3.31, p = 0.005, Table S3). Native Pollinator Visitation and Richness In support of our first hypothesis, fertilization increased native pollinator visitation to exotic asters and produced shifts in native pollinators from native to exotic flowers under combined fencing and fertilization treatments (Fig. 1 , Table S4). Combined herbivore exclusion and fertilization treatments, however, did not affect native pollinator visitation per capita (Table S4). Native pollinator visitation to exotic asters increased under fertilization when herbivores were excluded (z=-3.08, p = 0.011, Table S4) but not when herbivores were present (z= -2.38, p = 0.082, Table S4). Native pollinator visitation of exotic wildflowers did not differ by treatment when all exotic wildflower families were combined (Table S4). The joint effects of herbivore exclusion via fencing and fertilization increased native pollinator visitation to exotic asters (z=-2.69, p = 0.036, Table S4). This interaction between fencing and fertilization and provenance was not significant for pollinator visitation per capita (Table S4). Contrary to our first hypothesis, native pollinator morphospecies richness did not differ by treatment (Table S4). In further support of our first hypothesis, exotic wildflower abundance was positively correlated with native pollinator visitation to exotic wildflowers (z = 3.88, p < 0.001, Table S5) and native wildflower abundance was positively correlated with native pollinator visitation to native wildflowers (z = 4.42, p < .0001) (Figure S3, Table S5). Native Pollinator Community Composition In support of our second hypothesis, eutrophication and herbivore exclusion did not shift pollinator composition and function, despite variation in native pollinator composition by wildflower provenance. Fertilization and herbivore exclusion did not directly alter native pollinator composition (PERMANOVA, F = 1.0329, p = 0.431) and functional group abundance (e.g. long-tongued bees, short-tongued bees, flies, etc) (mvabund, deviance = 37.45, p = 0.063), but pollinator composition did vary with wildflower species (PERMANOVA, F = 2.1585, p = 0.001), wildflower provenance (PERMANOVA, F = 2.0865, p = 0.004), and wildflower taxonomic family (PERMANOVA, F = 1.9917, p = 0.001) (Figure S5). In further support of our second hypothesis, tall, bright, nectar-producing, abundant exotic asters (e.g., Carduus and Centaurea species) acted as core plant-pollinator network hubs that attracted both generalist and specialist polliantors and thus contributed strongly to the network’s nested structure (Fig. 2 ). Exotic asters were visited by all the main functional groups of native pollinators: butterflies and moths, long-tongued bees, short-tongued bees, beetles, flies, and wasps. At our wettest site, Hopland, the native wildflower western blue-eyed grass Sisyrinchium bellum (Iridaceae) most strongly contributed to the nested structure of the plant-pollinator network (nestedness contribution (nc) = 2.66), followed by exotic Italian thistle Carduus pycnocephalus (Asteraceae) (nc = 1.50). At our intermediate rainfall site McLaughlin, the highly invasive wildflower yellow star-thistle Centaurea solstitialis (Asteraceae) contributed most strongly to the nested structure of the network (nc = 2.76) followed by the exotic Geranium dissectum (Geraniaceae) (nc = 1.06). At our driest site Sierra Foothills, the exotic thistle C. pycnocephalus (Asteraceae) (nc = 3.47) had the strongest individual contribution to nestedness followed by the native aster mountain dandelion Agoseris heterophylla (Asteraceae) (nc = 2.34). Exotic Pollinator Visitation In support of our third hypothesis, exotic pollinators (western honeybee Apis mellifera and leafcutter bee Megachile apicalis ) were strongly associated with exotic asters. Within the fencing by fertilization plots, native wildflowers only received two honeybee ( A. mellifera ) visits from a single individual bee, while exotic wildflowers received 215 visits from honeybees (Fig. 3 ). In the absence of large herbivores, eutrophication increased honeybee visitation (z=-2.59, p = 0.047, Table S6) but not when herbivores were present (z=-0.02, p = 1.0000, Table S6) (Fig. 3 ). Honeybees showed strong preference for foraging on exotic Asteraceae species (mvabund, deviance = 36.415, p = 0.013). Likewise, the specialist exotic leafcutter bee ( M. apicalis ) foraged exclusively on the exotic aster Centaurea solstitialis . Leaf cutter bees made one visit to C. solstitialis under ambient conditions, 11 visits to C. solstitialis under herbivore exclusion, 1 visit to C. solstitialis under eutrophication, and 1 visit to C. solstitialis under combined herbivore exclusion and eutrophication. Discussion We tested the paradox that exotic wildflowers may compensate pollinators for declines in native wildflowers under eutrophication, herbivore exclusion and their interaction. Consistent with this paradox, we found that native pollinator abundance on exotic asters increased in eutrophied plots and that exotic asters supported diverse pollinator functional groups, contributing strongly to the nested structure of local plant-pollinator networks by attracting both generalist and specialist pollinators. The abundance of exotic bees increased with exotic wildflowers, suggesting the potential for invasional meltdowns as a cascading response to ongoing environmental change. Eutrophication and herbivore exclusion decreased native wildflower diversity and abundance, consistent with prior findings (Song et al. 2011 , Borer et al. 2014b, You et al. 2017 , Johnson et al. 2020 , Valliere et al. 2020 , Nelson et al. 2025 ). Exotic wildflowers, especially asters, however, were robust to the presence of large mammalian herbivores but increased under fertlization, as has been shown in past studies (Liu and Stiling 2006 , Seabloom et al. 2009 ). Native wildflowers increased under herbivore exclusion in the absence of eutrophication in contrast to prior findings (Bråthen et al. 2021 ). This contrast suggests that fertilization more so than defaunation drives wildflower invaisons. Indeed, the traits that make exotic wildflowers effective invaders – fast-growing, tall stature, low palatability and high seed set – may make them effective competitors against grasses in light-limited, fertilized landscapes (MacDougall and Turkington 2005 , Coleman and Levine 2007 , Ordonez et al. 2010 ). Consequently, eutrophication as well as its combined effect with herbivore exclusion increased total native pollinator visitation to exotic wildflowers relative to native wildflowers. Under these global changes, pollinators may forage in response to increased abundances of exotic relative to native wildflowers (Sih and Baltus 1987 , Lopezaraiza–Mikel et al. 2007 ), higher nectar production and reward nutritional quality in exotic flowers relative native flowers (Harmon-Threatt and Kremen 2015 ), and greater apparency of taller, exotic flowers (Ordonez et al. 2010 ). The lack of a response of per-plant visitation to eutrophication and herbivore exclusion may suggest the former. Consistent with findings from uninvaded systems (Veen et al. 2024 ), large mammalian herbivory increased pollinator visitation to native wildflowers and decreased pollinator visitation to exotic wildflowers under eutrophication, indicating that large herbivores may rescue plant-pollinator interactions from global changes, perhaps due to consumption of exotic plant species (Price et al. 2022 , Nelson et al. 2025 ). Contrary to prior findings in relatively uninvaded systems (Carvalheiro et al. 2014 , David et al. 2019 , Veen et al. 2024 ), eutrophication and herbivore loss did not alter pollinator composition. Thus, exotic wildflowers may maintain pollinator community composition and function under other global changes. This result, however, may be scale-dependent. Our findings reflect behavioral changes in pollinator foraging, as do past studies (Valdovinos et al. 2023 , Russo et al. 2023 , Veen et al. 2024 ). At landscape scales, however, these global changes may drive plant-pollinator interaction turnover (Kaiser-Bunbury et al. 2017 , Carvalheiro et al. 2020 , Filazzola et al. 2020 , Parra-Tabla and Arceo-Gómez 2021 ). Scale may explain why we do not find differences in pollinator richness under global changes in contrast to landscape-level declines in richness (van Klink et al. 2015 , Kaiser-Bunbury et al. 2017 , Russo et al. 2019 , Filazzola et al. 2020 , Wang et al. 2022 ) (but see (Veen et al. 2024 )). Consistent with prior studies (Russo et al. 2019 , Parra-Tabla and Arceo-Gómez 2021 ), nectar-rich, exotic wildflowers, particularly asters, acted as core network hubs for pollinators by attracting both generalist and specialist pollinators. Under eutrophication, pollinators may visit a greater proportion of exotic wildflowers, decreasing interacton diversity. Exotic asters increased exotic pollinator visitation, suggesting potential invasional meltdowns (Simberloff and Von Holle 1999 , Braga et al. 2018 ). This finding is consistent with prior work demonstrating that exotic asters and exotic pollinators can co-invade as mutualists (Barthell et al. 2001 ) but contrasts with theoretical predictions (Dritz et al. 2023 ). Our results suggest that eutrophication more so than herbivore exclusion may have cascading effects on plant-pollinator interactions through increasing invasional meltdowns. Indeed, observed increases in exotic honeybee visitation under eutrophication may alter native wildflower reproduction and native pollinator visitation. Despite their importance to agriculture (DeGrandi-Hoffman 2003 , Breeze et al. 2011 , Fikadu 2019 ) (but see (Kremen et al. 2002 , Garibaldi et al. 2013 )), evidence remains mixed as to whether honeybees enhance (Hung et al. 2018 , Stanley et al. 2020 ) or decrease native wildflower seed set (Magrach et al. 2017 , Page et al. 2021 , Page and Williams 2023 ). Apparent competition between honeybees and native pollinators for shared flowers can displace native bees (Page and Williams 2023 , Travis et al. in press). Thus, where environmental change increases exotic wildflowers, exotic pollinators may mediate the resulting impacts on native pollinators and plants. Taken together, our work shows that from the perspective of pollinators, the cause of invasion can determine the impact, but, importantly, exotic wildflowers, especially exotic asters, can compensate for reductions in native wildflowers. However, exotic wildflowers increase visits by exotic pollinators, which could lead to declines in native pollinators, presenting a paradox for pollinator conservation in rapidly changing environments. More generally, these findings emphasize the tradeoffs in the positive and negative effects of exotic species on ecological communities (Graves and Shapiro 2003 , Zarnetske et al. 2010 , Biel et al. 2017 ). Ultimately, this paradox raises the question of whether conservation efforts should value species based on their ecological traits and functions rather than provenance (Lundgren et al. 2024 ). Declarations Open Research Statement: Data for this paper have been made available to the public via the Envionmental Data Initiative. Data can be cited as follows: Nelson, R.A., E.W. Seabloom, and E.T. Borer. 2025. Data for "Pollinator Conservation Paradox: Exotic Wildflowers Support Native Pollinators Under Global Changes" by Nelson, Seabloom and Borer 2025, California grasslands, 2023-2024 ver 1. Environmental Data Initiative. https://doi.org/10.6073/pasta/1bda9f7c9f8c60bfca3eb8063dc76b78 (Accessed 2025-05-07). Code is accessible to the public via Zenodo at: 10.5281/zenodo.16322076. Conflict of Interest Statement: The authors have no competing interests to declare. Funding: Funding for this project was supported by the Nature Conservancy’s Oren Pollak Memorial Research Fund for grassland research, the Hopland Scholars Fund, the UC Davis Jastro & Shields graduate research award, the Irene Brown Memorial Fund for women in environmental science, and the Davis Botanical Society student scholarship. Funding was also provided by the National Science Foundation (NSF-DEB-1042132, NSF-DEB-1234162, and NSF-DEB-1831944), and the Institute on the Environment (DG-0001-13), and the Minnesota Supercomputer Institute hosted the project data. Author Contributions: Conceptualization: RAN, ES, ETB. Methodology: RAN, ES, ETB. Software: RAN. Validation: RAN, ES, ETB. Formal analysis: RAN. Investigation: RAN. Resources: RAN, ES, ETB. Data curation: RAN. Writing- Original draft preparation: RAN. Writing- Reviewing and Editing: RAN, ETB, ES. Visualization: RAN, ES, ETB. Supervision: RAN, ES, ETB. Project Administration: RAN, ES, ETB. Funding Acquisition: RAN, ES, ETB. Acknowledgements: We thank I. Wolter, K.T. Lynch, S. Tantuico, N.T. Tat, A. Karp, A. Appelgate, Z. Schneider and R. Li for assistance in the field; I. Slette, P. Wilfahrt, and A. Krause for assistance with data; and C. Koehler and P. Aigner (University of California McLaughlin Reserve), J. Bailey and G. Solberg (Hopland UC Research and Extension Center), and M. Flavell and A. Northup-Warner Sierra Foothills UC Research and Extension Center) for maintaining field sites and facilities. We provided data from this project to J. Cancela for Cancela et al. in prep . T. Zavortink (Bohart Museum of Entomology) identified insect voucher specimens. R. Dirzo, S. Harrison, and N. Williams provided feedback on this project. Two anonymous reviewers provided additional feedback on the manuscript. References Barthell, J. F., Randall, J. M., Thorp, R. W. and Wenner, A. M. 2001. Promotion of seed set in yellow star-thistle by honey bees: evidence of an invasive mutualism. - Ecological Applications 11: 1870–1883. Bartomeus, I., Vilà, M. and Santamaría, L. 2008. Contrasting effects of invasive plants in plant–pollinator networks. - Oecologia 155: 761–770. Bascompte, J. and Jordano, P. 2007. 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(Ciska), Heinen, R., Wang, D., Jiang, M., Jin, H. and Bakker, E. S. Large mammalian herbivores affect arthropod food webs via changes in vegetation characteristics and microclimate. - Journal of Ecology in press. Additional Declarations The authors declare no competing interests. Supplementary Files ParadoxSupplement71725.docx Cite Share Download PDF Status: Posted Version 1 posted You are reading this latest preprint version Research Square lets you share your work early, gain feedback from the community, and start making changes to your manuscript prior to peer review in a journal. As a division of Research Square Company, we’re committed to making research communication faster, fairer, and more useful. We do this by developing innovative software and high quality services for the global research community. Our growing team is made up of researchers and industry professionals working together to solve the most critical problems facing scientific publishing. Also discoverable on Platform About Our Team In Review Editorial Policies Advisory Board Help Center Resources Author Services Accessibility API Access RSS feed Manage Cookie Preferences © Research Square 2026 | ISSN 2693-5015 (online) Privacy Policy Terms of Service Do Not Sell My Personal Information {"props":{"pageProps":{"initialData":{"identity":"rs-7188448","acceptedTermsAndConditions":true,"allowDirectSubmit":true,"archivedVersions":[],"articleType":"Research Article","associatedPublications":[],"authors":[{"id":489268073,"identity":"bb576549-db2e-4f03-9f98-ec41b4d10c47","order_by":0,"name":"Rebecca Nelson","email":"data:image/png;base64,iVBORw0KGgoAAAANSUhEUgAAAZAAAAAyAQMAAABI0h/eAAAABlBMVEX///8AAABVwtN+AAAACXBIWXMAAA7EAAAOxAGVKw4bAAAA4ElEQVRIiWNgGAWjYBACAxiDD0R8QBPEr4UNiBlnkKyFmYcYLebszQ8//GC4I8fGf/jYY9u2usQG9uZtEvi0WPYcM5bsYXhmzCaRlm6c23Y4sYHnWBleLQY3ctgYeBgOJ7ZJ8JhJ57YdSGyQyDHDr+X+GzbGPwyH69v4z5hJW4IcJv+GgJYbPGxAXx9OYGPIMZNmbGMG2sKDX4tlT5qxtIzBYcM2ibQ0yZ5zh43beNKKLfBpMWc//PDjm4rD8vzAEJP4UVYn289+eOMNfFqgzkNisxFWPgpGwSgYBaOAEAAAL/A/ayJozrAAAAAASUVORK5CYII=","orcid":"https://orcid.org/0000-0002-9574-0241","institution":"Utah State University; University of California Davis","correspondingAuthor":true,"prefix":"","firstName":"Rebecca","middleName":"","lastName":"Nelson","suffix":""},{"id":489268228,"identity":"2f734eeb-52d0-40fa-9f3b-8e978e3359b8","order_by":1,"name":"Eric Seabloom","email":"","orcid":"","institution":"University of Minnesota","correspondingAuthor":false,"prefix":"","firstName":"Eric","middleName":"","lastName":"Seabloom","suffix":""},{"id":489268229,"identity":"0a7b270f-cf6e-4e08-b4b4-4dbbd8ce1a0c","order_by":2,"name":"Elizabeth Borer","email":"","orcid":"","institution":"University of Minnesota","correspondingAuthor":false,"prefix":"","firstName":"Elizabeth","middleName":"","lastName":"Borer","suffix":""}],"badges":[],"createdAt":"2025-07-22 15:07:09","currentVersionCode":1,"declarations":{"humanSubjects":false,"vertebrateSubjects":false,"conflictsOfInterestStatement":false,"humanSubjectEthicalGuidelines":false,"humanSubjectConsent":false,"humanSubjectClinicalTrial":false,"humanSubjectCaseReport":false,"vertebrateSubjectEthicalGuidelines":false},"doi":"10.21203/rs.3.rs-7188448/v1","doiUrl":"https://doi.org/10.21203/rs.3.rs-7188448/v1","draftVersion":[],"editorialEvents":[],"editorialNote":"","failedWorkflow":false,"files":[{"id":87357855,"identity":"d7b911bd-824d-4c13-bbe4-d2465b82696e","added_by":"auto","created_at":"2025-07-23 05:31:56","extension":"png","order_by":1,"title":"Figure 1","display":"","copyAsset":false,"role":"figure","size":237281,"visible":true,"origin":"","legend":"\u003cp\u003eThe effects of herbivore exclusion via fencing by fertilization on (A) floral abundance, (B) aster (Asteracae) abundance, (C) native pollinator visitation, pollinator abundance here refers to total number of native pollinator visits and (D) native pollinator visitation per capita (Native Pollinator Visitation/Floral Abundance). Teal refers to exotic wildflower species, while orange refers to native wildflower species. Black intervals show the mean and standard error. Each dot represents a value for a unique combination of plot and year sampled.\u003c/p\u003e","description":"","filename":"floatimage1.png","url":"https://assets-eu.researchsquare.com/files/rs-7188448/v1/32a22d26a4ed32aee9270345.png"},{"id":87357851,"identity":"69b88621-3e5d-4f4d-a977-c408622f6bfb","added_by":"auto","created_at":"2025-07-23 05:31:55","extension":"png","order_by":2,"title":"Figure 2","display":"","copyAsset":false,"role":"figure","size":534862,"visible":true,"origin":"","legend":"\u003cp\u003eFloral visitation networks by wildflower provenance for (A) Hopland, (B) McLaughlin and (C) Sierra Foothills in order of increasing site aridity. Each box on the top represents a plant species, while each box on the bottom represents a floral visitor functional group. The lines between boxes represent a given plant-pollinator interaction. The width of the boxes and lines is proportional to the frequency observed in the network. Boxes shaded in gold represent asters (Asteraceae). Boxes shaded in coral represent European honeybee, the exotic pollinator species in the data set (honeybees). Data used for networks are from plots from all four fencing by fertilization treatments.\u003c/p\u003e","description":"","filename":"floatimage2.png","url":"https://assets-eu.researchsquare.com/files/rs-7188448/v1/123cf8b7f7878d862b686e6e.png"},{"id":87357845,"identity":"ff7921ee-8650-4342-93e5-c087595417ad","added_by":"auto","created_at":"2025-07-23 05:31:55","extension":"png","order_by":3,"title":"Figure 3","display":"","copyAsset":false,"role":"figure","size":100871,"visible":true,"origin":"","legend":"\u003cp\u003eExotic pollinator (western Honeybee, \u003cem\u003eApis mellifera\u003c/em\u003e) visitation by fencing and fertilization treatments and wildflower provenance. Honeybee visits to exotic wildflower species are represented in green, while hoeneybee visits to native wildflower species are represented in orange. Black lines show mean and standard error. Each dot represents a unique combination of plot and year for which honeybees were present.\u003c/p\u003e","description":"","filename":"floatimage3.png","url":"https://assets-eu.researchsquare.com/files/rs-7188448/v1/06a51a8408f7d0e06e34a0bb.png"},{"id":87360601,"identity":"0f9c20a3-1ba9-41bd-b4dd-9a34888f913e","added_by":"auto","created_at":"2025-07-23 05:47:57","extension":"pdf","order_by":0,"title":"","display":"","copyAsset":false,"role":"manuscript-pdf","size":1370979,"visible":true,"origin":"","legend":"","description":"","filename":"manuscript.pdf","url":"https://assets-eu.researchsquare.com/files/rs-7188448/v1/f90548d7-9511-469c-83c5-d80b844c3dfd.pdf"},{"id":87359536,"identity":"23de8de0-7095-4b47-b7db-39bd4f688b85","added_by":"auto","created_at":"2025-07-23 05:39:55","extension":"docx","order_by":1,"title":"","display":"","copyAsset":false,"role":"supplement","size":3116013,"visible":true,"origin":"","legend":"","description":"","filename":"ParadoxSupplement71725.docx","url":"https://assets-eu.researchsquare.com/files/rs-7188448/v1/c5096f16fde856720f4530f1.docx"}],"financialInterests":"The authors declare no competing interests.","formattedTitle":"\u003cp\u003e\u003cstrong\u003ePollinator conservation paradox: exotic wildflowers support native pollinators under global changes\u003c/strong\u003e\u003c/p\u003e","fulltext":[{"header":"Introduction","content":"\u003cp\u003eInvasive wildflowers pose a paradox for conservation. While invasive wildflowers can decrease the diversity and abundance of native wildflower species (Brown et al. \u003cspan citationid=\"CR13\" class=\"CitationRef\"\u003e2002\u003c/span\u003e, Powell et al. \u003cspan citationid=\"CR75\" class=\"CitationRef\"\u003e2011\u003c/span\u003e, \u003cspan citationid=\"CR76\" class=\"CitationRef\"\u003e2013\u003c/span\u003e, Vil\u0026agrave; et al. \u003cspan citationid=\"CR98\" class=\"CitationRef\"\u003e2011\u003c/span\u003e, Goodell and Parker \u003cspan citationid=\"CR38\" class=\"CitationRef\"\u003e2017\u003c/span\u003e), they can benefit native pollinators, including imperiled insect species (Harmon-Threatt and Kremen \u003cspan citationid=\"CR41\" class=\"CitationRef\"\u003e2015\u003c/span\u003e, Gibson et al. \u003cspan citationid=\"CR37\" class=\"CitationRef\"\u003e2019\u003c/span\u003e, Russo et al. \u003cspan citationid=\"CR78\" class=\"CitationRef\"\u003e2019\u003c/span\u003e). Evidence for this paradox remains mixed (Stout and Tiedeken \u003cspan citationid=\"CR87\" class=\"CitationRef\"\u003e2017\u003c/span\u003e, Kaiser-Bunbury et al. \u003cspan citationid=\"CR52\" class=\"CitationRef\"\u003e2017\u003c/span\u003e, Szigeti et al. \u003cspan citationid=\"CR88\" class=\"CitationRef\"\u003e2020\u003c/span\u003e, Gaiarsa and Bascompte \u003cspan citationid=\"CR34\" class=\"CitationRef\"\u003e2022\u003c/span\u003e). Prior efforts to tackle this paradox have begun with wildflower invasion as the focal global change (Carvalheiro et al. \u003cspan citationid=\"CR16\" class=\"CitationRef\"\u003e2008\u003c/span\u003e, Molina-Montenegro et al. \u003cspan citationid=\"CR67\" class=\"CitationRef\"\u003e2008\u003c/span\u003e, Bartomeus et al. \u003cspan citationid=\"CR2\" class=\"CitationRef\"\u003e2008\u003c/span\u003e, Parra-Tabla and Arceo-G\u0026oacute;mez \u003cspan citationid=\"CR73\" class=\"CitationRef\"\u003e2021\u003c/span\u003e, Valdovinos et al. \u003cspan citationid=\"CR93\" class=\"CitationRef\"\u003e2023\u003c/span\u003e). Yet, pollinators are declining under many global changes (Grixti et al. \u003cspan citationid=\"CR40\" class=\"CitationRef\"\u003e2009\u003c/span\u003e, Potts et al. \u003cspan citationid=\"CR74\" class=\"CitationRef\"\u003e2010\u003c/span\u003e, Dirzo et al. \u003cspan citationid=\"CR24\" class=\"CitationRef\"\u003e2014\u003c/span\u003e, Carvalheiro et al. \u003cspan citationid=\"CR18\" class=\"CitationRef\"\u003e2020\u003c/span\u003e, Janousek et al. \u003cspan citationid=\"CR50\" class=\"CitationRef\"\u003e2023\u003c/span\u003e, Edwards et al. \u003cspan citationid=\"CR28\" class=\"CitationRef\"\u003e2025\u003c/span\u003e), and plant invasions that alter the quality of the environment for pollinators can be driven by interacting factors (Bradley et al. \u003cspan citationid=\"CR9\" class=\"CitationRef\"\u003e2010\u003c/span\u003e, Yang et al. \u003cspan citationid=\"CR101\" class=\"CitationRef\"\u003e2022\u003c/span\u003e). This reality raises the question: do individual or interacting anthropogenic changes that favor exotic wildflowers compensate pollinators for losses of native wildflowers?\u003c/p\u003e\u003cp\u003eTwo such anthropogenic changes driving plant invasions and global declines in native wildflowers are eutrophication (Song et al. \u003cspan citationid=\"CR85\" class=\"CitationRef\"\u003e2011\u003c/span\u003e, You et al. \u003cspan citationid=\"CR102\" class=\"CitationRef\"\u003e2017\u003c/span\u003e, Li et al. \u003cspan citationid=\"CR59\" class=\"CitationRef\"\u003e2022b\u003c/span\u003e, MacDougall et al. \u003cspan citationid=\"CR64\" class=\"CitationRef\"\u003e2024\u003c/span\u003e, Nelson et al. \u003cspan citationid=\"CR69\" class=\"CitationRef\"\u003e2025\u003c/span\u003e) and defaunation (Borer et al. 2014b, Br\u0026aring;then et al. \u003cspan citationid=\"CR11\" class=\"CitationRef\"\u003e2021\u003c/span\u003e, Price et al. \u003cspan citationid=\"CR77\" class=\"CitationRef\"\u003e2022\u003c/span\u003e, Borer and Risch \u003cspan citationid=\"CR6\" class=\"CitationRef\"\u003e2024\u003c/span\u003e, Zhu et al. in press), the loss of large wild mammalian herbivores (Dirzo et al. \u003cspan citationid=\"CR24\" class=\"CitationRef\"\u003e2014\u003c/span\u003e). Under both eutrophication and defaunation, losses in wildflower abundance and diversity resulting from increased grass competition (DeMalach et al. \u003cspan citationid=\"CR23\" class=\"CitationRef\"\u003e2017\u003c/span\u003e, You et al. \u003cspan citationid=\"CR102\" class=\"CitationRef\"\u003e2017\u003c/span\u003e, Eskelinen et al. \u003cspan citationid=\"CR29\" class=\"CitationRef\"\u003e2022\u003c/span\u003e) can decrease pollinator visitation (van Klink et al. \u003cspan citationid=\"CR95\" class=\"CitationRef\"\u003e2015\u003c/span\u003e, Filazzola et al. \u003cspan citationid=\"CR32\" class=\"CitationRef\"\u003e2020\u003c/span\u003e, Veen et al. \u003cspan citationid=\"CR97\" class=\"CitationRef\"\u003e2024\u003c/span\u003e) and change pollinator community composition (van Klink et al. \u003cspan citationid=\"CR95\" class=\"CitationRef\"\u003e2015\u003c/span\u003e, David et al. \u003cspan citationid=\"CR21\" class=\"CitationRef\"\u003e2019\u003c/span\u003e, Carvalheiro et al. \u003cspan citationid=\"CR18\" class=\"CitationRef\"\u003e2020\u003c/span\u003e, Filazzola et al. \u003cspan citationid=\"CR32\" class=\"CitationRef\"\u003e2020\u003c/span\u003e, Wang et al. \u003cspan citationid=\"CR100\" class=\"CitationRef\"\u003e2022\u003c/span\u003e). Unlike loss of large herbivores, eutrophication can also alter floral reward chemistry, with mixed effects on nectar and pollen attractiveness and quality (Ceulemans et al. \u003cspan citationid=\"CR19\" class=\"CitationRef\"\u003e2017\u003c/span\u003e, Majetic et al. \u003cspan citationid=\"CR66\" class=\"CitationRef\"\u003e2017\u003c/span\u003e, Russo et al. \u003cspan citationid=\"CR79\" class=\"CitationRef\"\u003e2023\u003c/span\u003e). Moreover, large mammalian herbivores can rescue native wildflowers and their pollinators from declines under eutrophication (Veen et al. \u003cspan citationid=\"CR97\" class=\"CitationRef\"\u003e2024\u003c/span\u003e). These past findings have largely focused on uninvaded ecosystems (van Klink et al. \u003cspan citationid=\"CR95\" class=\"CitationRef\"\u003e2015\u003c/span\u003e, Carvalheiro et al. \u003cspan citationid=\"CR18\" class=\"CitationRef\"\u003e2020\u003c/span\u003e, Wang et al. \u003cspan citationid=\"CR100\" class=\"CitationRef\"\u003e2022\u003c/span\u003e, Veen et al. \u003cspan citationid=\"CR97\" class=\"CitationRef\"\u003e2024\u003c/span\u003e). Under eutrophication and defaunation, invasive wildflowers can increase or maintain their abundances (Hayes and Holl \u003cspan citationid=\"CR45\" class=\"CitationRef\"\u003e2003\u003c/span\u003e, Liu and Stiling \u003cspan citationid=\"CR60\" class=\"CitationRef\"\u003e2006\u003c/span\u003e, Seabloom et al. \u003cspan citationid=\"CR82\" class=\"CitationRef\"\u003e2015\u003c/span\u003e, Lannes et al. \u003cspan citationid=\"CR54\" class=\"CitationRef\"\u003e2016\u003c/span\u003e). Herbivory also can counteract increases in invasive plants under eutrophication (Gianoli and Salgado-Luarte \u003cspan citationid=\"CR36\" class=\"CitationRef\"\u003e2017\u003c/span\u003e, Li et al. \u003cspan citationid=\"CR58\" class=\"CitationRef\"\u003e2022a\u003c/span\u003e). Thus, eutrophication and defaunation can lead to invasion, but their interaction can cause differing outcomes for pollinators, underscoring the need to consider the processes that lead to plant invasion.\u003c/p\u003e\u003cp\u003eParadoxically, invasive wildflowers may support pollinators in these rapidly changing landscapes. Pollinators adjust their foraging to track changes in floral abundance (Sih and Baltus \u003cspan citationid=\"CR83\" class=\"CitationRef\"\u003e1987\u003c/span\u003e) and floral rewards (Ito et al. \u003cspan citationid=\"CR49\" class=\"CitationRef\"\u003e2021\u003c/span\u003e), which may be altered in different ways by nutrient supply (Burkle and Irwin \u003cspan citationid=\"CR15\" class=\"CitationRef\"\u003e2010\u003c/span\u003e, Vaudo et al. \u003cspan citationid=\"CR96\" class=\"CitationRef\"\u003e2022\u003c/span\u003e) or herbivores (Moreira et al. \u003cspan citationid=\"CR68\" class=\"CitationRef\"\u003e2019\u003c/span\u003e). If exotic wildflowers become more abundant than native wildflowers under eutrophication or defaunation, native pollinators may visit a greater proportion of exotic wildflowers relative to native ones. Increases in exotic wildflowers under eutrophication and herbivore exclusion may maintain pollinator composition. Indeed, some exotic wildflower species can functionally replace native species as core network hubs (Russo et al. \u003cspan citationid=\"CR78\" class=\"CitationRef\"\u003e2019\u003c/span\u003e, Parra-Tabla and Arceo-G\u0026oacute;mez \u003cspan citationid=\"CR73\" class=\"CitationRef\"\u003e2021\u003c/span\u003e) that stabilize the network\u0026rsquo;s structure by supporting both generalist and specialist pollinators (Valdovinos et al. \u003cspan citationid=\"CR91\" class=\"CitationRef\"\u003e2009\u003c/span\u003e). However, exotic wildflowers can decrease specialist pollinators, decreasing pollinator diversity (Kaiser-Bunbury et al. \u003cspan citationid=\"CR52\" class=\"CitationRef\"\u003e2017\u003c/span\u003e). Alternatively, eutrophication and defaunation could interactively determine exotic vs native floral abundance (Burkepile et al. \u003cspan citationid=\"CR14\" class=\"CitationRef\"\u003e2017\u003c/span\u003e, Li et al. \u003cspan citationid=\"CR58\" class=\"CitationRef\"\u003e2022a\u003c/span\u003e) and consequently pollinator visitation.\u003c/p\u003e\u003cp\u003eIncreases in exotic wildflowers under eutrophication and defaunation could promote invasional meltdowns where exotic wildflowers co-invade with their exotic pollinators (Simberloff and Von Holle \u003cspan citationid=\"CR84\" class=\"CitationRef\"\u003e1999\u003c/span\u003e, Barthell et al. \u003cspan citationid=\"CR1\" class=\"CitationRef\"\u003e2001\u003c/span\u003e), or alternatively have contrasting effects on exotic pollinators. Exotic and native plants can indirectly affect each other\u0026rsquo;s fitness both positively and negatively through competition for shared pollinators (Molina-Montenegro et al. \u003cspan citationid=\"CR67\" class=\"CitationRef\"\u003e2008\u003c/span\u003e, Etter et al. \u003cspan citationid=\"CR30\" class=\"CitationRef\"\u003e2022\u003c/span\u003e, Dritz et al. \u003cspan citationid=\"CR27\" class=\"CitationRef\"\u003e2023\u003c/span\u003e), a well-documented form of apparent competition (Holt \u003cspan citationid=\"CR47\" class=\"CitationRef\"\u003e1977\u003c/span\u003e, Goodell and Parker \u003cspan citationid=\"CR38\" class=\"CitationRef\"\u003e2017\u003c/span\u003e). Moreover, exotic pollinators, particularly the western honeybee (\u003cem\u003eApis mellifera\u003c/em\u003e) in its introduced range, can displace native pollinators on shared floral resources, decreasing native wildflower seed set (Goodell and Parker \u003cspan citationid=\"CR38\" class=\"CitationRef\"\u003e2017\u003c/span\u003e, Page and Williams \u003cspan citationid=\"CR71\" class=\"CitationRef\"\u003e2023\u003c/span\u003e).\u003c/p\u003e\u003cp\u003eThe California Floristic Province provides an opportunity to examine this paradox. Although it is a hotspot for plant-pollinator biodiversity and specialization (Thorp and Leong \u003cspan citationid=\"CR89\" class=\"CitationRef\"\u003e1998\u003c/span\u003e, Frankie et al. \u003cspan citationid=\"CR33\" class=\"CitationRef\"\u003e2009\u003c/span\u003e, Harrison \u003cspan citationid=\"CR43\" class=\"CitationRef\"\u003e2013\u003c/span\u003e, Harrison et al. \u003cspan citationid=\"CR44\" class=\"CitationRef\"\u003e2018\u003c/span\u003e), the California Floristic Province experiences high rates of plant invasions especially invasive asters (Levine and D\u0026rsquo;Antonio \u003cspan citationid=\"CR57\" class=\"CitationRef\"\u003e1999\u003c/span\u003e, Harrison \u003cspan citationid=\"CR42\" class=\"CitationRef\"\u003e1999\u003c/span\u003e, Seabloom et al. \u003cspan citationid=\"CR80\" class=\"CitationRef\"\u003e2006\u003c/span\u003e), defaunation (Hayes and Holl \u003cspan citationid=\"CR45\" class=\"CitationRef\"\u003e2003\u003c/span\u003e), eutrophication (Valliere et al. \u003cspan citationid=\"CR94\" class=\"CitationRef\"\u003e2020\u003c/span\u003e), and introduced honeybees (Barthell et al. \u003cspan citationid=\"CR1\" class=\"CitationRef\"\u003e2001\u003c/span\u003e, Page and Williams \u003cspan citationid=\"CR71\" class=\"CitationRef\"\u003e2023\u003c/span\u003e, Travis et al. in press).\u003c/p\u003e\u003cp\u003eTo test the paradox that exotic wildflowers may support native pollinators under high nutrient conditions and altered herbivory, we replicated herbivore exclusion and nutrient addition experiments across three central California grasslands. We hypothesized that: increases in exotic wildflowers, particularly asters, and decreases in native wildflowers under eutrophication and herbivore exclusion will (1) increase native pollinator visitation and richness on exotic wildflowers relative to native wildfowers; (2) maintain native pollinator community composition via exotic asters acting as core network hubs for both specialist and generalist pollinators, but (3) increase exotic pollinator visitation via invasional meltdowns.\u003c/p\u003e"},{"header":"Materials and Methods","content":"\u003cp\u003e\u003cb\u003eStudy Sites\u003c/b\u003e: This research occurred at three grassland sites in California that spanned a 150 km longitudinal gradient with cool wet winters and hot, dry summers: the Sierra Foothill Research \u0026amp; Extension Center (39.24 N, -121.28 W, Dec 2023 precipitation: 121.77 cm), the University of California McLaughlin Reserve (38.86 N, -122.41 W, Dec 2023 precipitation: 164.07 cm), and the Hopland Research \u0026amp; Extension Center (39.01 N, -123.06 W, Dec 2023 precipitation: 196.27 cm). The exotic species pool contained primarily annual grasses and wildflowers from the Mediterranean region of Europe (Heady \u003cspan citationid=\"CR46\" class=\"CitationRef\"\u003e1977\u003c/span\u003e, Seabloom et al. \u003cspan citationid=\"CR80\" class=\"CitationRef\"\u003e2006\u003c/span\u003e).\u003c/p\u003e\u003cp\u003e\u003cb\u003eNutrient Network Experiments\u003c/b\u003e: All three sites contained replicates of a randomized block, factorial experiment that use identical methods to manipulate fertilization and herbivore presence as part of the globally replicated Nutrient Network (Borer et al. \u003cspan citationid=\"CR7\" class=\"CitationRef\"\u003e2014a\u003c/span\u003e, Seabloom et al. \u003cspan citationid=\"CR82\" class=\"CitationRef\"\u003e2015\u003c/span\u003e). Each site contained at least 3 randomized blocks of 5 m x 5m plots assigned to a factorial combination of the following treatments: Fencing (Control and Fenced) and Fertilization (Control and Fertilized with macro and micronutrients, NPK\u0026micro;). This design created four distinct fencing and fertilization treatments: Control (unfenced and unfertilized), Fence (fenced and unfertilized), NPK\u0026micro;, (unfenced and fertilized with NPK\u0026micro;), and NPK\u0026micro;,\u0026thinsp;+\u0026thinsp;Fence (fenced and fertilized with NPK\u0026micro;). The fencing treatment was designed to exclude large, nonclimbing mammals. The fences were 230 cm tall with the lower 90 cm composed of 1 cm wire mesh, which included a 30 cm wide, outward-facing flange stapled to the ground to prevent digging. The upper portion of the fence is composed of strands of barbless wire. The fertilization treatment is composed of a combination of macro- and micro-nutrients (10 g N m\u003csup\u003e\u0026minus;\u0026thinsp;2\u003c/sup\u003e yr\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e as timed-release urea; 10 g P m\u003csup\u003e\u0026minus;\u0026thinsp;2\u003c/sup\u003e yr\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e as triple-super phosphate 10 g K m\u003csup\u003e\u0026minus;\u0026thinsp;2\u003c/sup\u003e yr\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e as potassium sulfate; and 100 g m\u003csup\u003e\u0026minus;\u0026thinsp;2\u003c/sup\u003e yr\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e of a micronutrient mix 6% Ca, 3% Mg, 12% S, 0.1% B, 1% Cu, 17% Fe, 2.5% Mn, 0.05% Mo, and 1% Zn). The macronutrients were applied annually, and the micronutrients were applied a single time at the start of the experiment. The primary large herbivores excluded at the sites include cattle and deer. Treatments started in 2008. Hopland and McLaughlin had three experimental blocks with these treatments, while Sierra Foothill had five experimental blocks. Our experimental design captured changes in the behavioral visitation of pollinators within a given site rather than population-level changes in pollinators.\u003c/p\u003e\u003cp\u003e\u003cb\u003ePlant-Pollinator Observations\u003c/b\u003e:\u003c/p\u003e\u003cp\u003eWe surveyed each plot for floral visitors (20-minute observations/plot) and wildflower abundance (# floral units/wildflower species/plot of non-gramminoid, herbaceous angiosperms) (Veen et al. \u003cspan citationid=\"CR97\" class=\"CitationRef\"\u003e2024\u003c/span\u003e) at least three times over the course of the growing season (April-July) for two years, a minimum of 120 observer hours per plot per year in 2023 and 2024 across the full flowering season (see Supplemental Methods). Floral visitation surveys occurred only on calm sunny or partly sunny days and if flowering wildflowers were present. A floral visit occurred if an animal contacted the reproductive parts of the flower. When multiple floral visitors were present at the same time, we were able to simultaneously record visitation. We recorded the morphospecies identity of the floral visitor and the plant species visited. Floral abundance was estimated for each plot within ten days of the date a given pollinator survey occurred. We estimated floral abundance by counting the total number of floral units, defined as a 1 cm x 1 cm part of the flower, per flowering plant species in a given plot (Veen et al. \u003cspan citationid=\"CR97\" class=\"CitationRef\"\u003e2024\u003c/span\u003e).\u003c/p\u003e\u003cp\u003eVoucher specimens of each insect morphospecies were collected, pinned, and identified to species with the help of the Bohart Museum of Entomology (Davis, California, USA). We used these species data to classify pollinators into distinct morphospecies (see Supplemental Materials). When collecting specimens during our surveys, we paused the timer and resumed timing once collection was complete. For sensitive taxa (Papilionidae and \u003cem\u003eBombus\u003c/em\u003e spp.), we took photographs which were used to identify species. We ceased recording visits and collected vouchers once an individual left a plot to minimize the impact of collection.\u003c/p\u003e\u003cdiv id=\"Sec3\" class=\"Section2\"\u003e\u003ch2\u003eData Analysis\u003c/h2\u003e\u003cp\u003eAll data analyses were performed in R (v. 4.4.1). Models were checked for normality of residuals and overdispersion where appropriate.We performed rarefaction analyses for pollinator visitation data with the \u0026lsquo;rarecurve\u0026rsquo; function in the \u0026lsquo;vegan\u0026rsquo; package (Figure S4) (Dixon \u003cspan citationid=\"CR25\" class=\"CitationRef\"\u003e2003\u003c/span\u003e).\u003c/p\u003e\u003cp\u003eTo test our first hypothesis, we used a generalized linear mixed effects models to test whether fencing and fertilization treatments affected exotic wildflower floral abundance, native wildflower floral abundance, native wildflower plant species richness, pollinator morphospecies richness on native vs exotic wildflowers, raw native pollinator visitation to native wildflowers (total pollinator visits), native pollinator visitation to native wildflowers per floral unit (per capita) (total pollinator visits/# floral units), native pollinator visitation to exotic wildflowers, and native pollinator visitation to exotic wildflowers per floral unit (per capita). To calculate per capita visitation values, we divided pollinator visitation by plot-level floral abundance of the plant species visited. For abundance data, we used negative binomial models due to overdispersion, while for richness data we used Poisson distributions. For per capita response variables, we used Gaussian distributions. We included a main effect of experimental treatments and a random effect of experimental block nested within site, and dropped the nested effect of block in cases where we had convergence issues (see Supplement). We used the \u003cem\u003eemmeans\u003c/em\u003e package to compute estimated marginal least-square means for pairwise combinations of treatments (Lenth et al. \u003cspan citationid=\"CR56\" class=\"CitationRef\"\u003e2024\u003c/span\u003e).\u003c/p\u003e\u003cp\u003eWe used generalized linear models to examine whether native pollinator visitation (# visits by native pollinator morphospecies) and native pollinator morphospecies richness were explained by native plant species richness, native floral abundance, and exotic floral abundance using negative binomial mixed effects models with a random effect tof site (Figure S2).\u003c/p\u003e\u003cp\u003eTo test our second hypothesis regarding pollinator composition, we compared how native pollinator community composition differed among treatment, plant species, plant family and local provenance using the \u003cem\u003evegan\u003c/em\u003e package in R to run PERMANOVAs with Bray-Curtis dissimilarity indices (Dixon \u003cspan citationid=\"CR25\" class=\"CitationRef\"\u003e2003\u003c/span\u003e). We grouped native pollinator data by pollinator functional group (long-tongued bees, short-tongued bees, beetles, butterflies and moths, kleptoparasitic bees, and flies) using literature-based taxonomic and morphological information (LeBuhn \u003cspan citationid=\"CR55\" class=\"CitationRef\"\u003e2013\u003c/span\u003e), and then compared how treatment affected native pollinator functional group abundance using differential multivariate generalized linear models with the \u003cem\u003emvabund\u003c/em\u003e package (Wang et al. \u003cspan citationid=\"CR99\" class=\"CitationRef\"\u003e2012\u003c/span\u003e).\u003c/p\u003e\u003cp\u003eTo further test our the effects of plant provenance on network structure, for each site, we calculated bipartite networks of plant-pollinator interactions and measured the individual contribution of each wildflower species to network nestedness using the \u003cem\u003ebipartite\u003c/em\u003e package in R (Dormann et al. \u003cspan citationid=\"CR26\" class=\"CitationRef\"\u003e2009\u003c/span\u003e). Nestedness is the degree to which generalist plants (and pollinators) interact with both generalist and specialist pollinators (and plants) (Bascompte and Jordano \u003cspan citationid=\"CR3\" class=\"CitationRef\"\u003e2007\u003c/span\u003e, Bastolla et al. \u003cspan citationid=\"CR4\" class=\"CitationRef\"\u003e2009\u003c/span\u003e). Nestedness is a core property of mutualistic networks associated with structural stability (Bastolla et al. \u003cspan citationid=\"CR4\" class=\"CitationRef\"\u003e2009\u003c/span\u003e, Valdovinos et al. \u003cspan citationid=\"CR92\" class=\"CitationRef\"\u003e2016\u003c/span\u003e). By measuring the individual contribution of each plant species to network nestedness, we can determine whether or not invasive vs native plant species act as core hubs in the network attract diverse pollinators.\u003c/p\u003e\u003cp\u003eTo test our third hypothesis regarding invasional meltdowns, we examined how visitation of the two exotic pollinator species we observed responded to eutrophication and herbivore exclusion. We fit a negative binomial generalized linear effects model to test for the effects of treatments on visitation by the exotic western honeybee \u003cem\u003eApis mellifera\u003c/em\u003e wildflowers. We also counted visits by the exotic leafcutter bee \u003cem\u003eMegachile apicalis\u003c/em\u003e by treatment and plant species. Becase we only recorded 14 total visits from \u003cem\u003eM. apicalis\u003c/em\u003e, we did not have a sufficient sample size to fit a similar model for \u003cem\u003eM. apicalis.\u003c/em\u003e We used multivariate abundance analyses with the \u003cem\u003emvabund\u003c/em\u003e package to examine whether exotic pollinator visitation differed by wildflower family.\u003c/p\u003e\u003c/div\u003e"},{"header":"Results","content":"\u003cp\u003eIn total, we recorded 517 individual floral visitors to 31 wildflower species and 62 pollinator morphospecies across all sites and years and made 820 observations of floral abundances to 72 wildflower species across all sites and years (Tables S1-S2). At McLaughlin, we observed 31 pollinator morphospecies visiting 11 wildflower species, out of 29 total flowering wildflower species. At Sierra Foothills, we observed 32 pollinator morphospecies visiting 17 out of 42 total flowering wildflower species. At Hopland, we observed 28 pollinator morphospecies visiting 10 wildflower species out of 39 total flowering wildflower species. Common floral visitors included bee, syrphid fly, bombyliid fly and beetle species. We had 4 species of exotic asters that were visited by pollinators in our data: \u003cem\u003eCarduus pycnocephalus\u003c/em\u003e, \u003cem\u003eSenecio vulgaris\u003c/em\u003e, \u003cem\u003eSonchus asper\u003c/em\u003e, and \u003cem\u003eCentaurea solstitialis\u003c/em\u003e. We observed two exotic pollinator species: western honeybee \u003cem\u003eApis mellifera\u003c/em\u003e and the leafcutter bee \u003cem\u003eMegachile apicalis\u003c/em\u003e. We observed one imperiled pollinator species Crotch\u0026rsquo;s bumble bee (\u003cem\u003eBombus crotchii\u003c/em\u003e), visiting both native wildflowers and exotic asters.\u003c/p\u003e\u003cp\u003e\u003cb\u003eWildflower Communities\u003c/b\u003e\u003c/p\u003e\u003cp\u003eFertilization decreased native wildflower floral abundance and diversity but increased exotic aster floral abundance, as measured by the number of floral units (Fig.\u0026nbsp;\u003cspan refid=\"Fig1\" class=\"InternalRef\"\u003e1\u003c/span\u003e, Figure S1). Native wildflower species richness significantly declined under the combined effects of fertilization and herbivore exclusion relative to the unfenced controls (z\u0026thinsp;=\u0026thinsp;3.18, p\u0026thinsp;=\u0026thinsp;0.008, Table S3). When herbivores were excluded, native wildflower richness and native wildflower abundance significantly declined under fertilization (richness: z\u0026thinsp;=\u0026thinsp;2.99, p\u0026thinsp;=\u0026thinsp;0.015; abundance: z\u0026thinsp;=\u0026thinsp;3.33, p\u0026thinsp;=\u0026thinsp;0.005, Table S3). When herbivores were present, however, fertilization did not affect native wildflower floral abundance (Table S3) nor native wildflower species richness (Table S3). While the floral abundance of exotic wildflowers for all floral families combined together did not differ by treatment (Table S3), floral abundance of asters (Asteraceae), the most common family of exotic wildflower, increased under fertilization (z=-3.31, p\u0026thinsp;=\u0026thinsp;0.005, Table S3) as well as its combined effects with herbivore exclusion (z=-4.23, p\u0026thinsp;=\u0026thinsp;0.0001, Table S3) (Fig.\u0026nbsp;\u003cspan refid=\"Fig1\" class=\"InternalRef\"\u003e1\u003c/span\u003eB). The increase in exotic asters under fertilized conditions occurred both when herbivores were excluded (z=-3.66, p\u0026thinsp;=\u0026thinsp;0.001, Table S3) and present (z=-3.31, p\u0026thinsp;=\u0026thinsp;0.005, Table S3).\u003c/p\u003e\u003cp\u003e\u003c/p\u003e\u003cp\u003e\u003cb\u003eNative Pollinator Visitation and Richness\u003c/b\u003e\u003c/p\u003e\u003cp\u003eIn support of our first hypothesis, fertilization increased native pollinator visitation to exotic asters and produced shifts in native pollinators from native to exotic flowers under combined fencing and fertilization treatments (Fig.\u0026nbsp;\u003cspan refid=\"Fig1\" class=\"InternalRef\"\u003e1\u003c/span\u003e, Table S4). Combined herbivore exclusion and fertilization treatments, however, did not affect native pollinator visitation per capita (Table S4). Native pollinator visitation to exotic asters increased under fertilization when herbivores were excluded (z=-3.08, p\u0026thinsp;=\u0026thinsp;0.011, Table S4) but not when herbivores were present (z= -2.38, p\u0026thinsp;=\u0026thinsp;0.082, Table S4). Native pollinator visitation of exotic wildflowers did not differ by treatment when all exotic wildflower families were combined (Table S4). The joint effects of herbivore exclusion via fencing and fertilization increased native pollinator visitation to exotic asters (z=-2.69, p\u0026thinsp;=\u0026thinsp;0.036, Table S4). This interaction between fencing and fertilization and provenance was not significant for pollinator visitation per capita (Table S4). Contrary to our first hypothesis, native pollinator morphospecies richness did not differ by treatment (Table S4).\u003c/p\u003e\u003cp\u003eIn further support of our first hypothesis, exotic wildflower abundance was positively correlated with native pollinator visitation to exotic wildflowers (z\u0026thinsp;=\u0026thinsp;3.88, p\u0026thinsp;\u0026lt;\u0026thinsp;0.001, Table S5) and native wildflower abundance was positively correlated with native pollinator visitation to native wildflowers (z\u0026thinsp;=\u0026thinsp;4.42, p\u0026thinsp;\u0026lt;\u0026thinsp;.0001) (Figure S3, Table S5).\u003c/p\u003e\u003cp\u003e\u003cb\u003eNative Pollinator Community Composition\u003c/b\u003e\u003c/p\u003e\u003cp\u003eIn support of our second hypothesis, eutrophication and herbivore exclusion did not shift pollinator composition and function, despite variation in native pollinator composition by wildflower provenance. Fertilization and herbivore exclusion did not directly alter native pollinator composition (PERMANOVA, F\u0026thinsp;=\u0026thinsp;1.0329, p\u0026thinsp;=\u0026thinsp;0.431) and functional group abundance (e.g. long-tongued bees, short-tongued bees, flies, etc) (mvabund, deviance\u0026thinsp;=\u0026thinsp;37.45, p\u0026thinsp;=\u0026thinsp;0.063), but pollinator composition did vary with wildflower species (PERMANOVA, F\u0026thinsp;=\u0026thinsp;2.1585, p\u0026thinsp;=\u0026thinsp;0.001), wildflower provenance (PERMANOVA, F\u0026thinsp;=\u0026thinsp;2.0865, p\u0026thinsp;=\u0026thinsp;0.004), and wildflower taxonomic family (PERMANOVA, F\u0026thinsp;=\u0026thinsp;1.9917, p\u0026thinsp;=\u0026thinsp;0.001) (Figure S5).\u003c/p\u003e\u003cp\u003eIn further support of our second hypothesis, tall, bright, nectar-producing, abundant exotic asters (e.g., \u003cem\u003eCarduus\u003c/em\u003e and \u003cem\u003eCentaurea\u003c/em\u003e species) acted as core plant-pollinator network hubs that attracted both generalist and specialist polliantors and thus contributed strongly to the network\u0026rsquo;s nested structure (Fig.\u0026nbsp;\u003cspan refid=\"Fig2\" class=\"InternalRef\"\u003e2\u003c/span\u003e). Exotic asters were visited by all the main functional groups of native pollinators: butterflies and moths, long-tongued bees, short-tongued bees, beetles, flies, and wasps. At our wettest site, Hopland, the native wildflower western blue-eyed grass \u003cem\u003eSisyrinchium bellum\u003c/em\u003e (Iridaceae) most strongly contributed to the nested structure of the plant-pollinator network (nestedness contribution (nc)\u0026thinsp;=\u0026thinsp;2.66), followed by exotic Italian thistle \u003cem\u003eCarduus pycnocephalus\u003c/em\u003e (Asteraceae) (nc\u0026thinsp;=\u0026thinsp;1.50). At our intermediate rainfall site McLaughlin, the highly invasive wildflower yellow star-thistle \u003cem\u003eCentaurea solstitialis\u003c/em\u003e (Asteraceae) contributed most strongly to the nested structure of the network (nc\u0026thinsp;=\u0026thinsp;2.76) followed by the exotic \u003cem\u003eGeranium dissectum\u003c/em\u003e (Geraniaceae) (nc\u0026thinsp;=\u0026thinsp;1.06). At our driest site Sierra Foothills, the exotic thistle \u003cem\u003eC. pycnocephalus\u003c/em\u003e (Asteraceae) (nc\u0026thinsp;=\u0026thinsp;3.47) had the strongest individual contribution to nestedness followed by the native aster mountain dandelion \u003cem\u003eAgoseris heterophylla\u003c/em\u003e (Asteraceae) (nc\u0026thinsp;=\u0026thinsp;2.34).\u003c/p\u003e\u003cp\u003e\u003c/p\u003e\u003cp\u003e\u003cb\u003eExotic Pollinator Visitation\u003c/b\u003e\u003c/p\u003e\u003cp\u003eIn support of our third hypothesis, exotic pollinators (western honeybee \u003cem\u003eApis mellifera\u003c/em\u003e and leafcutter bee \u003cem\u003eMegachile apicalis\u003c/em\u003e) were strongly associated with exotic asters. Within the fencing by fertilization plots, native wildflowers only received two honeybee (\u003cem\u003eA. mellifera\u003c/em\u003e) visits from a single individual bee, while exotic wildflowers received 215 visits from honeybees (Fig.\u0026nbsp;\u003cspan refid=\"Fig3\" class=\"InternalRef\"\u003e3\u003c/span\u003e). In the absence of large herbivores, eutrophication increased honeybee visitation (z=-2.59, p\u0026thinsp;=\u0026thinsp;0.047, Table S6) but not when herbivores were present (z=-0.02, p\u0026thinsp;=\u0026thinsp;1.0000, Table S6) (Fig.\u0026nbsp;\u003cspan refid=\"Fig3\" class=\"InternalRef\"\u003e3\u003c/span\u003e). Honeybees showed strong preference for foraging on exotic Asteraceae species (mvabund, deviance\u0026thinsp;=\u0026thinsp;36.415, p\u0026thinsp;=\u0026thinsp;0.013). Likewise, the specialist exotic leafcutter bee (\u003cem\u003eM. apicalis\u003c/em\u003e) foraged exclusively on the exotic aster \u003cem\u003eCentaurea solstitialis\u003c/em\u003e. Leaf cutter bees made one visit to \u003cem\u003eC. solstitialis\u003c/em\u003e under ambient conditions, 11 visits to \u003cem\u003eC. solstitialis\u003c/em\u003e under herbivore exclusion, 1 visit to \u003cem\u003eC. solstitialis\u003c/em\u003e under eutrophication, and 1 visit to \u003cem\u003eC. solstitialis\u003c/em\u003e under combined herbivore exclusion and eutrophication.\u003c/p\u003e\u003cp\u003e\u003c/p\u003e"},{"header":"Discussion","content":"\u003cp\u003eWe tested the paradox that exotic wildflowers may compensate pollinators for declines in native wildflowers under eutrophication, herbivore exclusion and their interaction. Consistent with this paradox, we found that native pollinator abundance on exotic asters increased in eutrophied plots and that exotic asters supported diverse pollinator functional groups, contributing strongly to the nested structure of local plant-pollinator networks by attracting both generalist and specialist pollinators. The abundance of exotic bees increased with exotic wildflowers, suggesting the potential for invasional meltdowns as a cascading response to ongoing environmental change.\u003c/p\u003e\u003cp\u003eEutrophication and herbivore exclusion decreased native wildflower diversity and abundance, consistent with prior findings (Song et al. \u003cspan citationid=\"CR85\" class=\"CitationRef\"\u003e2011\u003c/span\u003e, Borer et al. 2014b, You et al. \u003cspan citationid=\"CR102\" class=\"CitationRef\"\u003e2017\u003c/span\u003e, Johnson et al. \u003cspan citationid=\"CR51\" class=\"CitationRef\"\u003e2020\u003c/span\u003e, Valliere et al. \u003cspan citationid=\"CR94\" class=\"CitationRef\"\u003e2020\u003c/span\u003e, Nelson et al. \u003cspan citationid=\"CR69\" class=\"CitationRef\"\u003e2025\u003c/span\u003e). Exotic wildflowers, especially asters, however, were robust to the presence of large mammalian herbivores but increased under fertlization, as has been shown in past studies (Liu and Stiling \u003cspan citationid=\"CR60\" class=\"CitationRef\"\u003e2006\u003c/span\u003e, Seabloom et al. \u003cspan citationid=\"CR81\" class=\"CitationRef\"\u003e2009\u003c/span\u003e). Native wildflowers increased under herbivore exclusion in the absence of eutrophication in contrast to prior findings (Br\u0026aring;then et al. \u003cspan citationid=\"CR11\" class=\"CitationRef\"\u003e2021\u003c/span\u003e). This contrast suggests that fertilization more so than defaunation drives wildflower invaisons. Indeed, the traits that make exotic wildflowers effective invaders \u0026ndash; fast-growing, tall stature, low palatability and high seed set \u0026ndash; may make them effective competitors against grasses in light-limited, fertilized landscapes (MacDougall and Turkington \u003cspan citationid=\"CR63\" class=\"CitationRef\"\u003e2005\u003c/span\u003e, Coleman and Levine \u003cspan citationid=\"CR20\" class=\"CitationRef\"\u003e2007\u003c/span\u003e, Ordonez et al. \u003cspan citationid=\"CR70\" class=\"CitationRef\"\u003e2010\u003c/span\u003e).\u003c/p\u003e\u003cp\u003eConsequently, eutrophication as well as its combined effect with herbivore exclusion increased total native pollinator visitation to exotic wildflowers relative to native wildflowers. Under these global changes, pollinators may forage in response to increased abundances of exotic relative to native wildflowers (Sih and Baltus \u003cspan citationid=\"CR83\" class=\"CitationRef\"\u003e1987\u003c/span\u003e, Lopezaraiza\u0026ndash;Mikel et al. \u003cspan citationid=\"CR61\" class=\"CitationRef\"\u003e2007\u003c/span\u003e), higher nectar production and reward nutritional quality in exotic flowers relative native flowers (Harmon-Threatt and Kremen \u003cspan citationid=\"CR41\" class=\"CitationRef\"\u003e2015\u003c/span\u003e), and greater apparency of taller, exotic flowers (Ordonez et al. \u003cspan citationid=\"CR70\" class=\"CitationRef\"\u003e2010\u003c/span\u003e). The lack of a response of per-plant visitation to eutrophication and herbivore exclusion may suggest the former. Consistent with findings from uninvaded systems (Veen et al. \u003cspan citationid=\"CR97\" class=\"CitationRef\"\u003e2024\u003c/span\u003e), large mammalian herbivory increased pollinator visitation to native wildflowers and decreased pollinator visitation to exotic wildflowers under eutrophication, indicating that large herbivores may rescue plant-pollinator interactions from global changes, perhaps due to consumption of exotic plant species (Price et al. \u003cspan citationid=\"CR77\" class=\"CitationRef\"\u003e2022\u003c/span\u003e, Nelson et al. \u003cspan citationid=\"CR69\" class=\"CitationRef\"\u003e2025\u003c/span\u003e).\u003c/p\u003e\u003cp\u003eContrary to prior findings in relatively uninvaded systems (Carvalheiro et al. \u003cspan citationid=\"CR17\" class=\"CitationRef\"\u003e2014\u003c/span\u003e, David et al. \u003cspan citationid=\"CR21\" class=\"CitationRef\"\u003e2019\u003c/span\u003e, Veen et al. \u003cspan citationid=\"CR97\" class=\"CitationRef\"\u003e2024\u003c/span\u003e), eutrophication and herbivore loss did not alter pollinator composition. Thus, exotic wildflowers may maintain pollinator community composition and function under other global changes. This result, however, may be scale-dependent. Our findings reflect behavioral changes in pollinator foraging, as do past studies (Valdovinos et al. \u003cspan citationid=\"CR93\" class=\"CitationRef\"\u003e2023\u003c/span\u003e, Russo et al. \u003cspan citationid=\"CR79\" class=\"CitationRef\"\u003e2023\u003c/span\u003e, Veen et al. \u003cspan citationid=\"CR97\" class=\"CitationRef\"\u003e2024\u003c/span\u003e). At landscape scales, however, these global changes may drive plant-pollinator interaction turnover (Kaiser-Bunbury et al. \u003cspan citationid=\"CR52\" class=\"CitationRef\"\u003e2017\u003c/span\u003e, Carvalheiro et al. \u003cspan citationid=\"CR18\" class=\"CitationRef\"\u003e2020\u003c/span\u003e, Filazzola et al. \u003cspan citationid=\"CR32\" class=\"CitationRef\"\u003e2020\u003c/span\u003e, Parra-Tabla and Arceo-G\u0026oacute;mez \u003cspan citationid=\"CR73\" class=\"CitationRef\"\u003e2021\u003c/span\u003e). Scale may explain why we do not find differences in pollinator richness under global changes in contrast to landscape-level declines in richness (van Klink et al. \u003cspan citationid=\"CR95\" class=\"CitationRef\"\u003e2015\u003c/span\u003e, Kaiser-Bunbury et al. \u003cspan citationid=\"CR52\" class=\"CitationRef\"\u003e2017\u003c/span\u003e, Russo et al. \u003cspan citationid=\"CR78\" class=\"CitationRef\"\u003e2019\u003c/span\u003e, Filazzola et al. \u003cspan citationid=\"CR32\" class=\"CitationRef\"\u003e2020\u003c/span\u003e, Wang et al. \u003cspan citationid=\"CR100\" class=\"CitationRef\"\u003e2022\u003c/span\u003e) (but see (Veen et al. \u003cspan citationid=\"CR97\" class=\"CitationRef\"\u003e2024\u003c/span\u003e)).\u003c/p\u003e\u003cp\u003eConsistent with prior studies (Russo et al. \u003cspan citationid=\"CR78\" class=\"CitationRef\"\u003e2019\u003c/span\u003e, Parra-Tabla and Arceo-G\u0026oacute;mez \u003cspan citationid=\"CR73\" class=\"CitationRef\"\u003e2021\u003c/span\u003e), nectar-rich, exotic wildflowers, particularly asters, acted as core network hubs for pollinators by attracting both generalist and specialist pollinators. Under eutrophication, pollinators may visit a greater proportion of exotic wildflowers, decreasing interacton diversity. Exotic asters increased exotic pollinator visitation, suggesting potential invasional meltdowns (Simberloff and Von Holle \u003cspan citationid=\"CR84\" class=\"CitationRef\"\u003e1999\u003c/span\u003e, Braga et al. \u003cspan citationid=\"CR10\" class=\"CitationRef\"\u003e2018\u003c/span\u003e). This finding is consistent with prior work demonstrating that exotic asters and exotic pollinators can co-invade as mutualists (Barthell et al. \u003cspan citationid=\"CR1\" class=\"CitationRef\"\u003e2001\u003c/span\u003e) but contrasts with theoretical predictions (Dritz et al. \u003cspan citationid=\"CR27\" class=\"CitationRef\"\u003e2023\u003c/span\u003e). Our results suggest that eutrophication more so than herbivore exclusion may have cascading effects on plant-pollinator interactions through increasing invasional meltdowns.\u003c/p\u003e\u003cp\u003eIndeed, observed increases in exotic honeybee visitation under eutrophication may alter native wildflower reproduction and native pollinator visitation. Despite their importance to agriculture (DeGrandi-Hoffman \u003cspan citationid=\"CR22\" class=\"CitationRef\"\u003e2003\u003c/span\u003e, Breeze et al. \u003cspan citationid=\"CR12\" class=\"CitationRef\"\u003e2011\u003c/span\u003e, Fikadu \u003cspan citationid=\"CR31\" class=\"CitationRef\"\u003e2019\u003c/span\u003e) (but see (Kremen et al. \u003cspan citationid=\"CR53\" class=\"CitationRef\"\u003e2002\u003c/span\u003e, Garibaldi et al. \u003cspan citationid=\"CR35\" class=\"CitationRef\"\u003e2013\u003c/span\u003e)), evidence remains mixed as to whether honeybees enhance (Hung et al. \u003cspan citationid=\"CR48\" class=\"CitationRef\"\u003e2018\u003c/span\u003e, Stanley et al. \u003cspan citationid=\"CR86\" class=\"CitationRef\"\u003e2020\u003c/span\u003e) or decrease native wildflower seed set (Magrach et al. \u003cspan citationid=\"CR65\" class=\"CitationRef\"\u003e2017\u003c/span\u003e, Page et al. \u003cspan citationid=\"CR72\" class=\"CitationRef\"\u003e2021\u003c/span\u003e, Page and Williams \u003cspan citationid=\"CR71\" class=\"CitationRef\"\u003e2023\u003c/span\u003e). Apparent competition between honeybees and native pollinators for shared flowers can displace native bees (Page and Williams \u003cspan citationid=\"CR71\" class=\"CitationRef\"\u003e2023\u003c/span\u003e, Travis et al. in press). Thus, where environmental change increases exotic wildflowers, exotic pollinators may mediate the resulting impacts on native pollinators and plants.\u003c/p\u003e\u003cp\u003eTaken together, our work shows that from the perspective of pollinators, the cause of invasion can determine the impact, but, importantly, exotic wildflowers, especially exotic asters, can compensate for reductions in native wildflowers. However, exotic wildflowers increase visits by exotic pollinators, which could lead to declines in native pollinators, presenting a paradox for pollinator conservation in rapidly changing environments. More generally, these findings emphasize the tradeoffs in the positive and negative effects of exotic species on ecological communities (Graves and Shapiro \u003cspan citationid=\"CR39\" class=\"CitationRef\"\u003e2003\u003c/span\u003e, Zarnetske et al. \u003cspan citationid=\"CR103\" class=\"CitationRef\"\u003e2010\u003c/span\u003e, Biel et al. \u003cspan citationid=\"CR5\" class=\"CitationRef\"\u003e2017\u003c/span\u003e). Ultimately, this paradox raises the question of whether conservation efforts should value species based on their ecological traits and functions rather than provenance (Lundgren et al. \u003cspan citationid=\"CR62\" class=\"CitationRef\"\u003e2024\u003c/span\u003e).\u003c/p\u003e"},{"header":"Declarations","content":"\u003cp\u003e\u003cstrong\u003eOpen Research Statement:\u0026nbsp;\u003c/strong\u003eData for this paper have been made available to the public via the Envionmental Data Initiative. Data can be cited as follows: Nelson, R.A., E.W. Seabloom, and E.T. Borer. 2025. Data for \u0026quot;Pollinator Conservation Paradox: Exotic Wildflowers Support Native Pollinators Under Global Changes\u0026quot; by Nelson, Seabloom and Borer 2025, California grasslands, 2023-2024 ver 1. Environmental Data Initiative. https://doi.org/10.6073/pasta/1bda9f7c9f8c60bfca3eb8063dc76b78 (Accessed 2025-05-07). Code is accessible to the public via Zenodo at: 10.5281/zenodo.16322076.\u003c/p\u003e\u003cp\u003e\u003ch2\u003eConflict of Interest Statement:\u003c/h2\u003e\u003cp\u003eThe authors have no competing interests to declare.\u003c/p\u003e\u003c/p\u003e\u003ch2\u003eFunding:\u003c/h2\u003e\u003cp\u003eFunding for this project was supported by the Nature Conservancy\u0026rsquo;s Oren Pollak Memorial Research Fund for grassland research, the Hopland Scholars Fund, the UC Davis Jastro \u0026amp; Shields graduate research award, the Irene Brown Memorial Fund for women in environmental science, and the Davis Botanical Society student scholarship. Funding was also provided by the National Science Foundation (NSF-DEB-1042132, NSF-DEB-1234162, and NSF-DEB-1831944), and the Institute on the Environment (DG-0001-13), and the Minnesota Supercomputer Institute hosted the project data.\u003c/p\u003e\u003ch2\u003eAuthor Contributions:\u003c/h2\u003e\u003cp\u003eConceptualization: RAN, ES, ETB. Methodology: RAN, ES, ETB. Software: RAN. Validation: RAN, ES, ETB. Formal analysis: RAN. Investigation: RAN. Resources: RAN, ES, ETB. Data curation: RAN. Writing- Original draft preparation: RAN. Writing- Reviewing and Editing: RAN, ETB, ES. Visualization: RAN, ES, ETB. Supervision: RAN, ES, ETB. Project Administration: RAN, ES, ETB. Funding Acquisition: RAN, ES, ETB.\u003c/p\u003e\u003ch2\u003eAcknowledgements:\u003c/h2\u003e\u003cp\u003eWe thank I. Wolter, K.T. Lynch, S. Tantuico, N.T. Tat, A. Karp, A. Appelgate, Z. Schneider and R. Li for assistance in the field; I. Slette, P. Wilfahrt, and A. Krause for assistance with data; and C. Koehler and P. Aigner (University of California McLaughlin Reserve), J. Bailey and G. Solberg (Hopland UC Research and Extension Center), and M. Flavell and A. Northup-Warner Sierra Foothills UC Research and Extension Center) for maintaining field sites and facilities. We provided data from this project to J. Cancela for Cancela et al. \u003cem\u003ein prep\u003c/em\u003e. T. Zavortink (Bohart Museum of Entomology) identified insect voucher specimens. R. Dirzo, S. Harrison, and N. Williams provided feedback on this project. Two anonymous reviewers provided additional feedback on the manuscript.\u003c/p\u003e"},{"header":"References","content":"\u003col\u003e\n\u003cli\u003eBarthell, J. F., Randall, J. M., Thorp, R. W. and Wenner, A. M. 2001. 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Previous work has framed wildflower invasions as outcomes of global change, but less is known about how interacting anthropogenic drivers influence both invasion and pollination. In particular, it remains unclear whether exotic wildflowers compensate for native floral losses under ongoing environmental change. To address this, we tested whether exotic wildflowers support native pollinators under two drivers of wildflower decline: eutrophication and defaunation. Using a factorial global change experiment at three sites in the highly invaded California floristic region, we tested whether increases in exotic wildflowers (1) sustain pollinator visitation and richness, (2) maintain pollinator composition and function, and (3) facilitate co-invasion by exotic pollinators. We found that eutrophication promoted exotic asters, which served as visually prominent, attractive hubs in plant-pollinator networks. These asters supported both generalist and specialist native pollinators but also increased visitation by exotic pollinators, raising the risk of invasional meltdown. Our results suggest that exotic wildflowers can buffer pollinator communities against global change, but may do so while shifting pollinator composition toward non-native species.\u003c/p\u003e","manuscriptTitle":"Pollinator conservation paradox: exotic wildflowers support native pollinators under global changes","msid":"","msnumber":"","nonDraftVersions":[{"code":1,"date":"2025-07-23 05:31:49","doi":"10.21203/rs.3.rs-7188448/v1","editorialEvents":[{"type":"communityComments","content":0}],"status":"published","journal":{"display":true,"email":"[email protected]","identity":"researchsquare","isNatureJournal":false,"hasQc":true,"allowDirectSubmit":true,"externalIdentity":"","sideBox":"","snPcode":"","submissionUrl":"/submission","title":"Research Square","twitterHandle":"researchsquare","acdcEnabled":true,"dfaEnabled":false,"editorialSystem":"","reportingPortfolio":"","inReviewEnabled":false,"inReviewRevisionsEnabled":true}}],"origin":"","ownerIdentity":"ac982c3e-06a0-4bf7-9f9a-5ee10c2018f3","owner":[],"postedDate":"July 23rd, 2025","published":true,"recentEditorialEvents":[],"rejectedJournal":[],"revision":"","amendment":"","status":"posted","subjectAreas":[],"tags":[],"updatedAt":"2025-07-23T05:31:49+00:00","versionOfRecord":[],"versionCreatedAt":"2025-07-23 05:31:49","video":"","vorDoi":"","vorDoiUrl":"","workflowStages":[]},"version":"v1","identity":"rs-7188448","journalConfig":"researchsquare"},"__N_SSP":true},"page":"/article/[identity]/[[...version]]","query":{"redirect":"/article/rs-7188448","identity":"rs-7188448","version":["v1"]},"buildId":"8U1c8b4HqxoKbykW_rLl7","isFallback":false,"isExperimentalCompile":false,"dynamicIds":[84888],"gssp":true,"scriptLoader":[]}