Can beef be bee-friendly? Using native warm-season grasses and wildflowers in pastures
to conserve bees
Jennie F. Wagner
Thesis submitted to the faculty of the Virginia Polytechnic Institute and State University
in partial fulfillment of the requirements for the degree of
Master of Life Sciences
In
Horticulture
Megan O’Rourke, Committee Chair
Benjamin F. Tracy
Gabriel Pent
Roger Schürch
April 21st, 2020
Blacksburg, VA
Keywords: bees, wildflowers, native warm-season grasses, pastures, conservation, cattle
Can beef be bee-friendly? Using native warm-season grasses and wildflowers in pastures
to conserve bees
Jennie F. Wagner
ABSTRACT
Over the past several decades, native and managed bee populations have decreased in the
United States and worldwide. Although bee decline is attributable to several factors,
habitat loss is the primary driver. Simultaneously, cattle producers in the eastern U.S.
rely primarily on cool-season forages that peak in biomass production in late spring,
leading to a lack of forage in the summer months and increasing the costs of cattle
production. Seeding pastures with a mix of native warm-season grasses and native
wildflowers could increase forage availability while also increasing available resources
for bees. In this study, a mix of three native warm-season grasses (NWSGs) and 15
wildflower species was planted at the Virginia Tech Shenandoah Valley Agricultural
Research and Extension Center (SVAREC). The objectives of this project were to
document the establishment and species composition of NWSG + wildflower pasture
mixtures, compare the attractiveness of wildflowers and weedy species to bees, and
compare the bee community between NWSG + wildflower pastures and more typical
cool-season grass pastures. The wildflowers in the NWSG + wildflower pastures
dominated over grasses. All wildflower species that established were attractive to bees,
as were some weedy species. The NWSG + wildflower treatments had the highest
abundance of bees collected, with an average of 14.8 bees collected per pasture per
sampling date in 2018, and an average of 12.4 bees collected per pasture per sampling
date in 2019. These results indicate that with modification of establishment methods so
that more grasses are present, this pasture system could be beneficial from both a cattle
production and bee conservation standpoint.
Can beef be bee-friendly? Using native warm-season grasses and wildflowers in pastures
to conserve bees
Jennie F. Wagner
GENERAL AUDIENCE ABSTRACT
Over the past several decades, there has been a decline in bee populations in the U.S. and
around the world. Bees play an important role in pollinating many food crops, including
most fruits and vegetables. Habitat loss is the biggest contributor to their decline. There
are also issues with cattle production in the eastern U.S. Most farmers rely on grasses
that are the most productive in the late spring and early summer, meaning that by mid-
and late summer, there is little grass available for cattle. Planting pastures with native
grasses designed to be the most productive in the late summer and native wildflowers
could increase food available for cattle as well as provide more pollen and nectar for
bees. In this experiment, we planted a mix of three grasses and 15 wildflowers. We
documented how well the grasses and wildflowers established. We also examined how
attractive wildflowers and weeds were to bees and compared the number and types of
bees collected between the new pastures and traditional pastures. We found that the
wildflowers, instead of the grasses, dominated the pastures. All wildflowers that
established, as well as some weeds, attracted bees and provided resources. Higher
numbers of bees were collected in the pastures with wildflowers than standard grass
pastures, but there were not necessarily more bee species present. These results suggest
that, with some modifications, planting native grasses and wildflowers in pastures could
help conserve bees as well as benefit cattle farmers.
iv
Acknowledgements
I’m taking this space to acknowledge the perseverance and determination that it took on
my part to finish this degree. I owe myself a huge thank you for gritting my teeth and
gutting it out when all I wanted to do was curl up in a ball and quit. I’m excited to move
on and start a job I’m going to love.
On a less self-indulgent note, there are a few people who deserve my sincere thanks. My
co-advisor, Dr. Ben Tracy, has been nothing but supportive throughout this entire
process. Dr. Megan O’Rourke, also my co-advisor, provided useful ideas and
knowledge.
I also owe thanks to my committee members, Dr. Roger Schürch and Dr. Gabe Pent, who
provided guidance on experimental design, methodology, and statistical analyses. The
crew at the Shenandoah Valley AREC, David Fiske, Colby, Brian, and Chris, made field
work a lot more fun and were always happy to help when needed. My lab members,
Velva, Mike, Chris, Gina, and Shayan, helped with field work and data analysis and
provided moral support.
Dr. Donna Westfall-Rudd has been a beacon of hope in the last year. Donna continually
reminded me that I had value and gave me confidence when I had none. Because of her
help, I landed a great job that I couldn’t be more excited about.
My community of friends in Blacksburg and beyond is made of special people. To
Alaina, thanks for making me laugh and understanding my anxiety about basically
everything. I can’t wait for another road trip with you! Cristina, I’m so thankful to have
a strong, supportive friend like you. I’m so excited for us both to move on to bigger and
better things! Sarah, thanks for the endless laughs and ridiculousness. I’m looking
forward to starting our farming commune one day. Sandy, I’m glad we’ve been able to
share the trials and tribulations of graduate school together, albeit from afar. Bennett,
thanks for being an emotional support friend and covert spy in another building. And,
last but not least: Jess, Corinne, and Harper, you boss ladies are real keepers. I couldn’t
have done any of the last two and a half years without y’all. Thanks for all the miles
(including my first marathon!), memories, adventures, laughs, chocolate, and French
fries. Oh, and the toilet paper during the apocalypse.
FINALLY, this monologue of mine wouldn’t be complete without thanking my mom,
dad, and brother. We’ve been through a lot in the last six months, but we somehow still
manage to love and support each other. I could say more, but there’s a good chance I’ll
cry while I’m typing, and I can’t have that. I love you guys.
v
Table of Contents
Literature Review................................................................................................................ 1
The “insect apocalypse” and decline of bees .................................................................. 1
Causes of bee decline ...................................................................................................... 2
Effects of bee decline ...................................................................................................... 5
Solutions to bee decline................................................................................................... 7
The importance of grasslands in the United States ......................................................... 8
Use of pastures for bee conservation............................................................................. 10
Summary ....................................................................................................................... 13
Introduction ....................................................................................................................... 15
Materials and Methods ...................................................................................................... 19
Study site ....................................................................................................................... 19
Experimental design and cattle stocking management ................................................. 19
NWSG + wildflower plot establishment ....................................................................... 21
Measurements................................................................................................................ 24
Data analysis ................................................................................................................. 25
Results ............................................................................................................................... 27
Establishment of native warm-season grass + wildflower mix ..................................... 27
Bee observations ........................................................................................................... 31
Bee community data ...................................................................................................... 43
Discussion ......................................................................................................................... 52
Establishment of native warm-season grass + wildflower mix ..................................... 52
Bee observations ........................................................................................................... 54
Bee community data ...................................................................................................... 58
Conclusions ....................................................................................................................... 62
References ......................................................................................................................... 63
1
Literature Review
The “insect apocalypse” and decline of bees
Global insect decline made headlines in 2018 as scientists and popular media warned of a
coming “insect apocalypse” (Jarvis 2018). Bees are one group of insects cited as part of the
decline (Winfree et al. 2009, Hallmann et al. 2017, Powney et al. 2019). Over the past several
decades, native and managed bee populations have decreased in the United States and worldwide
(Biesmeijer et al. 2006, Potts et al. 2010, 2016). Companies like Bayer and General Mills have
made honeybees (Apis mellifera L.) the focus of publicity campaigns because of widespread
population declines (Bayer 2019, Cheerios 2019). There is no sign that the decline is slowing:
2018 saw the largest winter loss of honeybee colonies in the U.S. since 2008 (Bee Informed
Partnership 2019). Additionally, some formerly widespread native bee species in the U.S. are
now listed as federally endangered (USFWS 2019), and at least 23 bee and flower-visiting wasp
species in the United Kingdom have gone extinct in the last 150 years (Ollerton et al. 2014).
Bee decline is attributable to several factors, including parasites, pathogens, pesticide
exposure, and habitat loss (Vanbergen and Insect Pollinators Initiative 2013, Ollerton et al. 2014,
Goulson et al. 2015, DiBartolomeis et al. 2019). The resulting loss of pollination services is a
cause of agricultural concern (Potts et al. 2016, Winfree et al. 2018), as animal pollinators
directly or indirectly affect up to 75% of the global food supply (Klein et al. 2007). Using
wildflower plantings to increase available habitat can combat bee decline (Blaauw and Isaacs
2014a, Venturini et al. 2017, Paterson et al. 2019), although the size of habitats may not be
enough to effectively bolster bee populations. Using pastureland to both feed livestock and
provide bee habitat may offer a solution, resulting in large-scale bee conservation.
2
Causes of bee decline
Parasites and pathogens
Parasites and pathogens affect honeybees and native bees, although their occurrence in
honeybees is more often studied (Goulson et al. 2015). Diseases and parasites, especially those
that are non-native to the host range, can transfer from honeybees to native bees. For example,
Nosema ceranae, a parasite that originated in the Asian honeybee (Apis cerana), spread to the
European honeybee (A. mellifera) and is now able to infect bumblebees (Graystock et al. 2013).
Other pests and pathogens have been introduced with managed bee colonies that can harm native
bee populations (Colla et al. 2006, Schmid-Hempel et al. 2014). Varroa mite, a honeybee
parasite, is host-specific to bees in the genus Apis (Goulson et al. 2015). However, Varroa mite
transmits deformed wing virus to honeybees, which can then transmit the virus to wild bees
through contact with the same flowers (Wilfert et al. 2016). Managed and native bees are
especially susceptible to pathogens and pests when weakened by pesticide exposure (Potts et al.
2010).
Pesticides
Pesticide use is one of the most debated causes of bee decline (Goulson et al. 2015).
Insecticides, acaricides, fungicides, and herbicides can all have negative effects on bees.
Insecticides, especially neonicotinoids, have been shown to have sublethal and occasionally
lethal effects on honey and bumblebee colonies (Sanchez-Bayo and Goka 2014, Pisa et al. 2014).
Bees can be exposed to neonicotinoids via dust, pollen, and nectar (Goulson 2013). When
exposed to neonicotinoids, honeybees exhibit slower return to foraging sites (Yang et al. 2008)
and increased mortality when returning to the colony from foraging sites (Henry et al. 2012).
Similarly, bumblebees that came in contact with neonicotinoids returned to their colonies with
3
pollen less frequently, and when they did return, brought back less pollen (Feltham et al. 2014).
Field exposure to neonicotinoids can also decrease bumblebee colony growth and reproduction
(Rundlöf et al. 2015). One-time exposure is damaging, but continued exposure over time and
exposure in combination with fungicides can increase the threat to honey and bumblebees
(Sanchez-Bayo and Goka 2014). The negative effects of neonicotinoids may be even more
potent for non-bumblebee native bees. Current research is sparse, but early work suggests that
exposure to neonicotinoids reduces native bee species richness (Main et al. 2020) and increases
population extinction rates (Woodcock et al. 2016). Native solitary bees are at least as sensitive
as honeybees to neonicotinoids (Pisa et al. 2014). One study found reduced nesting in solitary
bees exposed to neonicotinoid seed treatments (Rundlöf et al. 2015). By nature of their non-
communal population dynamics, solitary bee foragers may be less resilient to neonicotinoid
exposure (Henry et al. 2012).
Organophosphates and pyrethroids are other commonly used classes of insecticide that
impact bees. Organophosphates (Gregorc et al. 2018b, 2018a, Tomé et al. 2020) and
organophosphate alternatives like novaluron (Fine et al. 2017) decrease honeybee colony health
by impacting workers and immature bees. Additionally, pyrethroids, another class of
insecticides and acaricides, pose risks to honeybees under laboratory conditions (Sanchez-Bayo
and Goka 2014, Tomé et al. 2020, Qi et al. 2020). However, if properly applied, exposure to
pyrethroids at field concentrations may have very little effect on honeybees (Pokhrel et al. 2018).
While research on pyrethroid exposure in other types of bees is scarce, adult alfalfa leafcutting
bees (Megachile rotundata) were found to be less susceptible than honeybees to most
pyrethroids (Piccolomini et al. 2018).
4
Commonly used fungicides can have non-target effects on bees as well. Boscalid, a
fungicide applied to control floral fungi in apple orchards, can inhibit honeybee forager flight
efficiency (Liao et al. 2019). Iprodione (Carneiro et al. 2020) and picoxystrobin (Batista et al.
2020) are broad-spectrum fungicides that have been shown to cause midgut cell death in
honeybees, which could have broader implications for overall hive health. Chlorothalonil,
another broad-spectrum fungicide, can decrease honeybee queen egg-laying efficiency (Walsh et
al. 2020) and can decrease the health and survival of immature honeybees (Tomé et al. 2020).
Fungicides can also affect non-honeybee species. In studies with the non-native mason bee
Osmia cornifrons, female offspring number and weight were negatively affected when exposed
to higher levels of fungicides in pollen (Centrella et al. 2020).
Like insecticides and fungicides, herbicides can have sublethal effects on bees.
Honeybees exposed to field levels of glyphosate had more indirect flights to their colonies and
spent more time navigating home, which decreased foraging efficiency (Balbuena et al. 2015,
Farina et al. 2019). Glyphosate may have similar effects on non-honeybees. Glyphosate was
found to be lethal to Melipona quadrifasciata larvae, a native Brazilian stingless bee (Seide et al.
2018). Additionally, herbicide application has contributed to the loss of bee habitat and can
therefore affect bee abundance and diversity. In a study with genetically modified (GM),
glyphosate-resistant canola, there was lower weed diversity and abundance in GM versus
conventional fields, which decreased available bee habitat and possibly affected bee abundance
(Morandin and Winston 2005). Similarly, Bohan et al. (2005) found that GM canola had fewer
dicot weeds than conventional canola, leading to fewer bees. Another study showed an increase
in solitary bee species richness on organic farms due to a lack of herbicide use and increased
insect-pollinated forbs (Happe et al. 2018).
5
Habitat loss
Habitat loss leads to species decline at a local and regional scale in several taxa (Staude et
al. 2018, Horváth et al. 2019, Isbell et al. 2019). Unsurprisingly, habitat loss, especially in
extreme cases, is cited as the primary driver of managed and native bee decline (Winfree et al.
2009, Potts et al. 2010). Habitat area and connectivity are both crucial to long-term conservation
of biodiversity (Hanski 2007, McGuire et al. 2016, Neokosmidis et al. 2018, Pavageau et al.
2018, Thompson et al. 2019). Human-induced changes in land-use have been shown to
negatively affect bee species richness (Senapathi et al. 2015, Vray et al. 2019). For example,
bumblebees and other wild bees in the United Kingdom (Goulson et al. 2005, Baude et al. 2016)
and the Netherlands (Scheper et al. 2014) have likely declined because of the loss of preferred
flower species. Similarly, in the midwestern United States, a comparison between historical and
present-day records reveals that bees and their preferred forbs are now spatially separated due to
habitat loss, leading to the decline of bee populations (Burkle et al. 2013). While habitat loss
negatively affects the entire bee community in a locale, it disproportionately impacts some bee
species. Rare, specialist pollinators suffer the most from habitat loss because they rely on only a
few flower species that may be more vulnerable to land-use changes (Biesmeijer et al. 2006,
Powney et al. 2019).
Effects of bee decline
Genetic diversity
Species decline due to habitat loss can have long-lasting effects on bee community
resilience. Shrinking bee populations have less overall genetic diversity, decreasing their ability
to recover from stress (Willi et al. 2006). Bee phylogenetic diversity has also been shown to
decrease at a rate higher than what would be expected based on loss of species richness alone
6
(Grab et al. 2019). Given the myriad environmental stressors that bees face, a loss of genetic
diversity could be the nail in the coffin for struggling populations. Decreased genetic diversity
can result in the loss of traits from the population that could provide adaptive mechanisms to
environmental stressors.
Ecosystem services
Bee decline and the resultant loss of pollination services is of major agricultural concern
(Potts et al. 2016). Pollinators affect up to 75% of the global food supply (Klein et al. 2007) and
up to 80% of non-crop plants depend on pollinators in some way (Potts et al. 2010). In 2005, the
total value of pollination worldwide was $191.3 billion, amounting to 9.5% of the total value of
global agricultural production used for human food (Gallai et al. 2009). Among animal
pollinators, bees are the most frequent floral visitors (Neff and Simpson 1993). Habitat loss has
already affected pollination services in and around agricultural areas. For example, agricultural
intensification and the resultant loss of natural area threatened wild bee pollinators of nine crops
across four continents (Klein et al. 2007). Similarly, a meta-analysis found that as farm distance
from natural habitat increased, bee richness and visitation decreased (Ricketts et al. 2008).
Honeybees have been widely used for agricultural pollination but, as previously discussed,
colony numbers have declined sharply (National Research Council 2007). Native bees can
prevent agricultural production losses due to honeybee declines (Greenleaf and Kremen 2006,
Winfree et al. 2007, Lentini et al. 2012, Button and Elle 2014). However, the status of native bee
populations is at least as tenuous as honeybees. Without drastic conservation action, lost
pollination services could severely impact agricultural production and global plant diversity
(Potts et al. 2016).
7
Solutions to bee decline
Existing bee conservation efforts
Conservation of natural or wilderness areas prevents species extinction (Di Marco et al.
2019) and could help bee populations recover. Establishment of wildflower habitats has been
shown to be an effective method of conserving native bees (Blaauw and Isaacs 2014b, Kremen
and M’gonigle 2015, Venturini et al. 2017, Paterson et al. 2019). Wildflower plantings can be
useful bee habitat across a variety of landscapes. In urban areas, home gardens and city
greenspaces provide bee habitat (Goulson et al. 2002a, McFrederick and Lebuhn 2006, Osborne
et al. 2008). Some states have roadside wildflower plantings that provide habitat on state rights-
of-way, at rest areas, and in park-and-ride areas (NCDOT 2019, ODOT 2019, VDOT 2019). The
USDA recognized the need for bee habitat as part of the policy initiatives included in the 2008,
2014, and 2018 Farm Bills (USDA-NRCS 2019a). Traditionally, Natural Resource Conservation
Service (NRCS) programs have encouraged farmers to conserve bees with wildflower habitats
planted on marginal farmland (Vaughan and Skinner 2015). Additionally, the Xerces Society
suggests that peripheral farm areas like fencerows and ditches, poor-quality farmland, and cover
crops planted under perennial orchards can be used as bee habitat (Vaughan et al. 2015).
Problems with existing conservation efforts
While all efforts to conserve bees are important, many suffer from a lack of landscape
continuity and provide small patches of habitat. Large areas of bee habitat are important for
conservation and agricultural pollination services. In natural, montane meadows, there is a
positive correlation between meadow size and richness of flower-visiting species (Jones et al.
2019). The same correlation holds true in managed, agricultural systems. A study in Michigan
found that larger wildflower patch sizes (100 m2) increased wild bee density and subsequent
8
wildflower seed set as compared to patches between 1 and 30 m2 (Blaauw and Isaacs 2014c).
Another recent study found that in agricultural areas with low surrounding semi-natural habitat,
bees visited oilseed rape next to large wildflower patches (>1.5 ha) more frequently than oilseed
rape next to small wildflower patches (<1.5 ha) (Krimmer et al. 2019). Similarly, around tomato
fields, there was an increase in bee abundance when natural area increased (Franceschinelli et al.
2017). However, since current bee conservation efforts target wildflower plantings on marginal
farm areas, large wildflower patches may be difficult to create. Furthermore, wildflower patches
planted in marginal and peripheral farmlands may lose plant diversity and therefore conservation
value over time (Wesche et al. 2012). As wildflower patches are left to succession and gradually
convert to forested land, there is a decrease in total flower cover, resulting in a decrease in
flower-visiting insects like bees (Walcher et al. 2019). Wildflower patches planted in an
agricultural system with some form of disturbance management may last longer, therefore
retaining conservation value over time.
The importance of grasslands in the United States
Grasslands are defined as large land areas where grasses are the dominant vegetation
(USDA-NIFA 2020). Similarly, rangelands are large open areas that include grasses, forbs, and
shrubs (USDA-NIFA 2020). Both grasslands and rangelands are used primarily for grazing.
Globally, these biomes supply a variety of ecosystem services (Deru et al. 2017, Bengtsson et al.
2019), including water regulation (Sirimarco et al. 2018), carbon storage (Balasubramanian et al.
2020), erosion control (Wilson et al. 2014), and pollination (Johnson et al. 2009).
In the United States, grasslands and rangelands comprise 265 million hectares of land
(USDA-ERS 2017) and the majority of that land is used to graze cattle (USDA 2017). In the
eastern U.S., most pastures are continuously stocked (di Virgilio et al. 2019) and primarily
9
consist of cool-season grasses (Hoveland 2000). Cool-season grasses have a bimodal growth
pattern: they produce a large flush of biomass in the spring, when temperatures and moisture are
suitable, and produce another, smaller flush of biomass in the fall if there is adequate rainfall
(Moser and Hoveland 1996). Tall fescue (Schedonorus arundinaceus Schreb) is one of the most
common cool-season pasture grasses in the eastern U.S. (Nelson et al. 2017).
Most tall fescue is infected with a fungal endophyte, Epichloë coenophiala (Morgan-
Jones and W. Gram) (Bacon et al. 1977), that confers drought tolerance and resilience to
intensive grazing (Hoveland 1993). While the endophyte is beneficial for the plant, it produces
an ergot alkaloid that, when consumed, is toxic to cattle (Hoveland et al. 1980, Roberts and
Andrae 2004). The ergot alkaloid is a vasoconstrictor that reduces blood flow to extremities and
the surface of the skin, causing heat stress in the summer (Hannah et al. 1990). Heat stress
decreases cattle weight gain (Thompson and Stuedemann 1993) and reproductive success
(Schuenemann et al. 2005), both of which can be an economic concern to producers (Hoveland
1993).
When given access to alternative forages, cattle have been shown to limit their grazing of
endophyte-infected tall fescue (Maresh Nelson et al. 2019). Offering access to alternative
forages can dilute toxins from endophyte-infected tall fescue, increasing weight gain and
reproductive success (Coblentz et al. 2006, Drewnoski et al. 2009). Native warm-season grasses
(NWSGs) are a viable, non-toxic alternative to tall fescue. NWSGs, like big bluestem
(Andropogon gerardii Vitman), little bluestem (Schizachyrium scoparium Nash), and indiangrass
(Sorghastrum nutans L.), have a C4 photosynthetic pathway and peak in biomass production in
mid- to late summer when cool-season grass productivity is low (Brown 1999, Tracy et al. 2010).
NWSGs can provide supplemental summer forage when cool-season grasses are dormant (Moore
10
et al. 2004, Tracy et al. 2010, Burns and Fisher 2013, Keyser et al. 2016, Backus et al. 2017),
increasing summer weight gain without changing beef quality (Kurve et al. 2015). Stockpiled
NWSGs can even be selectively grazed in the winter (Tilhou et al. 2019), adding to their
economic benefits.
Use of pastures for bee conservation
Pastures and rangeland could be used for bee conservation in a land-sharing system
(Baude et al. 2016). A land-sharing approach increases the amount of wildlife present on
farmland (Green et al. 2005), providing an opportunity for agricultural production and
conservation of biodiversity (Ekroos et al. 2016). This approach can then increase agricultural
yields (Hipólito et al. 2018) and help mitigate bee decline (Kovacs-Hostyanszki et al. 2017).
However, there is little research about how adding native grassland plants to pastures and
managed grasslands affects bee populations. To date, most research has focused on the effects of
grazing management and grazing intensity on bees. For example, white clover (Trifolium
repens) has been shown to increase under extensively managed grazing (Pavlu et al. 2003),
increasing available bee resources. Moderate grazing had no effect on wild bee abundance,
species richness, and bee community composition, and in some cases positively affected
availability of floral and nesting resources (Shapira et al. 2019). In contrast, intensively
managed pastures negatively affect bee populations. Management-intensive grazing negatively
affected bee abundance (Lazaro et al. 2016, Smith et al. 2016) and bee body size (Smith et al.
2016). Intensively managed pastures likely have smaller and less diverse bee communities
because any flowers present during grazing are damaged or grazed. Paddocks that are not grazed
during peak forb bloom have been shown to have higher bumblebee abundance, density, and
species richness than paddocks that are grazed (Scohier et al. 2013, Enri et al. 2017). However,
11
Sjodin et al. 2008 did not find any significant differences in flower availability or bee diversity
between extensively and intensively managed pastures.
Bee conservation benefits of NWSGs in pastures
NWSGs benefit wildlife conservation as well as cattle production. NWSGs have been
shown to provide nesting habitat for grassland birds (Giuliano and Daves 2002, Harper et al.
2015, Monroe et al. 2016) and could provide nesting habitat for bees. NWSGs are bunch grasses
(Harper et al. 2007), that create patches of bare ground between plants, which are optimal nest
sites for ground-nesting bees. One study found that pastures planted with NWSGs had higher
abundances of native bees and honeybees than pastures planted with non-native grasses
(Bhandari et al. 2018). This may be because of increased nesting resources present in NWSG
pastures. Solitary stem-nesting bees, like bees of the genus Ceratina, may benefit from hollow
NWSGs stems that can be used for nests (Michener 2007).
Bee conservation benefits of native forbs in pastures
Native and managed grasslands provide nesting and food resources for bees (Morandin et
al. 2007, Wick et al. 2016, Shapira et al. 2019, Bendel et al. 2019). Augmenting existing
pastures with forage legumes and forbs increases resources to some extent. Woodcock et al.
(2014) and Orford et al. (2016) found that adding blooming forage legumes and forbs can
increase the abundance and species richness of several pollinators, including bees. However,
Frankie et al. (2005) and Tuell et al. (2008) suggest that native perennial flowers have more bee
conservation value than non-native annuals typically used in pollinator conservation plantings
and pastures. Frankie et al. (2005) suggest that non-native ornamental annuals have not been
bred to provide adequate nectar and pollen to native and honeybees, decreasing the attractiveness
of the flowers. Tuell et al. (2008) state that the seasonal succession of native blooming flowers
12
available over the course of the growing season provided more resources to bees than a non-
native annual planting would have.
Designing seed mixes for native pasture plantings
Given the lack of research on using native plants in pastures, designing appropriate seed
mixes could prove challenging. Plant biodiversity should be a top priority when designing
conservation-oriented pasture mixes; wildflower habitats with higher plant diversity attracted
more bees (Potts et al. 2003, Gill et al. 2014). Bees differ in their use of floral resources and
there are seasonal shifts in bee activity (Rundlöf et al. 2014, Bendel et al. 2019), so habitats with
higher plant diversity will likely attract more bees. Additionally, increased plant diversity fuels
increased bee diversity, which helps maintain the plant diversity in a positive feedback loop
(Fontaine et al. 2006).
Diversity of blooming plants in pasture mixtures is important, but there is variability in
the bee conservation value that different wildflower species provide. When designing seed
mixes, the flower species should be considered carefully to maximize bee conservation potential.
Rowe et al. (2018) found that common milkweed (Asclepias syriaca L.), bee balm (Monarda
fistulosa L.), and showy goldenrod (Solidago speciosa Nutt.) attracted the most native and
honeybees out of 51 plant species tested. Tuell et al. (2008) found similar results, showing that
several native forbs, including grey-headed coneflower (Ratibida pinnata Vent.), yellow giant
hyssop (Agastache nepetoides L. Kuntze.), showy goldenrod (Solidago speciosa Nutt.), and
lanceleaf coreopsis (Coreopsis lanceolata L.) were either highly or moderately attractive to bees.
Planting a biodiverse seed mix that contains several species of highly attractive wildflowers is
likely to maximize the bees conserved in a pasture system. However, when designing seed
mixes for pasture systems, the effects on cattle should also be considered. Plants that are toxic to
13
cattle, like common milkweed (A. syriaca L.) (USDA-ARS 2018) should not be included.
Additional research on the attractiveness and conservation value of diverse wildflower mixes
planted in pasture systems is necessary to best design future seed mixes.
Summary
Native and managed bee populations are declining around the globe. Although multiple
factors contribute to decline, habitat loss is the primary driver. Habitat loss to urbanization and
agriculture has led to a loss of bee abundance and species richness globally. In addition to
ecological implications, the effects of habitat loss can severely affect agriculture. Up to 75% of
the global food supply depends on pollination, and bees are the primary animal pollinator.
Adding bee habitat to simplified landscapes can bolster existing bee communities,
increasing abundance and species richness. Landowners in urban areas and rural areas alike
have added wildflower plantings to fallow land or marginal areas. However, these plantings are
small and fragmented. Larger wildflower plantings have been shown to attract more bees.
Pastures and rangeland, which covers 265 million hectares across the United States, could
provide enough land area for large-scale wildflower plantings, while still providing resources for
cattle.
If used in a land-sharing system, pastures and rangeland could be used for large-scale bee
conservation (Baude et al. 2016). A land-sharing approach provides an opportunity for
agricultural production and conservation of biodiversity (Ekroos et al. 2016) and can help
mitigate bee decline (Kovacs-Hostyanszki et al. 2017). A pasture designed for cattle production
and bee conservation will mirror a grassland plant community, where grasses dominate but forbs
fill other ecological niches. NWSGs will provide forage for cattle, especially during the summer
months (Moore et al. 2004, Tracy et al. 2010, Backus et al. 2017), and wildflowers will provide
14
food and nesting resources for bees. Together, NWSGs and native wildflowers can provide
extensive benefits for bee populations.
The new NRCS Conservation Stewardship Program Grasslands Conservation Initiative is
providing incentives for producers to manage their pastureland to conserve bees and other
wildlife (USDA-NRCS 2019b). However, there is little research about how using native grasses
and wildflowers in pastures affects bee populations and cattle production. If producers are to
adopt NWSG and wildflower conservation plantings, research on their use and benefits, both to
bees and cattle, is sorely needed. My research project will examine several aspects of NWSG-
wildflower plantings and their use in pasture systems. I will address plant community
establishment, attractiveness of different wildflower species to bees, and the overall bee
conservation value of wildflowers in pastures. My research should help inform future work and
help assess the benefits and feasibility of adding NWSGs and wildflowers to cattle production
systems.
15
Introduction
Over the last several decades, native and managed bee populations have steadily declined
(Biesmeijer et al. 2006, Winfree et al. 2009, Potts et al. 2010, 2016, Hallmann et al. 2017,
Powney et al. 2019). Bees are important pollinators of non-crop plants (Potts et al. 2010), but
their decline is also of agricultural concern (Potts et al. 2016, Winfree et al. 2018), as up to 75%
of global food production depends on animal pollinators in some way (Klein et al. 2007). While
bees are impacted by parasites, pathogens, and pesticides, the decline is primarily driven by
habitat loss (Winfree et al. 2009, Potts et al. 2010, Vanbergen and Insect Pollinators Initiative
2013, Ollerton et al. 2014, Goulson et al. 2015, DiBartolomeis et al. 2019).
Habitat loss and human-induced changes in land-use negatively affect bee species
richness (Senapathi et al. 2015, Vray et al. 2019). Additionally, habitat loss can have long-
lasting effects on bee community resilience. Shrinking bee populations lose genetic (Willi et al.
2006) and phylogenetic (Grab et al. 2019) diversity, which can result in the loss of traits that
could provide adaptive mechanisms to environmental stressors.
Previous research has shown that wildflower habitats in agricultural systems aid in native
bee conservation (Blaauw and Isaacs 2014b, Kremen and M’gonigle 2015, Venturini et al. 2017,
Paterson et al. 2019). Currently, pollinator conservation programs offered by the USDA-NRCS
encourage farmers to plant pollinator habitats on marginal or fallow farmland (Vaughan and
Skinner 2015). The non-profit Xerces Society suggests that peripheral farm areas like fencerows
and ditches, poor-quality farmland, and cover crops planted under perennial orchards can be used
as bee habitat (Vaughan et al. 2015). While other cultural practices such as minimizing
insecticide application and planting pollinator-friendly cover crops can also benefit bees,
permanent wildflower habitats will probably provide the greatest conservation benefits.
16
Furthermore, existing conservation efforts lack landscape continuity and provide relatively small
patches of habitat. Larger wildflower patches in agricultural systems have been shown to
increase bee visitation (Blaauw and Isaacs 2014c, Krimmer et al. 2019). Incorporating
wildflower plantings into a land-sharing system on large areas of existing agriculturally
productive land may increase bee habitat without removing land from production.
Pastures are agricultural land that, if enhanced, could serve as habitat for bees and other
pollinators. Pastures and rangeland comprise 265 million hectares (55%) of the land area in the
United States (USDA-ERS 2017) and the majority of that land is used to graze cattle (USDA
2017). If used in a land-sharing system, pastures and rangeland could be used both for cattle
production and large-scale bee conservation (Baude et al. 2016). A land-sharing approach
provides an opportunity for agricultural production and conservation of biodiversity (Ekroos et
al. 2016) and can help mitigate bee decline (Kovacs-Hostyanszki et al. 2017). In the eastern
U.S., pastures for cattle production primarily consist of cool-season grasses (Hoveland 2000) and
have few blooming forbs or weeds. Most cattle pastures are managed with continuous stocking,
an extensive system where a herd has access to an entire pasture during the grazing season, or
with rotational stocking, a more intensive system where a herd is rotated through a series of
paddocks over a given time period (Allen et al. 2011).
Several studies have shown that native and managed grasslands provide food and nesting
resources for bees (Morandin et al. 2007, Wick et al. 2016, Shapira et al. 2019, Bendel et al.
2019). The type of grazing management affects the availability of these resources, subsequently
impacting the bee community. White clover (Trifolium repens L.), a blooming legume, has been
shown to increase under continuous stocking (Pavlu et al. 2003), which could indirectly increase
available resources for bees. Moderate grazing in Mediterranean rangelands had no effect on
17
wild bee abundance, species richness, and bee community composition, and in some cases
positively affected availability of floral and nesting resources (Shapira et al. 2019). In France,
paddocks that were not grazed during peak forb bloom had higher bumblebee abundance,
density, and species richness (Scohier et al. 2013, Enri et al. 2017). In contrast, a lack of floral
resources in intensively grazed pastures has been shown to negatively affect bee abundance
(Lazaro et al. 2016, Smith et al. 2016) and bee body size (Smith et al. 2016). However, Sjodin et
al. (2008) found no significant differences in flower availability or bee diversity between
intensively and extensively grazed pastures.
While grazing management likely affects floral resources and bee abundance (Scohier et
al. 2013, Lazaro et al. 2016, Smith et al. 2016, Enri et al. 2017, Shapira et al. 2019), there has
been little prior research on how native wildflower habitats planted in pastures affect bee
populations. Tuell et al. (2008) suggest that native perennial flowers have more bee conservation
value than the non-native annuals typically used in pollinator conservation plantings and
pastures. Some studies have examined intentional establishment of pollinator plants in pastures
and found that additional forbs and legumes increased and sustained pollinator species richness
(Woodcock et al. 2014, Orford et al. 2016).
Native warm-season grasses (NWSGs), such as big bluestem (Andropogon gerardii
Vitman), little bluestem (Schizachyrium scoparium Nash), and Indiangrass (Sorghastrum nutans
L.), offer livestock producers in the eastern U.S. an alternative forage to use in their grazing
systems. NWSGs have a C4 photosynthetic pathway and are most productive during the mid- to
late summer when common cool-season grasses, like tall fescue (Schedonorus arundinaceus
Schreb), are dormant (Brown 1999, Tracy et al. 2010). NWSGs can provide supplemental
summer forage when cool-season grass productivity is low, allowing producers to extend the
18
grazing season (Moore et al. 2004, Tracy et al. 2010, Burns and Fisher 2013, Keyser et al. 2016,
Backus et al. 2017). Additionally, there is evidence that suggests that NWSGs provide pollinator
resources that are not as readily available in non-native grass pastures (Bhandari et al. 2018).
NWSGs are bunch grasses (Harper et al. 2007), creating patches of bare ground between plants
that are optimal nest sites for ground-nesting bees. Therefore, combining NWSGs with
wildflowers in pastures could provide nesting and food resources for bees, while also providing
forage for cattle production. If stands of NWSGs and wildflowers can be successfully
established and sustained, they would provide an opportunity for widespread bee conservation on
pasture and rangeland. Data collected from this study will provide information on establishment
methods for biodiverse pastures to optimize bee conservation and add information about the bee
conservation potential of NWSG + wildflower pastures used in cattle production.
To test the feasibility of this idea, pastures containing NWSGs and wildflowers were
established at the Virginia Tech Shenandoah Valley Agricultural Research and Extension Center
(SVAREC) in 2017. The NWSG + wildflower pastures were grazed by beef cows starting in
2018. The objectives of my research project were to: 1) document the establishment and species
composition of NWSG + wildflower pasture mixtures, 2) evaluate the attractiveness of
wildflowers included in the mix to bees, and 3) compare the bee community composition
between NWSG + wildflower pastures and more typical cool-season grass pastures. The
following hypotheses were tested: 1) the resulting plant community in the NWSG + wildflower
pasture mix would closely reflect the seed mix planted; 2) bees would be more attracted to plants
seeded in the NWSG + wildflower pasture than to naturalized pasture weeds; and 3) bee
abundance and species richness will be highest in NWSG + wildflower pastures compared with
more typical cool-season grass pastures.
19
Materials and Methods
Study site
The study was conducted on existing pastureland at the Virginia Tech Shenandoah Valley
Agricultural Research and Extension Center (SVAREC) in Raphine, VA from 2018-2019. The
study site was in central Virginia (37◦55’56” N latitude, 79◦12’51” W longitude, elevation: 530
m). The region has a humid continental climate, with an average monthly high temperature
ranging from 7.9℃ in January to 30.7℃ in July. On average, SVAREC receives 1389 mm of
precipitation annually. The soils are classified as well-drained Frederick and Christian silt
loams, with slopes ranging from 2% to 25%. Prior to the start of the experiment, soils were
tested for pH, P, and K and were amended based on recommendations from the Virginia Tech
Soil Testing Lab. Pastures were dominated by cool-season pasture grasses such as tall fescue
(Schedonorus arundinaceus Schreb.), orchardgrass (Dactylis glomerata L.), and Kentucky
bluegrass (Poa pratensis L.).
Experimental design and cattle stocking management
The experiment was conducted on 58-ha of pastureland formerly used for grazing
experiments at SVAREC (Tracy and Bauer 2019). Pastureland was divided into nine 6.5 ha
experimental units (Figure 1) and assigned to three stocking treatments with three replications.
Treatment 1 consisted of rotational stocking where cattle groups were moved through eight
equal-sized (0.8-ha) using stocking periods of 4 d with a fixed 28 d rest period for each paddock.
20
Figure 1: Layout of the three stocking treatments at SVAREC: rotationally stocked cool-season
grasses, rotationally stocked NWSGs and wildflowers, and continuously stocked cool-season
grasses. Each treatment was replicated three times.
Treatment 2 used the same rotational stocking scheme but with one paddock planted to a
NWSG + wildflower mixture (described below). Treatment 3 was continuous stocking that
represented a “business-as-usual” control treatment to reflect grazing practices most farmers
follow in the eastern U.S. Continuous stocking is defined by one uninterrupted stocking period
with a large “paddock” with no control of cattle movements across the pasture. Eight mature
beef cows were randomly assigned to each experimental unit in early May 2018 to impose
stocking treatments and grazing continued through October of each year. The seasonal stocking
duration for rotational treatments was the same as the continuous treatment, ~ 200 d.
21
NWSG + wildflower plot establishment
Establishment of the NWSG + wildflower paddocks was initiated during the fall of 2016.
The designated paddock was sprayed with Roundup (Bayer USA, Parsippany, NJ) to remove
existing cool-season pasture grasses and seeded with a barley cover crop. The cover crop was
burned down with Roundup in April 2017 and residue was left on the plots. Plots were then
seeded with a mix of three NWSGs and 15 wildflower species (Table 1) at a rate of 13.5 kg/ha in
early June 2017. The seed mix was 70:30 (by seed weight) of NWSGs: wildflowers. Seeds were
planted using a Truax no-till seed drill (Truax Company, New Hope, MN) specifically designed
for NWSGs and wildflowers. Seeds were obtained from Ernst Seeds (Meadville, PA) and
wildflower species were selected based on USDA-NRCS recommendations and resistance to the
herbicide imazapic.
22
Table 1: NWSGs and wildflowers seeded at SVAREC in May 2017. Flower species were selected based on their resistance to the herbicide
imazapic. The mix was planted at a rate of 12 lb/acre (13.5 kg/ha) and is a 70:30 ratio of NWSGs to wildflowers based on seed weight.
Recommended seeding rates and seeds per pound were based on information in the Virginia Plant Establishment Guide (Hall et al. 2011) and
information from Ernst Conservation Seeds, Inc. (Meadville, PA).
Scientific name Common name Mix proportion Seeds per lb lb/acre (kg/ha) seeding rate Seeds per acre
Andropogon gerardii Big bluestem 0.17 165,000 2.040 (2.295) 336,600
Schizachyrium scoparium Little bluestem 0.35 175,000 4.200 (4.725) 735,000
Sorghastrum nutans Indiangrass 0.13 260,000 1.560 (1.755) 405,600
Chamaecrista fasciculata Partridge pea 0.04 65,000 0.480 (0.540) 31,200
Baptisia australis Blue false indigo 0.01 25,600 0.120 (0.135) 3,072
Desmanthus illinoensis Illinois bundleflower 0.02 85,000 0.240 (0.270) 20,400
Rudbeckia hirta Black-eyed susan 0.03 1,575,760 0.360 (0.405) 567,274
Chrysanthemum leucanthemum Oxeye daisy 0.02 340,000 0.240 (0.270) 81,600
Chrysanthemum maximum Shasta daisy 0.02 436,000 0.240 (0.270) 104,640
Coreopsis lanceolata Lanceleaf coreopsis 0.03 221,000 0.360 (0.405) 79,560
Echinacea purpurea Purple coneflower 0.04 115,664 0.480 (0.540) 55,519
Ratibida pinnata Grey-headed coneflower 0.02 427,500 0.240 (0.270) 102,600
Gaillardia pulchella Annual gaillardia 0.03 238,144 0.360 (0.405) 85,732
Gaillardia aristata Common gaillardia 0.03 186,000 0.360 (0.405) 66,960
Linum perenne Perennial blueflax 0.04 328,000 0.480 (0.540) 157,440
Agastache foeniculum Anise hyssop 0.004 1,440,000 0.048 (0.054) 69,120
Liatris spicata Marsh blazingstar 0.01 100,000 0.120 (0.135) 12,000
Helianthus maximiliani Maximilian sunflower 0.03 196,360 0.360 (0.405) 70,690
23
24
Measurements
Plant species composition
Percent cover of plant species was estimated using a modified Daubenmire
method (Daubenmire 1959). Percent cover of planted wildflower and NWSG species,
weedy species, and bare ground was visually estimated in 0.25-m2 square quadrats in
NWSG + wildflower pastures on May 23rd and July 16th in 2018 and on June 16th and
August 21st in 2019. Weedy species included all species that were not sown in the
mixtures, such as cool-season grasses, clovers, and common pasture weeds. Percent
cover of cool-season grasses, clovers, weedy species, and bare ground was visually
estimated in rotationally and continuously stocked cool-season grass pastures in August
and September 2018 and June and August 2019. Measurements were taken at 10
randomly selected points in the interior (~10 m away from the fence line) of each
experimental unit.
Bee observations and collections
Bee observations were used to gather information on which flower species were
the most attractive to bees. Beginning in May 2018 and 2019, bee landings on all
blooming flower species, including weedy species, in all treatments were observed. After
assessing the flowers blooming on a given day, six random patches of each blooming
flower species were observed for one minute each. The size and bloom count of each
patch were visually estimated and recorded. Bees that landed on blooms were divided
into five easily identifiable morphological groups: honeybees, bumble and carpenter bees,
small bees, green bees, and other large bees (Winfree et al. 2007). In the NWSG +
wildflower pastures, observations continued bi-weekly until late September to capture the
25
bloom period of all flowers planted in the mix. In both treatments containing cool-season
grasses, observations were performed once per month, beginning in May and continuing
through September. Weeds blooming in cool-season grass pastures did not change
frequently, so monthly observations were adequate.
Bees were collected in each treatment to assess differences among bee
communities. Collections began in June 2018 and 2019 and continued monthly until
September. Two sets of traps were placed in each experimental unit consisting of one
blue vane trap (SpringStar, Inc., Woodinville, WA, USA) and one each of yellow, blue,
and white bee bowls. Bee bowls that were four inches in diameter were painted with
fluorescent blue, yellow, and white paint (Guerra Paint and Pigment, New York City,
NY, USA) according to methods described by Droege (2015). Traps were filled
approximately half-way with soapy water and positioned in each pasture treatment for 24
hours, after which traps were collected and bees frozen before processing. Bees were
then washed, dried, and pinned following guidelines in Droege (2015). Bees were
identified to genus and species by Sam Droege, wildlife biologist at the USGS Patuxent
Wildlife Research Center (Laurel, MD).
Data analysis
Statistical analysis was performed in R version 3.5.1 (R, 2018). Plant community
data was summarized and graphed using ggplot2. The bee observation data used to
gauge flower attractiveness was converted to presence/absence data to minimize noise.
Individual years were analyzed separately. A global ANOVA with month, flower, and a
month by flower interaction effect was run for each year. There was a significant
interaction effect, so each month was analyzed separately. Tukey’s mean separation (α =
26
0.1) was performed if results were significant. The vegan package was used to analyze
the bee community collected during trapping. A Bray-Curtis distance matrix was
calculated from the bee species data and was visualized using non-metric multi-
dimensional scaling (NMDS). A permutational MANOVA (PERMANOVA) was used to
test for dispersion effects among date and treatment within the Bray-Curtis distance
matrix. Trapped bee abundance data were analyzed by each month. A generalized linear
model with a Poisson distribution was used to test for differences among treatments.
Tukey’s pairwise comparisons (α = 0.05) were performed if results were significant. The
Shannon index and the Simpson index were calculated each year for bee community data.
A linear mixed effect model with date as a random effect was used to test for the
significance of treatment on Shannon and Simpson indices. Tukey’s pairwise
comparisons were conducted if results were significant.
27
Results
Establishment of native warm-season grass + wildflower mix
Overall plant species composition changed substantially from 2018 to 2019.
Average percent cover of bare ground was similar in 2018 (14%) and 2019 (15%) (Figure
2). The average cover of weedy species (species not sown as part of the NWSG +
wildflower mix, such as thistle, white clover, fleabane, and cool-season grasses)
increased from 41% in 2018 to 63% in 2019. Coverage of all three NWSG species was
minimal in both years. Black-eyed susan was the dominant wildflower species in 2018,
representing an average of 22% of total plant coverage. However, in 2019, it represented
less than one percent of average percent cover. Lanceleaf coreopsis, perennial blueflax,
and anise hyssop showed similar trends, all declining to less than one percent cover in
2019. In contrast, the coverage of grey-headed coneflower increased from 2018 to 2019.
In 2018, grey-headed coneflower was on average 6% of the plant community and
increased to 14% in 2019.
28
Figure 2: Average percent cover of bare ground, weedy species, and established NWSG
and wildflower species in 2018 (blue bars) and 2019 (red bars).
Although the NWSG + wildflower mix consisted of 18 sown species, only 11
were present in 2018 (Figure 2). Oxeye daisy occurred infrequently and was not sampled
in 2018 plant community surveys. It was present in enough quantity to use for bee
observations. At the start of the growing season in May 2018, most species established
disproportionately to the amount planted in the seed mix (Figure 3). Big bluestem
accounted for 17% of the original seed mix by weight, but only 0.5% of the pasture
consisted of this species. Indiangrass and little bluestem were 13% and 35%,
respectively, of the original seed mix. In 2018, indiangrass composed only 0.7% of the
pastures and little bluestem was not sampled. In contrast, several wildflower species
were observed in larger amounts than what the sown mix predicted. For example, black-
eyed susan accounted for 3% of the seed mix but covered 23% of the pasture. Lanceleaf
coreopsis was 2% of the initial seed mix and accounted for 8% of the plant community in
29
2018. Six wildflower species were not present in any measurable amounts in May 2018,
including partridge pea, blue false indigo, shasta daisy, indian blanket, blanketflower, and
marsh blazingstar.
Figure 3: Percent by seed weight of sown versus percent frequency of occurrence of
established native warm-season grasses and wildflowers in May 2018. Green bars
represent the percentage of the seed mix represented by each wildflower species. Orange
bars represent the frequency of occurrence of a wildflower species in May 2018.
In 2019, 10 sown species were observed (Figure 2). Perennial blueflax and oxeye
daisy were not sampled in 2019 plant community surveys, although both were present in
sufficient quantities for bee observations. Like 2018, the sown species abundance did not
reflect the seed mix in 2019 (Figure 4). Big bluestem composed 2% of the pasture in
June 2019, an increase from 2018. Indiangrass composed 0.2% of the pasture in 2019, a
slight decrease from 2018. Little bluestem was not sampled during the first sampling
30
date in 2019. There were more drastic changes in the wildflower community early in the
2019 growing season. Black-eyed susan decreased in coverage to 1%. Grey-headed
coneflower became the dominant wildflower, composing an average of 16% of the
pasture plant community. In June 2019, the same six wildflower species (partridge pea,
blue false indigo, shasta daisy, indian blanket, blanketflower, and marsh blazingstar)
were absent from the pastures as in 2018.
Figure 4: Percent by seed weight of sown versus percent frequency of occurrence of
established native warm-season grasses and wildflowers in June 2019. Green bars
represent the percentage of the seed mix represented by each wildflower species. Orange
bars represent the frequency of occurrence of a wildflower species in June 2019.
31
Bee observations
Attractiveness of blooming flower species
In the 2018 and 2019 growing seasons, 17 plant species, both sown and weedy,
were observed. A total of 442 plants in 2018 and 312 plants in 2019 were observed to
record bee visitations. Among sown species, anise hyssop was visited the most
frequently by bees in 2018, while grey-headed coneflower was visited the most often in
2019 (Table 2). Among weedy species, milkweed (Asclepias syriaca L.) and pokeweed
(Phytolacca americana L.) attracted the most bees in 2018 and 2019, respectively. Bull
thistle (Cirsium vulgare (Savi) Ten), anise hyssop, maximilian sunflower, and purple
coneflower were consistently attractive in both years.
32
Table 2: Results of bee observations from 2018 and 2019. Values are means that
represent proportion of observations when a bee was present on a bloom (± 1 SE). The
planted and weedy species where a bee was observed the most frequently in both years
are bolded.
Flower Type Observations
2018 2019
Anise hyssop Sown 0.64 ± 0.07 0.42 ± 0.15
Black-eyed susan Sown 0.37 ± 0.09 0.44 ± 0.12
Grey-headed coneflower Sown 0.33 ± 0.11 0.61 ± 0.12
Lanceleaf coreopsis Sown 0.43 ± 0.11 0.33 ± 0.14
Maximilian sunflower Sown 0.43 ± 0.09 0.50 ± 0.15
Oxeye daisy Sown 0.25 ± 0.09 0.17 ± 0.11
Perennial blueflax Sown 0.31 ± 0.08 0.17 ± 0.17
Purple coneflower Sown 0.55 ± 0.09 0.42 ± 0.10
Buttercup Weedy 0.08 ± 0.04 0.08 ± 0.08
Chicory Weedy 0.67 ± 0.21 0.17 ± 0.11
Fleabane Weedy 0.18 ± 0.12 0.13 ± 0.06
Horsenettle Weedy 0.17 ± 0.11 0.16 ± 0.11
Milkweed Weedy 0.83 ± 0.11 0.50 ± 0.22
Pokeweed Weedy 0.33 ± 0.14 0.67 ± 0.21
Queen Anne’s lace Weedy 0.33 ± 0.21 0.22 ± 0.10
Thistle Weedy 0.56 ± 0.06 0.61 ± 0.07
White clover Weedy 0.26 ± 0.06 0.19 ± 0.06
33
Bee observations in 2018
In 2018, date (F = 16.2; df = 4; P < 0.001) and type of flower (F = 2.3; df = 16; P
= 0.003) both explained a significant amount of the variability in flower visitation.
However, there was a significant interaction effect (F = 2.5; df = 21; P < 0.001), so each
sampling date was analyzed separately.
A total of 100 plants from five different blooming species were observed in May
2018. Three species (lanceleaf coreopsis, oxeye daisy, and perennial blueflax) were
sown; two (buttercup (Ranunculus spp. L.) and white clover (Trifolium repens L.)) were
weedy. Flower species did significantly affect the proportion of observations where a bee
was observed (F = 6.2; df = 4; P < 0.001). Of the five flowers blooming, lanceleaf
coreopsis was significantly more attractive than the others, and bees were observed on it
during 100% of observations (Figure 5).
Twelve blooming species were observed in June 2018 (Figure 6A). Six species
were sown (anise hyssop, black-eyed susan, lanceleaf coreopsis, oxeye daisy, perennial
blueflax, and purple coneflower), and the remainder were weedy (fleabane (Erigeron
annuus L.), horsenettle (Solanum carolinense L.), milkweed, pokeweed, thistle, and
white clover). Ninety-nine plants were observed. In July 2018, 113 plants from twelve
species were blooming (Figure 6B). The same sown species that were observed in June
were observed in July, in addition to maximilian sunflower. Five weedy species were
observed: chicory (Cichorium intybus L.), fleabane, milkweed, thistle, and white clover.
In both June and July, observations indicated no difference in flower attractiveness
among species (P > 0.10) (Figure 6).
34
Figure 5: Proportion of observations where a bee was recorded on a blooming flower
species in May 2018. Sown wildflower species are represented by maroon bars and
weedy species are represented by orange bars. Differences in letters indicate significant
post-hoc Tukey’s pairwise comparisons (α = 0.1).
35
Figure 6: Proportion of observations where a bee was recorded on all blooming flower
species in June (A) and July (B) 2018. Sown wildflower species are represented by
maroon bars and weedy species are represented by orange bars. N.s. indicates that no
post-hoc Tukey’s pairwise comparisons were significantly different.
36
There were 102 plants observed from nine different flower species in August.
The sown species included anise hyssop, black-eyed susan, grey-headed coneflower,
maximilian sunflower, and purple coneflower. Pokeweed, Queen Anne’s lace (Daucus
carota L.), thistle, and white clover were also observed. The proportion of bees observed
was affected by flower species (F = 3.05; df = 8; P = 0.004). Of the nine flowers
blooming, anise hyssop was most visited, while black-eyed susan and pokeweed were the
least (Figure 7A). Only 23 plants from four species (anise hyssop, maximilian sunflower,
horsenettle, and white clover) were observed in September. No significant differences in
visitation rates were found in September, although anise hyssop continued to attract bees
(F = 2.16; df = 3; P = 0.13) (Figure 7B).
37
Figure 7: Proportion of observations where a bee was recorded on all blooming flower
species in August (A) and September (B) 2018. Sown wildflower species are represented
by maroon bars and weedy species are represented by orange bars. Differences in letters
indicate significant post-hoc Tukey’s pairwise comparisons (α = 0.1). N.s. indicates that
no post-hoc Tukey’s pairwise comparisons were significantly different.
38
Bee observations in 2019
During 2019, date (F = 5.9; df = 4; P < 0.001) and type of flower (F = 2.8; df =
16; P < 0.001) explained a significant (α = 0.1) amount of the variability in flower
visitation. There was no significant interaction effect (F = 0.66; df = 11; P = 0.77), but
months were analyzed separately to maintain consistency between years.
Bee landings on 41 plants of five flower species (oxeye daisy, perennial blueflax,
buttercup, fleabane, and white clover) were recorded in May 2019 (Figure 8A). There
were no significant differences in the attractiveness of each flower species (F = 0.15; df =
4; P = 0.96). Eighty-three plants were observed in June. Nine total species were
blooming, four of which were sown (black-eyed susan, lanceleaf coreopsis, oxeye daisy,
and purple coneflower) and five of which were weedy (fleabane, horsenettle, milkweed,
thistle, and white clover). Thistle was significantly more attractive than any other
blooming flower species, while the fewest number of bees were recorded on white clover
and fleabane (F = 2.0; df = 8; P = 0.06) (Figure 8B). Bees were present on milkweed and
black-eyed susan during half of observations.
39
Figure 8: Proportion of observations where a bee was recorded on all blooming flower
species in May (A) and June (B) 2019. Sown wildflower species are represented by
maroon bars and weedy species are represented by orange bars. Differences in letters
indicate significant post-hoc Tukey’s pairwise comparisons (α = 0.1). N.s. indicates that
no post-hoc Tukey’s pairwise comparisons were significantly different.
40
Figure 9: Proportion of observations where a bee was recorded on all blooming flower
species in July (A) and August (B) 2019. Sown wildflower species are represented by
maroon bars and weedy species are represented by orange bars. Differences in letters
indicate significant post-hoc Tukey’s pairwise comparisons (α = 0.1). N.s. indicates that
no post-hoc Tukey’s pairwise comparisons were significantly different.
The most flower species were blooming in July, when 125 plants were observed.
The majority of the blooming flower species were weedy and included chicory, fleabane,
horsenettle, pokeweed, Queen Anne’s lace, thistle, and white clover. The four blooming
sown species were anise hyssop, black-eyed susan, grey-headed coneflower, and purple
coneflower. Grey-headed coneflower and thistle attracted significantly more bees than
any of the other blooming flower species (F = 2.7; df = 10; P = 0.005) (Figure 9A). As
the growing season continued into August, bee visitations did not differ among the flower
species (F = 1.3; df = 5; P = 0.29) (Figure 9B). Six flower species were blooming,
41
including grey-headed coneflower, maximilian sunflower, purple coneflower, chicory,
Queen Anne’s lace, and thistle. A total of 53 plants were observed. Maximilian
sunflower was the only flower species blooming in September, and only six plants were
observed. Bees were recorded during 67% of observations.
Bee observations among pasture treatments
White clover and thistle were the only two pasture weeds consistently observed in
all three pasture treatments (Figure 10). Thistle was sparsely distributed, so was not
always recorded in percent cover measurements.
Figure 10: Average percent cover of forage grasses, bare ground, white clover, thistle,
and weedy species in 2018 (blue bars) and 2019 (red bars) in the rotationally stocked (A)
and continuously stocked (B) treatments.
42
Of the two weeds, white clover attracted the fewest number of bees in both 2018
and 2019. In 2018, there were no significant differences in the proportion of bees
observed on white clover among the three treatments (F = 0.07; df = 2; P = 0.93) (Table
3). White clover in the NWSG + wildflower treatment attracted the fewest bees, with
bees recorded during 20% of observations. The proportion of bees observed on white
clover in 2019 was significantly different among treatments (F = 3.1; df = 2; P = 0.06).
Unlike 2018, white clover in the NWSG + wildflower treatment attracted the most bees
in 2019. Bees were recorded during 33% of observations.
Thistle was more attractive than white clover in both years. It attracted more bees
in the continuously stocked and NWSG + wildflower treatments. Fewer bees were
observed on thistle than white clover in the rotationally stocked treatment, but that is
likely because fewer thistles were present in those pastures. In 2018, the treatment the
thistles were in had a nearly significant effect on the proportion of bees attracted to the
plant (F = 2.3; df = 2; P = 0.11). Bees were observed on thistle during 63% of
observations in the NWSG + wildflower treatment and during 58% of observations in the
continuously stocked treatment (Table 3). Treatment had a significant effect on the
proportion of bees attracted to thistle in 2019 (F = 3.8; df = 2; P = 0.03). Bees were
observed on thistle during 73% and 61% of observations in the NWSG + wildflower and
continuously stocked treatments, respectively. Like 2018, thistles in the rotationally
stocked treatment attracted the fewest number of bees of any treatment.
43
Table 3: Proportion white clover and thistle patches visited by bees during observations
in three different pasture management treatments in 2018 and 2019. Different letters
indicate significant differences in means between treatments within years and flower
species (α = 0.1).
White clover Thistle
2018 2019 2018 2019
Continuously stocked 0.27A 0.05B 0.58A 0.61A
Rotationally stocked 0.25A 0.08AB 0.17B 0.17B
NWSG + wildflower 0.20A 0.33A 0.63A 0.73A
Bee community data
Community similarity among bee genera
A total of 363 and 393 bee specimens were collected in 2018 and 2019,
respectively. There were 12 bee genera identified in 2018, including Agapostemon, Apis,
Augochlora, Bombus, Ceratina, Halictus, Hylaeus, Lasioglossum, Melissodes, Melitoma,
Osmia, Peponapis, and Svastra. In 2019, 17 bee genera were identified: Anthrophora,
Apis, Augochlora, Augochlorella, Augochloropsis, Bombus, Ceratina, Halictus,
Lasioglossum, Megachile, Melissodes, Melitoma, Melitta, Peponapis, and Xylocopa.
In 2018, the NMDS (k = 4, stress = 0.08) ordination showed no obvious variations
in the bee genera present in any treatments (Figure 11A). A PERMANOVA found that
bee genera present varied by date (P = 0.001). The most distinct cluster of bee genera is
the September sampling date, present in the lower left of the ordination (Figure 11B).
The cluster representing the bee genera present in August (Apis, Augochlora, Bombus,
Halictus, Hylaeus, Melissodes, Lasioglossum, and Peponapis) were the farthest apart
from the cluster of bee genera present in September (Agapostemon, Apis, Bombus,
Halictus, Hylaeus, Melissodes, Lasioglossum), indicating that the genera and quantity of
44
late-season bees foraging in September may differ drastically from the bees sampled a
month prior.
Figure 11: A non-metric multidimensional scaling (NMDS) (k = 4, stress = 0.08)
ordination of bee samples collected in 2018. Each point represents the bee community
collected in each experimental unit. The ordination is overlaid with treatment (A) and
date (B).
45
In 2019, the NMDS (k = 4, stress = 0.09) ordination did not show distinct clusters
of bee samples present in any treatment (Figure 12A). Like 2018, a PERMANOVA
showed that date significantly predicted the bee genera present in 2019 (Figure 12B).
Both June (Agapostemon, Apis, Augochlorella, Augochloropsis, Bombus, Ceratina,
Halictus, Lasioglossum, Melitta, Peponapis, and Xylocopa,) and September (Apis,
Bombus, Ceratina, Halictus, Lasioglossum, and Melissodes) had distinct clusters of bee
genera, as did July (Agapostemon, Andrena, Anthrophora, Augochlorella, Augochlora,
Bombus, Halictus, Lasioglossum, Megachile, Melitoma, and Melissodes) and August
(Agapostemon, Apis, Augochloropsis, Bombus, Halictus, Lasioglossum, Melissodes, and
Peponapis). The PERMANOVA confirmed that date explained a significant portion of
the variation in the bee genera present (P = 0.001). There was, however, a significant
interaction effect between date and treatment in the model (P = 0.043), indicating that any
differences among treatments depended on the date sampled.
46
Figure 12: An NMDS ordination (k = 4, stress = 0.09) of bee samples collected in 2019.
Each point represents the bee community collected in each experimental unit. The
ordination is overlaid with treatment (A) and date (B).
Community diversity
In 2018, bee abundance showed significant treatment by date interaction effect (P
< 0.001), so sampling dates were analyzed separately (Table 4). The NWSG +
wildflower treatment had the highest abundance of bees in all but the June sampling date.
In July and August, the NWSG + wildflower treatment had significantly more bees
present than the continuously and rotationally stocked treatments.
47
Bee abundance in 2019 differed by treatment (P = 0.02) and date (P < 0.001),
although there was no significant interaction effect (P = 0.14). The NWSG + wildflower
treatment again had the highest number of bees present in all but the June sampling date.
Although bees were numerically more abundant in the NWSG + wildflower treatment in
July, August, and September, the abundances were not statistically different from the
continuously and rotationally stocked treatments.
Table 4: The average abundance among three pasture treatments in each collection
month of 2018 and 2019. Letters indicate significant differences (α = 0.05) and represent
pairwise comparisons among treatments within months and years.
June July August September
2018 2019 2018 2019 2018 2019 2018 2019
Continuously
stocked
14.3A 6.7A 12.0B 11.7A 14.0B 12.3B 1.3A 5.3A
Rotationally
stocked
4.7B 6.3A 13.3B 14.3A 6.3C 13.0AB 2.7A 5.0A
NWSG +
wildflower
8.7AB 3.7A 23.3A 19.0A 24.0A 20.7A 3.3A 6.3A
Table 5: The average Shannon and Simpson index values calculated for genera among
three pasture treatments in 2018 and 2019. Letters indicate significant differences (α =
0.05) and represent pairwise comparisons among treatments within years.
Shannon Index Simpson Index
2018 2019 2018 2019
Continuously
stocked
0.73A 0.80B 0.40A 0.44B
Rotationally
stocked
1.09A 1.23A 0.60A 0.64A
NWSG +
wildflower
1.09A 1.23A 0.58A 0.67A
48
The Shannon and Simpson indices showed similar trends in 2018 and 2019. In
2018, the Shannon (P = 0.05) and Simpson (P = 0.03) indices varied significantly with
treatment. There were no significant differences among treatments, however. In 2018,
the Shannon index values were identical in the rotationally stocked and NWSG +
wildflower treatments and the Simpson index values in the two treatments differed very
little from one another. Treatment again explained a significant amount of the variability
in the Shannon (P = 0.005) and Simpson (P = 0.005) indices in 2019. Like 2018, the
Shannon index values in the rotationally stocked and NWSG + wildflower treatments
were identical, and the Simpson index values were almost equivalent. In contrast to
2018, the Shannon and Simpson indices were significantly lower in the continuously
stocked treatment in 2019. The differences are likely driven by higher richness in the
rotationally stocked and NWSG + wildflower treatments in 2019.
49
Species-level diversity in 2018
A total of 39 bee species were identified in 2018, almost all of which are native to
North America (Table 6). Fourteen species of Lasioglossum, a native genus of sweat
bees, were collected, making it the most common genus present. The European honeybee
(Apis mellifera) was the only non-native bee species collected. Of the 363 total
specimens, only 19 were honeybees, most of which were collected in the NWSG +
wildflower treatment.
The NWSG + wildflower treatment supported the most bees (n = 174) and bee
species (n = 28) in 2018. Eight species were unique to the NWSG + wildflower
treatment: Bombus auricomus, Bombus griseocolis, Hylaeus affinis/modestus, Hylaeus
mesillae, Lasioglossum callidum, Lasioglossum leucozonium, Lasioglossum pruinosum,
and Melissodes denticulatus. The rotationally stocked treatment supported the second-
most diverse community of bees (26 species), but only 78 total bees were collected.
There were five species of bee that were only found in the rotationally stocked treatment,
including Ceratina dupla, Lasioglossum bruneri, Lasioglossum quebecense, Melitoma
taurea, and Osmia bucephala. The continuously stocked treatment supported the fewest
bee species (22) and 111 total bees were collected. Only two species were unique to the
treatment: Bombus citrinus and Lasioglossum imitatum.
50
Table 6: Bee species identified from collections done in 2018. Values represent the
number of bee species collected in each grazing treatment in 2018. Species found only in
one treatment are bolded.
Species Continuously
stocked
Rotationally
stocked
NWSG + wildflower
Agapostemon texanus 1 1 3
Agapostemon virescens 16 4 5
Apis mellifera 3 2 14
Augochlora pura 1 2 0
Augochlora aurata 1 2 0
Bombus auricomus 0 0 1
Bombus bimaculatus 1 1 0
Bombus citrinus 1 0 0
Bombus fervidus 3 1 0
Bombus griseocollis 0 0 1
Bombus impatiens 4 0 9
Bombus pensylvanicus 3 4 3
Ceratina dupla 0 1 0
Halictus ligatus 5 11 28
Halictus rubicundus 4 11 4
Hylaeus affinis/modestus 0 0 2
Hylaeus mesillae 0 0 1
Lasioglossum admirandum 11 1 5
Lasioglossum albipenne 4 0 4
Lasioglossum bruneri 0 1 0
Lasioglossum callidum 0 0 4
Lasioglossum hitchensi 0 1 3
Lasioglossum imitatum 1 0 0
Lasioglossum leucocomum 0 1 2
Lasioglossum leucozonium 0 0 1
Lasioglossum oceanicum 1 0 6
Lasioglossum pilosum 2 5 11
Lasioglossum pruinosum 0 0 1
Lasioglossum quebecense 0 1 0
Lasioglossum trigeminum 13 10 20
Lasioglossum versatum 30 5 15
Melissodes bimaculatus 0 1 5
Melissodes denticulatus 0 0 1
Melissodes desponsus 1 2 2
51
Melissodes trinodis 3 2 13
Melitoma taurea 0 1 0
Osmia bucephala 0 1 0
Peponapis pruinosa 2 5 9
Svastra obliqua 0 1 1
Totals 111 78 174
Species richness 22 26 28
52
Discussion
Establishment of native warm-season grass + wildflower mix
Although NWSGs were planted as 70% of the seed mix, they composed less than
5% of the final plant community in both 2018 and 2019. During their first year of
establishment, NWSGs invest in root growth rather than above-ground biomass (Keyser
et al. 2011, 2012). They are easily outcompeted by other, faster growing grasses and
broadleaf plants (Harper et al. 2007). Lack of establishment in the pastures was likely
caused by competition from both the wildflowers included in the mix and weeds present
in pasture seedbanks. Because the NWSGs in our experiment were planted
simultaneously with wildflowers, and because they devote most of their resources to root
system development, they may not have had enough time to develop adequate above-
ground biomass to shade out competing wildflowers and weedy species. Grazing in the
NWSG + wildflower pastures may have also promoted weed competition and decreased
NWSG establishment success (Pywell et al. 2011, Bonin et al. 2014).
The initial presence of NWSGs in May 2018 and June 2019 was comparable to
the seed mix even though the percent cover was not. This further supports the hypothesis
that the NWSGs were outcompeted by wildflowers and weeds over the course of the
growing season. The grasses came up and were present early in the season but did not
have enough time to develop above- and below-ground biomass before wildflowers and
weeds closed the canopy. Other studies have found that NWSGs peak in productivity
three to four years after establishment (Bonin and Tracy 2012, Tracy and Bonin 2013,
Bonin et al. 2014). The NWSGs in the pastures may need more time to establish before
they reach their peak productivity.
53
The species composition of sown wildflowers present in the aboveground plant
community changed from 2018 to 2019. In 2018, the second year after planting, black-
eyed susan was the dominant wildflower. However, by 2019, it had decreased
substantially. Black-eyed susan can be both an annual and biennial (USDA-NRCS 2020)
and has been shown to dominate during the first and second years after establishment,
tapering off during the third year (Christiansen 1994, Bonin and Tracy 2012, Tracy and
Bonin 2013). Furthermore, in wildflower plantings, prior research has shown that
annuals are eventually replaced by perennials (De Cauwer et al. 2005). Grey-headed
coneflower, a perennial (USDA-NRCS 2020), was the dominant wildflower in 2019. Its
dominance is likely attributable to succession where, as perennials establish, they
outcompete and replace annuals. Grey-headed coneflower has also been shown to be
highly competitive, explaining its dominance in our plots (Knee and Thomas 2002,
Rischette and Norlan 2017).
Seven wildflower species did not establish successfully: blanketflower, blue false
indigo, Illinois bundleflower, indian blanket, marsh blazingstar, partridge pea, and shasta
daisy. Some of those species, including blanketflower, blue false indigo, indian blanket,
and shasta daisy, were likely outcompeted by more aggressive species. Some
wildflowers that were included in the mix were not well-suited to the habitat. Selecting
species that are well-suited to the habitat is a key factor in successfully establishing a
wildflower planting (Aldrich and Norcini 2006). For example, marsh blazingstar is most
often found in marshy environments and moist woods and prairies (Lady Bird Johnson
Wildflower Center 2020). The pasture sites used were on well-drained, upland soils and
did not hold enough water to create a suitable environment for marsh blazingstar. Other
54
wildflowers, like Illinois bundleflower and partridge pea, are leguminous forbs (USDA-
NRCS 2020). Illinois bundleflower has been shown to be selectively grazed by cattle
(Berg 1990), so what was present may have been eaten by the cattle that grazed our
pastures. Other studies have found that bundleflower does not persist long-term (Jackson
1999). In contrast, partridge pea is considered toxic to cattle (USDA-NRCS 2020), but
the seeds are commonly eaten by song- and ground-nesting birds (Lady Bird Johnson
Wildflower Center 2020). The partridge pea planted in our pastures could have failed to
re-seed and subsequently establish because of seed predation.
Bee observations
Attractiveness of blooming flower species
After observing bee landings on seventeen different blooming flower species, we
identified both sown wildflower species and weedy species that are attractive to bees. Of
the fifteen sown wildflower species, all eight that established were attractive. The bloom
periods of the sown species were staggered over the course of the growing season so that
floral resources were available from May through September, therefore maximizing bee
conservation value (Tuell et al. 2008). Most of the wildflower species bloomed during
June and July when other nectar and pollen resources are lacking (Koh et al. 2016, Heller
et al. 2019). If designing a diverse wildflower mix to plant in pastures, we would
recommend planting the eight species that established: lanceleaf coreopsis, perennial
blueflax, oxeye daisy, anise hyssop, black-eyed susan, purple coneflower, grey-headed
coneflower, and maximilian sunflower.
Several wildflower species, like anise hyssop, lanceleaf coreopsis, oxeye daisy,
perennial blueflax, and purple coneflower attracted fewer bees in 2019 than in 2018.
55
These wildflower species occurred less frequently in 2019 than 2018. Previous research
has shown that wildflower patches with minimal management lose plant diversity, flower
cover, and conservation value over time (Wesche et al. 2012, Walcher et al. 2019). Our
pastures had limited grazing or management, possibly explaining the decline in plant
diversity. Because several wildflower species were more sparsely distributed than the
year prior, they attracted fewer bees and their conservation value decreased. Other
wildflower species, like grey-headed coneflower, attracted more bees in 2019 than 2018.
Grey-headed coneflower was the most common wildflower species in 2019. During its
bloom period, it dominated the 1.8-ha pastures, making the pastures highly visible to
bees. Large wildflower patches (>1.5 ha) have been shown to attract greater numbers of
bees than small patches (Krimmer et al. 2019).
Weedy species also attracted bees. Thistle and milkweed were especially
attractive in both years. Thistles have been found to emit fragrances attractive to
pollinators (Theis 2006, Theis et al. 2007) and produce high-protein pollen that attracts
bees (Russo et al. 2019). Milkweeds are attractive to both native bees and honeybees
(Stoepler et al. 2012, James et al. 2016). While neither species is desirable in pastures as
they can reduce forage quality (Grekul and Bork 2004, USDA-ARS 2018), they are both
useful for bee conservation. If a farmer wants to conserve bees but is unable to convert
their pastures to a NWSG + wildflower mix, minimizing control of thistle and milkweed
may be an economically viable alternative. Several studies have shown that minimizing
herbicide applications in agricultural systems can increase bee habitat and diversity
(Bohan et al. 2005, Morandin and Winston 2005, Happe et al. 2018). The conservation
benefits of leaving milkweed in pastures extend beyond bees as well. Loss of milkweeds
56
in agricultural fields is thought to have severely negatively impacted monarch butterflies
(Pleasants and Oberhauser 2013). Leaving more milkweed in pastures could therefore
benefit monarch populations. While it may not be desirable to stop managing for thistle
and milkweed altogether in pastures, reducing herbicide applications and leaving some
plants could have significant insect conservation benefits.
Although we were able to discern some differences between the attractiveness of
flower species, there are limitations to our interpretations of the data. First, the presence
or absence of a bee on a flower was used as a proxy for attractiveness. However, there
were several factors that potentially affected attractiveness that were not included in
analyses because of how data were collected. The flowers were planted together in a mix
in the pasture and could therefore not be directly compared to one another. The
experiment was not designed as a choice experiment that could directly compare
attractiveness of two flower species. Second, bee species have different foraging habits
and can compete with one another for floral resources (Goulson et al. 2002b, Wojcik et
al. 2018). The use of a flower by one bee species may have therefore precluded a
different species from using it. This adds another layer of complexity to the data and
again prevents us from making clear comparisons of attractiveness among flower species.
Finally, the attractiveness of some flowers may have been inflated in our analyses. For
example, in May 2018, a bee was recorded on lanceleaf coreopsis during 100% of
observations. It is important to note that lanceleaf coreopsis was only observed three
times during that month, so the rate of visitation may be skewed because of a small
sample size. However, lanceleaf coreopsis is still a valuable resource for bees, and
planting it or other wildflowers in pastures is beneficial for bees.
57
Bee observations among pasture treatments
Thistle and white clover were the only two blooming flower species observed in
all treatments. Comparing attractiveness of these two weedy species among the three
treatments can give information about the overall bee conservation value of the habitat
context.
White clover, a common pasture weed, attracted very few bees, regardless of the
treatment. Other studies have found similarly low densities of bees on white clover
flowers. Two studies from New Zealand both reported observing fewer than 10 bees per
1000 white clover flowers (Palmer-Jones et al. 1962, Goodwin et al. 2011). Low
visitation densities could be because white clover is a low-growing legume with small
blooms (Pederson and Brink 2000). Every pasture treatment contained plants with tall
growth habits, and since white clover has small blooms and is low-growing, it may not
have been visible to bees flying overhead (Dafni et al. 1997). White clover did attract
slightly more bees in the NWSG + wildflower treatment, likely because there were other
blooming flowers in the vicinity. The increased visitation rates to white clover in the
NWSG + wildflower treatment indicate that the treatment has potential spillover benefits
to other plants that require insect pollination (Krimmer et al. 2019).
Thistle was highly attractive to bees no matter the pasture treatment. In cool-
season grass pastures, thistle was one of the few blooming weeds available. Thistle
plants are usually at least equivalent in height to surrounding grasses, and their blooms
contrast with the surrounding greenery, making them highly visible to bees. In the
NWSG + wildflower treatment, thistle was highly attractive likely because it was tall and
visible among the other blooming wildflowers. The proximity of other blooming flower
58
species also likely encouraged visitation to thistle plants. Thistle has also been found to
produce fragrances (Theis 2006, Theis et al. 2007) and high-protein pollen (Russo et al.
2019) that increases its attractiveness to pollinators. Overall, in all three treatments,
thistle was an important and highly visited resource. Although it is not desirable from a
forage production standpoint, reducing spot-spraying of herbicides and leaving even a
few thistle plants in a pasture could be a cost-effective and practical way to conserve
bees.
Bee community data
Abundance, community similarity, and diversity
In both years, pasture treatments had very little effect on the bee genera present.
The effect of treatment on bee community similarity and diversity may have been
reduced because of overlap between plant species present in each treatment. For
example, thistle was present in every treatment. Since the plant communities in the
NWSG + wildflower treatment and the two cool-season grass treatments were not totally
distinct, the effect of treatment on the bee community was likely reduced. Treatment
appears to affect bee abundances to some extent, although there was an interaction effect
with sampling date in 2018. The NWSG + wildflower treatment attracted higher
numbers of bees during July and August compared to other months in both years.
Pollinator abundance has been found to positively correlate with the number of
wildflower blooms (Blaauw and Isaacs 2014a, Angelella et al. 2019), which could
explain the higher bee abundances in the NWSG + wildflower treatment. Additionally,
the mid- to late summer months are when pollinators traditionally experience a pollen
dearth (Koh et al. 2016, Heller et al. 2019). The NWSG + wildflower pastures were
59
blooming and had pollen and nectar resources available, which may explain why more
bees were present during those dates.
Sampling date strongly influenced the diversity of bee genera sampled in this
study. Bee species vary in their emergence and peak activity times during the growing
season (Hofmann et al. 2019). Temporal food availability can also affect the number and
kinds of bees present (Persson et al. 2015). In 2018, the bee genera collected during the
September sampling date were distinctly different from the bee genera collected during
August. Peponapis, a genus of squash bees, usually goes dormant in September
(Mathewson 1968), so it was absent in September 2018 likely because it was nearing the
end of its active season. Peponapis was also absent in September 2019, helping
distinguish that sampling date from the others in 2019.
Despite using multiple passive trapping methods, sampling would have likely
been more accurate with the addition of an active sampling method like hand-netting.
The brightly colored passive traps were in stark contrast to the green and brown of the
grasses in the cool-season grass treatments. Because the traps stood out compared to the
surrounding foliage in those treatments, the number and diversity of bees captured in
those treatments may be falsely inflated. Bees flying overhead that may not have
otherwise landed in the cool-season grass pastures may have seen the traps and landed in
the treatment solely because they were attracted to the traps. Using active sampling
methods such as hand netting could have improved the accuracy of the bee community
sampled and would have allowed us to collect only bees that were foraging in the
treatment. Passive sampling methods, like the blue vane traps and bee bowls used in this
experiment, have been shown to be a less effective measure of the bee community than
60
active sampling methods, like hand netting (Nemésio and Vasconcelos 2014). Another
study found distinct bee communities when comparing blue vane traps and hand netting,
suggesting that that blue vane traps can complement, but not replace, hand netting (Gibbs
et al. 2017). Using a variety of passive and active sampling methods is the best way to
accurately represent the bee community (Rhoades et al. 2017), and would have improved
the accuracy of our sampling.
Species-level data in 2018
Of the 38 bee species identified in 2018, 37 were native. The only non-native
species was the European honeybee (A. mellifera). Nineteen honeybees were collected,
14 of which were in the NWSG + wildflower treatment. There were no known honeybee
hives within a 0.5 mile radius, which could have influenced the number of honeybees
observed.
The NWSG + wildflower treatment did have the most diverse community of bees,
supporting 28 total species. Eight of those species were found only in that treatment:
Bombus auricomus, Bombus griseocollis, Hylaeus affinis/modestus, Hylaeus mesillae,
Lasioglossum callidum, Lasioglossum leucozonium, Lasioglossum pruinosum, and
Melissodes denticulatus. A higher diversity of blooming flower species was available in
the NWSG + wildflower treatment, so the treatment probably supported a greater
diversity of bee species. Prior studies have found similar results. For example,
Papanikolaou et al. 2017 found that increased plant species richness and functional
diversity was positively correlated with bee functional diversity. Potts et al. 2003 showed
that bee species richness was closely related to floral diversity and nectar resource
diversity in a post-fire Mediterranean landscape. Similarly, in another Mediterranean
61
landscape, floral diversity and the variety of available nectar resources were found to be
key parameters for bee community structure (Potts et al. 2006).
62
Conclusions
Although cool-season grass pastures are generally lacking in floral resources,
management strategies can increase the abundance and diversity of bees present.
Planting a pasture mix designed for bee conservation, like the mix planted in our NWSG
+ wildflower treatment, can attract more bees and a greater diversity of bees than typical
cool-season pastures in this region. However, the pasture mix is not ideal if the
management goal is to improve summer forage availability for cattle along with resources
for pollinators. The wildflowers were present in much higher quantities than expected
and outcompeted the NWSGs. Since there were so few grasses, the pastures were not
viable for cattle production. Different planting methods, to be tested in future
experiments, could increase the proportion of grasses present while still maintaining
wildflowers, benefitting both cattle and bees.
Even cool-season grass pastures that were not oversown with wildflowers
contained useful floral resources for bees. Thistles and milkweed are two common
weedy species that are valuable for all pollinators, including bees. While neither weedy
species is desirable in a pasture, if a farmer can minimize spot-spraying to maintain a few
thistle and milkweed plants, bee populations in pasture systems could be enhanced with
minimal changes to pasture management. Overall, adding floral resources to pastures,
whether as sown wildflower species or as weeds, benefits bee populations and could help
combat widespread loss of bee habitat and subsequent bee decline.
63
References
Aldrich, J. H., and J. G. Norcini. 2006. Establishment of Native Wildflower Plantings by
Seed. University of Florida Extension ENH968.
Allen, V. G., C. Batello, E. J. Berretta, J. Hodgson, M. Kothmann, X. Li, J. McIvor, J.
Milne, C. Morris, A. Peeters, and M. Sanderson. 2011. An international terminology
for grazing lands and grazing animals. Grass and Forage Sciences:2–28.
Angelella, G. M., L. Stange, H. L. Scoggins, and M. E. O’Rourke. 2019. Pollinator
Refuge Establishment and Conservation Value: Impacts of Seedbed Preparations,
Seed Mixtures, and Herbicides. HortScience 54:445–451.
Backus, W. M., J. C. Waller, G. E. Bates, C. A. Harper, A. Saxton, D. W. Mcintosh, J.
Birckhead, and P. D. Keyser. 2017. Management of native warm-season grasses for
beef cattle and biomass production in the Mid-South USA. Journal of Animal
Science 95:3143–3153.
Bacon, C. W., J. K. Porter, J. D. Robbins, and E. S. Luttrell. 1977. Epichloe typhina from
toxic tall fescue grasses. Applied and Environmental Microbiology 34:576–581.
Balasubramanian, D., W. J. Zhou, H. L. Ji, J. Grace, X. L. Bai, Q. H. Song, Y. T. Liu, L.
Q. Sha, X. H. Fei, X. Zhang, J. Bin Zhao, J. F. Zhao, Z. H. Tan, and Y. P. Zhang.
2020. Environmental and management controls of soil carbon storage in grasslands
of southwestern China. Journal of Environmental Management 254.
Balbuena, M. S., L. Tison, M. L. Hahn, U. Greggers, R. Menzel, and W. M. Farina. 2015.
Effects of sublethal doses of glyphosate on honeybee navigation. Journal of
Experimental Biology 218:2799–2805.
Batista, A. C., C. E. da C. Domingues, M. J. Costa, and E. C. M. Silva-Zacarin. 2020. Is a
64
strobilurin fungicide capable of inducing histopathological effects on the midgut and
Malpighian tubules of honey bees? Journal of Apicultural Research.
Baude, M., W. E. Kunin, N. D. Boatman, S. Conyers, N. Davies, M. A. K. Gillespie, R.
D. Morton, S. M. Smart, and J. Memmott. 2016. Historical nectar assessment reveals
the fall and rise of floral resources in Britain. Nature 530:85–88.
Bayer. 2019. Bee Care. https://beecare.bayer.com/home.
Bee Informed Partnership. 2019. National Management Survey.
https://research.beeinformed.org/survey/.
Bendel, C. R., K. C. Kral-O’Brien, T. J. Hovick, R. F. Limb, and J. P. Harmon. 2019.
Plant-pollinator networks in grassland working landscapes reveal seasonal shifts in
network structure and composition. Ecosphere 10:e02569.
Bengtsson, J., J. M. Bullock, B. Egoh, C. Everson, T. Everson, T. O’Connor, P. J.
O’Farrell, H. G. Smith, and R. Lindborg. 2019. Grasslands-more important for
ecosystem services than you might think. Ecosphere 10:e02582.
Berg, W. 1990. Native forb establishment and persistence in a grass-forb seeding in the
southern Plains. Pages 179–182 Proceedings of the 12th North American Prairie
Conference.
Bhandari, K., C. West, S. Longing, C. Brown, P. Green, and E. Barkowsky. 2018.
Pollinator abundance in semiarid pastures as affected by forage species. Crop
Science 58:2665–2671.
Biesmeijer, J. C., S. P. M. Roberts, M. Reemer, R. Ohlemüller, M. Edwards, T. Peeters,
A. P. Schaffers, S. G. Potts, R. Kleukers, C. D. Thomas, J. Settele, and W. E. Kunin.
2006. Parallel declines in pollinators and insect-pollinated plants in Britain and the
65
Netherlands. Science 313:351–354.
Blaauw, B. R., and R. Isaacs. 2014a. Flower plantings increase wild bee abundance and
the pollination services provided to a pollination-dependent crop. Journal of Applied
Ecology 51:890–898.
Blaauw, B. R., and R. Isaacs. 2014b. Flower plantings increase wild bee abundance and
the pollination services provided to a pollination-dependent crop. Journal of Applied
Ecology 51:890–898.
Blaauw, B. R., and R. Isaacs. 2014c. Larger patches of diverse floral resources increase
insect pollinator density, diversity, and their pollination of native wildflowers. Basic
and Applied Ecology 15:701–711.
Bohan, D. A., C. W. H. Boffey, D. R. Brooks, S. J. Clark, A. M. Dewar, L. G. Firbank,
A. J. Haughton, C. Hawes, M. S. Heard, M. J. May, J. L. Osborne, J. N. Perry, P.
Rothery, D. B. Roy, R. J. Scott, G. R. Squire, I. P. Woiwod, and G. T. Champion.
2005. Effects on weed and invertebrate abundance and diversity of herbicide
management in genetically modified herbicide-tolerant winter-sown oilseed rape.
Proceedings of the Royal Society B: Biological Sciences 272:463–474.
Bonin, C. L., R. Lal, B. F. Tracy, C. L. Bonin, and B. F. Tracy. 2014. Evaluation of
Perennial Warm-Season Grass Mixtures Managed for Grazing or Biomass
Production. Crop Science 54:2373–2385.
Bonin, C. L., and B. F. Tracy. 2012. Diversity influences forage yield and stability in
perennial prairie plant mixtures. Agriculture Ecosystems & Environment 162:1–7.
Brown, R. H. 1999. Agronomic Implications of C 4 Photosynthesis. Pages 473–207 in R.
F. Sage and R. K. Monson, editors. C4 Plant Biology. Academic Press, San Diego,
66
CA.
Burkle, L. A., J. C. Marlin, and T. M. Knight. 2013. Plant-Pollinator Interactions over
120 Years: Loss of Species, Co-Occurrence, and Function. Science 339:1611–1615.
Burns, J. C., and D. S. Fisher. 2013. Steer performance and pasture productivity among
five perennial warm-season grasses. Agronomy Journal 105:113–123.
Button, L., and E. Elle. 2014. Wild bumble bees reduce pollination deficits in a crop
mostly visited by managed honey bees. Agriculture, Ecosystems and Environment
197:255–263.
Carneiro, L. S., L. C. Martínez, W. G. Gonçalves, L. M. Santana, and J. E. Serrão. 2020.
The fungicide iprodione affects midgut cells of non-target honey bee Apis mellifera
workers. Ecotoxicology and Environmental Safety 189.
De Cauwer, B., D. Reheul, K. D’hooghe, I. Nijs, and A. Milbau. 2005. Evolution of the
vegetation of mown field margins over their first 3 years. Agriculture, Ecosystems &
Environment 109:87–96.
Centrella, M., L. Russo, N. Moreno Ramírez, B. Eitzer, M. van Dyke, B. Danforth, and
K. Poveda. 2020. Diet diversity and pesticide risk mediate the negative effects of
land use change on solitary bee offspring production. Journal of Applied Ecology.
Cheerios. 2019. #Bring Back the Bees. https://www.cheerios.com/bring-back-the-bees/.
Christiansen, P. A. 1994. Establishment of prairie species by overseeding into burned
roadside vegetation. Pages 167–169 Proceedings of the 14th North American Prairie
Conference: Prairie Biodiversity.
Coblentz, W. K., K. P. Coffey, T. F. Smith, D. S. Hubbell, D. A. Scarbrough, J. B.
Humphry, B. C. McGinley, J. E. Turner, J. A. Jennings, C. P. West, M. P. Popp, D.
67
H. Hellwig, D. L. Kreider, and C. F. Rosenkrans. 2006. Using orchardgrass and
endophyte-free fescue versus endophyte-infected fescue overseeded on
bermudagrass for cow herds: I. Four-year summary of forage characteristics. Crop
Science 46:1919–1928.
Colla, S. R., M. C. Otterstatter, R. J. Gegear, and J. D. Thomson. 2006. Plight of the
bumble bee: Pathogen spillover from commercial to wild populations. Biological
Conservation 129:461–467.
Dafni, A., & Peter, and G. G. Kevan. 1997. Flower size and shape: implications in
pollination. Israel Journal of Plant Sciences 45:20–211.
Daubenmire, R. 1959. A canopy-coverage method of vegetational analysis. Northwest
Science 33:43–64.
Deru, J. G. C., J. Bloem, R. De Goede, H. Keidel, H. Kloen, M. Rutgers, J. Van Den
Akker, L. Brussaard, and N. Van Eekeren. 2017. Soil ecology and ecosystem
services of dairy and semi-natural grasslands on peat. Applied Soil Ecology 125:26–
34.
DiBartolomeis, M., S. Kegley, P. Mineau, R. Radford, and K. Klein. 2019. An
assessment of acute insecticide toxicity loading (AITL) of chemical pesticides used
on agricultural land in the United States. PLoS ONE 14.
Drewnoski, M. E., E. J. Oliphant, M. H. Poore, J. T. Green, and M. E. Hockett. 2009.
Growth and reproductive performance of beef heifers grazing endophyte-free,
endophyte-infected and novel endophyte-infected tall fescue. Livestock Science
125:254–260.
Droege, S. 2015. The Very Handy Manual : How to Catch and Identify Bees and Manage
68
a Collection. Beltsville, MD, USA.
Ekroos, J., A. M. ödman, G. K. S. Andersson, K. Birkhofer, L. Herbertsson, B. K. Klatt,
O. Olsson, P. A. Olsson, A. S. Persson, H. C. Prentice, M. Rundlöf, and H. G.
Smith. 2016. Sparing land for biodiversity at multiple spatial scales. Frontiers in
Ecology and Evolution 3:1–11.
Enri, S. R., M. Probo, A. Farruggia, L. Lanore, A. Blanchetete, and B. Dumont. 2017. A
biodiversity-friendly rotational grazing system enhancing flower-visiting insect
assemblages while maintaining animal and grassland productivity. Agriculture
Ecosystems & Environment 241:1–10.
Farina, W. M., M. S. Balbuena, L. T. Herbert, C. M. Goñalons, and D. E. Vázquez. 2019,
October 1. Effects of the herbicide glyphosate on honey bee sensory and cognitive
abilities: Individual impairments with implications for the hive. MDPI AG.
Feltham, H., K. Park, and D. Goulson. 2014. Field realistic doses of pesticide
imidacloprid reduce bumblebee pollen foraging efficiency. Ecotoxicology 23:317–
323.
Fine, J. D., C. A. Mullin, M. T. Frazier, and R. D. Reynolds. 2017. Field Residues and
Effects of the Insect Growth Regulator Novaluron and Its Major Co-Formulant N -
Methyl-2-Pyrrolidone on Honey Bee Reproduction and Development. Journal of
Economic Entomology 110:1993–2001.
Fontaine, C., I. Dajoz, J. Meriguet, and M. Loreau. 2006. Functional Diversity of Plant-
Pollinator Interaction Webs Enhances the Persistence of Plant Communities. PLoS
Biology 4:129–135.
Franceschinelli, E. V, M. A. S. Elias, L. L. Bergamini, · Carlos, M. Silva-Neto, and E. R.
69
Sujii. 2017. Influence of landscape context on the abundance of native bee
pollinators in tomato crops in Central Brazil. Journal of Insect Conservation 21:715–
726.
Frankie, G. W., R. W. Thorp, M. Schindler, J. Hernandez, B. Ertter, and M. Rizzardi.
2005. Ecological Patterns of Bees and Their Host Ornamental Flowers in Two
Northern California. Page Journal of the Kansas Entomological Society.
Gallai, N., J. M. Salles, J. Settele, and B. E. Vaissière. 2009. Economic valuation of the
vulnerability of world agriculture confronted with pollinator decline. Ecological
Economics 68:810–821.
Gibbs, J., N. K. Joshi, J. K. Wilson, N. L. Rothwell, K. Powers, M. Haas, L. Gut, D. J.
Biddinger, and R. Isaacs. 2017. Does passive sampling accurately reflect the bee
(apoidea: Anthophila) communities pollinating apple and sour cherry orchards?
Environmental Entomology 46:579–588.
Gill, K. A., R. Cox, and M. E. O’Neal. 2014. Quality Over Quantity: Buffer Strips Can be
Improved With Select Native Plant Species. Environmental Entomology 43:298–
311.
Giuliano, W. M., and S. E. Daves. 2002. Avian response to warm-season grass use in
pasture and hayfield management. Biological Conservation 106:1–9.
Goodwin, R., H. Cox, M. Taylor, L. Evans, and H. McBrydie. 2011. Number of honey
bee visits required to fully pollinate white clover (Trifolium repens) seed crops in
Canterbury, New Zealand. New Zealand Journal of Crop and Horticultural Science
39:7–19.
Goulson, D. 2013. An overview of the environmental risks posed by neonicotinoid
70
insecticides. Journal of Applied Ecology 50:977–987.
Goulson, D., M. E. Hanley, J. S. Ellis, and M. E. Knight. 2005. Causes of rarity in
bumblebees. Biological Conservation 122:1–8.
Goulson, D., W. O. . Hughes, L. C. Derwent, and J. C. Stout. 2002a. Colony growth of
the bumblebee, Bombus terrestris, in improved and conventional agricultural and
suburban habitats. Oecologia 130:267–273.
Goulson, D., E. Nicholls, C. Botias, and E. L. Rotheray. 2015. Bee declines driven by
combined stress from parasites, pesticides, and lack of flowers. Science
347:1255957–1255957.
Goulson, D., J. C. Stout, and A. R. Kells. 2002b. Do exotic bumblebees and honeybees
compete with native flower-visiting insects in Tasmania?
Grab, H., M. G. Branstetter, N. Amon, K. R. Urban-mead, M. G. Park, J. Gibbs, E. J.
Blitzer, K. Poveda, G. Loeb, and B. N. Danforth. 2019. Agriculturally dominated
landscapes reduce bee phylogenetic diversity and pollination services. Science
363:282–284.
Graystock, P., K. Yates, B. Darvill, D. Goulson, and W. O. H. Hughes. 2013. Emerging
dangers: Deadly effects of an emergent parasite in a new pollinator host. Journal of
Invertebrate Pathology 114:114–119.
Green, R. E., S. J. Cornell, J. P. W. Scharlemann, and A. Balmford. 2005. Farming and
the fate of wild nature. Science 307:550–555.
Greenleaf, S. S., and C. Kremen. 2006. Wild bees enhance honey bees’ pollination of
hybrid sunflower. Proceedings of the National Academy of Sciences of the United
States of America 103:13890–13895.
71
Gregorc, A., M. Alburaki, N. Rinderer, B. Sampson, P. R. Knight, S. Karim, and J.
Adamczyk. 2018a. Effects of coumaphos and imidacloprid on honey bee
(Hymenoptera: Apidae) lifespan and antioxidant gene regulations in laboratory
experiments. Scientific Reports 8.
Gregorc, A., M. Alburaki, B. Sampson, P. R. Knight, and J. Adamczyk. 2018b. Toxicity
of selected acaricides to honey bees (Apis mellifera) and varroa (varroa destructor
anderson and trueman) and their use in controlling varroa within honey bee colonies.
Insects 9.
Grekul, C. W., and E. W. Bork. 2004. Herbage Yield Losses in Perennial Pasture Due to
Canada Thistle ( Cirsium arvense ) . Weed Technology 18:784–794.
Hall, G., J. B. Daniel, B. Glennon, S. Gordon, and C. Wentz. 2011. Virginia Plant
Establishment Guide. USDA-NRCS.
Hallmann, C. A., M. Sorg, E. Jongejans, H. Siepel, N. Hofland, H. Schwan, W.
Stenmans, A. Mü Ller, H. Sumser, T. Hö Rren, D. Goulson, and H. De Kroon. 2017.
More than 75 percent decline over 27 years in total flying insect biomass in
protected areas.
Hannah, S. M., J. A. Paterson, J. E. Williams, M. S. Kerley, and J. L. Miner. 1990.
Effects of increasing dietary levels of endophyte-infected tall fescue seed on diet
digestibility and ruminal kinetics in sheep. Journal of animal science 68:1693–1701.
Hanski, I. 2007. Metapopulation dynamics. Nature 396:2005–2007.
Happe, A.-K., F. Riesch, V. Rösch, R. Gallé, T. Tscharntke, and P. Batáry. 2018. Small-
scale agricultural landscapes and organic management support wild bee
communities of cereal field boundaries. Agriculture, Ecosystems & Environment
72
254:92–98.
Harper, C. A., G. E. Bates, M. P. Hansbrough, M. J. Gudlin, J. P. Gruchy, and P. D.
Keyser. 2007. Native Warm-Season Grasses: Identification, Establishment and
Management for Wildlife and Forage Production in the Mid-South. University of
Tennessee Extension Institute of Agriculture, Knoxville, TN.
Harper, C. A., J. L. Birckhead, P. D. Keyser, J. C. Waller, M. M. Backus, G. E. Bates, E.
D. Holcomb, and J. M. Brooke. 2015. Avian habitat following grazing native warm-
season forages in the mid-south United States. Rangeland Ecology and Management
68:166–172.
Heller, S., N. K. Joshi, T. Leslie, E. G. Rajotte, and D. J. Biddinger. 2019. Diversified
Floral Resource Plantings Support Bee Communities after Apple Bloom in
Commercial Orchards. Scientific Reports 9.
Henry, M., M. Béguin, F. Requier, O. Rollin, J. F. Odoux, P. Aupinel, J. Aptel, S.
Tchamitchian, and A. Decourtye. 2012. A common pesticide decreases foraging
success and survival in honey bees. Science 336:348–350.
Hipólito, J., D. Boscolo, and B. F. Viana. 2018. Landscape and crop management
strategies to conserve pollination services and increase yields in tropical coffee
farms. Agriculture, Ecosystems and Environment 256:218–225.
Hofmann, M. M., C. M. Zohner, and S. S. Renner. 2019. Narrow habitat breadth and late-
summer emergence increases extinction vulnerability in Central European bees.
Proceedings of the Royal Society B 286:1–8.
Horváth, Z., R. Ptacnik, C. F. Vad, and J. M. Chase. 2019, June 1. Habitat loss over six
decades accelerates regional and local biodiversity loss via changing landscape
73
connectance. Blackwell Publishing Ltd.
Hoveland, C. S. 1993. Importance and economic significance of the Acremonium
endophytes to performance of animals and grass plant. Agriculture, Ecosystems &
Environment 44:3–12.
Hoveland, C. S. 2000. Achievements in Management and Utilization of Southern
Grasslands. Journal of Range Management 53:17.
Hoveland, C. S., R. L. Haaland, C. C. King, W. B. Anthony, E. M. Clark, J. A. McGuire,
L. A. Smith, H. W. Grimes, and J. L. Holliman. 1980. Association of Epichloe
Typhina Fungus and Steer Performance on Tall Fescue Pasture. Agronomy Journal
72:1064.
Isbell, F., D. Tilman, P. B. Reich, and A. T. Clark. 2019. Deficits of biodiversity and
productivity linger a century after agricultural abandonment. Nature Ecology and
Evolution 3:1533–1538.
Jackson, L. L. 1999. Establishing Tallgrass Prairie on Grazed Permanent Pasture in the
Upper Midwest. Restoration Ecology 7:127–138.
James, D., L. Seymour, G. Lauby, and K. Buckley. 2016. Beneficial Insect Attraction to
Milkweeds (Asclepias speciosa, Asclepias fascicularis) in Washington State, USA.
Insects 7:30.
Jarvis, B. 2018. The Insect Apocalypse is Here: What does it mean for the rest of life on
Earth? https://www.nytimes.com/2018/11/27/magazine/insect-apocalypse.html.
Johnson, S. D., L. F. Harris, and Ş. Procheş. 2009. Pollination and breeding systems of
selected wildflowers in a southern African grassland community. South African
Journal of Botany 75:630–645.
74
Jones, J. A., R. Hutchinson, A. Moldenke, V. Pfeiffer, E. Helderop, E. Thomas, J. Griffin,
and A. Reinholtz. 2019. Landscape patterns and diversity of meadow plants and
flower-visitors in a mountain landscape. Landscape Ecology 34:997–1014.
Keyser, P. D., E. D. Holcomb, C. M. Lituma, G. E. Bates, J. C. Waller, C. N. Boyer, and
J. T. Mulliniks. 2016. Forage Attributes and Animal Performance from Native Grass
Inter-Seeded with Red Clover. Agronomy Journal 108:373.
Keyser, P., C. Harper, and G. Bates. 2012. Competition Control in Native Warm-season
Grasses Grown for Livestock Forage in the Mid-South. University of Tennessee
Extension Bulletin SP731-E:1–12.
Keyser, P., C. Harper, G. Bates, J. Waller, and E. Doxon. 2011. Establishing Native
Warm-Season Grasses for Livestock Forage in the Mid-South. University of
Tennessee Extension Bulletin SP731-B:8.
Klein, A.-M., B. E. Vaissiere, J. H. Cane, I. Steffan-Dewenter, S. A. Cunningham, C.
Kremen, and T. Tscharntke. 2007. Importance of Pollinators in Changing
Landscapes for World Crops. Proceedings: Biological Sciences 274:303–313.
Knee, M., and L. C. Thomas. 2002. Light utilization and competition between Echinacea
purpurea , Panicum virgatum and Ratibida pinnata under greenhouse and field
conditions. Ecological Research 17:591–599.
Koh, I., E. V. Lonsdorf, N. M. Williams, C. Brittain, R. Isaacs, J. Gibbs, and T. H.
Ricketts. 2016. Modeling the status, trends, and impacts of wild bee abundance in
the United States. Proceedings of the National Academy of Sciences 113:140–145.
Kovacs-Hostyanszki, A., A. Espindola, A. J. Vanbergen, J. Settele, C. Kremen, and L. V
Dicks. 2017. Ecological intensification to mitigate impacts of conventional intensive
75
land use on pollinators and pollination. Ecology Letters 20:673–689.
Kremen, C., and L. K. M’gonigle. 2015. Small-scale restoration in intensive agricultural
landscapes supports more specialized and less mobile pollinator species.
Krimmer, E., E. A. Martin, J. Krauss, A. Holzschuh, and I. Steffan-Dewenter. 2019. Size,
age and surrounding semi-natural habitats modulate the effectiveness of flower-rich
agri-environment schemes to promote pollinator visitation in crop fields.
Agriculture, Ecosystems and Environment 284.
Kurve, V., P. Joseph, J. B. Williams, H. T. Boland, S. K. Riffell, T. Kim, and M. W.
Schilling. 2015. The effect of feeding native warm-season grasses during the stocker
phase on meat composition, quality characteristics, and sensory properties of loin
steaks from forage-finished cattle. Journal of Animal Science 93:2576–2586.
Lady Bird Johnson Wildflower Center. 2020. Plant Database.
https://www.wildflower.org/plants/result.php?id_plant=LISP.
Lazaro, A., T. Tscheulin, J. Devalez, G. Nakas, and T. Petanidou. 2016. Effects of
grazing intensity on pollinator abundance and diversity, and on pollination services.
Ecological Entomology 41:400–412.
Lentini, P. E., T. G. Martin, P. Gibbons, J. Fischer, and S. A. Cunningham. 2012.
Supporting wild pollinators in a temperate agricultural landscape: Maintaining
mosaics of natural features and production. Biological Conservation 149:84–92.
Liao, L. H., W. Y. Wu, A. Dad, and M. R. Berenbaum. 2019. Fungicide suppression of
flight performance in the honeybee (Apis mellifera) and its amelioration by
quercetin. Proceedings of the Royal Society B: Biological Sciences 286.
Main, A. R., E. B. Webb, K. W. Goyne, and D. Mengel. 2020. Reduced species richness
76
of native bees in field margins associated with neonicotinoid concentrations in non-
target soils. Agriculture, Ecosystems and Environment 287.
Di Marco, M., S. Ferrier, T. D. Harwood, A. J. Hoskins, and J. E. M. Watson. 2019,
September 26. Wilderness areas halve the extinction risk of terrestrial biodiversity.
Nature Publishing Group.
Maresh Nelson, S. B., J. J. Coon, W. H. Schacht, and J. R. Miller. 2019. Cattle select
against the invasive grass tall fescue in heterogeneous pastures managed with
prescribed fire. Grass and Forage Science 74:486–495.
Mathewson, J. A. 1968. Nest construction and life history of the Eastern Cucurbit Bee,
Peponapis pruinosa (Hymenoptera: Apoidea ). Journal of the Kansas Entomological
Society 41:255–261.
McFrederick, Q. S., and G. Lebuhn. 2006. Are urban parks refuges for bumble bees
Bombus spp. (Hymenoptera: Apidae)? Biological Conservation 129:372–382.
McGuire, J. L., J. J. Lawler, B. H. McRae, T. A. Nuñez, and D. M. Theobald. 2016.
Achieving climate connectivity in a fragmented landscape. Proceedings of the
National Academy of Sciences 113:7195–7200.
Michener, C. D. 2007. Bees of the World. 2nd edition. The Johns Hopkins University
Press, Baltimore, MD, USA.
Monroe, A. P., R. B. Chandler, L. W. Burger, and J. A. Martin. 2016. Converting exotic
forages to native warm-season grass can increase avian productivity in beef
production systems. Agriculture, Ecosystems and Environment 233:85–93.
Moore, K. J., T. A. White, R. L. Hintz, P. K. Patrick, and E. C. Brummer. 2004.
Sequential Grazing of Cool- and Warm-Season Pastures. Agronomy Journal
77
96:1103.
Morandin, L. A., and M. L. Winston. 2005. Wild bee abundance and seed production in
conventional, organic, and genetically modified canola. Ecological Applications
15:871–881.
Morandin, L. A., M. L. Winston, V. A. Abbott, and M. T. Franklin. 2007. Can
pastureland increase wild bee abundance in agriculturally intense areas? Basic and
Applied Ecology 8:117–124.
Moser, L. E., and C. S. Hoveland. 1996. Cool-Season Grass Overview. Pages 1–14 in L.
E. Moser, D. R. Buxton, and M. D. Casier, editors. Cool Season Forage Grasses.
Agronomy M. ASA, CSSA, SSSA, Madison. WI.
National Research Council. 2007. Status of Pollinators in North America. The National
Academies Press, Washington.
NCDOT. 2019. Pollinator Habitats. https://www.ncdot.gov/initiatives-
policies/environmental/wildflower/Pages/pollinator-habitats.aspx.
Neff, J. L., and B. B. Simpson. 1993. Bees, Pollination Systems and Plant Diversity.
Pages 143–167 Hymenoptera and Biodiversity. C.A.B. International.
Nelson, C. J., K. J. Moore, and M. Collins. 2017. Forages and Grasslands in a Changing
World. Pages 3–17 in M. Collins, C. J. Nelson, K. J. Moore, and R. F. Barnes,
editors. Forages, Volume 1: An Introduction to Grassland Agriculture. 7th edition.
John Wiley & Sons.
Nemésio, A., and H. L. Vasconcelos. 2014. Effectiveness of two sampling protocols to
survey orchid bees (Hymenoptera: Apidae) in the Neotropics. Journal of Insect
Conservation 18:197–202.
78
Neokosmidis, L., T. Tscheulin, J. Devalez, and T. Petanidou. 2018. Landscape spatial
configuration is a key driver of wild bee demographics. Insect Science 25:172–182.
ODOT. 2019. Roadside Wildflower Program. https://www.odot.org/beauty/wildflower/.
Ollerton, J., H. Erenler, M. Edwards, and R. Crockett. 2014. Extinctions of aculeate
pollinators in Britain and the role of large-scale agricultural changes. Science
346:1360–1362.
Orford, K. A., P. J. Murray, I. P. Vaughan, and J. Memmott. 2016. Modest enhancements
to conventional grassland diversity improve the provision of pollination services.
Journal of Applied Ecology 53:906–915.
Osborne, J. L., A. P. Martin, C. R. Shortall, A. D. Todd, D. Goulson, M. E. Knight, R. J.
Hale, and R. A. Sanderson. 2008. Quantifying and comparing bumblebee nest
densities in gardens and countryside habitats. Journal of Applied Ecology 45:784–
792.
Palmer-Jones, T., I. W. Forster, and G. L. Jeffery. 1962. Observations on the role of the
honey bee and bumble bee as pollinators of white clover (Trifolium repens Linn.) in
the Timaru district and Mackenzie country. New Zealand Journal of Agricultural
Research 5:318–325.
Papanikolaou, A. D., I. Kuhn, M. Frenzel, M. Kuhlmann, P. Poschlod, S. G. Potts, S. P.
M. Roberts, and O. Schweiger. 2017. Wild bee and floral diversity co-vary in
response to the direct and indirect impacts of land use. Ecosphere 8.
Paterson, C., K. Cottenie, and A. S. MacDougall. 2019. Restored native prairie supports
abundant and species-rich native bee communities on conventional farms.
Restoration Ecology 27:1291–1299.
79
Pavageau, C., C. Gaucherel, C. Garcia, and J. Ghazoul. 2018. Nesting sites of giant
honeybees modulated by landscape patterns. Journal of Applied Ecology 55:1230–
1240.
Pavlu, V., M. Hejcman, L. Pavlu, and J. Gaisler. 2003. Effect of rotational and
continuous grazing on vegetation of an upland grassland in the Jizerske Hory Mts.,
Czech Republic. Folia Geobotanica 38:21–34.
Pederson, G. A., and G. E. Brink. 2000. Seed production of white clover cultivars and
naturalized populations when grown in a pasture. Crop Science 40:1109–1114.
Persson, A. S., M. Rundlöf, Y. Clough, and H. G. Smith. 2015. Bumble bees show trait-
dependent vulnerability to landscape simplification. Biodiversity and Conservation
24:3469–3489.
Piccolomini, A. M., S. R. Whiten, M. L. Flenniken, K. M. O’Neill, and R. K. D.
Peterson. 2018. Acute Toxicity of Permethrin, Deltamethrin, and Etofenprox to the
Alfalfa Leafcutting Bee. Journal of Economic Entomology 111:1001–1005.
Pisa, L. W., V. Amaral-Rogers, L. P. Belzunces, J. M. Bonmatin, C. A. Downs, D.
Goulson, D. P. Kreutzweiser, C. Krupke, M. Liess, M. Mcfield, C. A. Morrissey, D.
A. Noome, J. Settele, N. Simon-Delso, J. D. Stark, J. P. Van Der Sluijs, H. Van
Dyck, and M. Wiemers. 2014. Effects of neonicotinoids and fipronil on non-target
invertebrates. Environmental Science and Pollution Research 22:68–102.
Pleasants, J. M., and K. S. Oberhauser. 2013. Milkweed loss in agricultural fields because
of herbicide use: Effect on the monarch butterfly population. Insect Conservation
and Diversity 6:135–144.
Pokhrel, V., N. A. DeLisi, R. G. Danka, T. W. Walker, J. A. Ottea, and K. B. Healy.
80
2018. Effects of truck-mounted, ultra low volume mosquito adulticides on honey
bees (Apis mellifera) in a suburban field setting. PLoS ONE 13.
Potts, S. G., J. C. Biesmeijer, C. Kremen, P. Neumann, O. Schweiger, and W. E. Kunin.
2010. Global pollinator declines: trends, impacts and drivers. Trends in Ecology &
Evolution 25:345–353.
Potts, S. G., V. Imperatriz-Fonseca, H. T. Ngo, M. A. Aizen, J. C. Biesmeijer, T. D.
Breeze, L. V. Dicks, L. A. Garibaldi, R. Hill, J. Settele, and A. J. Vanbergen. 2016,
December 8. Safeguarding pollinators and their values to human well-being. Nature
Publishing Group.
Potts, S. G., T. Petanidou, S. Roberts, C. O’Toole, A. Hulbert, and P. Willmer. 2006.
Plant-pollinator biodiversity and pollination services in a complex Mediterranean
landscape. Biological Conservation 129:519–529.
Potts, S. G., B. Vulliamy, A. Dafni, G. Ne’eman, and P. Willmer. 2003. Linking bees and
flowers: how do floral communities structure pollinator communities? Page
Ecology.
Powney, G. D., C. Carvell, M. Edwards, R. K. A. Morris, H. E. Roy, B. A. Woodcock,
and N. J. B. Isaac. 2019. Widespread losses of pollinating insects in Britain. Nature
Communications 10.
Pywell, R., J. Bullock, K. Walker, J. Tallowin, and E. Warman. 2011. Grazing effects on
wildflower establishment in restored grassland. Aspects of Applied Biology 108:37–
44.
Qi, S., X. Niu, D. hui Wang, C. Wang, L. Zhu, X. Xue, Z. Zhang, and L. Wu. 2020.
Flumethrin at sublethal concentrations induces stresses in adult honey bees (Apis
81
mellifera L.). Science of the Total Environment 700.
Rhoades, P., T. Griswold, L. Waits, N. A. Bosque-Pérez, C. M. Kennedy, and S. D.
Eigenbrode. 2017. Sampling technique affects detection of habitat factors
influencing wild bee communities. Journal of Insect Conservation 21:703–714.
Ricketts, T. H., J. Regetz, I. Steffan-Dewenter, S. A. Cunningham, C. Kremen, A.
Bogdanski, B. Gemmill-Herren, S. S. Greenleaf, A. M. Klein, M. M. Mayfield, L.
A. Morandin, A. Ochieng’, and B. F. Viana. 2008. Landscape effects on crop
pollination services: are there general patterns? Ecology Letters 11:499–515.
Rischette, A. C., and J. E. Norlan. 2017. Germination characteristics of a native non-
indigenous prairie forb in prairie plantings. Ecological Restoration 35:296–298.
Roberts, C., and J. Andrae. 2004. Tall Fescue Toxicosis and Management. Crop
Management 3:0.
Rowe, L., D. Gibson, D. Landis, J. Gibbs, and R. Isaacs. 2018. A comparison of drought-
tolerant prairie plants to support managed and wild bees in conservation programs.
Environmental Entomology 47:1128–1142.
Rundlöf, M., G. K. S. Andersson, R. Bommarco, I. Fries, V. Hederström, L. Herbertsson,
O. Jonsson, B. K. Klatt, T. R. Pedersen, J. Yourstone, and H. G. Smith. 2015. Seed
coating with a neonicotinoid insecticide negatively affects wild bees. Nature
521:77–80.
Rundlöf, M., A. S. Persson, H. G. Smith, and R. Bommarco. 2014. Late-season mass-
flowering red clover increases bumble bee queen and male densities. Biological
Conservation 172:138–145.
Russo, L., A. D. Vaudo, C. J. Fisher, C. M. Grozinger, and K. Shea. 2019. Bee
82
community preference for an invasive thistle associated with higher pollen protein
content. Oecologia 190:901–912.
Sanchez-Bayo, F., and K. Goka. 2014. Pesticide Residues and Bees – A Risk
Assessment. PLoS ONE 9:e94482.
Scheper, J., M. Reemer, R. Van Kats, W. A. Ozinga, G. T. J. Van Der Linden, J. H. J.
Schaminée, H. Siepel, and D. Kleijn. 2014. Museum specimens reveal loss of pollen
host plants as key factor driving wild bee decline in the Netherlands. Proceedings of
the National Academy of Sciences of the United States of America 111:17552–
17557.
Schmid-Hempel, R., M. Eckhardt, D. Goulson, D. Heinzmann, C. Lange, S. Plischuk, L.
R. Escudero, R. Salathé, J. J. Scriven, and P. Schmid-Hempel. 2014. The invasion of
southern South America by imported bumblebees and associated parasites. Journal
of Animal Ecology 83:823–837.
Schuenemann, G. M., M. E. Hockett, J. L. Edwards, N. R. Rohrbach, K. F. Breuel, and F.
N. Schrick. 2005. Embryo development and survival in beef cattle administered
ergotamine tartrate to simulate fescue toxicosis. Reproductive biology 5:137–50.
Scohier, A., A. Ouin, A. Farruggia, and B. Dumont. 2013. Is there a benefit of excluding
sheep from pastures at flowering peak on flower-visiting insect diversity? Journal of
Insect Conservation 17:287–294.
Seide, V. E., R. C. Bernardes, E. J. G. Pereira, and M. A. P. Lima. 2018. Glyphosate is
lethal and Cry toxins alter the development of the stingless bee Melipona
quadrifasciata. Environmental Pollution 243:1854–1860.
Senapathi, D., L. G. Carvalheiro, J. C. Biesmeijer, C. A. Dodson, R. L. Evans, M.
83
McKerchar, D. R. Morton, E. D. Moss, S. P. M. Roberts, W. E. Kunin, and S. G.
Potts. 2015. The impact of over 80 years of land cover changes on bee and wasp
pollinator communities in England. Proceedings of the Royal Society B 282.
Shapira, T., Z. Henkin, A. Dag, and Y. Mandelik. 2019. Rangeland sharing by cattle and
bees: moderate grazing does not impair bee communities and resource availability.
Ecological Applications.
Sirimarco, X., M. P. Barral, S. H. Villarino, and P. Laterra. 2018. Water regulation by
grasslands: A global meta-analysis. Ecohydrology 11.
Sjodin, N. E., J. Bengtsson, and B. Ekbom. 2008. The influence of grazing intensity and
landscape composition on the diversity and abundance of flower-visiting insects.
Journal of Applied Ecology 45:763–772.
Smith, G. W., D. M. Debinski, N. A. Scavo, C. J. Lange, J. T. Delaney, R. A. Moranz, J.
R. Miller, D. M. Engle, and A. L. Toth. 2016. Bee Abundance and Nutritional Status
in Relation to Grassland Management Practices in an Agricultural Landscape.
Environmental Entomology 45:338–347.
Staude, I. R., E. Vélez-Martin, B. O. Andrade, L. R. Podgaiski, I. I. Boldrini, M.
Mendonça, V. D. Pillar, and G. E. Overbeck. 2018. Local biodiversity erosion in
south Brazilian grasslands under moderate levels of landscape habitat loss. Journal
of Applied Ecology 55:1241–1251.
Stoepler, T. M., A. Edge, A. Steel, R. L. O’Quinn, and M. Fishbein. 2012. Differential
pollinator effectiveness and importance in a milkweed (Asclepias , Apocynaceae)
hybrid zone. American Journal of Botany 99:448–458.
Theis, N. 2006. Fragrance of Canada thistle (Cirsium arvense) attracts both floral
84
herbivores and pollinators. Journal of Chemical Ecology 32:917–927.
Theis, N., M. Lerdau, and R. A. Raguso. 2007. The challenge of attracting pollinators
while evading floral herbivores: Patterns of fragrance emission in Cirsium arvense
and Cirsium repandum (Asteraceae). International Journal of Plant Sciences
168:587–601.
Thompson, F. N., and J. A. Stuedemann. 1993. Pathophysiology of fescue toxicosis. Page
Ecosystems and Environment.
Thompson, S. E. D., R. A. Chisholm, and J. Rosindell. 2019. Characterising extinction
debt following habitat fragmentation using neutral theory. Ecology Letters 22:2087–
2096.
Tilhou, N. W., R. L. G. Nave, J. T. Mulliniks, and Z. D. McFarlane. 2019. Winter grazing
stockpiled native warm-season grasses in the Southeastern United States. Grass and
Forage Science 74:171–176.
Tomé, H. V. V., D. R. Schmehl, A. E. Wedde, R. S. M. Godoy, S. V. Ravaiano, R. N. C.
Guedes, G. F. Martins, and J. D. Ellis. 2020. Frequently encountered pesticides can
cause multiple disorders in developing worker honey bees. Environmental Pollution
256.
Tracy, B. F., and R. B. Bauer. 2019. Evaluating mob stocking for beef cattle in a
temperate grassland. PLOS ONE 14:e0226360.
Tracy, B. F., and C. L. Bonin. 2013. Yield potential of native warm-season grasses grown
in mixture. Virginia Cooperative Extension Publicatio.
Tracy, B. F., M. Maughan, N. Post, and D. B. Faulkner. 2010. Integrating Annual and
Perennial Warm-season Grasses in a Temperate Grazing System. Crop Science
85
50:2171.
Tracy, B. F., and M. A. Sanderson. 2000. Patterns of plant species richness in pasture
lands of the northeast United States. Plant Ecology 149:169–180.
Tuell, J. K., A. K. Fiedler, D. Landis, and R. Isaacs. 2008. Visitation by Wild and
Managed Bees (Hymenoptera: Apoidea) to Eastern U.S. Native Plants for Use in
Conservation Programs. Environmental Entomology 37:707–718.
USDA-ARS. 2018. Milkweed (Asclepias spp.). https://www.ars.usda.gov/pacific-west-
area/logan-ut/poisonous-plant-research/docs/milkweed-asclepias-spp/.
USDA-ERS. 2017. Major Land Uses. https://www.ers.usda.gov/data-products/major-
land-uses/.
USDA-NIFA. 2020. Glossary. https://nifa.usda.gov/glossary#G.
USDA-NRCS. 2019a. How NRCS is helping pollinators.
https://www.nrcs.usda.gov/wps/portal/nrcs/main/national/plantsanimals/pollinate/hel
p/.
USDA-NRCS. 2019b. Conservation Stewardship Program.
https://www.nrcs.usda.gov/wps/portal/nrcs/main/national/programs/financial/csp/.
USDA-NRCS. 2020. Plants Database. https://plants.sc.egov.usda.gov/java/.
USDA. 2017. U.S. National Level Data. Census of Agriculture.
USFWS. 2019. Rusty patched bumble bee (Bombus affinis).
https://ecos.fws.gov/ecp0/profile/speciesProfile?sId=9383.
Vanbergen, A. J., and Insect Pollinators Initiative. 2013. Threats to an ecosystem service:
Pressures on pollinators. Wiley Blackwell.
Vaughan, M., J. Hopwood, E. Lee-Mäder, M. Shepherd, C. Kremen, A. Stine, and S. H.
86
Black. 2015. Farming for Bees: Guidelines for Providing Native Bee Habitat on
Farms. Portland, OR.
Vaughan, M., and M. Skinner. 2015. Using 2014 farm bill programs for pollinator
conservation. Page Biology Technical Note, 2nd ed. Washington, DC.
VDOT. 2019. Pollinator Habitat Program.
https://www.virginiadot.org/programs/pollinator_habitat_program.asp.
Venturini, E. M., F. A. Drummond, A. K. Hoshide, A. C. Dibble, and L. B. Stack. 2017.
Pollination reservoirs for wild bee habitat enhancement in cropping systems: a
review. Agroecology and Sustainable Food Systems 41:101–142.
di Virgilio, A., S. A. Lambertucci, and J. M. Morales. 2019, November 1. Sustainable
grazing management in rangelands: Over a century searching for a silver bullet.
Elsevier B.V.
Vray, S., O. Rollin, P. Rasmont, M. Dufrêne, D. Michez, and N. Dendoncker. 2019. A
century of local changes in bumblebee communities and landscape composition in
Belgium. Journal of Insect Conservation 23:489–501.
Walcher, R., R. I. Hussain, L. Sachslehner, A. Bohner, I. Jernej, J. G. Zaller, A.
Arnberger, and T. Frank. 2019. Long-term abandonment of mountain meadows
affects bumblebees, true bugs and grasshoppers: A case study in the Austrian ALPS.
Applied Ecology and Environmental Research 17:5887–5908.
Walsh, E. M., S. Sweet, A. Knap, N. Ing, and J. Rangel. 2020. Queen honey bee (Apis
mellifera) pheromone and reproductive behavior are affected by pesticide exposure
during development. Behavioral Ecology and Sociobiology 74.
Wesche, K., B. Krause, H. Culmsee, and C. Leuschner. 2012. Fifty years of change in
87
Central European grassland vegetation: Large losses in species richness and animal-
pollinated plants. Biological Conservation 150:76–85.
Wick, A. F., B. A. Geaumont, K. K. Sedivec, and J. R. Hendrickson. 2016. Grassland
Degradation. Pages 257–276 Biological and Environmental Hazards, Risks, and
Disasters. Elsevier Inc.
Wilfert, L., G. Long, H. C. Leggett, P. Schmid-Hempel, R. Butlin, S. J. M. Martin, and
M. Boots. 2016. Deformed wing virus is a recent global epidemic in honeybees
driven by Varroa mites. Science 351:594–597.
Willi, Y., J. Van Buskirk, and A. A. Hoffmann. 2006. Limits to the Adaptive Potential of
Small Populations. Annual Review of Ecology, Evolution, and Systematics 37:433–
458.
Wilson, G. L., B. J. Dalzell, D. J. Mulla, T. Dogwiler, and P. M. Porter. 2014. Estimating
water quality effects of conservation practices and grazing land use scenarios.
Journal of Soil and Water Conservation 69:330–342.
Winfree, R., R. Aguilar, D. P. Vázquez, G. LeBuhn, and M. A. Aizen. 2009. A meta-
analysis of bees’ responses to anthropogenic disturbance. Ecology 90:2068–2076.
Winfree, R., J. R. Reilly, I. Bartomeus, D. P. Cariveau, N. M. Williams, and J. Gibbs.
2018. Species turnover promotes the importance of bee diversity for crop pollination
at regional scales. Science 359:791–793.
Winfree, R., N. M. Williams, J. Dushoff, and C. Kremen. 2007. Native bees provide
insurance against ongoing honey bee losses. Ecology Letters 10:1105–1113.
Wojcik, V. A., L. A. Morandin, L. Davies Adams, and K. E. Rourke. 2018. Floral
Resource Competition Between Honey Bees and Wild Bees: Is There Clear
88
Evidence and Can We Guide Management and Conservation? Environmental
Entomology 47:822–833.
Woodcock, B. A., N. J. B. Isaac, J. M. Bullock, D. B. Roy, D. G. Garthwaite, A. Crowe,
and R. F. Pywell. 2016. Impacts of neonicotinoid use on long-term population
changes in wild bees in England. Nature Communications 7:12459.
Woodcock, B. A., J. Savage, J. M. Bullock, M. Nowakowski, R. Orr, J. R. B. Tallowin,
and R. F. Pywell. 2014. Enhancing floral resources for pollinators in productive
agricultural grasslands. Biological Conservation 171:44–51.
Yang, E. C., Y. C. Chuang, Y. L. Chen, and L. H. Chang. 2008. Abnormal Foraging
Behavior Induced by Sublethal Dosage of Imidacloprid in the Honey Bee
(Hymenoptera: Apidae). Journal of Economic Entomology 101:1743–1748.