THE INTERACTIVE EFFECTS OF DROUGHT AND PLANT INVASION ON PINUS

ELLIOTTII AND PINUS TAEDA

By

JULIENNE E. NESMITH

A THESIS PRESENTED TO THE GRADUATE SCHOOL

OF THE UNIVERSITY OF FLORIDA IN PARTIAL FULFILLMENT

OF THE REQUIREMENTS FOR THE DEGREE OF

MASTER OF SCIENCE

UNIVERSITY OF FLORIDA

2016

© 2016 Julienne E. NeSmith

To my grandmother, family, and friends for their longstanding support and encouragement

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ACKNOWLEDGMENTS

I thank members of the Flory Lab for helpful discussions and revisions on earlier versions

of the manuscript. Statistical support was provided in part by James Colee, a consultant with

University of Florida Institute of Food and Agricultural Sciences.

5

TABLE OF CONTENTS

page

ACKNOWLEDGMENTS ...............................................................................................................4

LIST OF TABLES ...........................................................................................................................7

LIST OF FIGURES .........................................................................................................................8

ABSTRACT ...................................................................................................................................10

CHAPTER

1 LITERATURE REVIEW .......................................................................................................12

Overview .................................................................................................................................12 Climate Change ......................................................................................................................15 Plant Invasions ........................................................................................................................17 Southeastern US Pine Forests .................................................................................................20 Climate Change: Effects on Southeastern Pine Forests ..........................................................23 Plant Invasions: Effects on Southeastern Pine Forests ...........................................................25 Research Needs .......................................................................................................................28

2 THE EFFECTS OF DROUGHT AND PLANT INVASION ON PINE SEEDLINGS .........32

Introduction .............................................................................................................................32 Materials and Methods ...........................................................................................................34 Study Species ..........................................................................................................................34 Experimental Design ..............................................................................................................35 Data Collection .......................................................................................................................36 Statistical Analysis ..................................................................................................................38

3 FINDINGS ..............................................................................................................................39

Results.....................................................................................................................................39 Discussion ...............................................................................................................................41

APPENDIX

A NUMBER OF BRANCHES ...................................................................................................54

B NUMBER OF WEBWORM NESTS .....................................................................................55

C RELATIONSHIPS BETWEEN SOIL VOLUMETRIC WATER CONTENT AND

PINE SEEDLING RESPONSE ..............................................................................................56

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D RELATIONSHIPS BETWEEN PHOTOSYNTHETICALLY ACTIVE RADIATION

AND PINE SEEDLING RESPONSE ....................................................................................57

E RELATIONSHIPS BETWEEN RESIDENT SPECIES COVER AND PINE

SEEDLING RESPONSE ........................................................................................................58

F RELATIONSHIPS BETWEEN COGONGRASS COVER AND PINE SEEDLING

RESPONSE ............................................................................................................................59

LIST OF REFERENCES ...............................................................................................................60

BIOGRAPHICAL SKETCH .........................................................................................................73

7

LIST OF TABLES

Table page

3-1 The scientific names and functional types of twelve native understory species planted

in 2013 and the most common resident species established in the plots by 2015. ............46

3-2 Results of mixed model ANOVAs testing the fixed effects of drought, invasion, and

their interaction on slash and loblolly pine survival, relative growth rate of height .........47

3-3 Mean and SE of final height, diameter, biomass, and survival of slash and loblolly

pine seedlings under drought and invasion treatments. .....................................................48

8

LIST OF FIGURES

Figure page

1-1 Conceptual diagram illustrating three potential scenarios for the magnitude of

independent (additive) or interactive (synergistic and offsetting) negative effects as a

result of multiple stressors acting on ecological communities. .........................................31

3-1 Mean ± SE of soil moisture (percent volumetric water content) averaged over 2015

a) and by month b) in plots exposed to ambient or drought conditions and with

resident species only or resident species invaded by Imperata cylindrica ........................49

3-2 Mean ± SE of light availability (photosynthetically active radiation) above the

vegetation canopy at 0.5 m and at ground level averaged over 2015 (a, b) and by

month (c, d) in plots exposed to drought and invasion treatments. ...................................49

3-3 Mean ± SE percent survival of slash a) and loblolly b) pine seedlings exposed to

drought and invasion treatments. .......................................................................................50

3-4 Mean ± SE of relative growth rates of height of slash a) and loblolly b) pine

seedlings exposed to drought and invasion treatments. .....................................................51

3-5 Mean ± SE of relative growth rates of diameter of slash a) and loblolly b) pine

seedlings under drought and invasion treatments. .............................................................52

3-6 Mean ± SE biomass of slash a) and loblolly b) pine seedlings grown under drought

and invasion treatments......................................................................................................53

A-1 Mean ± SE of number of limbs of slash a) and loblolly b) pine seedlings exposed to

drought and invasion treatments. .......................................................................................54

B-1 Count of pine webworm (Pococera robustella) nests on slash a) and loblolly b) pine

seedlings exposed to drought and invasion treatments. .....................................................55

C-1 Relationships between soil volumetric water content and slash (top) and loblolly

(bottom) pine seedling survival a), natural-log-transformed relative growth rates of

height b) and diameter c), and biomass d). ........................................................................56

D-1 Relationships between photosynthetically active radiation and slash (top) and

loblolly (bottom) pine seedling survival a), natural-log-transformed relative growth

rates of height b) and diameter c), and biomass d). ...........................................................57

E-1 Relationships between resident species cover and slash (top) and loblolly (bottom)

pine seedling survival a), natural-log-transformed relative growth rates of height b)

and diameter c), and biomass d). .......................................................................................58

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F-1 Relationships between cogongrass cover and slash (top) and loblolly (bottom) pine

seedling survival a), natural-log-transformed relative growth rates of height b) and

diameter c), and biomass d). ..............................................................................................59

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Abstract of Thesis Presented to the Graduate School

of the University of Florida in Partial Fulfillment of the

Requirements for the Degree of Master of Science

THE INTERACTIVE EFFECTS OF DROUGHT AND PLANT INVASION ON PINUS

ELLIOTTII AND PINUS TAEDA

By

Julienne E. NeSmith

December 2016

Chair: Stephen Luke Flory

Major: Interdisciplinary Ecology

Climate change and non-native invasive species are two predominant drivers of global

environmental change, yet little is known about how they might interact to affect native

communities and ecosystems. Drought and plant invasions are intensifying in forests worldwide,

including ecologically and economically important pine forests of the southeastern United States.

Together, these stressors may exert additive, synergistic, or offsetting effects on native species,

but such outcomes are difficult to predict. We used a factorial common garden experiment to

determine how simulated drought, invasion by Imperata cylindrica (cogongrass), and their

interaction affected seedling survival and performance (relative growth rates of height and

diameter, biomass) of two native pine species, Pinus elliottii var. densa (South Florida slash

pine) and Pinus taeda (loblolly pine). In general, loblolly pine outperformed slash pine over the

course of the experiment, but the magnitudes of each species’ responses to the treatments were

similar, with the two stressors often exhibiting additive negative effects on pine seedling

performance. Drought, but not invasion, was associated with lower relative growth rates in

height of both pine species while drought and invasion had an additive negative effect on

diameter for both species. Given the suppressive effects of drought demonstrated here and

projections for increased drought in the Southeast US under climate change, land managers must

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select appropriate field sites for plantations, and should pursue pine species and varieties with

improved drought tolerance. Moreover, our results demonstrate experimentally the dramatic

effects of cogongrass invasion on pine seedlings, supporting efforts of land-owners and property

managers to remove this noxious invasive species. To predict the outcome of drought and

invasion effects on forest stand dynamics, additional temporally and spatially robust measures of

the conditions that mediate pine responses to these stressors are needed.

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CHAPTER 1

LITERATURE REVIEW

Overview

Climate change (Scheffers et al., 2016) and non-native species invasions (D'Antonio and

Vitousek, 1992; Dukes and Mooney, 1999; Mack et al., 2000) are primary global change drivers

with complex implications for native ecosystems and diversity (Walther et al., 2002). Climate

change and biological invasions have been researched extensively, and independently these

drivers typically have negative ecological effects (Walther et al., 2002; Parmesan 2006; Pyšek et

al., 2012). For example, rising sea levels under climate change may alter the hydrology and

salinity of intertidal zones (Middleton and Souter 2016), thereby displacing native species.

Likewise, invasive plants, can enhance fire intensity and suppress native tree regeneration (Flory

et al., 2015). However, relatively little is known about how interactions between climate change

and invasive species may affect native species (Sala et al., 2000; Mora et al., 2007; Halpern et

al., 2008a, b). Additional research on how climate change and invasions may co-occur and affect

native plant and animal communities and ecosystem functions is important for predicting their

combined impacts and developing management plans to mitigate their effects (Burgiel and Hall

2014).

Stress in ecological systems is defined by any abiotic or biotic factor that affects resource

availability for species, thereby limiting individual physiology, performance, or survival, or

affects ecosystem processes or productivity (Hoffmann and Hercus, 2000; Freedman, 2016).

Stressors can be persistent or occur as distinct events but, by definition, their occurrence

ultimately harms or alters the structure and function of ecosystems (Borics et al., 2013;

Freedman 2016). Extreme temperatures (Pörtner and Knust, 2007; Wingfield, 2013), wind

(Wingfield, 2013), and precipitation events (Knapp et al., 2008) that result in too much or too

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little moisture (Freedman, 2016) are examples of weather and climatic stressors that can be

detrimental when their magnitude exceeds the limits of tolerance of an organism or ecosystem.

Biological stressors stem from natural (e.g., native insect outbreaks) or anthropogenic (e.g., non-

native species introductions) sources and can include trophic interactions such as herbivory,

damage from pests and pathogens, and competition from abundant and dominant invasive

species (Freedman, 2016). Most previous studies have concentrated on the independent effects of

abiotic and biotic stressors (Sala et al., 2000; Dávalos et al., 2014), yet interactions between them

are common in the environment. In addition, their combined effects potentially are complex and

difficult to predict (Folt et al., 1999; Dávalos et al., 2014; Ramegowda and Senthil-Kumar 2014;

Gassmann et al 2016).

There are three general possible outcomes for how a combination of biotic and abiotic

stressors may act together to alter native ecosystems. The first possibility is that the two stressors

may additively affect ecosystems such that their combination is equal in effect size to the sum of

how each stressor alone would have impacts (Figure 1a; Folt et al., 1999; Breitburg and Riedel

2005). For example, we might describe the effects of a simultaneous insect outbreak and drought

event on pine tree survival as additive if each stressor alone reduced survival by 10% and

together the sum of their effects reduced survival by 20%. Second, stressors may combine in a

way that results in a synergistic interaction and a stronger negative effect on native species than

if the two stressors occurred independently (Figure 1b; Breitburg and Riedel, 2005; Folt et al.,

1999). In this case, if the effects of an insect outbreak and drought on pine trees are greater than

each acting alone (30% reduction in survival instead of 20% as in the additive scenario) we

would describe their effects as synergistic. The third possibility is that the two stressors may act

on each other antagonistically so that one stressor offsets the effect of the other and their

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combination actually has less of an impact than the sum of each stressor occurring independently

(Figure 1c; Breitburg and Riedel, 2005; Folt et al., 1999). For example, a drought might suppress

insect populations, suppressing the effects of an outbreak such that pine tree survival is only

reduced by 15% instead of 20% (additive scenario) or 30% (synergistic scenario). Determining if

multiple stressors have additive, synergistic, or offsetting effects on native species and

ecosystems is critical for developing management responses but quantitative studies on multiple

stressors have thus far been lacking.

Interactions among abiotic and biotic stressors are prevalent in the environment and will

likely become more abundant in the future under climate change and globalization. Interactions

among multiple stressors are challenging to study because their effects can be difficult to isolate,

and often manifest across multiple ecological pathways, span a range of spatial and temporal

scales, and have several different outcomes (National Research Council, 2007; Pendleton et al.,

2016; Todgham and Stillman, 2013). In particular, it might be plausible to manipulate a single

abiotic factor, such as CO2 in a Free Air CO2 Enrichment experiment (Norby and Zak, 2011),

but simultaneously applying a second stressor to the system, such as a fungal pathogen or

invasive plant, can be logistically difficult and expensive (but see Belote et al., 2004). Therefore,

current studies on multiple stressors have been limited in scale and scope, for example, restricted

to greenhouse or laboratory conditions, brief time scales, or few species (Ramegowda and

Senthil-Kumar, 2014). Thus, to effectively quantify how changes in abiotic conditions under

climate change (e.g., drought) will interact with biotic stressors (e.g., plant invasions),

experiments that manipulate each stressor alone and in combination under field conditions are

needed (Pendleton et al., 2016).

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Climate Change

The Intergovernmental Panel on Climate Change defines climate change as “…a

statistically significant variation in either the mean state of the climate or in its variability,

persisting for an extended period (typically decades or longer).” These changes can be natural or

directly and indirectly caused by anthropogenic activity (IPCC, 2014; Scheffers et al., 2016). In

recent decades, the term climate change has become synonymous with human-induced climatic

change as the consensus of peer-reviewed scientists (97%) assert it is caused and accelerated by

human impacts on the environment, primarily through the use of fossil fuels and land use change

such as deforestation (Cook et al., 2016). The main effects of climate change are increases in

CO2 levels in the atmosphere (IPCC, 2014; Scheffers et al., 2016), changes in regional and

global temperature averages (Barker et al., 2007; IPCC, 2014), and shifts in the frequency and

intensity of precipitation events (IPCC, 2014), including more frequent and prolonged drought.

These patterns of changes in abiotic conditions are expected to increase in severity over time.

The most conservative estimates suggest that global mean CO2 levels will increase from

370 µmol mol−1, measured in 2001, to 540 µmol mol−1 by 2100, while the worst case scenarios

indicate that CO2 levels will exceed 970 µmol mol−1 by 2100 (Houghton et al., 2001; Nowak et

al., 2004). As a result, global mean temperatures are expected to increase by as much as 2°C to

4°C above pre-industrial levels by 2100, with corresponding effects on sea level rise and other

associated environmental changes (IPCC, 2014). However, the effects of climate change may be

reduced if adaptation and mitigation strategies are broadly implemented such as reducing the

production of CO2 or increasing carbon sequestration through changes in land use (IPCC, 2014).

As a whole, climate change is expected to significantly disturb native ecosystems and

alter their structure and function in complex ways (Mooney et al., 2009; Vose et al., 2012). For

example, climate change-induced shifts in temperature and precipitation regimes are creating

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novel abiotic conditions and pushing the limits of species ranges and ecological communities

across the globe (Scheffers et al., 2016). Higher global mean temperature (IPCC, 2007) is

reducing sea ice in the Arctic (Maslanik et al., 1996; Serreze et al., 2000) and lessening the

earth’s reflectivity (Perovich et al., 2007), leading to forest expansion northward into the tundra

(Hassol, 2004). Additionally, increases in the extent and severity of drought in some regions

(Houghton et al., 2001; Easterling et al., 2000; Hoerling and Kumar 2004) are resulting in native

species die-off (Allen et al., 2010; Breshears et al., 2005, Clark et al., 2016; Thomas et al., 2004,

Vose et al., 2012) and displacement (Clark et al., 2016; Lenoir et al., 2008). A recent meta-

analysis found that species distributions have shifted to higher elevations at a rate of over 12 m

per decade and to higher latitudes at a rate of nearly 17 km per decade (Chen et al., 2011). A

review by Scheffers et al. (2016) found supporting evidence of range shifts of North American

plant species (Wolf et al., 2016), mountainous stream-dwelling fish in France (Comte and

Grenouillet, 2013), and insects in Borneo (Chen et al., 2009). Collectively, the effects of climate

change on native species and ecosystems will likely be severe, with particularly strong effects on

forest systems.

Climate is the primary force in shaping biomes across the globe, including the

distribution, diversity, and functions of forests (Prentice, 1990; Hansen et al., 2001). Changes in

climate will invariably impact the extent of most biomes and the species distributions, biological

diversity and species composition, and the structure and function of ecosystems within them

(IPCC, 2007; Secretariat of the Convention on Biological Diversity, 2010; Scheffers et al.,

2016). It has been estimated that more than half of terrestrial flora and fauna species reside in

forests and tropical rain forests in particular, which contain perhaps the highest biodiversity in

the world (Seppala et al., 2009). Forest biomes cover approximately 31% of global land area and

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account for more than two-thirds of net primary production by all terrestrial ecosystems

(Secretariat of the Convention on Biological Diversity, 2010). Alteration of forest biomes by

natural or anthropogenic disturbances can be particularly alarming in part because of such high

biodiversity levels, the length of time required for mature tree stands to develop, and the reliance

of humans on ‘forest goods and services’ such as food, carbon sequestration, timber products,

and non-timber forest products (MEA, 2005). Because newly planted forests take a long time to

develop, may incorporate only a single tree species, and often have lower biodiversity values

than mature forests (Secretariat of the Convention on Biological Diversity, 2010), the

consequences of forest loss can be both immediate, and also span much longer-term time frames.

Some immediate threats to the integrity, diversity, function, and productivity of forests include

changes in climate such as the intensity, frequency, and duration of precipitation events,

windstorms, and drought, as well as other disturbances such as fire, insect pest and pathogen

outbreaks, and invasions of non-native plant species (Dale et al., 2001).

Plant Invasions

Non-native invasive species are a primary biotic stressor that can greatly impact native

ecosystems (Mack et al., 2000; NISC 2014). Plant invasions are characterized by plant

populations that have established and spread widely outside of their native range (Richardson,

2000) and subsequently suppress or alter the function or biodiversity of the ecosystem in which

they’ve been introduced (Vitousek et al., 1996; Wilcove et al., 1998; Executive Order 13112,

1999; Simberloff, 2000; National Research Council, 2007). By definition, invasive species are

the result of transportation of propagules by humans outside of their native ranges. For example,

non-native plants have been intentionally introduced for use as feed crops, horticulture, livestock

forage, ornamentals in landscaping, wildlife food sources, and erosion control (MacDonald,

2004; Lockwood et al., 2007). Plant species also have been accidentally introduced as

18

contaminants in shipping containers, with livestock, as ballast, and by various other means. Both

intentional and unintentional introductions have increased substantially with advances in

technology and globalization of travel and trade (van Kleunen et al., 2015). As a result, Van

Kleunen et al. (2015) found that 13,168 plant species, or approximately four percent of all

vascular plant species on earth, have become naturalized outside their native ranges due to the

movement of species by humans. The United States bears the brunt of non-native plant

introductions with approximately 6,000 total different naturalized species. California, Florida,

and states in the Northeast US, are the top host states with approximately 1700 species each (van

Kleunen et al., 2015). Most naturalized species cause relatively little harm (Williamson and

Fitter, 1996) but it has been suggested as a general rule that approximately one in ten naturalized

species result in harmful invasions (Richardson and Pyšek, 2006; Williamson and Fitter, 1996).

The ecological, economic, and social impacts of non-native plant invasions can be

extremely costly. Invasive plant species can impact native communities and ecosystems through

a wide range of mechanisms, including direct and indirect competition for limiting resources,

alteration of habitat conditions, changes in mutualisms such as mycorrhizal symbiosis, and

changes in disturbances regimes such as the frequency and intensity of fire (Levine et al., 2003;

Brooks et al., 2004, Pyšek et al., 2012). As a result, plant invasions can cause changes in

community dynamics and succession, ecosystem processes, and wildlife habitat suitability (Vose

et al., 2012). For example, when plant invaders compete with native species for light, water, and

other limiting resources (Wilcove et al., 1998), they can reduce the diversity and shift the

composition of species in herbaceous plant communities (Hejda et al., 2009) or inhibit the

natural colonization and performance of native trees (Reinhart et al., 2005). Invasive plants may

also alter ecosystem processes such as the distribution and availability of water (Richardson et

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al., 2007), fire regimes (Brooks et al., 2004), and carbon and nutrient cycling (Ehrenfeld, 2003).

Altogether, the ecological effects of plant invasions can be severe.

In North America, public and private sectors spend billions of dollars annually to prevent,

manage, and eradicate invasive plant populations. Laroche (1999) reported that between the

years 1991 and 1998, the cost of control and management of Melaleuca quinquenervia by the

South Florida Water Management District amounted to $13 million dollars. In the seminal book

about invasive species, Strangers in Paradise (Simberloff et al., 1997), Center et al., noted that

Florida spends on average, $14.5 million dollars per year on the control of hydrilla, a noxious

aquatic invader. In addition to the costs of control, income losses from limited aquatic-recreation

activities in hydrilla-infested lakes are greater than $10 million dollars annually (Center et al.,,

1997). Other aquatic invasive plants such as phragmites, water lettuce, and water hyacinth can

clog waterways and inhibit water flow (OTA 1993) and result in localized flooding and damage

to water treatment industry equipment (Pimental et al., 2005). In addition, terrestrial invasive

plants can compete with agricultural crops and cause yield losses, and combined with the input

costs for control measures, have been estimated to cost the U.S. approximately $27 billion

dollars annually (Pimental 1997; Pimental et al., 2005). After the 1960s, the focus of public

forests expanded into ecosystem services beyond timber production, such as carbon storage,

watershed protection, wildlife habitat and diversity, and recreational activities (Glück, 2000,

Pearce, 2001). Water quality and rangeland grazing quality also can be negatively affected by

non-native plant invasions and the indirect effects on trophic systems and food chains can

negatively affect the livelihood of fishers, hunters, and bird watchers (Charles and Dukes, 2007),

recreational activities worth $7.8 billion dollars/year (FDEP, 2001).

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Southeastern US Pine Forests

Southeastern pine forests constitute an extensive ecoregion of the United States that is

threatened by climate change, plant invasions, and potentially their combined effects. Covering

92 million acres that spans from Louisiana, Mississippi, Alabama, Georgia, and Florida (Oswalt

et al., 2012), southeastern pine forest is the largest conifer forest ecoregion east of the

Mississippi River and the second largest conifer ecoregion in the contiguous United States,

following the forests of the North Central Rockies (Dinerstein et al., 2016). Historically, this

ecoregion was dominated by longleaf pine (Pinus palustris), however, less than 3% of old

growth longleaf pine forests remain due to extensive logging, tree decline following harvests for

turpentine production, and major land-use changes (Van Lear et al., 2005), such as tree clearing

for development, agriculture, and grazing. The primary native conifer tree species in the

ecoregion include slash (Pinus elliottii var. elliottii), loblolly (Pinus taeda), sand (Pinus clausa),

spruce (Pinus glabra), shortleaf (Pinus echinata), and pond (Pinus serotina) pine. Currently,

slash, loblolly, longleaf, and shortleaf pine are the most common pine species that either

naturally occur or are planted in the region for timber production or restoration, covering

approximately 45% of forested land. Across southeastern pine forests, 42% of land area is

managed specifically for timber and commercial purposes, and loblolly and slash pine are the

primary commercial species (Wear and Gries, 2002; Wear and Gries, 2012). South Florida slash

pine (Pinus elliottii var. densa), a variety of slash pine (Pinus elliottii var. elliottii) endemic to

central and south Florida, is present on approximately 121,410 hectares, but is less widely

distributed and planted less often than Pinus elliottii var. elliottii (Sheffield and Bechtold, 1981).

Nevertheless, the two varieties of the species co-occur and hybridize naturally in central Florida

(Lohrey and Kossuth, 1990). Unlike Pinus elliottii var. elliottii, South Florida slash pine has a

distinct ‘grass’ seedling stage, thick bark, and a thick taproot, rendering it more functionally

21

similar to longleaf pine than the common slash pine variety (Pinus elliottii var. elliottii) in terms

of early life history morphology and fire tolerance (Bethune, 1966; Langdon and Schultz, 1973;

Fowells, 1965; Christman, 2011).

Southeastern pine forests have been extensively researched (Huntington et al., 2000;

Johnston and Crossley, 2002; Kim, 2001; Sun et al., 2005) and their ecological and economic

importance is well established. The World Wildlife Fund (2016) declared the biological diversity

in the ecoregion as unrivaled in North America. In particular, southeastern pine forests are within

the top ten US ecoregions in species richness of amphibians, birds, and reptiles, and top ten in

endemism of amphibians, butterflies, mammals, and reptiles. Additionally, Dr. John Kartesz,

Director of the Biota of North America Program (BONAP) of the North Carolina Botanical

Garden, reported that the region has some of the highest levels of plant endemism in North

America with over 3,400 native shrub and herbaceous species endemic to the ecoregion (World

Wildlife Fund, 2016). Although land clearing, fire suppression, and other anthropogenic

activities have eliminated or reduced the quality of southeastern pine forests in many areas, large

forest tracts and various sized habitat fragments remain, but altogether they comprise less than

12 percent of the original ecoregion. Despite fragmentation and management of only a quarter of

the remaining forested land area, many animals and migratory birds still find refuge in

southeastern pine forests (Ware et al., 1993). For example, gopher tortoises, keystone species

whose burrows provide cover for approximately 400 species of wildlife (Cox et al., 1994), can be

found residing in pine plantations and remaining forest fragments. In addition, these forests

provide habitat for bald eagles and other vulnerable and endangered species such as fox squirrels

and red-cockaded woodpeckers. Although natural and improved varieties of slash and loblolly

pine have largely replaced longleaf in southeastern pine forests today, the ecoregion still

22

maintains important ecological functions as a wildlife refuge (Gilman and Watson, 2006; Bremer

and Farley, 2010) and provides critical ecosystem services. Forests help stabilize soil, filter

water, and cycle nutrients, thereby reducing sedimentation and nutrient inputs into nearby water

bodies (Phillips, 1989; de Groot et al., 2002). Southeastern pine forests also act as a major carbon

sink (Turner et al., 1995; Alavalapati; 2007), sequestering approximately 10% of US forest

carbon (Birdsey, 1992). For example, natural loblolly pine stands store approximately 120,000

lbs/acre of carbon. However, planted pine stands store less carbon, approximately 80,000

lbs/acre, because the stands are generally comprised of younger trees (Birdsey, 1992). On

average, the economic value of carbon stored in forests associated with the Forest Stewardship

Program is $3,154/hectare while carbon values on adjacent non-affiliated forest lands are

$2,622/hectare (Kreye et al., 2014).

The economic importance of planted and managed southeastern pine forest is unrivaled

among forest ecoregions as it supplies 62% of the timber harvested in the U.S. (Smith et al.,

2009), 16% of global industrial wood, and more timber than any other country (Prestemon and

Abt, 2002; Wear and Gries, 2002). Loblolly and slash pine are two of the fastest growing and

therefore most productive southern pine species (Sheffield and Knight, 1982; Sheffield et al.,

1983; Schultz, 1997). Loblolly pine, the most commonly planted and commercially important

timber pine in the southern United States, is dominant on approximately 11.7 million hectares

and comprises more than half of the region’s standing pine volume (Baker and Langdon, 1990).

Slash pine is the second most commonly planted timber species in the south. Together, these

species are a primary source for paper, lumber, and pulpwood (Schultz, 1997), and are used to

stabilize soil during mine reclamation, collectively generating billions of dollars in revenue for

regional economies (FDACS, 2015). For example, 50% of land area in the state of Florida is

23

covered by natural or managed forestlands, with the greatest concentration in the north and

central regions. Of the seven million hectares of forested land, over 6.2 million are timberlands

that support economic activities. In 2013, these managed lands generated more $16 billion

dollars and provided over 80,000 jobs to Florida residents (Nowak, 2015). Valuation of

ecosystem services provided by these forests is also highly regarded. Kreye et al., (2014)

determined that an acre of forest can generate ecosystem services worth between $264 and

$13,442 per year in terms of gas and climate regulation, pollination, habitat, and water regulation

and supply. Furthermore, conclusions from two questionnaires disseminated by Kreye et al.,

(2014) and filled out by public land management decision-makers at the local to federal level and

private landowners who participate in forestry education programs, showed that decision-makers

believed it was their responsibility to ensure that recreation, aesthetics, and habitat and natural

resources from forests were intact and available to society. Similarly, 84.5% of private

landowners reported aesthetics as important while more than 70% reported environmental

quality for recreation opportunities and quality of drinking water as important (Kreye et al.,

2014). Separately, a study by Ernst et al., (2004), demonstrated that a 10% increase in forest land

within water recharge zones could lower chemical and treatment costs by an average of 20%. In

total, the direct and indirect economic value of ecosystem services provided by southeastern pine

forests is substantial.,

Climate Change: Effects on Southeastern Pine Forests

The implications of climate change for both the ecological integrity of southeastern

United States ecosystems and human dependence on their services are vast and alarming. In

addition to the projections of increased global average temperatures, temperatures in the interior

regions of the southeast are expected to warm by 0.5°C to 1°C more than coastal regions (Kunkel

et al., 2013; Carter et al., 2014). The southeast is already experiencing an increase in the number

24

of days and nights with temperatures above 35°C and 24°C, respectively, since 1970 (Kunkel et

al., 2013). Heat stress and higher temperatures can have adverse effects on human health (Luber

et al., 2014), the spread of mosquitos (Carter et al., 2014), dairy and livestock production (West,

2003), crop productivity (Hatfield et al., 2008; Hatfield and Takle, 2014), and the risk of wildfire

(Gramley 2005). Drought is also a primary concern for the region, as multiple scientific studies

project increases in drought due to climate change for the southeastern United States and

significantly greater frequency and intensity of drought in Florida (Karl et al., 2009; Wang et al.,

2010).

Drought is a primary climate change factor that can have significant effects on forest

ecosystems in various ways (Allen, 2010). When water is a limiting element in the environment

for extended periods of time, species have to adjust their growth strategies in order to preserve

energy for maintenance. Anderegg et al., (2016) found that physiological traits, including the

hydraulic safety margin, stomatal activity, and wood density, work independently and

interactively to influence tree species mortality and response to drought. At the species level,

drought-stress related defenses, such as the closing of stomata or leaf shedding, often come at the

expense of transpiration, photosynthesis, or growth (Hsiao, 1973; McDowell et al., 2008) and in

severe scenarios, may result in mortality due to carbon starvation or hydraulic failure (McDowell

et al., 2008). At the ecosystem scale, drought stress in forests, particularly in regions marked by

high temperatures and long growing seasons, can weaken resistance to stressors like insects

(McDowell et al., 2008; Adams et al., 2009; Breshears et al., 2009; Clark et al., 2016), pathogens

(T.E. Kolb et al., 2016), fire (Dale et al., 2001; Clark et al., 2016), and plant invasions (Dale et

al., 2001). Forest fires in the Amazon for example, are especially injurious and kill more trees

during drought because fallen leaves, branches, and other litter is both drier and more abundant,

25

creating fuel conditions that are more susceptible to and facilitative of higher intensity fires

(Brando et al., 2014). While the impacts of drought on plant communities are widely recognized

as negative, their effects on tree mortality and stand dynamics are difficult to predict because

drought stress increases susceptibility to, and thus is often accompanied by other stressors (Allen

2010; Clark et al., 2016), including non-native plant invasions (Dale et al., 2001). Furthermore,

these direct and indirect effects are difficult to extrapolate between organism and stand scale

observations (Clark et al., 2016).

Plant Invasions: Effects on Southeastern Pine Forests

Slash and loblolly pine forests cover millions of hectares in the southeast and are

frequently subjected to non-native plant invasions (Simberloff et al., 1997; Miller et al., 2002).

The total area of non-native plant infestations is not well quantified (Miller et al., 2002), however

it has been reported that all federal park and forest land in the region have some level of

infestation by an invasive plant species (Hamel and Shade, 1985; Hester, 1991). Altogether, the

southeastern states of Louisiana, Mississippi, Alabama, Georgia, and Florida contain on average,

more than 1,000 naturalized non-native plants, with Florida hosting a disproportionate number

compared to the other states (van Kleunen et al., 2015). Of the 1300 or so non-native plant

species in Florida (FLEPPC, 2007; van Kleunen et al., 2015), approximately 900 species have

become established in natural areas (Frank and McCoy, 1995; Frank et al., 1997, Simberloff et

al., 1997; Pimental et al., 2005), 124 are listed as invasive, and 92 typically invade forests

(FLEPPC, 2007). In total, these infestations span more than 400,000 hectares of public land

(FDEP, 2006). Florida is inundated with invasive species, primarily due to its amenable climate

and importance as a key transportation hub for North America (van Kleunen et al., 2015). For

example, approximately 85% of all non-native plants imported to the United States enter through

the state of Florida (Simberloff, 1994). Total costs of management and economic impacts from

26

invasive species in the southeast have yet to be well quantified however, reports for state level

spending and costs for individual species are more common. For example, Florida spent over $37

million dollars on invasive plant management in the fiscal year 2005 – 2006 (Langeland, 2013)

and estimates for economic losses associated with Melaleuca quinquenervia (melaleuca)

invasion alone range from $168 million to two billion dollars over a 20-year period (Serbesoff-

King, 2003).

The ecological effects of invasive plants have been seriously detrimental for southeastern

states and include changes in fire and other disturbance regimes and displacement of endangered

species (Serbesoff-King, 2003). Altogether, approximately 42% of all threatened and endangered

species are in decline due to invasive species (Pimental et al., 2005). For example, melaleuca

was at one time spreading through South Florida forests and grasslands at a rate of

approximately 11,000 hectares/year (Campbell, 1994), at the expense of native vegetation and

wildlife (OTA, 1993). However, the invasion has since been curtailed by a well-organized

integrated pest management program (Silvers et al., 2007). Sapium sebiferum (Chinese tallow) is

a shade tolerant tree that is invading forests and competing with native tree species in nine

southern states including Louisiana, Mississippi, and Florida (Bruce et al., 1997; Miller, 1997;

McCormick and Leslie 2005). In a water gradient experiment, Butterfield et al., (2004)

demonstrated that Chinese tallow significantly outperformed three ecologically important native

tree species, loblolly pine, Nyssa aquatica (water tupelo), and N. sylvatica (black gum), under

flooded and drought conditions. More broadly, the high tannin content of Chinese tallow can

alter litter composition, and thus decomposition processes, including temporal nutrient releases

in the form of nitrogen pulses, and subsequent shifts in decomposer and microbial communities

(McCormick and Leslie, 2005). Another noxious tree species, Ligustrum sinense (Chinese privet)

27

is inhibiting regeneration of bottomland pine-hardwood forests (Miller, 1998). Of the variety of

invasive plant species found in southeastern pine forests, perhaps the most significant ecological

problems are generated by invasions of the non-native grass Imperata cylindrica (cogongrass).

Cogongrass, a perennial C4 rhizomatous grass native to Asia, is especially pervasive in

pinelands where it can inhibit pine establishment (Daneshgar et al., 2008) and cause a range of

other ecological impacts. Cogongrass is one of the most prolific and aggressively spreading plant

invaders in the southern United States. The two epicenters of invasion are Alabama, where it was

accidently introduced within packing material in 1912, and west central Florida, where it was

intentionally introduced as a forage crop in the 1920s (Hubbard, 1944; Dozier et al., 1998).

Although the species was introduced as a potential forage, high silicate content in the leaves

renders cogongrass almost useless as a forage crop beyond the juvenile stages of development

(Dozier et al., 1998). Since its initial introduction, cogongrass has spread into ecosystems

throughout the region including undisturbed natural areas, fallow agriculture lands, urban areas

and roadsides, and managed agriculture and timber lands (Dozier et al., 1998). The species now

covers hundreds of thousands of hectares across the region (Schmitz and Brown, 1994; Estrada

& Flory, 2015).

Cogongrass spreads vegetatively throughout the southeast US via rhizomes but

supposedly only produces viable seed outside of Florida (G. MacDonald, personal

communication). Across the region it forms dense monocultures at the expensive of native

species abundance, distribution, and diversity (Estrada and Flory, 2015), causing significant

ecosystem-level effects. Cogongrass commonly invades after disturbances caused by, for

example, fire, agricultural operations such as mowing, and mine reclamation (Lippincott, 2000;

Holzmueller and Jose, 2012). It establishes and spreads across a wide variety of environmental

28

conditions including under drought, shade, or full sun, and is reportedly fire tolerant (Patterson,

1980; Bryson et al., 2010). Cogongrass is highly flammable and facilitates longer, hotter burning

fires that can kill mature pine trees and scorch typically fire resistant species (Lippincott, 2000;

Platt and Gottschalk, 2001). It reportedly competes strongly with native plant species for water

and nutrients, resulting in reduced species diversity and shifts in species composition (Kuusipalo

et al., 1995; MacDonald 2004), although little empirical evidence is currently available (Estrada

and Flory, 2015).

Research Needs

Altogether, the pine forest ecoregion of the southeastern United States provides

significant contributions to the economy, environment, and social wellbeing of citizens both

globally and locally, but are under threat from climate change, plant invasions, and their

combined effects. These forests produce approximately 16% of the global wood supply

(Prestemon and Abt, 2002; Wear and Gries, 2002), sequester 10% of the country’s carbon

(Birdsey, 1992), and are valued for other ecosystem services including drinking water quality,

wildlife habitat, recreational use, and aesthetics (Kreye et al., 2014). The ecoregion generates

millions of dollars in revenue from tourism activities in the region, such as wildlife viewing and

hunting, and provides food and cover for numerous flora and fauna, including threatened and

endangered species. Unfortunately, the rate of forest cover loss in the region is one of the highest

in the world (Hansen et al., 2010) due in part to urbanization, climate change induced

disturbances and impacts, and invasions by non-native plant species. However, there have been

some successful programs that focus on preventing and managing plant invaders, and have in

some cases resulted in successful control of invasive plant species (e.g., melaleuca). In addition,

standard practices that cope with aspects of climate change involve careful selection of seed

source and the use of drought-hardy or cold-hardy strains where needed (Fowells, 1965;

29

Dorman, 1976), as well as seed strains that are more resistant to fusiform rust (Baker and

Langdon, 1990). There are also new artificial hybrids of loblolly pine and shortleaf pine to be

used in areas prone to fusiform rust (Kraus and LaFarge, 1977), and artificial crosses between

loblolly and pitch pine for enhanced cold-hardiness (Dorman et al., 1973).

Recent research has focused on the independent effects of climate change (e.g., drought

impacts on pine forests) and plant invasions (e.g., how competition for resources inhibits pine

performance, but more information is needed to help predict the longer-term effects of such

stressors. Moreover, there is a critical need for additional data to help uncover how multiple

stressors and novel interactions among stressors may affect critical habitats, and the survival and

performance of pine tree species in particular. In general, research needs to be conducted across

different pine tree developmental stages from seed germination to seedlings to adult, and across a

range of spatial and temporal scales. For example, experiments on stressor effects should be

conducted across latitudinal gradients where climatic conditions are likely to vary widely. More

locally, plant invasions and climate change factors (e.g., drought) may have contrasting effects

on pine tree survival and performance across different environmental conditions including

variation in light availability due to stand characteristics or different soil types. Greenhouse

studies, invader addition and removal studies, mesocosm, common garden, and field experiments

all can be designed to uncover independent and interactive effects of multiple stressors on native

pine species.

More broadly, such experiments require acknowledgment (and funding) from

government and other research entities, and cooperation among scientists, land management

practitioners, property owners, and federal and state agencies. Clearly, it is in the best interest for

land owners, scientists, and policy and management professionals to work together to address the

30

problem of climate change and plant invasions on forest ecosystems. For example, improved

education and awareness of civilians who benefit from healthy forest ecosystems and the

resources they provide can result in action, such as participation in invasive plant prevention and

management efforts. A better understanding of these issues can provide more insight and action

into preserving pine forest ecosystem integrity in the face of impacts by multiple stressors.

31

Figure 1-1. Conceptual diagram illustrating three potential scenarios for the magnitude of

independent (additive) or interactive (synergistic and offsetting) negative effects as a

result of multiple stressors acting on ecological communities. Here, “D” represents

drought and “I” represents invasion by a non-native species. The lower dashed line

provides a reference point for the additive effects scenario.

32

CHAPTER 2

THE EFFECTS OF DROUGHT AND PLANT INVASION ON PINE SEEDLINGS

Introduction

Climatic change, including shifts in temperature and precipitation regimes, is creating

novel abiotic conditions that can alter the structure and function of ecological communities. In

particular, the extent and severity of drought is intensifying in many ecosystems worldwide due

to climate change (Easterling et al., 2000; IPCC, 2001; Hoerling and Kumar 2004), resulting in

native species die-off (Thomas et al., 2004; Breshears et al., 2005; Vose et al., 2012) and

displacement (Lenoir et al., 2008). Non-native plant invasions are also threatening native

ecosystems (Vitousek et al., 1996) by modifying habitat conditions (Vose et al., 2012) and

competing with native species for limited resources such as nutrients, light, and water (Wilcove

et al., 1998). Native and invasive species will likely interact in new ways as novel abiotic

conditions caused by climate change, and drought specifically, shift species ranges (Woodward

et al., 1987; Hoffman et al., 1997; Pounds et al., 1999;) and alter plant community dynamics by

opening niches (Walther et al., 2002) and changing habitat suitability (Pimental et al., 2005).

Although drought and invasive species are primary abiotic and biotic stressors, little is known

about how they might interact to affect native ecosystems (Sala et al., 2000; Mora et al., 2007;

Halpern et al., 2008a, b).

There are three scenarios for how multiple global change stressors may interact to affect

native ecosystems. Drought and invasion may exert negative effects that manifest additively such

that the combination of stressors is equal to the sum of each acting independently (Folt et al.,

1999; Breitburg and Riedel, 2005). Alternatively, the two stressors may have synergistic

interactive effects, whereby together they yield stronger negative effects than would be predicted

based on each stressor acting in isolation (Folt et al., 1999; Breitburg and Riedel, 2005). Finally,

33

one stressor may act antagonistically with the other and offset the effects of that other stressor

(Folt et al., 1999; Breitburg and Riedel, 2005), resulting in less negative impacts than would be

expected compared to the additive scenario. For example, a plant invader with a dense canopy

may, under drought conditions, compete strongly against native species for both light and limited

soil water, leading to synergistic negative effects. Alternatively, the dense invasion might offset

drought stress by lowering ground-surface temperatures and air flow, thereby reducing

evapotranspiration. Given the increasing severity and extent of both drought and plant invasions,

the paucity of studies and the unpredictable nature of interactions between these stressors

represent a critical knowledge gap.

Drought and plant invasions are intensifying in forests worldwide, including ecologically

and economically important pine forests of the southeastern United States (Simberloff et al.,

1997; Wang et al., 2010). Drought stress can degrade forest health and resistance to stressors

such as fire, pathogens, insects, and plant invasions (Dale et al., 2001), particularly in regions

marked by high temperatures and long growing seasons (Aber, 2001). Both natural and planted

coastal plain forests in the southeastern US are largely dominated by natural or improved

varieties of Pinus elliottii (slash pine) and Pinus taeda (loblolly pine) forests. Because slash and

loblolly pine forests have largely replaced Pinus palustris (longleaf pine) across its historic

range, they can represent important ecological refuges (Gilman and Watson, 2006; Bremer and

Farley, 2010) and act as a major carbon sink in the US (Turner et al., 1995; Alavalapati, 2007).

These forests also largely drive the economy of the region as primary sources for lumber and

pulpwood (Prestemon and Abt, 2002). Southeastern US forests are increasingly invaded by non-

native plant species, including Imperata cylindrica (cogongrass), a perennial C4 grass native to

Asia that now covers hundreds of thousands of hectares across the region (Schmitz and Brown,

34

1994; Estrada & Flory, 2015) and can inhibit pine establishment (Daneshgar et al., 2008).

Cogongrass commonly invades after disturbances (Lippincott, 2000; Holzmueller and Jose,

2012), is reportedly drought and fire tolerant (Patterson, 1980; Bryson et al., 2010), and strongly

competes with native species for water and nutrients (Kuusipalo et al., 1995; MacDonald, 2004;

Estrada and Flory, 2015). Thus, I hypothesize that drought and cogongrass invasions will have

additive or synergistic negative effects on the survival and performance of pine seedlings.

To test my hypothesis, I used a factorial common garden experiment to evaluate the

independent and interactive effects of drought (simulated with rainout shelters) and cogongrass

invasion on slash and loblolly pine seedling survival and performance. Over one year I measured

seedling survival and several aspects of performance, including biomass and relative growth rate,

as well as variation in abiotic conditions such as soil moisture and light availability across the

drought and invasion treatments. My results demonstrate that both drought and invasion

significantly suppress the survival and performance of these ecologically and economically

important pine species, and together these two stressors have the potential to dramatically alter

southeastern US forests.

Materials and Methods

Study Species

Slash and loblolly pine forests occur naturally or are planted across much of the

southeastern US. They provide critical habitat for wildlife and generate billions of dollars in

revenue for regional economies (FDACS, 2015). Slash pine is moderately to highly drought

tolerant relative to other pines (Burns and Honkala, 1990; Gilman and Watson, 2006) and can

grow across a range of soil conditions from seasonally dry to wet soils near streams and swamps

and in hammocks and mesic flatwoods (Meyers and Ewel, 1990). Loblolly pine has low to

moderate drought tolerance (Burns and Honkala 1990; Gilman and Watson, 2006) and

35

predominantly occurs in poorly drained soils in mesic forests, floodplains, and hydric hammocks

(Meyers and Ewel, 1990). Both species have low to moderate shade tolerance and exhibit poor

establishment under competition (Burns and Honkala, 1990; Gilman and Watson, 2006). P.

elliottii var. elliottii is the most common and widely distributed P. elliottii variety, occurring

across the southeastern coastal plain from Louisiana to South Carolina, down to Central Florida

(USDA NRCS, 2006). South Florida slash pine (Pinus elliottii var. densa) has a more limited

range, being restricted to Central and South Florida, and has the distinction of having a grass

stage, unlike P. elliottii var. elliottii. The two varieties hybridize naturally where their ranges

overlap, producing offspring that are indistinguishable from either variety (Lohrey and Kossuth,

1990).

Experimental Design

We conducted the common garden field experiment at the Bivens Arm Research Site

(BARS) in Gainesville, FL (29.628489°N, -89.353370°W). In May 2012 we established 40 4 m

x 4 m plots and randomly assigned treatments to each of ten blocks. Treatments included 1)

ambient precipitation, resident species only; 2) ambient precipitation, resident species plus

cogongrass; 3) reduced precipitation via rainout shelters (hereafter referred to as “drought”

plots), with resident species only; and 4) drought plots, with resident species plus cogongrass.

We selected 12 native herbaceous understory species that occur in southeastern US pine forests

and planted three individuals of each species into a 6 x 6 grid design. The natives consisted of

seven grass and five forb species (Table 3-1). Plots were colonized by other native and

naturalized species from the seed bank and surrounding area, such as Bidens alba and Paspalum

notatum, thus we refer to plots without cogongrass as “resident” species plots. Cogongrass

rhizomes were collected from a nearby population and grown for three months in a greenhouse.

36

In June 2013 we planted nine cogongrass ramets in each “invaded” plot. All ramets survived

transplantation.

In February 2013, we constructed wooden, lean-to style rainout shelters over drought

treated plots. We used corrugated polycarbonate roofing (89% areal coverage and light

transmittance; Tufttex, Fredericksburg, VA) and aluminum gutters (Amorfill Aluminum,

Gainesville, FL) to capture and direct precipitation offsite. We diverted surface- and ground-

water flow from drought plots by lining the perimeter of each with ground-level aluminum

flashing (Amerimax aluminum flashing) and belowground (to one-meter depth) plastic sheeting

(20 mm thick; Global Plastic Sheeting Inc., Vista, CA). Control shelters were constructed over

non-drought plots and topped with shade cloth (22% shade; International Greenhouse Company),

which created comparable light levels in ambient and drought plots (mean ± SE percent light

reduction ambient: 33.4 ± 1.01 and drought: 31.1 ± 1.2; t (37) = 1.5, P = 0.14).

In January 2015, I planted four bareroot seedlings each of slash and loblolly pine into all

40 plots at 0.5 m spacing in an alternating arrangement. For this experiment, I used P. elliottii

var. densa (South Florida slash pine, hereafter slash pine) because seedlings were more

accessible when the experiment was initiated. Slash pine seeds were collected from a native

stand in Avon Park, Florida and grown for one year at Andrews Nursery in north central Florida.

Loblolly pine seeds were sourced from Livingston Parish, Louisiana and grown for one year at

Dwight Stansel Farm in Wellborn, FL. Seedlings that did not survive after ten weeks were

assumed to have died from transplant shock and were replaced.

Data Collection

To characterize the density of cogongrass invasion, we measured cogongrass cover in

February, July, and October of 2015. We divided each plot into a grid of quadrats and then

averaged them at the plot level for final analysis. We recorded percent cover of all species

37

present in the plots in July but only recorded the species covering 5% or more of each quadrat in

February and October. Plant canopies commonly overlapped and therefore total vegetation cover

in a plot often exceeded 100%. To determine how the drought and invasion treatments affected

abiotic conditions in the plots we quantified soil volumetric water content (HydroSense II;

Campbell Scientific, Logan, UT) and photosynthetically active radiation (PAR; ACCUPAR LP-

80; Decagon Devices, Pullman, WA). We measured soil moisture at a depth of 0-12 cm (n = 4

subsamples per plot) and data collection occurred every two weeks during the dry season

(December to April) and weekly during the wet season (May until December). We measured

light availability (PAR) at ground level and 0.5 m as well as above the vegetation canopy (~1.5

m) in each plot (n = 4 subsamples per plot).

To evaluate how slash and loblolly pine responded to drought and invasion, we quantified

survival to harvest (one growing season), relative growth rates (RGR) in height and diameter,

and biomass (dried to constant mass at 60 °C) at final harvest. To quantify growth rates, we

measured height to the apical meristem (mm) and root crown diameter (mm) two months after

pines were planted (March 2015) and again at final harvest (December 2015). We calculated

RGR according to Hunt (1982) as ln(W2) – ln(W1)/t2 – t1, where W2 and W1 are the final and

initial height and diameter, respectively, and t2 and t1 are the final and initial dates of

measurement. We analyzed the log-transformed growth rates (Hunt 1982) but present the

untransformed data to facilitate ecological interpretation. To determine aboveground biomass,

we clipped seedlings at the soil surface and then counted the number of limbs (Appendix A-1)

and pine webworm (Pococera robustella; Appendix B-1) nests. After separating the nests from

the seedlings we dried and weighed seedling biomass.

38

Statistical Analysis

Cogongrass percent cover was analyzed with mixed model ANOVA using the nlme

package in R version 3.2.3 (v.3.23, R Development Core Team). Fixed effects included drought,

date, and a drought x date interaction, with a random effect of plot nested with block. Soil

moisture, light availability, and pine performance (proportion of seedlings surviving per plot,

RGR of height and diameter, and biomass) were analyzed using mixed model ANOVA, with soil

moisture and light response models accounting for repeated measures. Response variables were

transformed as necessary (square root of soil moisture, light availability and biomass, and log of

RGR) to improve normality and homogeneity of variance based on inspection of residual-versus-

predicted and residual-versus-quantile plots. The fixed effects for soil moisture and light

availability were drought, invasion, date, and all interactions, with block as a random effect. For

pine responses to the treatments, species were analyzed individually except for overall survival.,

Percent survival by plot was analyzed using proc mixed in SAS (v. 9.4, SAS Institute). Height

and diameter RGR and biomass, which had unbalanced data sets due to unequal seedling

survival across the treatments at final harvest, were analyzed using proc glimmix with a

Gaussian distribution and logit link function. All post-hoc models included Tukey’s adjustment

for multiple comparisons.

39

CHAPTER 3

FINDINGS

Results

Over the three measurements during the 2015 growing season, cogongrass percent cover

was not affected by drought (mean ± SE drought: 54.7 ± 2.8; ambient: 57.7 ± 3.8; F (1, 9) = 3.0; p

= 0.119). Across the drought and ambient treatments in 2015, cogongrass percent cover

increased from 53% ± 2.7 in February to 75% ± 2.4 in October. On average, volumetric water

content in drought conditions was 48% lower than under ambient conditions (Figure 1a; F =

145(1, 701); p < 0.0001) and 37% higher in invaded plots relative to resident plots (Figure 1a; F =

7.3 (1,701); p = 0.007) under drought conditions. Significant temporal variation (Figure 1b; 2015, F

= 104 (18, 701); p < 0.0001) in soil moisture dynamics show that the effects of the invader

manifested mostly from July onward (Figure 1b). Overall, light availability was 22% greater at

0.5 m (Figure 2a) and 58% greater at ground level (Figure 2b) in resident species plots compared

to invaded plots. As with soil moisture, there was temporal variation in light availability (Figure

2c, F = 44.6 (7, 286); p < 0.0001; Figure 2d, F = 63.4 (7, 286); p < 0.0001) with the most pronounced

difference being that at ground level cogongrass reduced light all year whereas resident species

reduced light to the same extent as cogongrass only from July to October (Figure 2d).

Across all four treatments, slash pine survival (23% ± 3.8) was lower than that of loblolly

pine (62% ± 3.8; F (1, 63) = 53.9; p < 0.0001). Slash survival was 58% lower under drought (13.8

± 5.2% survival) than ambient (32.5 ± 7.5) conditions and 72% lower with the invader than with

resident species (Figure 3; Table 3-2). Invasion offset any effects of drought on slash pine

survival (Figure 3a; Table 3-2, significant drought x invasion interaction). Loblolly pine survival

was 38% lower under drought (47.5 ± 8.7) than ambient (76.3 ± 7.2) conditions and 26% lower

when growing with the invader (52.5 ± 9.8) than with resident species only (71.3 ± 6.1; Figure

40

3b; Table 3-2). In contrast to slash survival, invasion did not significantly offset the effect of

drought on loblolly survival (Table 3-2, no drought x invasion interaction).

Drought, but not invasion, was associated with lower relative growth rates in height of

both pine species (Figure 4a, b; Table 3-2; Table 3-3). Height RGR of slash pine was 47%

slower under drought (0.0379 ± 0.010 mm mm-1day-1) than ambient (0.0720 ± 0.019 mm mm-

1day-1) conditions and of loblolly was 42% slower under drought (0.0854 ± 0.0064 mm mm-1day-

1) than ambient (0.147 ± 0.0069 mm mm-1day-1) conditions. In terms of stem diameter, slash pine

only grew larger under baseline conditions of ambient precipitation with resident species, while

remaining stagnant when exposed to drought or invasion independently or in concert (Figure 5a,

Table 3-2; Table 3-3). Growth in loblolly diameter was 43% slower under drought (0.00666 ±

0.00077 mm mm-1day-1) than ambient (0.0177 ± 0.00082 mm mm-1day-1) precipitation and 39%

slower when growing with the invader (0.00693 ± 0.00074 mm mm-1day-1) compared to resident

(0.0114 ± 0.00085 mm mm-1day-1; Table 3-2; Table 3-3) species. For both species, drought and

invasion had an additive negative effect on diameter (Figure 5), with growth in slash and loblolly

reduced by 241% and 71%, respectively, under both stressors relative to baseline conditions

(Figure 5, Table 3-2; Table 3-3).

Slash pine biomass was 45% lower under drought than ambient conditions when growing

with resident species, but there was no effect of drought under the invasion treatment. In parallel,

there was only an effect of invasion on biomass under ambient precipitation, where presence of

invasion resulted in 53% lower slash biomass compared to resident vegetation (Figure 6a).

Despite the tendency for the invader to offset drought stress, the drought x invasion interaction

was not statistically significant, possibly due to low power associated with low seedling number

at harvest. In contrast to the slash pine results, both drought and invasion significantly affected

41

loblolly pine biomass (Figure 6b; Table 3-2). Average loblolly seedling biomass was only half as

much under drought (8.6 ± 2.1 g) compared to ambient (16.3 ± 2.4 g) conditions (Table 3-2).

Separately, invasion resulted in 36% less loblolly biomass (9.7 ± 2.1 g) compared to seedling

performance in resident species plots (15.2 ± 2.4 g). Thus, there was an additive negative effect

of drought and invasion on loblolly seedlings, resulting in 70% lower biomass under the

combined treatments relative to baseline ambient/resident conditions (Figure 6b).

Discussion

Climate change and plant invasions are two predominant drivers of global environmental

change, yet it is difficult to predict how these abiotic and biotic stressors will affect native

species. Here I demonstrate that experimental drought and invasion by an aggressive non-native

grass, both individually and in concert, significantly suppressed slash and loblolly pine seedlings.

In general, loblolly pine outperformed slash pine over the course of the experiment, consistent

with the findings of Shiver et al., (2000) who showed that loblolly had higher survival and

growth than slash across several sites in Georgia and northern Florida (Shiver et al., 2000).

However, I found that despite the higher performance of loblolly overall, the magnitudes of both

species’ responses to the treatments were similar (Figures 3, 4, 5, and 6). In both species, each

stressor alone inhibited seedling survival, while the combination of drought and invasion resulted

in 86% and 51% fewer surviving slash and loblolly seedlings, respectively, when compared with

pine survival under ambient conditions with resident species. In addition, seedling performance

metrics tied closely to juvenile and adult tree performance (McGrath and Duryea, 1994),

including RGR of stem diameter in both species, and first-year biomass in loblolly exhibited

additive negative effects under both stressors that led to critical reductions in growth relative to

baseline conditions. I found that interactive effects, where the two stressors in combination have

greater (synergistic) or lesser (offsetting) effects than expected, were uncommon, suggesting that

42

the effect of each stressor acting in isolation is to some degree predictive of their effect in the

presence of the other stressor. Overall, my results suggest that the combination of abiotic stress

from drought and biotic stress from a plant invader can severely affect slash and loblolly pine,

two of the most ecologically and economically important forest species in the southeastern US.

The majority of experiments that have investigated how low soil moisture affect slash

and loblolly pine have demonstrated strong detrimental effects on tree performance. My findings

not only support but greatly expand the inference space of these previous studies, which were

largely conducted from a timber production or maximum-yield perspective (Bongarten and

Teskey, 1987; Clark and Saucier, 1991; VanderSchaaf and South, 2003); were implemented in

field settings where site location was considered a proxy for drought (Shoulders, 1977); or were

conducted in a less realistic greenhouse setting where soil water was reduced by limiting

irrigation (Bongarten and Teskey, 1987). In contrast, I experimentally isolated the effect of

drought on pines in an ecologically relevant setting using diverse plant communities growing and

competing under field conditions. In this novel experimental context, I found that for both pines,

drought but not invasion drove a reduction in seedling height growth, possibly because seedlings

growing under lower light conditions in the dense invader canopy were cued to prioritize growth

in height. In contrast, the relative growth rates in stem diameter of both pines were strongly

inhibited by both drought and invasion, which could have important implications for stand

dynamics given that this trait is strongly linked to water-stress acclimation and resulting pine tree

survival (McGrath and Duryea, 1994). Biomass was the only performance metric for which the

species responded in a different way: slash biomass tended (although not statistically significant)

to only decline from drought when growing with resident species, while loblolly biomass was

reduced by drought regardless of the associated vegetation.

43

Typically, invasive grasses inhibit tree seedling establishment, growth, and survival by

reducing overstory light and moisture and nutrients in the soil solution (D’Antonio and Vitousek,

1992; Flory and Clay, 2010). Despite the apparent effects of cogongrass invasions on native

species and ecosystem functions, relatively few other studies have quantified impacts of

cogongrass invasion on native plant communities (Brewer, 2008; Daneshgar and Jose, 2009;

Estrada and Flory, 2015). I show here that invasion by cogongrass occurs rapidly in terms of

increased cover over a growing season, and that stands can form near-monocultures that greatly

reduce light availability. Surprisingly, however, cogongrass maintained high levels of soil

moisture in both the ambient and drought treatments, suggesting its potential to offset drought

stress to pines. While there was some evidence of an offsetting effect on slash pine survival and

biomass, cogongrass generally limited other resources (e.g., light as we have shown, or possibly

nutrient availability) such that pine seedlings did not benefit from the additional soil moisture

observed in invaded plots. These findings of cogongrass’ strong competitive ability mirror those

of Daneshgar et al., (2008) who conducted an observational study in plots with cogongrass,

native species, or no vegetation and measured survival, height, root collar diameter, and biomass

of planted loblolly pine seedlings. They found that cogongrass inhibited seedling survival and

suppressed seedlings for all growth responses compared to native vegetation or no vegetation

treatments. My expanded comparison with two species under experimental conditions shows that

for survival, invasion more strongly inhibited slash than loblolly pine, while for biomass, it more

strongly affected loblolly than slash. Regardless, it is clear that cogongrass invasions have

significant implications for pine tree seedling establishment and performance.

While some studies have evaluated how drought and competition individually affect pine

seedling performance, little is known about the combined effects of these stressors. I am aware

44

of only one study that has tested the combination of plant competition and low soil moisture on

slash and loblolly pine seedlings. Stransky and Wilson (1966) planted seedlings into plots with

and without turf grass competition and, after four months, erected rainout shelters to simulate

drought. They found little effect of low soil moisture without competition but the combination of

drought and competition with turf grass reduced slash and loblolly seedling survival by more

than 80%. In contrast to our results where we found higher soil moisture in invaded plots,

Stransky and Wilson (1966) reported lower soil moisture in plots with plant competition.

However, they compared bare ground to plots with plant competition whereas our comparison

was between resident species only and invaded plant communities. Regardless, the difference in

results between our study and Stransky and Wilson (1966) indicates plant responses to multiple

stressors may be context and system specific. More recently, Dávalos et al., (2014) evaluated the

effects of multiple stressors, including non-native plant invasion, on the survival and growth of

four rare plant species in the US, and concluded that interactions among stressors were present

yet unpredictable and require multifactor approaches to elucidate. Given the predicted increased

prevalence of drought and other climate change factors, and the spread of plant invaders (Van

Kleunen et al., 2015), natural and managed ecosystems are increasingly likely to be subjected to

multiple stressors operating outside of historic norms in terms of timing or severity.

In this experiment I evaluated South Florida slash pine (Pinus elliottii va. densa), which

is less widely distributed and less often planted than Pinus elliottii var. elliottii. The ranges of the

two varieties extend over mostly separate geographic regions, although they co-occur in Central

Florida. More importantly, they have distinct life histories where var. densa has a ‘grass’

seedling stage and var. elliottii does not. Thus, although the responses to drought and invasion I

observed are congruent with previous findings for other varieties, we urge caution in

45

extrapolating our specific results for var. densa to var. elliottii, or to other coastal plain pine

species or varieties. In addition, I focused on first-year seedling performance, which is known to

be particularly influential in the long-term growth patterns of slash and loblolly pine trees

(Stranksy and Wilson, 1966; Bongarten and Teskey, 1987; Clark and Saucier, 1989), but studies

that focus on earlier (seed) and later (juvenile and adult) pine life history stages are important.

Furthermore, studies that evaluate pine responses to multiple stressors across variable field sites

would provide more robust measures to predict the outcome of drought and invasion effects on

forest stand dynamics. Despite these important caveats, I found some generality in how loblolly

and South Florida slash pine respond to abiotic and biotic stressors.

Our drought by invasion factorial experiment is the first to demonstrate both the

independent and combined effects of multiple stressors on slash and loblolly pine seedling

survival and performance. The effect of drought on seedlings of both species was significant,

suggesting that land managers should carefully select field sites for plantations, and may benefit

from considering different pine varieties or those with improved drought tolerance in the face of

climate change. In addition, my results demonstrate experimentally the dramatic effects of

cogongrass invasion on pine seedlings, which should further motivate land owners and property

managers to remove this noxious invasive species. Additional work is needed to determine the

longer-term effects of drought and invasions on pine forests, but clearly both of these stressors,

and in particular their combination, may have profound consequences for southeastern US pine

forests.

46

Table 3-1. The scientific names and functional types of twelve native understory species planted in 2013 and the most common

resident species established in the plots by 2015.

Genus Species Functional group Genus Species Functional group

Planted native species (2013) Most common resident species (2015)

Andropogon brachystachyus grass Ambrosia artemisiifolia forb

Andropogon virginicus glaucus grass Aristida stricta grass

Aristida stricta grass Baccharis halimifolia shrub

Eragrostis elliotti grass Bidens alba forb

Eragrostis spectabilis grass Bothriochloa pertusa grass

Muhelenbergia capillaris grass Eragrostis spectabilis grass

Panicum anceps grass Eupatorium capillifolium forb

Carophephorus subtropicanus forb Muhlenbergia capillaris grass

Elephantopus elatus forb Paspalum notatum grass

Liatrus laevigata forb Pityopsis graminifolia forb

Pityopsis graminifolia forb Solidago fistulosa forb

Solidago fistulosa forb Urochloa maxima grass

47

Table 3-2. Results of mixed model ANOVAs testing the fixed effects of drought, invasion, and their interaction on slash and loblolly

pine survival, relative growth rate of height (RGR height) and diameter (RGR diameter), and aboveground biomass. P-

values less than or equal to 0.05 indicate significant differences (α = 0.05).

Fixed

effects

Source of

variation

Survival

(%)

RGR height

(mm/day)

RGR diameter

(mm/day)

Biomass (g)

d.f. F P d.f. F P d.f. F P d.f. F P

Slash pine Drought (D) 1, 27 8.89 0.0060 1, 25 3.91 0.0593 1, 25 4.11 0.0535 1, 10 1.64 0.2287

Invasion (I) 1, 27 17.43 0.0003 1, 25 0.78 0.3857 1, 25 6.53 0.0171 1, 10 0.80 0.3907

D*I 1, 27 4.78 0.0376 1, 25 0.66 0.6896 1, 25 0.51 0.4828 1, 10 1.02 0.3360

Loblolly Drought (D) 1, 27 11.87 0.0019 1, 84 17.5 < 0.0001 1, 84 12.4 0.0007 1, 25 14.26 0.0009

pine Invasion (I) 1, 27 5.05 0.0330 1, 84 3.04 0.0849 1, 84 11.3 0.0012 1, 25 5.89 0.0228

D*I 1, 27 2.72 0.1110 1, 84 0.01 0.9407 1, 84 1.0 0.3199 1, 25 0.19 0.6627

48

Table 3-3. Mean and SE of final height, diameter, biomass, and survival of slash and loblolly pine seedlings under drought and invasion

treatments.

Species Treatment Height (mm) Diameter (mm) Biomass (g) Survival

Drought Invaded mean SE mean SE mean SE mean SE

Slash pine ambient resident 32.9 4.57 7.4 0.70 11.6 2.80 52.5 8.98

ambient invaded 31.0 5.07 5.6 0.28 5.4 0.98 15.0 5.24

drought resident 22.1 2.22 5.9 0.40 4.6 0.72 20.0 4.74

drought invaded 17.7 2.28 6.9 0.07 5.3 1.01 7.5 5.06

Loblolly

pine ambient resident 69.3 3.80 7.8 0.55 18.9 2.93 92.5 3.62

ambient invaded 66.7 3.08 6.7 0.43 13.7 1.85 57.5 9.39

drought resident 54.2 6.22 6.2 0.59 11.4 2.49 50.0 7.91

drought invaded 47.9 3.72 4.6 0.32 5.7 1.12 45.0 8.51

49

Figure 3-1. Mean ± SE of soil moisture (percent volumetric water content) averaged over 2015 a)

and by month b) in plots exposed to ambient or drought conditions and with resident

species only or resident species invaded by Imperata cylindrica (cogongrass).

Figure 3-2. Mean ± SE of light availability (photosynthetically active radiation) above the

vegetation canopy at 0.5 m and at ground level averaged over 2015 (a, b) and by

month (c, d) in plots exposed to drought and invasion treatments.

50

Figure 3-3. Mean ± SE percent survival of slash a) and loblolly b) pine seedlings exposed to

drought and invasion treatments.

51

Figure 3-4. Mean ± SE of relative growth rates of height of slash a) and loblolly b) pine

seedlings exposed to drought and invasion treatments.

52

Figure 3-5. Mean ± SE of relative growth rates of diameter of slash a) and loblolly b) pine

seedlings under drought and invasion treatments.

53

Figure 3-6. Mean ± SE biomass of slash a) and loblolly b) pine seedlings grown under drought

and invasion treatments.

54

APPENDIX A

NUMBER OF BRANCHES

Figure A-1. Mean ± SE of number of limbs of slash a) and loblolly b) pine seedlings exposed to

drought and invasion treatments.

55

APPENDIX B

NUMBER OF WEBWORM NESTS

Figure B-1. Count of pine webworm (Pococera robustella) nests on slash a) and loblolly b) pine

seedlings exposed to drought and invasion treatments.

56

APPENDIX C

RELATIONSHIPS BETWEEN SOIL VOLUMETRIC WATER CONTENT AND PINE

SEEDLING RESPONSE

Figure C-1. Relationships between soil volumetric water content and slash (top) and loblolly

(bottom) pine seedling survival a), natural-log-transformed relative growth rates of

height b) and diameter c), and biomass d).

a) b)

c

)

d)

a) b)

c

)

d)

57

APPENDIX D

RELATIONSHIPS BETWEEN PHOTOSYNTHETICALLY ACTIVE RADIATION AND

PINE SEEDLING RESPONSE

Figure D-1. Relationships between photosynthetically active radiation and slash (top) and

loblolly (bottom) pine seedling survival a), natural-log-transformed relative growth

rates of height b) and diameter c), and biomass d).

a) b)

c)

d)

a)

b)

c)

d)

58

APPENDIX E

RELATIONSHIPS BETWEEN RESIDENT SPECIES COVER AND PINE SEEDLING

RESPONSE

Figure E-1. Relationships between resident species cover and slash (top) and loblolly (bottom)

pine seedling survival a), natural-log-transformed relative growth rates of height b)

and diameter c), and biomass d).

a) b)

c)

d)

d)

a)

b)

c)

59

APPENDIX F

RELATIONSHIPS BETWEEN COGONGRASS COVER AND PINE SEEDLING RESPONSE

Figure F-1. Relationships between cogongrass cover and slash (top) and loblolly (bottom) pine

seedling survival a), natural-log-transformed relative growth rates of height b) and

diameter c), and biomass d).

a) b)

c) d)

a) b)

c) d)

60

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BIOGRAPHICAL SKETCH

In 2013, Julienne E. NeSmith received a Bachelor of Science in Plant Science with a

concentration in Restoration Ecology, and minors in Agriculture and Natural Resource Law and

Wildlife Ecology and Conservation. In 2016, she earned a Master of Science in Interdisciplinary

Ecology, with a concentration in Forest Resource Conservation and a certificate in Sustainable

Development Practice. Overall, Julienne is interested in restoration ecology, invasion ecology,

community forest management, and sustainable development practice. Her graduate

research focuses on how multiple environmental stressors can interact to affect native plant

performance in southeastern U.S. pine ecosystems. In the future she hopes to explore how

ecological restoration, water use, and reduction in consumption contribute to achieving balance

between short term needs and long term ecological and environmental impacts.