the interactive effects of drought and plant...
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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.
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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
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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
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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
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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
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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
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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.