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

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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

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