Alteration of Ecosystem Nutrient Pools and Microbial Communities after Invasion of Melaleuca quinquenervia

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Alteration of Ecosystem Nutrient Pools and Microbial Communities after Invasion of Melaleuca quinquenervia
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Biomass ( jstor )
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Microbial biomass ( jstor )
Nitrogen ( jstor )
Nutrients ( jstor )
Phosphorus ( jstor )
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Copyright 2006 by Melissa R. Martin


To my parents and siblings.


iv ACKNOWLEDGMENTS I would like to express my gratitude to my committee members, Dr. Andrew V. Ogram and Dr. Michelle C. Mack, whose pati ent instruction has allowed me to gain a deeper understanding my chosen field of resear ch. Special thanks go to the faculty, staff, and students of the Soil and Water Scien ce Department, who have provided insight, expertise, and laughter over the past two year s. Deepest thanks go to my major advisor, Dr. James O. Sickman, whose encouragement has provided me prof essional opportunities I would have not dreamed possible. I am especially appreciativ e of the invaluable insights and long hours of field support of Dr. Philip W. Tipping and the staff of the USDA-ARS Invasive Plant Research Laboratory, including Eileen Pokorny, Ryan Pierce, Matthew Smart, Emily White, and Susan Keusch. Last but certainly not least, I thank my family and friends whose confidence in my abilities has always exceeded my own. Th rough their love and support all things are possible.


v TABLE OF CONTENTS page ACKNOWLEDGMENTS.................................................................................................iv LIST OF LIST OF FIGURES..........................................................................................................vii ii CHAPTER 1 INTRODUCTION........................................................................................................1 2 MATERIALS AND METHODS.................................................................................8 Site Description............................................................................................................8 Vegetation Sampling....................................................................................................9 Soil Sampling..............................................................................................................10 Soil Characteristics.....................................................................................................10 Nutrient Analyses.......................................................................................................11 Microbial Population Size..........................................................................................11 Microbial Population Activity....................................................................................13 Statistical Analyses.....................................................................................................14 3 RESULTS...................................................................................................................15 Vegetation Processes..................................................................................................15 Soil Processes.............................................................................................................17 Microbial Biomass......................................................................................................17 Microbial Activity......................................................................................................18 4 DISCUSSION.............................................................................................................27 Species Invasion and Aboveground Effects...............................................................27 Species Invasion and Belowground Effects...............................................................31 Recommendations for Future Study...........................................................................37 LIST OF REFERENCES...................................................................................................39 BIOGRAPHICAL SKETCH.............................................................................................46


vi LIST OF TABLES Table page 3-1 General soil characterisitics.....................................................................................23 3-2 Carbon concentration of litter and soil in the invaded and non-invaded sites.........24 3-3 Nitrogen concentration of litter and soil in the invaded and non-invaded sites.......24 3-4 Phosphorus concentration of litter and so il in the invaded and non-invaded sites..25 3-5 Total litterfall biomass, total litter po ol, and turnover time in the invaded and non-invaded sites......................................................................................................25 3-6 Microbial activity measured from diff erent soil depths in the invaded and noninvaded sites.............................................................................................................26


vii LIST OF FIGURES Figure page 3-1 Standing biomass of the i nvaded and non-invaded sites..........................................18 3-2 Weights of litterfall compone nts in the non-invaded sites.......................................19 3-3 Weights of litterfall compone nts in the invaded sites..............................................19 3-4 Litterfall biomass in the invaded and non-invaded sites..........................................20 3-5 Litter pool biomass in the invaded and non-invaded sites.......................................20 3-6 Root biomass in the i nvaded and non-invaded sites................................................21 3-7 Microbial biomass in the invaded and non-invaded sites........................................22


viii Abstract of Thesis Presen ted to the Graduate School of the University of Florida in Partial Fulfillment of the Requirements for the Degree of Master of Science ALTERATION OF ECOSYSTEM NUT RIENT POOLS AND MICROBIAL COMMUNITIES AFTER INVASION OF MELALEUCA QUINQUENERVIA By Melissa R. Martin May 2006 Chair: James O. Sickman Major Department: Soil and Water Science Ecosystem invasion by exotic plant species poses a significant th reat to community biodiversity, function, and stab ility. Although many studies have discussed alterations of aboveground communities, less is known of the impact of plant invasions on belowground systems. This study investigated changes in ecosystem structure in a pinecypress forest after invasion of Melaleuca quinquenervia and the implications for soil microbial community structure and function. Melaleuca quinquenervia is an invasive Australian tree that was introduced into the Florida Everglades in the early 20th century. Due to its fire adapted nature and copious seed production, melaleuca is able to out-compete and replace many native species thereby potentially a ltering the underlying soils. F our hypotheses were tested: 1) melaleuca litter will be lower in nutrient concentration compared to cypress; 2) decomposition rates will be lower in invaded so ils; 3) microbial pools will be smaller in the invaded soils; and 4) microbial activ ity will be lower in invaded soils.


ix Results revealed that in the early stages of ecosystem invasion by Melaleuca quinquenervia, both the quantity and nutrient concentra tion of litterfall was significantly reduced. Annual litterfall was 5.5 times highe r in the non-invaded si tes, with 1.9 times higher phosphorus concentration. The non-invade d plots had a significantly larger litter pool as compared to the invaded forest treatment, 2.4 0.5 kg m-2 and 0.62 0.1 kg m-2 respectively. In addi tion, microbial biomass nutrient pools were consistently lower in the invaded soils as compared to non-invaded soils. In contrast, there was no difference in the activity of the microbial popu lations between the two sites at either depth as measured by rates of soil oxygen demand and potent ially mineralizable nitrogen. Ecosystem invasion by melaleuca si gnificantly altered both aboveand belowground nutrient storage in this study. These changes may affect both native plant growth and water quality, as well as promote a nd maintain site dominance of melaleuca. A full-scale, multi-habitat investigation of the mechanisms that drive these changes would provide critical info rmation needed to develop effective management and restoration techniques.


1 CHAPTER 1 INTRODUCTION Globalization of world markets has increa sed in the exchange of international goods and services. However, with this risi ng economic mobility have come significant environmental changes including an increase in the concentration of atmospheric greenhouse gases from the industrial use of fo ssil fuels, and the rapid deforestation of rainforests in order to meet agricultura l and timber demands (Steffen et al. 2003). However, perhaps the most ecologically a nd economically devasta ting consequence of market globalization has been the anthr opogenic distribution of many species beyond their native ranges (Ehrenfeld 2005). In th e United States alone, an estimated 5,000 nonindigenous or exotic plants have been intr oduced and become esta blished resulting in costs of approximately 34 billion dollars annually (Pimentel et al. 2000). The investigation of exotic plant introduction and invasion is a v ital area of research due to the potentially significant economic and environmenta l consequences. According to Levin (1989) non-native or exotic species are introduced into new ecosystems three ways: 1) accidentally; 2) for a specific purpose but with accidental expansion beyond the intended range; and 3) de liberate introductions on a large scale. Accidental introductions can occur through contaminated agricultural shipments or during human and animal migration (Levine and DÂ’Antonio 2003). Ornamental plants, such as water hyacinth, Eichhornia crassipes (Mart.) Solms.-Laubach, were deliberately introduced on a limited scale bu t escaped into natural areas (Devine 1998). Others like the Asian vine Kudzu, Pueraria montana (Lour.) Merr., established in the Southeastern


2 United States after repeated introductions for the purpose of reducing erosion in agricultural fields (Forseth and Teramura 1986). While so me species persist after introduction, others fail to establish or are limited by factors such as temperature, precipitation, and resource availa bility. However, species that are better adapted to the new environmental conditions may establish permanent populations. After establishment, some exotic plants th rive in new habitats where they are free of top-down regulation from herbivores and pathogens which allow them to out-compete native plants (Mitchell and Power 2003). One mitigation method employed to overcome this enemy-release advantage is classical bi ological control where coevolved host specific insect herbivores from the plantÂ’s native range are reunited with the pest in its extant range (Mack et al. 2000). This practice has lead to the succ essful control of many exotic species including the floating aquatic plant Salvinia molesta Mitchell (Room et al. 1981) . In addition to freedom from herbivory, plan ts may expedite their establishment and domination of communities by th e release of allelopathic compounds. Allelopathy is defined as non-resource competition through the release of chemical compounds that negatively effect intra-specific plant produc tion (Hierro and Callaway 2003). Callaway and Aschehoug (2000) found that root exudates from populations of the Eurasian forb Centaurea diffusa decreased the biomass production of neighboring native grasses. Once established, exotic plant populations either remain localized or spread into new areas. Although exotic plants may dominate on a local scale, they need specialized dispersal mechanisms to invade new habitats. Factors that can limit the spread of exotic plants into new ecosystems include a lack of pollinators and seed dispersal mechanisms


3 (Bryson and Carter 2004). Species traits su ch as rapid growth, hi gh rates of sexual and asexual reproduction, and early maturation can al l be predictors of exotic plant spread (Bryson and Carter 2004). For example, Brazilian pepper, Schinus terebinthifolius Raddi, a woody plant that has i nvaded South Florida, flower s twice a year and produces large quantities of bright red fruits (Ferriter et al. 2005). The fruits are eaten by birds, raccoons, and opossums, which then disperse th e seeds in their stools (Ferriter et al. 2005). However, exotic plant ecology is not the sole predictor of ecosystem invasion. Ecosystems vary in their susceptibility to i nvasion by exotic species. Specific site factors such as resource availability, the pres ence of open ecological niches, the level and frequency of disturbance, vegetation densit y, and plant species diversity influence the degree to which exotic plants establish a nd invade (Sakai et al. 2001). A recent multihabitat analysis by de Grunchy et al. (2005) c oncluded that ecosystem disturbance of any kind (hiking trails, logging, farming, or develo pment) could open an ecosystem to exotic plant invasion. Badgery et al (2005) reported that high densities of native Australian grasses reduced ecosystem i nvasion by the exotic grass Nassella trichotoma (Nees) Hack. However, given a sufficient propagule pressure it is predicted that most ecosystems will become invaded (Sakai et al. 2001). Ecosystem invasion by exotic plant species poses a significant th reat to community diversity, function, and stability. Whethe r through competition for resources or allelopathic interference, exotic species invasion has been shown to decrease plant species diversity. Kohli et al . (2004) reported that in the no rthwestern Himalayas of India plants species richness was significantly re duced by the invasion of three exotic weed species. Another study found similar results in highly disturbed, abandoned agricultural


4 lands in New Jersey where four exotic spec ies reduced the coloni zation rates of native species, thereby reducing species richness (Yurkonis et al. 2005). Although it is clear that exotic species invasion alters native plant community structure, the resulting consequences for ecosystem func tion are less predictable. Although this issue is of global concern, th e invasion of exotic plants has been a significant problem in the state of Florida, especially South Florida, which is home to Everglades. The Everglades is an extens ive forest and graminoid wetland community that once occupied 4,000 square miles of th e state’s lower peninsula, over twice its present day land area (Gunderson 1994). This complex a nd greatly altered ecosystem has proved an ideal habitat for many exotic plant sp ecies. In 1999, in order to address this growing problem, the Florida Exotic Pest Pl ant Council compiled a li st of Florida’s one hundred and twenty-five most invasive exot ic plants and separated them into two categories (Doren 2002). Sixty-six plants rece ived a Category I designation, which indicates they are considered “highly disruptive to native plant communities” (Doren 2002). These 66 plants include 20 tree species, 15 vines, 14 shrubs, 7 forbs, 6 grasses, 3 floating aquatics, and 2 submerged aquatics (Doren 2002). One of the Category I invasive plants is Melaleuca quinquenervia (Cav.) Blake otherwise known as the pape r-bark tree, cajeput, punk tree, or white bottlebrush tree (hereafter referred to as melaleuca). Melale uca is a member of the Myrtacaea family, sub family Leptospermoidae. This tall evergreen tree historically occ upied tropical wetland sites throughout its native range along the eastern coast of Australia (Kaufaman and Smouse 2001). Melaleuca was first introduced into Florida in the early twentieth century, as an ornamental plant (Bodle et al. 1994). In addi tion to its ornamental uses,


5 melaleuca was planted along canals as erosi on control and large areas of land were seeded from the air in an effort to dry out the Everglades (Bodle et al. 1994). The exotic tree colonized and thrived in most natural areas of South Florida, in cluding bayhead tree islands, sawgrass prairies, pine flat woods, pastures, and cypress forests (Bodle et al. 1994). Melaleuca has several morphological adapta tions that make it well suited to the dynamic environmental condition of Florida. Me laleuca can tolerate sh ifts in temperature and pH, moderate salinity, and variable water regimes (Kaufaman and Smouse 2001). Melaleuca grows preferentially in USDA hardin ess zones of 9a and 10b with average low temperature between 20 F and 40 F, with light freezes (Turner et al. 1998). Myers (1983) reported that melaleuca seedling can germ inate in both acid and alkaline soils. In addition, dense and productive stands of melaleuca can be found in well drained, seasonally saturated, and permanently flooded sites (Bodle et al. 1994). Plant survival is enhanced by large numbers of arenchym enous surface roots produced shortly after flooding (Serbesoff-King 2003). In order to survive prolonged dry periods, trees produce sinker roots that extend down to the water table (Serbesoff-King 2003). Melaleuca is also a competitive rooter (Lopez-Zamora et al. 2004). Not only can melaleuca grow in variable water tables but it aggressively colonizes areas dominated by roots of native plants (Lopez-Zamora et al. 2004). Once established, melaleuca trees have a high growth rate and reach reproductive maturity at three years after which it may flow er two to five times a year (Bodle et al. 1994). It is estimated that a single Mela leuca tree can hold an estimated 5.6 million viable seeds in its canopy, which once released may remain viable in the soil for up to


6 three years (Rayachhetry et al. 2002 and Van et al. 1998). Seeds are held in capsules that open in response to desiccation, frost, or fire (Serbesoff-King 2003). In addition, high concentrations of essential oils found in matu re Melaleuca trees fuel canopy fires that can kill native vegetation. Following destruc tive canopy fires, Melaleuca will release massive amounts of seeds, which may result virtual monocultures of Melaleuca (Serbesoff-King 2003). By the early 1980s it became clear that Melaleuca was significantly altering the plant communities of the Flor ida Everglades. In 1980, a survey completed by the U.S. Forestry Service revealed that 470,000 acres of land was covered by “pure” Melaleuca stands (50% or greater) (Bodle et al. 1994). LaRoche and Ferriter (1992) es timated that once melaleuca populations reached a critical ecosystem concentration of two to five percent, ninety-five percent in festation occurred within 25 ye ars. Once an ecosystem is invaded by melaleuca, native plant communities are significantly altered (Mazzotti et al. 1997). DiStefano and Fisher (1983) found th at the relative density, frequency, and dominance of several native plant species wa s significantly diminished in sites invaded by melaleuca as compared to neighboring non-in vaded sites. This difference was explain in part by potential allelopathic properties of melaleuca leaf extracts that reduced seed germination and seedling growth of native pl ant species. In add ition, Mazzotti et al. (1981) found decreased diversity and abundance of wildlife, namely small mammals, in melaleuca stands as compared to native plant communities. Myers (1984) described one of South Fl orida’s most susceptible habitats for melaleuca invasion, the pine-cypress ecotone. This is the transition zone between upland pine-dominated sites and depressional cypressdominated swamps. In this area, pine and


7 cypress trees co-dominate but neither plant is able to grow to its full potential (Myers 1983). Based on a combination of greenhouse an d field experiments, Myers showed that in the pine-cypress ecotone bot h melaleuca seedling survival and growth rate was high, indicating the suitability of this area for melaleuca inva sion (Myers 1984). Ewel (1986) agreed with MyersÂ’ assessment and stated that ecosystem alteration after such introductions is inevitable and that only by gaining an understanding of the interactions between native and introduced species can we develop effective management strategies. The major objective of this work was to de termine if invasion of the exotic plant Melaleuca quinquenervia induced both aboveand belowground changes in a cypress dominated eco-tone forest. In particular, wh at were the consequences for nutrient pools and microbial population dynamics? Two overa rching hypotheses were tested. First, melaleuca litter will have lower nutrient concentr ations compared to cypress. I expected this because of differences in plant characte ristics: melaleuca is an evergreen tree and cypress is deciduous. Chabot and Hicks (1982) found that evergreen plants tend to have lower foliar nutrient concen trations compared to deciduous plants. Second I hypothesized that changes in nut rient concentrations of melaleuca litt er will: 1) lower decomposition rates in invaded soils; 2) lowe r microbial nutrient pools in the invaded soils; and 3) lower microbial activity in i nvaded soils. A better understanding of the belowground ecosystem alterations caused by exotic plant invasion will provide vital information needed to develop effective management and restoration techniques.


8 CHAPTER 2 MATERIALS AND METHODS Site Description The study site is located in the Belle Meade Tract of the Picayune Strand State Forest in Collier County, Florida. A 1998 survey completed by the United States Department of Agriculture, Natural Resource Conservation Service mapped this area as Pineda-Boca-Hallandale soil seri es. The soils are moderate ly to poorly drained sands over-lying limestone bedrock at a depth of approximately 1.4 m. The water table in this area fluctuates annually between greater than 15 cm below th e soil surface to approximately 25 cm above. The area has distin ct wet season from approximately July to December and a dry season from January to June. Vegetation in this area was a mixed pine -cypress forest with hardwood under-story and a few mature Melaleuca quinquenervia (hereafter ‘melaleuca’) trees. In August 1998, the area experienced a canopy fire that killed many of the native trees and precipitated a large release of seeds from melaleuca trees. This was followed by a concomitant pulse of recruitment of melaleu ca seedlings that exceeded densities of 100 plants per square meter. The burned area was thus transformed into a landscape of sparsely distributed mature melaleuca trees surrounded by a virtual monoculture of even aged saplings. Those areas not burned maintain ed their original structure and tree species composition.


9 Vegetation Sampling In February 2005, seven plots were esta blished in each of two areas: an area dominated by mature melaleuca trees and sa plings (hereafter re ferred to as “invaded plots”) and an area dominated by mature Cypress taxodium var nutans trees (hereafter referred to as “non-invaded plot s”). Each circular plot m easured 6 m in diameter and contained a mature tree in the cen ter. Aboveground biomass of the Cypress taxodium var. nutans in each plot was estimated from thre e separate regression equations using diameter at breast height (DBH) and total tree height (HT) (Mitsch and Ewel 1979). DBH of all cypress trees within the plot was measured with a diameter tape. The height of larger trees was estimated with a clinometer while shorter trees were measured directly with a tape measure. The general form of the cypress biomass regression equation used was: y = a * xb (where y = biomass, and x = (dbh * ht)-1/2). Aboveground biomass of mature melaleuca tr ees was estimated using diameter tape measurements of diameter outside bark (DOB) (Van et. al 2000). The regression equation used was: Loge (W) = -1.83 + 2.01 * Loge (DOB) (where W = total biomass). The biomass of the melaleuca under-story in each plot was measured directly by collecting all of the stan ding vegetation in two 1 m2 sub-plots. All biomass measurements were reported on a dry weight basis. To determine litterfall in each plot, four 0.25 m diameter litter traps were placed at four cardinal directions approxi mately 1 m from the base of the mature tree. Litterfall was collected every sixty days for eight months and separated into component parts. The organic litter pool in each treatment was sampled from two 0.1 m2 sub-plots. The litter layer was separated into an undecomposed and moderately decomposed Oi and Oe layer,


10 and a humified Oa layer. Litterfall and the l itter pool samples were air dried and reported on a dry weight basis. Soil Sampling To estimate belowground biomass in each circular plot two-5 cm diameter soil cores were taken at each plot at a distance of 1 m from the base of the mature tree. Each belowground biomass core was separated at two depths, 0-5 cm a nd 5-15 cm. The soil cores were sieved to separate fine roots, rinsed, air-dried, and th e mass reported on a dry weight basis. In February 2005, four-5 cm diameter so il cores were taken in each plot at a distance of 1 m from the base of the native or invasive mature tree. Each core was separated at two depths, 0-5 cm and 5-15 cm, and combined to form two composite soil samples per plot at each depth. Soil samples we re returned to the laboratory, sieved to remove roots and large plant debr is, homogenized, and kept at 4 C for a maximum of 10 days before microbial analysis. Soil Characteristics Percent moisture and bulk density of soils were determined by drying twenty to thirty gram sub-samples of field-mois t, sieved, and homogenized soil at 70 C for three days. Bulk density and percent moisture were determined on a wet soil weight basis and pH was measured on a 2:1 water:soil slu rry with an Accumet Research, AR50 dual channel pH/ion/conductivity meter. Soil or ganic matter was measured by loss on ignition from 0.2 to 0.5 g samples of dried and ground soil s, which were first measured into 50 ml beakers (Luczak et al. 1997). The beakers were put in a muffle furnace and brought to


11 250 C for 30 minutes. The furnace temper ature was then increased to 550 C for 4 hours. Organic matter content was calculated as the ma ss loss on ignition on a dry weight basis. Nutrient Analyses Dried and ground soil and plant material was analyzed for percent carbon and nitrogen on a Thermo-Electron, 1112 Series elemental analyzer. Total phosphorus was determined by a two-phase acid extraction after loss on ignition (Anderson 1976). The ash remaining in the 50 ml beaker was moiste ned with 2 to 3 ml distilled de-ionized water and then extracted with 20 ml of 6 N HCl. After all of the water was removed, the hot plate was placed on high for 30 minutes to completely dry samples. After cooling, 2.25 ml of 6N HCl was added to each beaker and the beakers placed on a hot plate until almost boiling. Extracts were then filte red through a #41 Whatman filter into 50 ml volumetric flasks. Flasks were brought to vol ume with distilled de-ionized water. Total phosphorus was measured with an automated ascorbic acid method on a Bran and Luebbe Auto Analyzer 3, Digital Colorimeter (Method 365.4; USEPA 1993). Microbial Population Size Microbial biomass carbon (MBC) and n itrogen (MBN) were measured by a chloroform-fumigation extraction method modi fied from Vance et al. (1987). Two replicates of 1 g field-moist soil samples were weighed into 50 ml centrifuge tubes. One of the duplicates was immediately extracted with 25 ml of 0.5M K2SO4, shaken for 1 hour, and then filtered through a Whatman #41 filter. The second replicate of soil underwent chloroform fumigation where 0.5 ml of chloroform was added to each tube and the tubes were placed in a dessicator wi th a beaker containing 30 ml chloroform and boiling chips. The dessicator was then vacuum-sealed fo r 24 hours. After fumigation, the dessicator was alternatively filled with air and evacuated ten times to remove all


12 residual chloroform. The tubes were then ex tracted as described above and all extracts were stored at 4 C until analysis for MBC. This wa s calculated as the difference between the total carbon in the fumigated and un-fu migated soil extracts as measured on a Shimadzu TOC 5050C, total organic carbon analy zer. An extraction coefficient of 0.45 was applied to all calculations of microbi al biomass carbon (Chen et al. 2005). To measure MBN, 10 ml of the 0.5M K2SO4 extract from the fumigated and unfumigated tubes, 2 ml of concentrated H2SO4, and one scoop of Kjelda hl salt catalyst (ca. 0.6 g) were placed in 40 ml digestion tubes. Tubes were placed on a digestion block and brought to 125 C for 2-3 hours, then 150 C for 2-3 hours until all water was removed, and finally at 380 C for 2-3 hours until samples were clear and digestion was complete. Deionized water was added to the acid digestat es to reach a final volume of 20 ml. MBN was calculated as the difference between the to tal Kjeldahl nitrogen in the fumigated and unfumigated acid digestates as measured on a Bran and Luebbe Auto Analyzer, Digital Colorimeter (Method 351.2; USEPA 1993). An extraction coefficient of 0.2 was applied to all calculations for microbial bi omass nitrogen (Chen et al. 2005). Microbial biomass phosphorus (MBP) was m easured with the same chloroformfumigation extraction method described above with the exception th at after fumigation, 25 ml of pH adju sted 0.5M NaHCO3 was used to extract the soil, tubes were shaken for 16 hours and extracts were filtered through a 0.45 m filter (Brookes et al. 1982). Five ml of the soil extracts were digested with 1 ml of 11 N H2SO4 and one scoop potassium persulfate (ca. 0.6 g) using the same di gestion block program described above. Deionized water was added to reach a final volume of 10 ml. MBP was calculated as the difference between the total phosphorus in fumi gated and unfumigated acid digestates as


13 measured with an automated ascorbic acid method on a Bran and Luebbe Auto Analyzer 3, Digital Colorimeter (Method 365.4; USEPA 1993). Microbial Population Activity Soil oxygen demand is a measurement of the dissolved oxygen consumed during a twenty-four hour dark incubati on (Malecki et al. 2004). Five to six grams of field moist soil were weighed into dark, biological oxyge n demand bottles with a magnetic stirrer. Bottles were then completely filled with oxyge n-saturated water, put on a stir plate and agitated for 30 minutes, and an initial measurement of dissolved oxygen was taken with an Accumet Research, AR40 dissolved oxygen meter and recorded after which bottles were incubated at room temperature for 24 hour s. After incubation the bottles were again agitated and a final dissolved oxygen readi ng was taken. Oxygen demand was reported as the difference between the final and initia l dissolved oxygen concentrations per g of soil per hour. Potentially mineralizable nitrogen (PMN ) was measured with a method modified from White and Reddy (2000). This method cal culates a mineralization potential based on the net production of ammonium during a 10 day incubation. In order to determine the initial amount of ammonium in the soil, a set of 1 g field-moist soil samples were weighed into 50 ml centrifuge tubes. The tubes received 25 ml of 0.5M K2SO4, shaken on a longitudinal shaker for one hour, and ex tracts were filtered through a #41 Whatman filter. An additional set of 1 g field-moist soil samples were weighed into 50 ml serum bottles with 5 ml of deionized water. Bottles were sealed with butyl rubber stoppers and aluminum crimps. The bottle head-space was purged with O2-free N2 gas for 2-5 minutes. Bottles were inc ubated in the dark at 40 C for 10 days then extracted as


14 described above with 25 ml of 0.5M K2SO4 to determine final ammonium concentration. Total ammonium was determined with an au tomated colorimetric method on a Bran and Luebbe Auto Analyzer 3, Digital Co lorimeter (Method 353.2; USEPA 1993). Statistical Analyses The experimental unit in this study was the mature tree at the center of each plot and not the entire melaleuca and cypress fore sts, thereby eliminating issues of pseudoreplication. Values of measur ed soil and vegetation characteri stics, as well as microbial population size and activity were calculated as a mean for each plot. All error values were reported as standard errors (S.E.). StudentÂ’s t tests were used to detect any differences between the measured parameters. Data that varied from normal distributions was analyzed nonparametrically with the Kruskal-Wallis test on rank sums. All statistical analyses were pref ormed using JMP 4.0 software.


15 CHAPTER 3 RESULTS Vegetation Processes The mean density (+ S.E.) of melaleuca saplings in the invaded plots was 40 23 saplings m-2. In the non-invaded plots there was an average of 16 2 Cypress taxodium var. nutans trees in each 28.3 m-2 plot. The mean height a nd DBH of the cypress trees were 360 29 cm and 5.5 0.6 cm, respectively. Using the regression equations developed by Mitsch and Ewel (1979), I es timated the average aboveground biomass of cypress to be 1.2 kg m-2 for the “mature tree” and 2.26 kg m-2 for all other understory trees, with a total of 3.46 kg m-2 of cypress biomass (Figure 31). I also used regression equations developed by Van et al. (2000) for the “mature” me laleuca tree to obtain an estimate of aboveground biomass of 4.5 kg m-2 (Figure 3-1). Direct measurements of understory biomass yielded a mean of 0.48 kg m-2, with an overal l total of 4.98 kg m-2 of melaleuca biomass (Figure 3-1). There wa s no difference in total aboveground biomass between the melaleuca and the cypr ess-dominated plots (Figure 3-1). Examination of the components of the litterfall data revealed that in each treatment area the dominant tree comprised the bul k of the material (Figures 3-2 and 3-3). Average litterfall was 42.4 ± 4.5 g m-2 yr-1 for the invaded plots and 234.2 ± 35.8 g m-2 yr-1 in the non-invaded plots, a 5.5 fold diffe rence (Table 3-5). There was a temporal component to litterfall differences among the treatments; differences among sites were first detected in July 05 and through Oct ober 05 (Figure 3-4). Overall, litterfall


16 production per unit tree biomass was approxima tely 7 times greater in the non-invaded plots compared to the invaded plots, 69.4 8.9 g kg-1 and 10.1 1.7 g kg-1 respectively. There were no differences in the amounts of carbon and nitrogen concentrations in the litterfall between sites (T able 3-2 and 3-3). However, in addition to the decreased quantity of the invaded plot li tterfall, total phosphorus con centrations average 47% less than in the non-invaded soil (p<0.05; Table 3-4). Th e non-invaded plots had a significantly larger total litter pool compared to the invaded plots (Figure 3-5). The Oi / Oe layer was 3 times larger in the non-invaded pl ots compared to the invaded plots, with values of 1.79 ± 0.11 kg m-2 and 0.57 ± 0.25 kg m-2 respectively (Table 3-5). The Oa layer was 12 times larger in the non-invaded plots compared to the invaded plots, with values of 0.6 ± 0.03 kg m-2 and 0.05 ± 0.02 kg m-2 respectively. The turnover time for each litter pool was not different between sites with a mean (+ S.E.) of 11.2 (+ 2.5) and 15.2 (+ 2.3) years in the non-invade d and invaded sites, resp ectively (Table 3-5). As with litterfall, there was a significant decrease in C, N and P concentrations of the invaded plot Oi / Oe layer relative to non-invaded soils (Tables 3-2, 3-3, and 3-4). The Oi / Oe layer in the invaded plots had 34% less total nitrogen concentration and 39% less total phosphorus concentration (Tables 3-3 and 3-4). Nutrient concentrations of the Oa layer in the invaded plots followed the same pa ttern with a 28% decrease in total nitrogen concentration and a 35% decrea se in total phosphorus concen tration as compared to the non-invaded plots (Tables 3-3 and 3-4). Root biomass was significantly higher in th e surface (0-5 cm) so il layer of invaded plots as compared to the noninvaded plots, with means (+ S.E.) of 210 ± 26.3 g m-2 and 118 ± 21.9 g * m-2, respectively (Figure 3-6). In addi tion, root biomass was significantly


17 higher in the subsurface (5-15 cm) soil layer of invaded plots as compared to the noninvaded plots, with means (+ S.E.) of 354 ± 32 * m-2 and 157 ± 17.2 g * m-2, respectively (Figure 3-6). Although the quant ity of root biomass was high er in the invaded soil, the concentration of nitrogen in the root biomass was 29% and 39% less in the surface and subsurface sample compared to the non-invade d soils. Phosphorus concentration of the root biomass was not measured. Soil Processes Percent moisture was higher in the surface (0-5 cm) as compared to the subsurface (5 – 15 cm) soil layers of both the invaded and non-invaded plots (Table 3-1). Percent moisture in the subsurface soil la yer of the melaleuca-dominated plots was significantly lower than all other soil layers . pH values followed the same pattern as percent moisture: highest in the surface soil layers of both treatments and lowest in the subsurface soil layer of the melaleuca-dominated plots. Bulk density was significantly higher in the surface soils of both treatments . The two surface soils had significantly different bulk densities with the cypress-plot s having a greater bulk density. Bulk density was not determined for the organic layer. Th e organic matter content of the surface soils was significantly higher than the subsurface so ils at both sites. There was no significant difference in organic matter or nutrient concen tration between sites at either soil depth (Tables 3-1, 3-2, 3-3, and 3-4). Microbial Biomass Microbial biomass was consistently lowe r in the invaded plot soils at both soil depths (Figure 3-7). Microbi al biomass carbon was 17% lower in the invaded soils for the 0-5 cm soil depth and 25% lower for the 5-15 cm depth (Figure 3-7a). Microbial biomass nitrogen and phosphorus followed a si milar pattern. MBN wa s 34% lower in the


18 invaded soils for the 0-5 cm soil depth and 48% lower for the 5-15 cm depth (Figure 37b). MBP was 39% lower in the invaded soil s for the 0-5 cm soil depth and 55% lower for the 5-15 cm depth (Figure 3-7c). In bot h treatments the amount of microbial biomass was lower in the deepest soil layer sampled (Figure 3-7). Microbial Activity There were no differences in microbi al activity between the experimental treatments at either depth (Table 3-6). In both treatments microbi al activity declined sharply with depth (Table 3-6). Ther e was also a 75% and 65% decrease in mineralization capacity in the subsurface so il of the invaded and non-invaded plots, respectively, and a 58% and 64% decrease in ox ygen demand in the subsurface soil of the invaded and non-invaded plots, respectively (Table 3-6). 0 2 4 6 8 "Mature tree""Under-story"TotalBiomass (kg m-2) Invaded Non-invaded Figure 3-1. Mean standing biomass ( S.E.) of the “mature trees”, “Under-story” vegetation, and total biomass as estima ted by published regression equations for the invaded and non-invaded sites.


19 0 50 100 150 200 250 Non-invadedBiomass (g m-2) Other leaf Other wood Pine Cypress leaf Melaleuca leaf Figure 3-2. Mean weights ( S.E.) of litterfall component s for the non-invaded sites. 0 5 10 15 20 25 30 35 40 45 InvadedBiomass (g m-2) Melaleuca seed Other wood Pine Cypress leaf Melaleuca leaf Figure 3-3. Mean weights ( S.E.) of litterfall component s for the invaded sites.


20 1 10 100 1000 10000 AMJJASONDBiomass (log (g m-2)) Invaded Non-invaded Figure 3-4. Mean litterfall biomass ( S.E.) of each sample collection for the invaded and non-invaded sites. 0.01 0.10 1.00 10.00 100.00 Oi and OeOaTotalBiomass (log (kg m-2)) Non-invaded Invaded Figure 3-5. Mean litter pool biomass ( S.E.) of the Oi and Oe, Oa , and total litter pools for the invaded and non-invaded sites.


21 0 100 200 300 400 500 0-5cm 5-15cmBiomass (g m-2 cm-1) Non-invaded Invaded Figure 3-6. Mean root biomass ( S.E.) of 0-5cm and 5-15cm so il depths for the invaded and non-invaded sites.


22 Figure 3-7. Mean microbial biomass carbon (a), nitrogen (b), and phosphorus (c) ( S.E.) of 0-5cm and 5-15cm soil depths for the invaded and non-invaded sites. a b c Soil Depth g C g-1 g N g-1 g P g-1 0 2 4 6 8 0-5 cm5-15 cm 0 250 500 750 1000 1250 1500 Non-invaded Invaded 0 50 100 150 200Microbial biomass


23 Table 3-1. Mean (+ S.E.) of soil variables measured from different depths in the invaded and non-invaded sites at Picayune State Forest, 2005. Variable Sample Location Non-invaded Invaded P pH 0-5 cm depth 5.4 + 0.2 5.3 + 0.3 0.75 5-15 cm depth 5.3 + 0.2 5 + 0.3** 0.02 ---------------% ---------------Moisture 0-5 cm depth 5.6 + 0.4 5.0 + 0.2 0.1 5-15 cm depth 5.3 + 0.2 4.1 + 0.2** 0.007 ------------mg cm-3 ------------Bulk Density 0-5 cm depth 1.3 + 0.1 1.1 + 0.1** 0.04 5-15 cm depth 1.5 + 0.02 1.5 + 0.1 0.1 ---------------% ---------------Organic Matter 0-5 cm depth 4.1 + 0.9 4.1 + 0.9 0.98 5-15 cm depth 1.7 + 0.3 1.3 + 0.3 0.22 (**Indicates significant P values)


24 Table 3-2. Mean (+ S.E.) of carbon content measured in the litter and soil of the invaded and non-invaded sites. Variable Sample Location Non-invaded Invaded P --------------g g-1--------------Carbon Litterfall 0.5 + 0.01 0.5 + 0.004 0.1 Litter (Oi and Oe) 0.4 + 0.02 0.5 + 0.1** 0.01 Litter (Oa) 0.4 + 0.1 0.4 + 0.04 0.8 -------------mg g-1 -------------0-5 cm depth 14.6 + 2.5 18.5 + 3.9 0.2 5-15 cm depth 6.5 + 1.4 4.2 + 0.8 0.2 (**Indicates significant P values) Table 3-3. Mean (+ S.E.) of nitrogen content measur ed in the litter and soil of the invaded and non-invaded sites. Variable Sample Location Non-invaded Invaded P --------------mg g-1--------------Nitrogen Litterfall 10.2 + 0.3 8.9 + 0.5 0.1 Litter (Oi and Oe) 18.3 + 0.4 12.1 + 0.4** 0.1 Litter (Oa) 20.8 + 0.8 15 + 0.5** 0.03 --------------mg g-1--------------0-5 cm depth 0.9 + 0.2 0.8 + 0.03 0.6 5-15 cm depth 0.4 + 0.1 0.3 + 0.03 0.1 (**Indicates significant P values)


25 Table 3-4. Mean (+ S.E.) of phosphorus content measured in the litter and soil of the invaded and non-invaded sites. Variable Sample Location Non-invaded Invaded P --------------g g-1--------------Phosphorus Litterfall 220.6 + 25.1 117.2 + 5.9** 0.02 Litter (Oi and Oe) 255.4 + 12.1 156.5 + 11.3** 0.01 Litter (Oa) 346.3 + 25.3 225.8 + 32.5** 0.03 --------------g g-1--------------0-5 cm depth 24.4 + 5.8 18.1 + 1.2 0.3 5-15 cm depth 7.8 + 0.5 6.7 + 0.4 0.1 (**Indicates significant P values) Table 3-5. Mean (+ S.E.) of total litterfall biomass, to tal litter pool, and turnover time in the invaded and non-invaded sites fo r the eight-month sampling period. Variable Non-invaded Invaded P ----------g m-2 ---------Litterfall 234.2 + 35.8 42.4 + 4.5** <0.001 ----------kg m-2 ---------Litter pool 2.4 + 0.5 0.6 + 0.1** <0.001 -----------year-----------Turnover time 11.2 + 2.5 15.2 + 2.3** 0.3 (**Indicates significant P values)


26 Table 3-6. Mean (+ S.E.) of microbial activity measured from different soil depths in the invaded and non-invaded sites. Variable Sample Location Non-invaded Invaded P ---------------g g-1 day -1---------------Potentially mineralizable 0-5 cm 1.04 + 0.2 1.4 + 0.2 0.2 nitrogen 5-15 cm 0.3 + 0.06 0.5 + 0.1 0.1 ---------------g g-1 hour -1---------------Soil Oxygen Demand 0-5 cm 2.4 + 0.4 2.2 + 0.2 0.7 5-15 cm 1.0 + 0.1 0.8 + 0.1 0.3 (**Indicates significant P values)


27 CHAPTER 4 DISCUSSION Species Invasion and Aboveground Effects Species alteration may reduce or eliminat e an ecosystemÂ’s ability to provide ecological goods and services, such as wa ste processing and car bon sequestration (Fenn et al. 2003). However, studies have found both positive and negative changes in the rates of ecosystem nutrient storage and cycling (Ehrenfeld 2003). Functional changes in plant structure and growth rate can alter aboveand belowground nut rient pool sizes (Ehrenfeld 2003). For example, grassland invasion by w oody plants has been shown to increase the storage of carbon in standing biomass (Jacks on et al. 2002). In addition, exotic plants may differ from native species in litt er nutrient concentration and relative decomposability (Ehrenfeld 2003). A sample of 30 invasive species from Hawaii was found to have higher foliar nutrient levels as compared to native plants, potentially altering the rate of ecosystem nutrient fluxes (Baruch and Gold stein 1999). In this study, invasion of a pine-cypress eco -tone by melaleuca resulted in changes in both forest structure and ecosystem nutrient pools. The pine-cypress eco-tonal habitats of sout hern Florida are natu rally fire and flood regulated (Myers 1984). While both tree sp ecies are fire-adapted, cypress is more tolerant of longer periods of inundation and tends to domin ate in wetter areas (Myers 1984). The non-invaded plant community obser ved in this study was a closed canopy mature pine-cypress forest that experienced a hydro-period of approximately six months. Cypress was the dominant tree species comprising on average 77 8.3 % of the woody


28 species present. Therefore, standing biom ass estimates for the non-invaded plots were made solely on the aboveground biomass of the cypress trees. On average each 28 m2 plot contained seven mature cypress trees th at consisted of 99.8% of the total biomass and seven cypress saplings that comprised the remain 0.2%. The community structure in the invaded pl ots was very different from the native plots and consisted of sparse mature melaleuca trees with an under-story of dense and even-aged saplings. On average each 28 m2 plot contained one mature melaleuca tree that consisted of 90% of the total biomass and 1131 melaleuca saplings that comprised the remaining 10%. Although the aboveground stan ding biomass did not differ significantly for the dominant species, this temporary al teration of plot structure had significant consequences for ecosystem nutrient pools. Av erage litterfall was 5.5 times greater in the non-invaded plots. This difference may be explained in part by differences in plant characteristics. Cypress is a de ciduous tree that sheds all of it s leaves in late fall to early winter (Ewel and Odum 1984). In this study there was a distinct seasonality to the litterfall in the non-invaded pl ots; 70% of the litterfall was produced from August to November. Litterfall measured in the non-inva ded plots was similar to values reported by Brown (1981) of 224 and 597 g dry weight m-2 year-1 in un-impacted seasonally flooded cypress forests with mature tr ee densities (DBH>2.5 cm) of 0.022 to 0.29 m-2 respectively. I measured an aver age litter product ion rate of 234.21 35.8 g dry weight m-2, with mature tree densities of 0.25 m-2 (Table 3-5). In contrast, melaleuca is an evergreen tr ee, which can hold its leaves in the canopy for two years (Van et al. 2002). Unlike the non-invaded plots, litte rfall in the invaded plots was lower than reported in previous studies. Van et al . (2002) reported litter


29 production in South Florida of 750 to 930 g dry weight m-2 year-1, with the highest values in a seasonally flooded site. An Australian study from melaleucaÂ’s native range reported values of litter production rangi ng from 675 to 809 g dry weight m-2 year-1 in seasonally inundated sites (Greenway 1994). I measur ed litter production of only 42.36 g dry weight m-2 in the eight-month sampling period (Tab le 3-5). These differences may be explained by differences in melaleuca forest structure. Both of the aforementioned studies cited were conducted in closed-canopy melaleuca stands with densities of mature trees (DBH>20 cm) of approximately 0.1 and 0.15 trees m-2 respectively. The melaleuca forest in the current study is in the early stages of fo rest development with only 0.04 mature trees m-2, and a dense under-story of 40 melaleuca saplings m-2. Another factor potentia lly influencing litte rfall in this study was the presence of two biological control agents released on melaleuca over the past decade: Oxyops vitiosa Pascoe (Coleoptera: Curculionidae) and Boreioglycaspis melaleucae Moore (Hemiptera: Psyllidae). Pratt et al. ( 2005) found that herbivory from O. vitiosa significantly altered melaleuca growth and reproduction. Alt hough not quantified, the melaleuca leaves collected during litterfall sampling in this study had noticeable insect feeding damage. High levels of herbivory c ould be responsible for the low litter production per unit standing biomass in the invaded site. Future re search is warranted not only to investigate changes in foliage production and mass-loss of litter but also potential plant nutrient allocation and defense responses to insect herbivory. Both the quantity and nutrient concentrati on of the litterfall in the invaded plots were lower compared to the non-invaded pl ots (Tables 3-2, 3-3, and 3-4). The total litterfall in the invaded plot s had 47% less total phosphorus c oncentration as compared to


30 the non-invaded plots and th erefore approximately 10 times more phosphorus per m-2 was added to the forest floor in the non-invaded area during the course of this study. The nutrient poor, sandy surface soils of these fore sts do not have a large capacity to store nutrients vital for ecosystem maintenance. Therefore reducing the quantity of nutrients returned to the soil through litterfall could have significan t consequences for forest productivity. This finding may represen t a potential mechanism through which melaleuca is able to promote and maintain ecosystem dominance. The litter pool was also significantly larger in the non-invaded pl ots with almost 4 times as much litter accumulated on the forest floor as compared to the invaded plots. As before in the invaded plots, there was also a significant decrease in the nutrient concentration of the litter pool. Both the Oi / Oe and the Oa litter pools of the invaded plots had significantly less nitrogen and phos phorus than the noninvaded litter pools (Tables 3-2, 3-3, and 3-4). Lower nutrient con centration and quantity of the litter layer in the invaded plots lead to changes in the aboveground storage of nutrients where three times less carbon, six times less nitrogen, and seven times less phosphorus were stored in the organic litter layer per m-2 compared to the non-inva ded plot. Reductions in aboveground nutrient inputs and storage cau sed by invasion may have significant consequences for long-term ecosystem function. I hypothesized that the lower nutrient concen tration in the melaleuca litter would decrease its turnover time in the invaded plots. In order to estimate turnover or residence time of the litter, was assumed that the ecosystem is in steady state (Chapin et al. 2002). However, despite differences in nutrient c oncentration, turnover ti me of the litter pools were not different between the two sites indi cating that other envi ronmental factors may


31 be controlling the rate of litter decom position such as organic carbon composition and soil temperature (Table 3-5). In addition, the lack of difference in turnover time, coupled with differences in litterfall production, explains the larger litter laye r present in the noninvaded plots. While there are no published values for mela leuca litter, the tu rnover time of the cypress litter in this study was longer than previously reported. Nessel and Bayley (1984) estimated turnover time of cypress litt er by the same method and found it to be 3 years compared to 11.2 years in this study. The reason for this difference is unknown but this study site was subjected to multiple high wind events caused by the close passage of several hurricanes, which probabl y increased estimates of the litter pool. Another factor that may have lead to an over-estimation of the litter pool is the seasonality of cypress litterfall. The li tter pool collection in this study was taken a mont h after the cypress leafdrop and before the period of seasonal i nundation. Repeated sampling throughout the year, to quantify seasonal variation in the litter pool, would provide a more accurate estimation of turnover time. In addition, th e sampling period for lit terfall collection was relatively short in this study better estimates of litter tur nover would be produced from a yearlong collection. These factors could have resulted in an over-estimation of litter turnover time in both invaded and non-invaded plots. Species Invasion and Belowground Effects Alteration of ecosystem structure has b een shown to reduce ecosystem stability. Exotic plants can alter res ource availability, fire regi mes, site hydrology, and light availability (Mack and DÂ’Antoni o 1998). The invasive plant Myrica faya Aiton, a nitrogen-fixing tree native to the Canary Islands, invaded Hawaii and increased the supply of soil nitrogen by ninety-fold (Vitous ek and Walker 1989). Increased nitrogen


32 availability paved the way for the invasion of additional plants with nitrogen needs greater than was supplied by the native soils (Vitousek and Walker 1989). In addition, exotic plants such as water hyacinth and gi ant salvinia have invaded every continent except Antarctica, with devastating conseque nces: fisheries have collapsed, waterways have become impassable, and drinking wate r supplies have been degraded (Gopal and Sharma 1981). Elucidation of the extent, du ration, and impact of the changes caused by invasive plants will help in developing more effective restoration and management techniques. One of the ways in which melaleuca is able to out-compete native plants and dominate ecosystems is through avoiding and withstanding root competition. LopezZamora et al. (2004) found that melaleuca plants developed signi ficantly higher root densities than native grass competitors and maintained high root densities throughout the soil profile independent of moisture conten t. DiStefano and Fisher (1983) found that melaleuca stands had higher root densities in the upper 20 cm of soil when compared to stands of both Pinus elliottii and Eucalyptus gradis. The current study supports these findings: fine root densities were significan tly higher at each soil depth in the invaded forest treatment (Figure 3-6). The aggressi ve growth and water to lerance of melaleuca roots may help the tree compete for nutri ents and maintain ecosystem dominance. Differences in canopy structure, litter laye r, and fine rooting depth between the two treatments may be responsible for the decrea sed soil moisture in the sub-surface horizon of the invaded plots (Table 3-1) . In the early stages of deve lopment, the soil surface in a melaleuca plot receives less shade than th e closed canopy of a mature cypress stand, which may increase rates of evaporation. In addition, the thicker orga nic litter layer on


33 the soil surface of the non-invaded plots w ould reduce evaporative water loss and inhibit the establishment of under-story plant spec ies that might reduce water resources. My study found significantly greater biomass of fine roots in the soil of the invaded plots, most likely the result of the dense melaleu ca-sapling under-story. In periods of water stress increased fine roots could result in higher rates of water transpiration. As hypothesized there were significant changes in microbial biomass nutrient concentrations between treatments. Values of microbial bioma ss carbon, nitrogen, and phosphorus were consistently lower in the invaded plots (Figure 3-7). Although the mechanisms by which melaleuca alters the so il environment were not explored in this study, differences in microbial pools between si tes points to a potential plant mediated change. Melaleuca alternafolia , a close relative of Melaleuca quinquenervia , is known to produce volatile essential oi ls with antimicrobial propert ies (Carson et a. 2006). Other studies have presented limited evidence that Melaleuca quinquenervia (hereafter referred to as “melaleuca”) may produce similar compounds that may affect ecosystem function. DiStefano and Fisher (1983) found that extr acts of melaleuca leaves reduced seed germination and seedling growth of native plan t species. In addition, fungal infection of plant embryos was significantly reduced by the melaleuca leachate treatment. The production and release of alle lopathic compounds by melaleuc a may be responsible for alterations in microbial biomass pools and is deserving of further study. Alteration of microbial bioma ss nutrient concentrations ma y also suggest shifts in microbial community composition. On average soil bacteria have C:N ratios of 4:1 to 10:1, while fungi range from 8:1 to 15:1 (Davet 2004). In this study, the C:N ratio of the microbial biomass was 11:1 and 9:1 in the surf ace soil and 9:1 and 7:1 in the subsurface


34 soil of the invaded and non-i nvaded plots respectively. Al though it is expected that fungal biomass would dominate in the acidic soils of these forests, plant mediated alteration of the rhizosphere may favor one species over another. Alteration of native microbial community composition may further decrease competition from native plants and therefore support melaleuca dominance. St udies have shown that through a series of aboveground-belowground feedback loops plan t community structure can drive changes in soil community size and composition (Wardle et al. 2004). For example, invasion of a California grassland by yellow starthistle, Centaurea solstitialis L., can decrease the concentration of the arbuscular mycorrhiz ae fungi phospholipid fatty acid biomarker, 16:1 5c (Batten et al. 2006). A loss of the a ssociated mycorrhizal associations could decrease the competitive ability of the native plant community. Another consequence of changes in microbial biomass is the potential alteration of the rates of turnover of soil nut rient pools. Microbial commun ities are responsible for the cycling of nutrients critical to plant growth (Chapin et al . 2002). However, before a nutrient is released into the ecosystem, internal needs of mi crobial communities must be met. For example, it is estimated that microbial communities will immobilize nitrogen in order to meet internal needs if the C:N molar ratio of the substrate is greater than 25:1 to 30:1 (Chapin et al. 2002). This figure is ba sed on two assumptions: a microbial substrate use efficiency of 40% and a C: N microbial biomass ratio of 10: 1 (Chapin et al. 2002). As suggested, melaleuca may alter microbial speci es composition, which in turn may change rates of ecosystem nutrient cycling. For ex ample, if fungal biomass were to become more prevalent in the invaded soils, substrat e use efficiency could be greater increasing the availability of carbon substrates (Dav et 2004). In addition, microbial communities


35 with higher C:N biomass ratios have lower requirements for nitrogen and will therefore mineralize nitrogen when it is at lower con centrations (Eviner and Chapin 2003). In low fertility soils this could alter the availability of nutri ents vital to ecosystem maintenance. In both the invaded and non-invaded plot s, microbial biomass carbon was 1.6 and 1.4 times higher in the surface soil, respectively, as compared to the subsurface. Several studies have reported a decrease in microb ial biomass with depth (Fierer et al. 2003; Lejon et al. 2005; Shi et al. 2006). In the current study the decrease in microbial biomass is most likely a result of the reduced quantit y and quality of carbon substrate at the lower soils depths. As organic matter is deco mposed more labile carbon compounds are preferentially degraded. The accumula tion of resistant carbon compounds in combination with processes of soil organic matter humification and the seasonal leaching of soluble compounds, can lead to an accumula tion of poorer quality carbon substrates at lower soil depths. Although there was a m easured decrease in quantity of organic material in the subsurface soil (Table 3-1), further testing is needed to determine the relative decomposability of car bon at each soil depth. In the sandy surface soils investigated in this study, microbial biomass accounted for a large proportion of the total nutrient storag e. For example, in the 5-15 cm soil depth microbial biomass phosphorus accounted for 34% of the total phosphorus in the noninvaded and 18% in the invaded sites. Th erefore, any reduction in the size of the microbial pools could increase nutrient mob ility and result in leaching losses. Other studies have reported similar findings. Vit ousek and Matson (1984) found that microbial immobilization of nitrogen was the primary control of nitrogen retention in forest soils after an ecosystem disturbance. Based on th e lower microbial pool size in the invaded


36 site, the top 15 cm of soil in the total area sampled (196 m2) will stabilize approximately 82 g less phosphorus by microbial immobilization as compared to the non-invaded sites. This is over 80 times as much phosphorus as is added to the invaded site by litterfall each year. If these changes lead to increased nutrient leaching, dr ainage waters could become enriched which may have significant cons equence for down-stream communities. Although the size of the micr obial pool was significantly lower in the invaded plots, microbial activity, as measured by so il oxygen demand, was not different between sites at either depth. Soil oxygen demand is a measure of the rate of oxygen consumption during chemical and microbial mediated oxi dation (Reddy et al. 1980) . Malecki (2004) reported that a seasonal decrease in micr obial biomass phosphorus corresponded with a slower rate of oxygen consumption. The curre nt study did not support this finding (Table 6). Microbial biomass was consistently lowe r in the invaded soil, however, there was no difference in the rate of oxygen consump tion between sites at either depth. In addition, it was hypothesized that the rate of soil nitr ogen supply as measured by potentially mineralizable nitrogen would be lo wer in invaded soils. This biological index is a measurement of the soil organic nitrogen that can be decomposed and made available for plant uptake. A 2000 study reported that rates of potentially mi neralizable nitrogen were significantly correlated with micr obial biomass pools (White and Reddy 1999). However, as before with soil oxygen demand, th ere were no significant differences in the levels of soil nitrogen supply between treatments at either de pth. These results point to a potential resilience of microbial activity. While it has been suggested that invasion of melaleuca may alter microbial community stru cture, microbial activity as measured by rates of soil oxygen demand and potentially mineralizable nitrogen appears to be


37 unaltered. However, further tes ting is needed to determine what factors control microbial activity in these site s in order to predict future implic ations of melaleuca invasion. The extensive and devastating effects of exotic plant invasion on aboveground communities have been well documented (DÂ’Antonio and Meyerson 2002, Crookes 2002, Mack and DÂ’Antonio 1998). However, fewer studies have examined how differences in aboveground plant communities may translate into alteration of belowground systems (Ehrenfeld 2003, Wardle et al. 2004, and M ack et al. 2000). With the widespread introduction and invasion of exo tic plants in Florida there is a vital need for studies that investigate alteration of basi c ecosystem structure and function. This study reveals significant changes in both aboveand be lowground communities after invasion of melaleuca in a South Florida pine-cypress eco-tone. Recommendations for Future Study The current study provides a solid foundation for future research investigating the ecosystem-level alterations caused by invasion by melaleuca. Several areas that deserve further study have been highlighted in th is document including: indirect effects of biological control agents; shifts in so il microbial community composition; and allelopathic effects of melaleuca biomass. While there are many ways to tackle each of these questions general recommen dations will be provided below. The two biological control agents mentioned, Oxyops vitiosa Pascoe (Coleoptera: Curculionidae) and Boreioglycaspis melaleucae Moore (Hemiptera: Psyllidae), were released in 1998 and 2002 respectively and ha ve become established throughout South Florida (USDA 2006). This provides a natura l laboratory for testing melaleuca plant responses to biological control agents. Ho wever, in order to quantify these effects controlled greenhouse and small plot studies need to be designed with varying levels of


38 insect density as well as insect excl usion methods. Analyses should include measurements of: melaleuca growth rate a nd reproduction capacity; re source partitioning in live plants; and mass and quality of litter produced. In order to measure shifts in the soil microbial community structure, phospholipid fatty acid profiles can be generated. Phos pholipids are present in the cell walls of microorganisms and are composed of a polar phosphate head connect ed to two non-polar, long-chain carbon, fatty acid tails. The lengt h and bonding structure of the two fatty acid tails can differentiate major microbial taxonom ic groups such as aerobic fungi, protozoa, and gram-positive bacteria (Zelles 1999). Therefore, changes in the structure of fatty acid profiles generated from environmental sa mples may indicate general shifts in the community structure and provide an overvie w of the diversity of the soil microbial community (Kourtev et al. 2003). Investigations of the alle lopathic properties of mela leuca biomass should begin with the identification of melaleuca specific compounds. Column chromatography coupled with mass spectrometry methods used to analyze chemicals in M. alternafolia biomass could be employed to extract and purify M. quinquenervia (melaleuca) compounds (Carson et al. 2006). If a speci fic compound can be identified, greenhouse pot studies and laboratory culture tests can be used to investigate in hibitory effects of melaleuca compounds on plant and microbial gr owth. These studies, as well as, fullscale, multi-habitat investigations of the mechanisms that drive these changes would provide critical information needed to de velop effective management and restoration techniques. .


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46 BIOGRAPHICAL SKETCH Melissa Rosemary Martin was born and ra ised in South Bend, Indiana where she developed a love of nature at an early age. As an undergraduate at the University of Notre Dame, Melissa participated in rese arch on the plant community structure and biogeochemistry of wetlands. After graduati on in 2002, Melissa accepted an internship through the Student Conservati on Association at the USDA-AR S Invasive Plant Research Laboratory in Fort Lauderdale, FL. Through this internship, she was introduced to research on the management and control of inva sive exotic plants. Melissa looks forward to the opportunity to continue her graduate studies at th e University of Florida, investigating ecosystem-level effects of the i nvasion of exotic plants in order to develop more effective management and restoration techniques.