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Phosphorus removal and soil stability within emergent and submerged vegetation communities in treatment wetlands

University of Florida Institutional Repository
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PHOSPHORUS REMOVAL AND SOIL STABILITY WITHIN EMERGENT AND SUBMERGED VEGETATION COMMUNITIES IN TREATMENT WETLANDS By KEVIN GRACE A THESIS PRESENTED TO THE GRAD UATE SCHOOL OF THE UNIVERSITY OF FLORIDA IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF MASTER OF SCIENCE UNIVERSITY OF FLORIDA 2003

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ii ACKNOWLEDGMENTS This research was made possible through moral, technical, and financial support from DB Environmental, Inc. of Rockledge FL, its staff, and founders Thomas DeBusk and Dr. Forrest Dier berg, whose encouragement has been invaluable. Additional support was provided by the Everglades Agricultural Area, Environmental Protection District and the Soil and Water Science Department staff and faculty. I extend thanks to Dr. John White and Dr. Ramesh Reddy for their guidance; and for the improvements they made to this thesis. I also appreciate the contributions of fr iends and fellow students. Thanks go to Patrick Owens and Scott Jackson for their expert assistance in the field. Finally, my father Don, sister Becca and brothe r Geoff are also an important part of anything I do, and we fondly remember my mother Nancy, always.

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iii TABLE OF CONTENTS page ACKNOWLEDGMENTS................................................................................................ii LIST OF TABLES..............................................................................................................v LIST OF FIGURES...........................................................................................................vi ABSTRACT.....................................................................................................................xi i CHAPTER 1 PHOSPHORUS DYNAMICS IN TREATMENT WETLANDS: A REVIEW......1 Introduction................................................................................................................1 Phosphorus in the Water Column...........................................................................5 Phosphorus Removal Processes...............................................................................6 Everglades Research Site Description...................................................................13 Need for Research....................................................................................................14 2 PHOSPHORUS STABILITY IN ACCRETED WETLAND SOILS.....................15 Introduction..............................................................................................................15 Materials and Methods...........................................................................................18 Results and Discussion............................................................................................31 Conclusions...............................................................................................................69 3 BIOMASS PHOSPHORUS STORAGES AND DYNAMICS..............................71 Introduction..............................................................................................................71 Materials and Methods...........................................................................................74 Results and Discussion............................................................................................84 Conclusions.............................................................................................................110 4 SYNTHESIS.............................................................................................................112

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iv Irradiance and Water Column Shading..............................................................114 Oxygen Supply.......................................................................................................117 Community Metabolism.......................................................................................119 Implications for STA Management.....................................................................122 Conclusions.............................................................................................................126 LIST OF REFERENCES...............................................................................................127 BIOGRAPHICAL SKETCH........................................................................................136

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v LIST OF TABLES Table page 2-1 Sequential extraction of phosphorus from wetland soils using a method adapted from Hieltjes and Lijklema (1980).........................................................24 2-2 Water column characteristics at 3 cm and 20 cm depths in duplicate mesocosms (2.2m L x 0.79m W x 0.4m D) dominated by emergent and submerged aquatic vegetation (EAV and SAV, respectively) on May 30, 2001........................................................................................................................... .40 2-3 Water column D.O., pH, and temperature profiles in the STA-1W Cell 1 surface waters at the time of peeper retrieval on June 27, and on September 29, 2002..................................................................................................46 2-4 Flux estimates from intact SAV and EAV soil cores kept under dark conditions for 35 days.............................................................................................60 3-1 Phosphorus removed from the water column over the 2.7-year study period and recovered in the vegetation ( Typha + Najas ) and soils upon termination of the study on August 20, 2001......................................................87 3-2 Dry matter and P concentrations in accrued soil and tissue storages (live and dead Typha shoots, below-ground Typha (roots and rhizomes), and Najas tissues) retrieved from mesocosms on August 20, 2001.........................89 3-3 Mean ( 1 s.d.) water quality parameter values measured in the Cell 4 outflow water used in the preliminary desiccation experiment......................92 3-4 Flux estimates from intact Typha -region soil cores and submerged Typha litter kept under dark conditions for 35 days...................................................103

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vi LIST OF FIGURES Figure page 1-1 The historic Everglades region of south Florida is now three distinct parcels, including the Everglades Agricultural Area (EAA), Water Conservation Areas (WCA) and Everglades National Park. Stormwater treatment areas (STA) are shown (in gray), including STA-1W, where field investigations took place.................................................................................3 1-2 General model of phosphorus storages in wetlands. Arrows indicate potential P exchange pathways between storages...............................................4 2-1 Stormwater Treatment Area 1W in Palm Beach county, Florida. Three flowpaths receive surface water drained from adjacent agricultural soils.....19 2-2 Aerial photograph of STA-1W Cell 1, first cell of the eastern flowpath, in November 2000........................................................................................................21 2-3 Schematic of porewater equilibrator used in estimates of P flux from STA-1W Cell 1 soils into the overlying water column.......................................29 2-4 Average bulk density values for accrued sediments at depth intervals below the sediment water interface.....................................................................32 2-5 Soil phosphorus pools determined through a sequential extraction procedure using 1.0 M NH4Cl, 0.1 M NaOH, and 0.5 M HCl..........................34 2-6 Inflow rates and stage level of water in the outflow region of Cell 1 compared to bottom elevations at the emergent and submerged aquatic vegetation (SAV) stations, prior to and during peeper and Hydrolab deployments............................................................................................................42 2-7 Water column temperature within Cell 1 emergent and SAV communities during the Hydrolab mo nitoring period (July 3 – 17, 2002).....43 2-8 Water column dissolved oxygen saturation levels within Cell 1 emergent and SAV communities during the hydrolab monitoring period (July 3 – 16, 2002)....................................................................................................................43

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vii 2-9 Water column specific conductance levels within Cell 1 emergent and SAV communities during the hydrolab monitoring period (July 3 16, 2002).......................................................................................................................... 44 2-10 Water column pH levels within Cell 1 emergent and SAV communities during the hydrolab monitori ng period (July 3 16, 2002)..............................44 2-11 Incident light (measured as photon flux) remaining at water depths of 0+, 35 and 65 cm in emergent and submerged macrophyte communities, as well as within open water reaches, of STA 1W, Cell 1, in September 2002....47 2-12 Stage level recorded during the seventeen day equilibration period for porewater samplers deployed in the outflow region of STA-1W Cell 1, during June 2002.....................................................................................................48 2-13 Vertical profiles of porewater pH values from soil collected from emergent and submerged vegetation...................................................................49 2-14 Vertical profiles of soluble reacti ve phosphorus concentrations in soil porewater collected from emergent and submerged vegetation....................51 2-15 Vertical profiles of dissolved organic P concentrations in soil porewater collected from emergent and submerged vegetation........................................52 2-16 Vertical profiles of dissolved iron concentrations in sediment porewater collected from emergent and submerged vegetation........................................54 2-17 Vertical profiles of alkalinity concentrations (as CaCO3) in sediment porewater collected from soils below emergent and submerged vegetation.................................................................................................................55 2-18 Vertical profiles of specific cond uctance values in sediment porewater collected from soils below emergent and submerged vegetation....................5 6 2-19 Vertical profiles of dissolved calcium concentrations in sediment porewater collected from soils below emergent and submerged vegetation.................................................................................................................57 2-20 Vertical profiles of calcium carbon ate saturation index (SI) values in porewater collected from soils beneath emergent and submerged vegetation.................................................................................................................58 2-21 Soluble reactive phosphorus (SRP) concentrations in the water columns of SAVand Typharegion soils and control columns (no soil) during 28day dark laboratory incubation............................................................................59

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viii 2-22 Mean dissolved calcium and alkalinity concentrations in the water columns of three treatments (see text for details) during 28-day dark laboratory incubation.............................................................................................61 2-23 Mean dissolved organic phosphorus concentrations in the water columns above SAV and Typha -region soils and in control (water only) columns during 28-day dark laboratory incubation..........................................................62 2-24 Dissolved organic carbon concentrations in water columns over Typha soils, SAV soils and control columns (no soil) after 4, 7, 10 and 14 days of dark laboratory incubation....................................................................................63 2-25 Water column SRP concentrations above a “background” concentration as determined before a 100 g spike was added to each column....................64 2-26 Phosphorus diffusive flux estimates based on 22 porewater equilibrators deployed in STA-1W Cell 1 (June-July 2002; closed markers) and Cell 4 (November-December 2001; open markers), as a function of distance through the entire wetland....................................................................................68 3-1 The historic Everglades region of south Florida is now three distinct parcels, including the Everglades Agricultural Area (EAA), Water Conservation Areas (WCA) and Everglades National Park. Stormwater treatment areas (STA) are shown (in gray), including STA-1W......................72 3-2 Stormwater Treatment Area 1 W in Palm Beach county, Florida. Three flowpaths receive surface water drained from adjacent agricultural soils and reduce the phosphorus load to Water Conservation Area 2A.................76 3-3 Aerial photograph of STA-1W Cell 1, first cell of the eastern flowpath, in November 2000 (courtesy SFWMD).....................................................................77 3-4 Schematic of incubation design, containing Typha litter and P-amended surface waters from WCA 3A...............................................................................82 3-5 Inflow and outflow TP concentrations for mesocosms dominated by Typha and operated from December 1998 through August 2001.....................85 3-6 Partitioning of recovered Najas and Typha tissue P within the inflow (In) and outflow (Out) halves of two mesocosms operated from December 1998 through August 2001.....................................................................................88 3-7 Dissolved organic phosphorus (DOP) and soluble reactive phosphorus (SRP) concentrations in floodwaters containing fresh Najas tissues and Typha litter................................................................................................................93

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ix 3-8 Soluble reactive phosphorus concentrations in the water columns of Typharegion intact soil cores with and without Typha litter amendments during 20-day dark laboratory incubation..........................................................95 3-9 Mean dissolved organic phosphorus concentrations in the water columns of four treatments (see text for details) during 28-day dark laboratory incubation.................................................................................................................97 3-10 Mean dissolved calcium and alkalinity concentrations in the water columns of four treatments (see text for details) during a 28-day dark laboratory incubation.............................................................................................98 3-11 Dissolved organic carbon concentrations in water columns of four treatments (see text for details) after 4, 7, 10 and 14 days of a dark laboratory incubation...........................................................................................100 3-12 Water column SRP concentrations above a “background” concentration as determined before a 100 g spike was added to each column..................102 3-13 Mean soluble reactive phosphorus (SRP) concentrations in triplicate 0.5 L flasks, containing Typha litter incubated in the dark under oxic or anoxic conditions for 5 days.............................................................................................104 3-14 Mean soluble reactive phosphorus (SRP) concentrations in triplicate 0.5 L flasks, containing Typha litter incubated in the dark under oxic or anoxic conditions...............................................................................................................106 3-15 Change in microbial biomass phosphorus (MBP) of Typha litter incubated for seven days........................................................................................................107 4-1 Conceptual diagrams of phosphorus calcium and oxygen levels in the water column in relation to vegetation type.....................................................124

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x 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 PHOSPHORUS REMOVAL AND SOIL STABILITY WITHIN EMERGENT AND SUBMERGED VEGETATION COMMUNITIES IN TREATMENT WETLANDS By Kevin Grace August 2003 Chair: John R. White Major Department: Soil and Water Science Phosphorus (P) removal by treatment wetlands is an integral part of Everglades protection and restoration. Th e effects of water column shading on P cycling and retention were explored in emergent ( Typha spp.) and submerged ( Najas guadalupensis ) vegetation communities within a Stormwater Treatment Area (STA-1W) wetland in south Florida. The physicochemical aquatic environments within these two vegetati on communities were hypothesized to differentially affect community metabolism, which in turn would affect biological P uptake rates and P stability in accrued soils. Muck soils beneath emergent aquatic vegetation (EAV) were P-depleted over 8 years of operation in STA-1W, while the muck soils beneath submerged aquatic vegetation (SAV) beds were P-en riched. The stability of P within newly

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xi accrued soils was dependent on macrophyte community type, and was likely increased through water column CaCO3 precipitation. Soils (0-4 cm layer) in SAV communities contained significantly more P in residual, Ca/Mg-bound and fulvic/humic acid-bound pools than soils in EAV. Because of similar pools of exchangeable and Al/Fe-bound P, the two soil types each released P to an oxygenated water column at similar flux rates. Typha litter and associated microbial biomass retained P mineralized from soils under oxic water column conditions, but retention was lower under anoxic conditions. Dense EAV stands accumu late oxygen demand, reduce light penetration and may have little microbial P uptake and retention capacity due to anoxic conditions. While Typha biomass persists as leaf litter and detritus to a greater extent than Najas tissues, the extensive Typha root system has the potential to hinder long-term storage by mobilizing P from enriched soils. Community metabolism was influenced by water column shading, which reduced CaCO3 precipitation and Ca-bound P pools; and reduced oxygen supply to microorganisms in EAV communities. Managing for SAV and eliminating dense EAV stands from treatment wetlands may reduce surface water TP concentrations. Phosphorus-enriched areas within the northern Everglades may also be contained or restored by increasing light penetration to the water column, in order to enhance soil P sorption capacity through CaCO3 precipitation and increase photosynthetic oxygen supply to the aquatic microbial communities.

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1 CHAPTER 1 PHOSPHORUS DYNAMICS IN TREATMENT WETLANDS – A REVIEW Introduction Phosphorus (P) availability limits primary productivity in many freshwater systems (Boers et al., 1998; Reddy et al., 1999). Cyanobacterial mats are periphyton communities that can maintain high productivity in low P environments due to rapid internal nutrient recycling (Wetzel, 1993). Cyanobacteria ( Schizothrix Scytonema ) within the mats are competitive with emergent macrophytes (i.e., cattails ( Typha spp.)) in waters with low P concentrations, however they are sensitive to increases in P levels. In contrast, Typha grows well in P-enriched soils, and can produce a large biomass annually (Toth, 1988; Davis, 1991; Davis, 1994). A shift from open water areas dominated by calcareous, cyanobacteria periphyton to monoculture stands of Typha in areas downstream of water control structures, has raised concerns among resource managers about the loss of ecological integrity of the oligotrophic Everglades marsh (Davis 1994). Constructed wetlands have become a practical treatment technology for removing P from surface waters. In south Florida, farmland in the Everglades Agricultural Area (EAA) has been taken out of cultivation and flooded to create wetlands for surface water treatment (Figure 1-1). The created wetlands, or

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2 stormwater treatment areas (STAs), were subsequently colonized by a mixture of emergent, submerged and floating wetland vegetation. Submerged aquatic vegetation (SAV) including Najas guadalupensis Spreng., and emergent cattail ( Typha latifolia L., and T. domingensis Pers.), communities coexist throughout the STAs. Floating macrophytes ( Eichhornia crassipes Pistia stratiotes ) and periphyton appear transient in the upstream and downstream reaches of the wetland flow path, respectively. Current water management objectives in south Florida for treating agricultural drainage waters (ADW) with treatment wetlands include achieving low effluent total phosphorus (TP) concentrations, maximizing phosphorus removal from the water column, and producing stable soils for long-term P storage. Essential to achieving these object ives is an understanding of P exchange rates between three major storages: soil, biomass and the water column (Figure 1-2). The water column is a compartment of variable P storage that must be minimized for effective P removal. The biomass of macrophytes and microorganisms influences several biogeoch emical cycles that control P exchange between compartments. Soils are viewed as the long-term storage compartment, though surface soils and associated P interact with the biomass and water column compartments.

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3 Figure 1-1. The historic Everglades region of south Florida is now three distinct parcels, including the Everglades Agricultural Area (EAA), Water Conservation Areas (WCA) and Evergl ades National Park. Stormwater treatment areas (STA) are shown (i n gray), including STA-1W, where field investigations took place. N E W S Everglades National Park WCA-1 WCA-3 WC A -2A FLORIDA, USA EAA Lake Okeechobee STA-1W

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4 Figure 1-2. General model of phosphorus storages in wetlands. Arrows indicate potential P exchange pathways between storages. Deep soils represent a long-term P storage compartment. Water column P removal and soil stability (P storage potential) can be understood through examination of biological processes (e.g., plant P uptake, tissue senescence, and microbial decomposition), chemical processes (e.g., adsorption/desorption from surfaces and precipitation/dissolution of P compounds), and physical processes (e.g ., particulate settling and resuspension, soil burial). Each of these processes with in the treatment system can influence the P retention capacity of the system. Biomass Water Column Dee p Soil Surface Soil

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5 Phosphorus in the Water Column Phosphorus Species Phosphorus exists in a variety of chemical forms within natural waters. Compounds can range from dissolved ion species (e.g. H2PO4 -, HPO4 2-) to large complex molecules. Because of the range in molecular size, it is important to describe two classes of phosphorus compounds, dissolved and particulate species. An operational definition of dissolved P is that which passes through a 0.45 m membrane filter, whereas the particulate P is retained. Particulates > 0.45 m in diameter are generally too large for direct uptake by plants, algae and microorganisms. However, particles can be transported by surface waters, and subsequently mineralized downstream. Phosphorus compounds also exist as organic or inorganic forms, though there can be considerable exchange between the two. Organic P is either derived from cell components or is soluble P that has been complexed by organic matter, and can be resistant to mineralization or dissociation. Inorganic P can be either dissolved phosphate ions or sorbed on to (or incorporated within) a mineral particle. Soluble reactive phosphorus (SRP) is the dissolved inorganic fraction that is readily bioavailable. Dissolved organic phosphorus (DOP) is less reactive and must be hydrolyzed prior to biological assimilation. The remaining fraction is particulate phosphorus (PP), which is retained by a 0.45 m membrane.

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6 Phosphorus Sources Phosphorus in any form can be imported into a wetland through surface inflow waters, precipitation, or groundwater seepage (allochthonous). It can also be transformed within the wetland from one form to another (autochthonous). Internal loading represents P released to the water column from biomass and soil storages. The allochthonous load can be sequestered during the growing season by plant uptake and then released as an internal load during periods of plant senescence. Substantial internal loads can also be attributed to soil P release. Both allochthonous (external) P loads to th e Everglades system and autochthonous (internal) loads from nutrient-impacted areas must be reduced in order to achieve management and restoration objectives. Phosphorus Removal Processes Biological Processes Assimilation Biological P uptake is almost exclusively in the soluble reactive form (SRP). Carignan (1982) correlated 97% of P uptake by macrophytes with changes in soil porewater and water column SRP concentrations. Bacteria and other microorganisms assimilate P and rapidly recycle it to the water column upon death (Currie and Kalff, 1984). Rates of SRP assimilation by macrophytes are slower and highly variable across species. Richardson and Marshall (1986) found that PO4 additions to emergent macrophyte enclosures were initially assimilated into the microbial biomass and detritus, rather than into the emergent biomass.

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7 Emergent macrophytes and some submerged plants obtain P from the soil porewater, while other submerged macrophytes obtain the nutrient from water directly through stem and leaf tissues. Phytoplankton assimilate P directly from the surrounding water column, and are dependent on mixing to circulate sufficient nutrients for growth. Epiphytes and benthic populations of algae, bacteria and fungi depend on water flow to carry necessary nutrients across the substrate surface, and reduce boundary la yers that can otherwise limit nutrient supply to slow rates of molecular diffusion. Phosphorus uptake rates depend on the microbial, algal or higher plant biomass. Macrophyte biomass P storag e increases throughout the growing season as biomass increases, though tissue P concentrations may be variable (Hill, 1979). Cattails can allocate 60% of the total biomass to below-ground tissue in low P environments, but only 40% in P-enriched environments (Miao and Sklar, 1998; Miao and DeBusk, 1999). Biomass P is then subject to recycle into the water column, or incorporation into the soil detritus/microorganism compartment where it is slowly buried into the deep sediments (Richardson and Marshall, 1986). Nutrient availability in the water column dictates the importance of soil nutrients. Rattray et al. (1991) suggested that in eutrophic, aquatic environments macrophyte growth was not necessarily limited by soil nutrient deficiencies. Using radioisotopes, however, Carignan and Kalff (1980) have shown that SAV rooted in eutrophic waters (167 g SRP L-1) would still obtain nearly all tissue P

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8 from the soils, and they implicate macrophytes as potential nutrient “pumps” that return P from the soils to the water column. In oligotrophic water, soil nutrition (especially P) can limit growth of rooted, submerged macrophytes (Rattray et al., 1991). Organic matter accumulation on mineral soils was shown to influence SAV species succession in lakes (Barko and Smart, 1983) with maximum growth rates in soils containing 5-20% organic matter. If either SAV species or growth rate affects P removal by SAV beds, organic matter accumulations would influence P removal capacity of the community. Despite the potential importance of soil P chemistry in rooted macrophyte nutrition, other aquatic plants such as phytoplankton, periphyton, floating plants, unrooted macrophyte species (e.g. Ceratophyllum demersum ), and SAV fragments depend entirely on nutrient uptake from the water column. Furthermore, water column P uptake beyond the requirements for maximum growth, referred to as “luxury uptake”, has been reported for a Ceratophyllum /periphyton complex growing in South Florida ADW (Pietro 1998). Currie and Kalff (1984) suggested similar maximum P uptake kinetics exist for bacteria and phytoplankton. However, they also observed a five-fold greater increase in bacteria intracellular P concentrations over phytoplankton during “P-limited” lab incubations. Algae are therefore capable of greater P uptake per mass in the water column, because its biomass is usually greater than that of bacteria (Currie and Kalff, 1984). However, under the same P-limited

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9 conditions, bacteria are able to incorporate P more readily into cellular tissue, which may affect the rate at which P becomes available again. Chemical Processes Adsorption Phosphorus adsorption onto mineral surfaces is often related to the presence of calcium carbonates and Fe an d Al hydroxides in soils (Patrick and Khalid, 1974; Richardson, 1985; Porter and Sanchez, 1992). Patrick and Khalid (1974) demonstrated the influence of iron chemistry on P sorption and release from flooded soils. In the oxidized water column, dissolved iron is low because Fe3+ dominates and is precipitated as ferric oxyhydroxides. Under reducing conditions after oxygen is consumed by chemical and biological oxidation, iron is solvent as the Fe2+ ion. In acid soils, the phosphorus adsorption potential can be estimated accurately from the extractable alumin um and iron contents in the soil (Richardson, 1985). In soils of higher pH, carbonates control P sorption capacity and availability. Due to relatively low Fe and Al contents of EAA histosols, P sorption isotherms were unrelated to Al levels and weakly correlated to Fe, but were significantly correlated with total Ca and free carbonates (Porter and Sanchez, 1992). Richardson and Vaithiyanathan (1995) examined P sorption along a gradient of soil-P enrichment in the northern Everglades. They found the P adsorption coefficient (a measure of P so rption capacity) was lower in Everglades

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10 histosols than for mineral wetland soils where iron regulates P sorption characteristics. In Everglades soils where anoxic, reducing conditions are common, P sorption is controlled by CaCO3 (Koch and Reddy, 1992) rather than iron hydroxides (Reddy et al., 1993; Richardson and Vaithiyanathan, 1995). Precipitation In addition to biological uptake and chemical adsorption processes, CaCO3 precipitation is an important P removal mechanism in SAV systems, either through CaCO3 sorption of P or direct coprecipitation (Gumbricht, 1993). CaCO3 compounds are not subject to dissolution and subsequent P release when reducing conditions develop, but are sensitive to change in CO3 2equilibria, pH and temperature. CaCO3(s) Ca2+ + CO3 2Ksp = 10-8.4 (at 25C) [1] In Fe and Al-dominated systems, Fe and Al-oxide precipitates play a similar role by incorporating phosphate from the water column into insoluble precipitates (Richardson, 1985). Insolubl e calcium phosphate minerals such as apatite can only form under high concentrations of P, well above the levels observed in most surface waters (Golterman, 1998). Porewater P concentrations are frequently higher than those in th e water column, however, and precipitation of metastable Ca-P minerals (e.g., trical cium phosphate) may occur within the soil environment.

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11 Physical Processes Settling Newly accrued wetland soil is a particulate matrix comprised of microbial biomass, macrophyte detritus, and inorganic solids settled from the water column. Settling of suspended particles o ccurs faster in emergent and submerged macrophyte communities than in open water (Madsen et al., 2001), due to reduced flow velocities and mixing. We tland substrata including standing shoots and litter in Typha stands, SAV biomass, and the soil surface collect flocculent material as it settles onto surfaces which expands the “active” surface into three dimensions. Recently settled particles are in dynamic exchange with the water column via diffusion gradients, decomposition and leaching, bioturbation and resuspension. They can also be buried deeper into the soil profile. Surface soil P is still potentially bioavailable to rooted macrophytes, benthic algae, and soil microbial populations. Surface soil is subject to resuspension by turbulent, high-velocity flow, as well as through bioturbation. Soil resuspen sion increases turbidity and decreases light penetration through the water column (Bloesch, 1995), which affects community metabolism and P uptake. Since P concentrations are generally lower in the water column than in the soil porewater, resuspension of sediment particles increases desorption from so lid phases into the bulk solution. Unvegetated reaches of a wetland flow path may dramatically alter physical soil stability. In addition to faster flow velocities and shorter hydraulic

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12 retention times, unvegetated areas lack roots to stabilize the soil. High flow conditions may scour the bottom, suspendi ng these soils in the water column for transport downstream. This process maintains deep-water conditions relative to nearby vegetated areas. Emergent and submerged macrophyte communities stabilize the water column and reduce soil resuspension (Gumbricht, 1993; Madsen et al. 2001), compared to open water sites. Burial Whether initial removal processes are biological, chemical, or physical, long-term storage of P in wetlands requires burial into deep sediments. Burial by organic matter accumulation occurs as the net result of a productive community metabolism – or whenever primary production exceeds respiration. Excessive external P loading to the Everglades marsh has resulted in long-term accumulation of P in the sediments and biomass storages (Reddy et al. 1993, Reddy et al. 1998). Using 137Cs dating techniques, Reddy et al. (1993) estimated P accumulation rates in Everglades soils from 0.11 to 1.14 g m-2 yr-1. They also identified higher nutrient retention by Typha than Cladium (sawgrass) communities. Long-term burial rates are also influenced by temperature, hydrology, and fire regime (Reddy et al. 1993), as each of these factors affects the rate at which accumulating organic matter is oxidized to CO2 and water. Decomposition of organic matter increases with temperature due to increased microbial metabolism. The warm temperatures in south Florida allow rapid

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13 decomposition, yet flooded conditions reduce the supply of oxygen to decomposers. Alternate electron acceptors (other than O2) are used in oxidation reactions under flooded conditions, and may control the overall rates of organic matter decomposition/ accumulation and associated P burial in wetlands (White and Reddy 2001). Periodic fires, often the result of lightning strikes, can rapidly oxidize organic matter and affect long-term rates of organic matter accumulation. Everglades Research Site Description Located between the EAA and Water Conservation Area (WCA) – 1, STA1W was the first of six STAs to begin flow-through operations in 1994. Wetland environments in STA-1W that were dominated by SAV and cattail communities were examined in this study with respect to phosphorus removal and retention into new soils. Currently managed as the Arthur R. Marshall Loxahatchee National Wildlife Refuge, WCA-1 represents the northern extent of the largely unaltered Everglades land. To the south, WCA-2 has been the focal point of much research on the impacts of P on wetland processes. A cl ear and well-documented transition from Cladium jamaicense prairies to Typha spp.-dominated wetlands exists south and west from canal discharge points. Data from this study are compared to the many relevant studies that were conducted within WCA-2A along the northern eutrophication gradient. Directly south of WCA-2, and between urban Miami and the Everglades National Park, is WCA-3. This region is a mosaic of pristine Everglades ecotypes,

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14 from wet sloughs dominated by Nymphaea spp. and Utricularia spp. to sawgrass ( Cladium ) ridge and tree island communities. Isolated stands of Typha exist across the landscape, and are associated with alligator holes. Need for Research Emergent and submerged vegetation communities differ in physical and chemical structure, yet the effects of these differences on P removal and soil stability characteristics within the treatment wetland context are largely unexplored. Until recently, the two vegetation types were each considered a part of the naturally-recruited wetland community, and little effort was made to separate the effects of one independent of the other. For example, the biomass storage potential of cattails has been viewed as an asset to wetland treatment, and many Typha wetlands function well as wastewater polishing areas (Kadlec and Knight 1996). However the contribution towards P removal and retention of the submerged macrophytes, epiphyton and microbial community, is unknown. In eutrophic, aquatic environments co ntinuously loaded with P, cattails are capable of creating large monocultur e stands (Davis 1991). Since, they are ubiquitous, aggressive colonizers, the differences between SAV and cattail communities in providing different P remo val and effluent TP concentrations are important. My research, therefore, compared emergent and submerged macrophyte communities with respect to P interactions within the water column and the development of new soils.

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15 CHAPTER 2 PHOSPHORUS STABILITY IN ACCRETED WETLAND SOILS Introduction Long-term retention of phosphorus (P ) in wetlands can occur through the accretion of new soil (Reddy et al., 1993). Treatment wetlands have been constructed for P removal from surface wa ters, and accrete soils to store P and protect downstream waters from eutr ophication (Kadlec and Knight, 1996). Stability of newly accrued soil-P is an important issue in such wetlands, as internal P cycling can elevate wate r column P above target outflow concentrations. Newly accreted surface so ils are in contact with both the water column and macrophyte roots, thus soil-P can be released into the water column. Soil P retention depends on the characterist ics of the wetland vegetation as well as the chemical environment within th e water column and surface soil layer. Aquatic photosynthesis elevates water column pH levels during daylight hours due to consumption of dissolved inorganic carbon species (i.e.CO2, HCO3 -) that equilibrate with solid carbonates. When water column calcium and alkalinity levels are sufficiently high, pH elevations lead to calcium carbonate supersaturation and precipitation (Otsuki and Wetzel, 1972). The precipitate, in turn, provides sorption sites for dissolved inorganic P, and results in a calcium-

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16 and phosphorus-enriched soil. At 25 C, the following equilibrium reactions govern CaCO3 solubility in water: CaCO3(s) + CO2(g) + H2O Ca2+ + 2HCO3 K = 10-6.03 [2] 2HCO3 2H+ + 2CO3 2Ksp = 10-10.33 [3] CaCO3(s) + CO2(g) + H2O Ca2+ + 2H+ + 2CO3 2K = 10-26.63 [4] Calcium carbonate chemistry may play an important role in wetland treatment of drainage water from the Everglades Agricultural Area (EAA), a region productive in sugarcane and wint er vegetables. The EAA has muck soils underlain with calcareous limestone (Gle ason and Stone, 1994). Shallow surface waters coupled with a high water table increases surface water and groundwater interactions, which elevate Ca and carb onate (alkalinity) concentrations in irrigation waters, agricultural drainage water (ADW), and in the soil itself. Excessive phosphorus (P) loading from EAA ADW discharges has been identified as the primary cause for an ob served eutrophication gradient in the northern Everglades. Changes observed near the discharge structures include increased water column, soil, and plant tissue P concentrations, and change in ecosystem function, relative to the interior marsh (Craft and Richardson, 1993; DeBusk et al., 1994; Reddy et al., 1993; Reddy et al., 1998). Phosphorus enrichment has led to increased Typha spp. above-ground biomass and shoot density (Grimshaw et al., 1997; Wu et al., 1997; Miao et al., 2000), and reduced PAR levels at the air-water interface, re lative to levels at nearby open slough

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17 sites. Reduced PAR may limit aquatic photosynthesis, which could reduce CaCO3 precipitation and soil P retention. Rates of P diffusion from soils to overlying water can be controlled by the strength of the ion activity gradient, temperature and soil porosity. Unequal distribution of phosphate, due to biologic al uptake mechanisms or sorption onto solid minerals such as CaCO3 and FeOOH (Golterman, 1998), can establish an ion activity (related to ion concentration) gradient. Moore and others (1991) found a diffusive P flux rate of 1.69 mg m-2 day-1 for sediments in a hypereutrophic freshwater lake (Lake Apopka, FL). The magnitude of internal nutrient flux can equal external loads to such a system, and maintain water column P above 30 g L-1 concentrations typical of eutrophic systems (Nurnberg, 1996). Internal loading can potentially impair treatment wetland performance. As P-enriched soils accumulate, the pote ntial for diffusive flux to mobilize soil-P into the water column may increase. Additionally, emergent macrophyte shade may limit water column photosynthesis and CaCO3 precipitation, relative to submerged macrophyte areas, thereby prod ucing soils of different P retention capacity and internal P flux rates. The influence of macrophyte vegetation shading of the water column on CaCO3 precipitation and accrued soil stability is currently unknown. The objectives of this study were to: Characterize the stability of P within accreted soils of emergent and submerged macrophyte stands through sequential P extraction,

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18 Determine the potential for CaCO3 formation in emergent and submerged macrophyte communities, both in the water column and the soil porewater, Estimate the potential internal P flux to the water column from newly accrued soils formed within emergent and submerged macrophyte stands. These research objectives will address the stability of P in soils formed in emergent and submerged macrophyte-dominated wetlands, which is essential to the long-term management of STAs as well as the nutrient-impacted northern Everglades. Materials and Methods This investigation was pursued through bench–scale incubations, in fieldoperated mesocosms, and in situ within the full-scale wetland environment. Bench-scale studies took place at DB Environmental, Inc., in Rockledge FL; and at the Wetland Biogeochemistry Laboratory at the University of Florida in Gainesville, FL. Outdoor mesocosms were located next to the STA-1W inflow canal on an experimental platform provided by DB Environmental. STA-1W Site Description Agricultural runoff from the Everglades Agricultural Area (EAA) Basin S5 (Figure 2-1) is pumped via canals to STA-1W, a full-scale (2699 ha) treatment wetland operated by the South Florida Water Management District (SFWMD) to reduce P loadings to the Everglades (SFWMD, 2003). Everglades muck soils were drained for agricultural production decades ago. Nearly 50, 000 acres of former ag land have recently been reflooded to create STAs, of which the

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19 Everglades Nutrient Removal project, now part of STA-1W, was the prototype. A detailed spatial analysis of phosphorus an d other constituents of the pre-existing farm soils in the STA 1W footprint was conducted prior to flooding (Reddy and Graetz, 1991). 4 2 5 B 5 A 1 3 Loxahatchee NWR Water Conservation Area 1 S TA 1W (2699 ha) NorthernFlowpath EasternFlowpath WesternFlowpath Figure 2-1. Stormwater Treatment Area 1W in Palm Beach county, Florida. Three flowpaths receive surface water drained from adjacent agricultural soils. The wetland functions to reduce water phosphorus concentrations prior to discharge into Water Conservation Area 2A. Cells 1, 2 and 3 are comprised of mixed emergent, submerged and floating vegetation, while Cells 4, 5A, and 5B are primarily submerged and floating vegetation. Arrows indicate general direction of flow. Cell 1 is the first cell of the eastern flowpath, and is comprised of emergent aquatic vegetation, or EAV, ( Typha spp.) in the inflow region and along the

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20 eastern edge (Figure 2-2). A mixture of emergent, floating, and submerged aquatic vegetation (SAV) occupy the do wnstream reaches of the cell. This distribution of community types has remained relatively constant in Cell 1 since the wetland was first flooded in 1994 (Newman and Pietro, 2001). The juxtaposition of areas dominated by subm erged and emergent vegetation in the Cell 1 outflow region suggests that the two community types have developed under similar hydraulic and nutrient lo adings. Eight years of flow-through operations in this wetland have been sufficient to allow accretion of soils in both community types, and to establish a record of long-term effectiveness of P sequestration from the water column. Two sampling sites were selected in the outflow region of Cell 1. One station (26.6292N, 80.4219 W) represented EAV, Typha domingensis while SAV species Najas guadalupensis and Ceratophyllum demersum occupied the second station (26.6278N, 80.4352 W). The longevity of the contrasting vegetation communities observed at the two locations was verified with aerial photos provided by SFWMD, as well as through personal communication with field personnel. Stability of recently accreted wetland soil-P was characterized through bulk density analysis and sequential ex tractions. Vertical profiles of water column chemistry were constructed for mesocosms dominated by emergent and submerged macrophytes and for emergent and submerged macrophyte communities in southeastern Cell 1 of STA-1W. Profiles of soil porewater

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21 constituents were also constructed for the STA soils. For each profile, the CaCO3 saturation index was calculated to determine whether conditions for precipitation were present. Figure 2-2. Aerial photograph of STA-1W Ce ll 1, first cell of the eastern flowpath, in November 2000 (courtesy SFWMD). Also shown is the location of the field station in emergent (pink) and submerged (grey) vegetation stands. Phosphorus diffusion flux rates acro ss the sediment-water interface were calculated based on porewater concentration gradients. Potential P flux from these soils was then determined experimentally using intact cores incubated under lab conditions. Cell 1Outflow InflowWCA-1 Loxahatchee NWR Field Station

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22 Soil Collection and Analysis Soil cores were retrieved from both field sites on July 3 and 4, 2002. Each core was retrieved by pushing acrylic core tubes (7 cm dia.) through the accrued soil layer into the underlying native farm muck to a minimum depth of 10 cm. The top end of the core was sealed with a #13 rubber stopper prior to soil core extraction. The horizon dividing new wetland soils from the native muck soil was determined by differences in color and texture, and the depth of accrued wetland soil was recorded. Cores were sectioned in the field at 2 cm depth intervals. The 0-2 cm, 2-4 cm, and 4-6 cm intervals were retained fo r analysis, along with the underlying muck. Like depth increments from five replicate cores were composited and homogenized before analysis to account for field variability. This procedure was repeated three times at both sites, and re sulted in triplicate samples for each of the four soil layers. A fresh 90 cc subsample of composite soil samples was dried to constant weight (65C) for bulk density. Flocculent surface (0-2 cm) samples were allowed to settle in graduated cylinders overnight (dark, 4 C) and the excess water was discarded prior to analyses. Sequential Extractions for Inorganic Phosphorus Pools Inorganic P pools in the soil surface intervals of 0-2 cm and 2-4 cm were characterized by sequential extraction (modified from Hieltjes and Lijklema, 1980) and contrasted to the underlying muck soils. Fresh soil samples were weighed (5g wet) into 50 mL centrifuge tube s. To each centrifuge tube, 40 mL of

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23 1M NH4Cl was added. The slurry was then shaken for two hours. Samples were then centrifuged for 15 minutes at 2800 rp m before the supernatant was decanted and filtered (0.45 m) for SRP analysis. The NH4Cl extraction was repeated, and the supernatant was added to that of the fi rst extraction before analysis. The tube and sample residue were then weighed, and 40 mL of the next extractant was added (Table 2-1). Sodium hydroxide (NaOH) and HCl extractions were shaken for 17 and 24 hr, respectively. Both SRP and TP analysis were performed on NaOH extractions, while only SRP analysis was performed on the HCl extracts. Soil residue remaining after HCl extraction was subjected to TP analysis. TP samples were filtered through Whatman 41 qualitative filters. Extractants were refrigerated at 4C until analysis. Soluble reactive phosphorus colorimetric analysis (potassium antimony tartrate, sulfuric acid, ammonium molybdate, ascorbic acid) was performed on a Spectronics Genesys 5 spectrophotometer. Total P analysis included a persulfate digestion and neutralization prior to SRP analysis. Using the residue weights recorded afte r each extraction, an estimation of P carry-over (the extractant volume and P mass carried over between extractions) was calculated and used to adjust each P pool. Flux Study Using Intact Soil Cores Three additional replicate intact soil cores were retrieved from the SAV and EAV communities for intact core incubations. Each soil core was topped off

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24 with site water, then sealed with rubbe r stoppers. An opaque shroud minimized solar heating and blocked light during transport to the lab. Table 2-1. Sequential extraction of phosphorus from wetland soils using a method adapted from Hieltjes and Lijklema (1980). Sediment P pool Extractant Shake Time Analysis Stability Exchangeable 1.0 M NH4Cl at pH 7 Two 2 hr periods SRP Readily bioavailable Feand Al-bound SRP Humic and Fulvic Acid-bound 0.1 N NaOH 17 hr TP-SRP Ca-bound 0.5 N HCl 24 hr SRP Residue Remainder digested for TP Recalcitrant Overlying water was replaced with 1.15 L filtered (0.45 m) STA-1W Cell 4 outflow water. The 30 cm water column was aerated and incubated in a dark water bath at ~22C for 28 days. Water samples (30 mL) were withdrawn at t = 0, 0.5, 1, 1.5, 2, 4, 6, 10, 14, 20, and 28 days, and analyzed for SRP and pH. Representative samples were also analyzed for TSP, dissolved calcium, total alkalinity, and DOC. Reflood water was added (30 mL) after each sampling to maintain water volume. Samples for SRP and TSP were filtered through a 0.45 m polyether sulfone filter immediately after sample collection. Persulfate digestion followed by neutralization was performed on TSP samples. SRP and TSP water determinations followed the ascorbic acid-molybdenum blue method (EPA 365.2;

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25 EPA 1979) using a Spectronics Genesys 5 spectrophotometer. Dissolved calcium was determined using flame atomic absoption spectroscopy (EPA 215.1; EPA 1979) on a Perkin-Elmer 3110. Alkalinity was titrated with 0.02N H2SO4 (EPA 310.1; EPA 1979). Dissolved organic carbon analysis was on acidified, filtered (0.45 m) samples, and measured with a Shimadzu TOC-5050A (Duisburg, Germany) TOC analyzer equipped with an ASI-5000A autosampler (5310-A; APHA 1992). Sample pH was recorded immediately following collection, using a 3 in 1 gel filled combination pH electrode and Corning 313 pH meter. Water bath temperature was continuously recorded by a StowAway Tidbit logging probe (Onset Computer; Bourne, MA) as well as monitored periodically with a thermometer. At the conclusion of the P Flux study, the water column of each core was spiked with 100 g P L-1 (as KH2PO4). The water volumes above each core differed slightly ( 5 mL) from the original water volume of 1.15 L added one month prior, likely due to different evaporation rates induced by the aerators. These differences were recorded but volumes were not adjusted at that time. Water samples were withdrawn t = 0, 4, 8, 24 and 53 hours after the amendment, and analyzed for SRP. Each core received 30 mL of unamended reflood water after sampling to maintain water volume, and was kept under an opaque shroud.

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26 Mesocosm Design and Background One SAV-dominated and two EAV-dominated mesocosms (4.2m L x 0.79m W x 1.0m D) were maintained for 2.7 years at a hydraulic loading rate of 10 cm day-1, as part of another study (DBEL, 2001). These systems were inoculated with Najas guadalupensis and Typha spp. plants, respectively, which were collected from within STA-1W Cell 4. Weekly monitoring of inflow and outflow TP concentrations, temperature and pH, and periodic monitoring of other constituents (TSP, SRP, dissolved Ca (dCa), total alkalinity (TA), specific conductance) was performed from December 29, 1998 through August 8, 2001. Shade Effects within EAV and SAV Mesocosm Communities In order to investigate the effects of shade on P removal in the EAV mesocosms, I examined the ambient light regime and calcium carbonate saturation index (SI). Duplicate SAV mesocosms with similar operational history were sampled for comparison. Incident light was measured at 2m above the water surface (above cattail canopy), and at the water surface. Two 4 spherical quantum sensors recorded available PAR during peak daylight hours (1000-1400), using a Li-COR LI-1000 data logger (Lincoln, Nebraska). Eight replicate one-second measurements of photon flux were averaged for each datu m value, and recorded when the value had stabilized to roughly within 1%. All comparisons were made to the simultaneous “ambient” light levels, to ad just for short-term temporal changes in incident radiation (e.g. change in cloud cover).

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27 Chemical profiles through mesocosm water columns were constructed from in situ measurements of dissolved oxygen (DO) concentrations, pH, specific conductance and temperature taken on May 30, 2001. Concurrently, surface (3 cm) and bottom (20 cm) water samples were collected and analyzed for SRP, dCa, TA concentrations. Water samples were withdrawn from the inflow and outflow regions of each tank (approximately 10 cm from end wall), as well as from each inflow stream. The SI of CaCO3 was calculated from water column and porewater chemistry profiles of submerged and emergent macrophyte communities. Specific conductance values provided an approximation for ionic strength. Hydrogen ion concentration was calculated from pH values. Calcium and alkalinity concentrations were used to calculate activity products of the Ca2+ CO3 2and HCO3 ions. Temperature values were used to adjust all solubility constants. The CaCO3 SI was then computed for water at the surface and at 20 cm depth according the following relationship: 0 3 23 2S HCO CaK H K HCO Ca SI [5] where: indicates the activity coefficient of Ca2+ and HCO3 [ ] indicates the concentration of Ca2+ HCO3 and H+ K is the acidity constant of HCO3 -, and KS0 is the solubility constant of CaCO3 at equilibrium.

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28 Diel monitoring of STA-1W Typha and Najas communities The mesocosm platforms provided a controlled environment for examining the influence of macrophyte s on the water column chemistry. However, the water column environment in a full-scale STA during operations (e.g. high flow events) may be different th an that observed at the small-scale. To investigate STA water column chemistry, diurnal monitoring of emergent and SAV communities took place in the southeastern region of STA-1W Cell 1 (Figure 2-2) in July 2002. Two Datasonde Hydrolab multiprobes were deployed to record pH, temperature, D.O. and specific conductance at 15 min intervals during a two-week period of high flows (up to 1200 cfs through STA 1W). Each device was suspended from a tripod to an initial probe depth of 10 cm below the water surface. Dissolved oxygen, temperature and pH profiles through the water column were recorded for the same SAV and emergent stations on July 27 and September 29, 2002, using Yellow Springs Instrument dissolved oxygen meter and Corning pH meter. In September, the water column light regime was also assessed in SAV and emergent macrophyte stands, as well as in open water reaches of STA 1W Cell 1. Incident light was recorded simultaneously above the water column (ambient level), in surface waters (3 cm), at mid-depth and 10 cm above the substrate, using the methodology describ ed for mesocosm light measurements.

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29 Flux Study Using Porewater Equilibrators Porewater was collected using porewater equilibrators, or peepers, modeled after Hesslein (1976) (Figure 2-3). Triplicate peepers were deployed in Cell 1 (~ 1 m apart) within emergent and submerged vegetation on June 10, 2002, and retrieved 17 days later. Three peepers at each station were inserted vertically through the accrued soil into the underlying muck soils to a depth of ~25 cm. Figure 2-3. Schematic of porewater equilibrat or used in estimates of P flux from STA-1W Cell 1 soils into the overlying water column. A 0.2 m polyether sulfone membrane (Supor 200) was inserted between a coarse particle filter and the 8 mL sampling cells. Some sample cells are above the soil-water interface, while others equilibrate with porewater below the soil surface. The soil-water interface bisected th e peeper such that some sampling chambers would equilibrate with soil porewater, and others would equilibrate with the water column above the interface. Specific conductance, SRP, TSP, Soil surface Coarse filter 0.2 m membrane Water Cover plate 8mLcell Soil surface Coarse filter 0.2 m membrane Water Cover plate Soil surface Coarse filter 0.2 m membrane Water Cover plate 8mLcell

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30 dissolved calcium and iron, and alkalinity concentrations and pH levels were measured above and below the soil-water interface. Dissolved iron determinations were made by a bathophenanthroline method, modified from APHA 3500-Fe D (APHA, 1992) for small sample volume and using a Spectronics Genesys 5 spectrophotometer. Diffusive fluxes of P (i.e. H2PO4 and HPO4 2-) at each coring location were calculated based on changes in SRP conc entration with depth. The concentration gradient of soil porewater constituents with distance can drive diffusive flux according to Fick’s first law, where: 91 1 47 0 dz dC D Js [6] where J = diffusive flux, mg m-2 s-1 = porosity Ds = the sediment diffusion coefficient, cm2 sec-1, and dC/dz is concentration change, g L-1, per depth interval cm. The final term accounts for sediment porewater tortuosity, a parameter that is linearly related to porosity (Sweerts et al., 1991). The concentration gradient dC/dz was determined using the slope of a linear regression through the +5 to –5 cm depth interval. Temperature was recorded during th e equilibration period with max/min thermometers deployed at the soil-water in terface adjacent to one peeper at each station. Alkalinity concentrations and pH levels were recorded in the field immediately after peeper retrieval. Th e thermodynamic potential for calcium

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31 carbonate and calcium phosphate saturat ion and mineral formation was then calculated for the water column-p orewater continuum at each site. Statistical analyses on experimental data were performed using MSExcel (v. 2000 Microsoft Corp.) ANOVA and t-te st macros. Error around mean values is presented as one standard deviation for replicate samples. Results and Discussion Soil Characterization STA-1W was previously farmland, with irrigation canals transecting the acreage. Since the land was flooded in 1993, cattail preference for shallow waters has encouraged growth along old canal banks. Dredged spoil from pre-flooding canal maintenance resulted in raised soil surfaces, and shallow water depths. This pattern is evident in aerial photographs taken of the wetland in November 2000, after six years of flow-through operation (Figure 2-2). Measured from water surface to the top of the litter (or accumulated soil) layer, free-water depths on June 10, 2002, ranged spatially in the outflow region of STA-1W Cell 1 from 0 to 56 cm. Additional water column was occupied by Typha leaf litter, with up to 77 cm from air to soil surface. Based on probing-rod sounding measurements, litter accrued to depths ranging from ~ 0 to 30 cm. Wetland soils had accrued above the native muck soils to variable depths beneath the emergent (6-14 cm) and submerged (4-20 cm) macrophyte communities. The accrued material was flocculent organic matter of lower bulk density than the underlying muck soils (F igure 2-4). Accrued soil bulk density in

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32 the two community types was not significantly different for any given depth (p > 0.05). Muck soils were near the surface in some cores, while other cores were characteristic of deeper (> 6 cm) soil accrual, which may explain wide variability in the 4-6 cm layer bulk density measur ements at both the SAV station (0.169 0.049 g cm-3) and Typha station (0.223 0.093 g cm-3). Figure 2-4. Average bulk density values for accrued sediments at depth intervals below the sediment water interface. The underlying farm muck was collected from below the horizon of the accrued sediment. Error bars indicate 1 s.d. for triplicate sample s, and each sample is a composite of five discrete soil samples. Soil Phosphorus Pools A sequential extraction was used to characterize the relative bioavailability of P within the surface so ils and the underlying muck soils. Reddy and Graetz (1991) used a similar sequential extraction, except 1M KCl was substituted for NH4Cl to characterize the readily available P pool. Soil TP 0.000.050.100.150.200.250.30 0-2 2-4 4-6 MuckSoil Layer (cm)Bulk Density (g cm-3) Submerged Emergent

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33 decreased significantly from surface soils (0-2 cm and 2-4 cm layers) in the emergent stand to the underlying muck soils(p < 0.05), while the SAV 2-4 cm soil layer showed a slight increase above the surface layer in each of three replicate composites (Figure 2-5). In the upper 4 cm, soils from the Typha community were P-enriched (584 mg P kg-1) relative to the native muck soils (334 mg P kg-1, p < 0.05), but enriched less than 0-4 cm SAV soils (813 mg P kg-1; p < 0.01). There was a significant difference in muck soil P levels beneath the SAV and the emergent communities. SAV-region muck soils were 2-3 times higher in P than the emergent muck soils (500 212 and 168 94 mg P kg-1, respectively). In comparison to the native farm (Knight’s Farm) soils (335 31 mg P kg-1) and soils 10 months after flooding (358 35 mg P kg-1), the muck in the emergent region has become depleted in P over the 7 years of operation, while submerged macrophyte muck has been P-enriched (Figure 2-5). The residual P pool was increased 10 months after flooding, and in recent samples from all depths. Variation in soil TP on Knight’s farm prior to flooding was greater than variation for other soil parameters (Reddy and Graetz, 1991), yet no soil-P value was reported lower than those observed be low the emergent stand in this study. Local variation in P distribution was a ccounted for with composite soil samples from multiple cores, taken several meters apart. The SAV and emergent stations were several 100 meters apart. Compositing vertically, however, may complicate the interpretation of observed differences between this study and that of Reddy and Graetz (1991).

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34 Figure 2-5. Soil phosphorus pools determ ined through a sequential extraction procedure using 1.0 M NH4Cl, 0.1 M NaOH, and 0.5 M HCl. The residual P pool remained associat ed with the soil through all three extractions. Knight’s Farm data from Reddy and Graetz, 1991. 02004006008001000 0-2 cm 2-4 cm Muck 0-2 cm 2-4 cm Muck SAV EAV Soil Phosphorus (mg P kg-1) Drained 0-30cm Flooded 0-5cm Knight's Farm Residue Ca/Mg Bound Fulvic/Humic Acid Fe/Al bound Readily Available

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35 Those investigators homogenized soils at 30 cm increments before flooding, and 5 cm increments after flooding. During ag ricultural production surface soils are mixed through tillage. However, wetland soils lower in the profile may have lower P concentrations than surface soils, and large depth intervals do not adequately describe surface soil chemistry. Thus, small depth increments were used in this study. Accrued soil-P occurred primarily (34-60%) as residual P, a highly recalcitrant form. Calcium-bound HCl-extrac table P was also a major pool in new wetland soils (15-45%), especially in th e SAV surface 0-4 cm soils (mean 33%, or 271 mg HCl-extractable P kg-1). This is a substantial increase from 8 mg HClextractable P kg-1 found in the drained surface (0-30 cm) soils (Reddy and Graetz, 1991). Significantly lower fractions of total soil P were associated with fulvic and humic acids in accrued soils (5-26%), relative to drained farm muck (58%) (p < 0.05). This pool of organic, moderately available P characterized by 0.1 M NaOH extraction was highest in the Typha surface 0-2 cm soils (19%, or 128 mg NaOHextractable organic P kg-1), and decreased with depth at both stations. The Feand Al-bound P pool was relatively small in accrued wetland soils, representing 2-8% of soil TP. Neve rtheless, emergent surface 0-2 cm soils contained 45 5 mg NaOH-extractable inorganic P kg-1, nearly twice that of SAV surface soils (26 11 mg kg-1). If such a pool was primarily P associated with Fehydroxides, it would be subject to mobilization during prolonged anoxic

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36 conditions. Release of the whole 0-2 cm pool into a 1-m deep water column, assuming equal soil bulk density of 0.1 g cm-3, would elevate TP concentrations by 52 g L-1 in SAV communities, and 90 g L-1 in the emergent region. Muck soils below emergent vegetation were three-fold lower in Feand Al-bound P (4 3 mg kg-1) than muck below SAV (12 5 mg kg-1). The NH4-extractable P pool (readily available fraction) was higher in the SAV accrued soils (56 8 mg kg-1) than in emergent accrued soils (32 10 mg kg1). Muck soils were low in NH4Cl-extractable P, with 5 2 mg kg-1 readily bioavailable. These values are within the range of values reported by Reddy and Graetz (1991) for farm soils prior to flooding (0-30 cm, 42 mg kg-1), and muck soils after eight years of submergence (0-5 cm, 2 mg kg-1). Accumulation of organic matter, P enrichment of surface soils and detritus, and subsequent humification of the organic material, have been viewed as positive wetland attributes for P removal treatment applications. Phosphorus can be removed from the ambient water and become concentrated in recalcitrant organic soil components. When the availability of soil-P was characterized by sequential extraction, however, it was shown that the “bioavailable” and “recalcitrant” P pools alike were reduced in the soil beneath emergent vegetation, relative to either pre-flooded conditions or a contrasting community type, namely submersed macrophytes ( Najas guadalupensis ) (Figure 2-5). It seems likely that recalcitrant P compounds may be susceptible to mobilization through biotic mechanisms such as organic acid mineralization or enzymatic hydrolysis.

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37 In contrast to the Najas community soils, the accrued soil beneath Typha was likely in close contact with, and under the influence of, extensive below ground biomass. Water Column Profile and CaCO3 Saturation Index Each mesocosm received a mean inflow P loading of 3.5 g P m-2 yr-1 (average inflow concentration = 96 g TP L-1) over the 2.7-year period of operation. Total P concentration reduction was significantly greater ( = 0.05) in the SAV mesocosm, with an average outflow concentration of 26 g L-1, than in the cattail mesocosms, whose outflows averaged 42 and 62 g L-1. Differences between cattail communities were not significant with respect to P removal performance. Calcium and alkalinity concentration reductions were also observed in SAV-mesocosm surface waters, but not in the cattail stands. Between February 15, 2000 and August 8, 2001, Ca levels in inflow waters (72 mg L-1) common to both SAV and cattail mesocosms were reduced to 47 mg L-1 at the outflow of the SAV mesocosm. Cattail mesocosm outflows averaged 69 and 70 mg Ca L-1. Likewise, alkalinity was reduced from 206 to 135 mg CaCO3 L-1 by SAV while cattail outflows averaged 202 and 204 mg CaCO3 L-1. On May 30, 2001, SAV mesocosm surface water outflow SRP concentrations were low (5 g L-1 ), relative to the inflow region (20-43 g L-1), and Post-BMP inflow waters (49-58 g L-1) (Table 2-2). In the Typha -dominated mesocosms, SRP concentration reductio ns were smaller, with surface outflow

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38 waters of 18 and 37 g L-1. Water at 20 cm depth had higher SRP concentration than surface waters in all mesocosms, perh aps a result of internal P loading from the soil to the bottom waters. Similar trends were seen in the Ca and alkalinity concentrations (Table 22). For example, surface outflow Ca levels (43 and 45 mg L-1) from SAV mesocosms were appreciably lower than those of the ADW inflow water (61 mg L-1). Water at 20 cm depth in both community types was nearly equal in Ca and alkalinity concentration, and similar to th e inflow waters (Table 2-2). In the SAV systems, constituent concentrations dec lined between inflow and outflow, but reductions were more apparent in surface water than at 20 cm depth. Calcium, alkalinity, pH and temperature levels in the Typha mesocosms were uniform internally, and showed little change with depth or with distance, compared to the SAV mesocosms (Table 2-2). Surfac e waters within the SAV mesocosms had elevated pH levels and temperatures, co mpared to the surface waters shaded by Typha Calcium carbonate precipitation was thermodynamically favored (SI>1) only in SAV surface waters (Table 2-2). Typha surface waters and waters at 20 cm depth in either community were unsaturated with respect to CaCO3, and dissolution of the mineral was favored (SI < 1). While CaCO3 formation may have occurred in the surface waters, the cooler temperatures and lower pH levels of waters at 20 cm depth created an unsaturated environment. Calcium and alkalinity concentrations were higher at depth than in surface waters, which

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39 suggests that bottom waters may be in equilibrium with settled precipitates in the surface soils. Calcium carbonate prec ipitates may slowly dissolved in the CaCO3 -undersaturated bottom water environment. The influence of shade on aquatic phot osynthesis was apparent in the D.O. concentrations of the two community typ es (Table 2-2). One replicate cattail mesocosm contained more Najas than the other, and had a more open Typha canopy. Increased light penetration was observed for this mesocosm, with an 878% reduction in incident irradiance at the water surface compared to 972% reduction within the more dense Typha community. SAV surface water D.O. concentrations were 16.2 mg O2 L-1 or greater, whereas Typha surface waters were 0.5 – 3.5 mg O2 L-1. Waters at depth were lower in D.O. than surface waters in both SAV and Typha due to lower aquatic productivity and greater distance from the air-water interface. Diel Water Quality Monitoring of Emergent and Submerged Communities Since it was created in 1994, surface water in Cell 1 has been maintained at an average stage of 3.64 0.16 m NGVD. During the 2002 deployments of the porewater equilibrators (June 1027) and hydrolabs (July 3 – 17), the cell stage was only slightly higher than average (3.89 m) (Figure 2-6). Prior months had declining water levels and low (<10 cfs) hydraulic loads, typical flows during the spring dry season.

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40 Table 2-2. Water column characteristics at 3 cm and 20 cm depths in duplicate mesocosm s (2.2m L x 0.79m W x 0.4m D) dominated by emergent and submerged aquatic vege tation (EAV and SAV, re spectively) on May 30, 2001. Water from STA-1W inflow canal was delivered to each mesocosm at 10 cm day-1 between December 1998 and August 2001. TimeDO CaCO3 SI AlkalinityDiss. CaSRPpHTemp Rep mg L-1mg CaCO3 L-1mg L-1g L-1C 113:2214859.6587.3631.0 213:2215060.5567.5933.3 113:4014261.1527.3733.5 213:4014862.7497.5634.3 110:003.50.713460.1147.3825.7 211:201.50.414658.2367.1827.3 111:4016.2 15 14459.7298.6031.5 212:0020+ 32 10841.6119.2631.9 110:242.30.814658.5187.4225.4 210:450.30.314759376.9925.1 112:2020+ 103 12242.759.9033.2 212:3520+ 105 11142.359.9133.9 110:150.50.314658.4207.0924.5 211:300.20.216459.5496.8924.7 111:500.80.617267.7567.2224.8 212:100.10.615459.1307.3325.0 110:280.50.315058.5257.0024.6 210:550.20.214059.4406.9424.2 112:280.10.717660.577.3025.2 212:550.50.716260.757.3425.4 Outflow Region EAV SAV Inflow 3 cm Inflow Region EAV SAV EAV SAV 20 cm Inflow Region EAV SAV Outflow Region EAV SAV

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41 Fluctuating water levels exposed the Hydrolab probes to air twice during the two-week deployment, and only a subset of the data (July 9 – July 17, 2002) was used in this discussion. Differences in bottom elevation between stations meant that water depth was on average 0.3 m deeper at the SAV station than at the nearby emergent station (Figure 2-6). Water quality monitoring in the Cell 1 outflow region revealed similar temperature, dissolved oxygen saturation and specific conductance levels between emergent and SAV communities (F igure 2-7, Figure 2-8, and Figure 2-9). The emergent community, however, had pH levels roughly 1 pH unit lower than the submerged community (Figure 2-10). Average ( one s.d.) pH level from July 9 – 16, 2002 was 8.05 0.17 in the SAV community, as compared to 7.10 0.34 in the emergent community. Such a difference in pH may result from inorganic carbon uptake by SAV for aquatic photosynthesis. Conversion of HCO3 and CO2 into organic cell components likely depleted these constituents despite the evidence of well-mixed conditions provided by other parameters. Calcium carbonate equilibria can buffer water column pH around 8.3. In the SAV community, sustained underwater photosynthesis would tend to increase local pH levels, drive CaCO3 precipitation, and create a CaCO3 –buffered environment. The lack of a diel pattern in pH leve ls suggests that during the high flow event in June, little photosynthesis was occu rring at either station. The high flows may have suspended bottom sediment, and water depths were substantially

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42 Figure 2-6. Inflow rates and stage level of water in the outflow region of Cell 1 compared to bottom elevations at the emergent and submerged aquatic vegetation (SAV) stations, prior to and during peeper and Hydrolab deployments. increased. The dissolved organic carbon in wetland surface waters can attenuate light in a few meters depth (Krause-Jensen and Sand-Jensen, 1998). While D.O. 2.5 3.0 3.5 4.0 4.5 5.0 4/15/16/17/18/19/110/111/1NGVD Stage (m)HydrolabsPee p ersSAV Bottom Elevation EAV Bottom Elevation Cell 1 Stage 0 200 400 600 800 1000 4/15/16/17/18/19/110/111/1Date (2002)Cell 1 Flow (cfs)Hydrolabs Peepers

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43 Figure 2-7. Water column temperature within Cell 1 emergent and SAV communities during the Hydrolab monitoring period (July 3 – 17, 2002). Figure 2-8. Water column dissolved ox ygen saturation levels within Cell 1 emergent and SAV communities during the hydrolab monitoring period (July 3 – 16, 2002). 20 24 28 32 36 40 7/97/107/117/127/137/147/157/16 DateTemperature (C) SAV EAV 0 20 40 60 80 100 7/97/107/117/127/137/147/157/16 DateDissolved Oxygen Saturation (%) SAV EAV

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44 Figure 2-9. Water column specific conduc tance levels within Cell 1 emergent and SAV communities during the hydrolab monitoring period (July 3 – 16, 2002). Figure 2-10. Water column pH levels within Cell 1 emergent and SAV communities during the hydrolab monitoring period (July 3 – 16, 2002). 800 900 1000 1100 1200 1300 7/97/107/117/127/137/147/157/16 DateSpecific Conductance (S cm-1) SAV EAV 6.0 6.5 7.0 7.5 8.0 8.5 9.0 7/97/107/117/127/137/147/157/16 DatepH SAV EAV

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45 concentrations did increase from 0 to 20% saturation during daylight hours at each station, this is lower than the le vels observed in SAV communities under more typical, quiescent conditions (DBEL, 2001) or as observed in the mesocosms (Table 2-2). Oxygen concentrations in June were lower than those observed on September 29, 2002 at the same Cell 1 stations (Table 2-3). In September, D.O. concentration profiles at the SAV station revealed supersaturated conditions in the SAV surface waters (10.6 mg L-1, 31.8C) and lower levels (0.65 mg L-1) near the bottom of the water column (0.65 m from water surface and 0.1 m above the sediment surface). Emergent-stand surface waters were less saturated than the SAV station with respect to D.O. (2.15 mg L-1, 31.5C), but waters at depth were similar (0.60 mg D.O. L-1). Surface water pH levels at that time were 8.29 and 7.66 for SAV and emergent stations, respectively. Light penetration into the water column is essential for aquatic photosynthesis. On September 29, 2002, belo w-surface light levels were reduced in emergent stands relative to open water and SAV beds (Figure 2-11). In the SAV and open water areas, light available just below the surface was substantially reduced (~65%) from ambient light 2 m above the surface. Light reflecting off the water surface may have increased the “ambient” measurements and decreased the submerged surface me asurements. Further light reduction within the open water column may have resulted from attenuation and scattering by dissolved organic matter or suspended particulate matter.

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46 Table 2-3. Water column dissolved ox ygen (D.O.) concentrations, pH, and temperature profiles in the STA-1W Cell 1 surface waters at the time of peeper retrieval on June 27, and on September 29, 2002. The Typha canopy drastically reduced the amount of light available just below the water surface. SAV tissues reduced available light by shading the waters below the leaf canopy. Interesting to note, however, is the near total light extinction (>99% attenuated) at bottom depths (62 65cm) in both SAV and emergent communities, regardless of the shallow water column (75 cm). The phototrophic benthos at the open water stations, in contrast, had 7.7 0.3 % (as mean 1 s.d.) of the incident photon flux available as light energy. EAV SAV Parameter Depth June Sept. June Sept. Dissolved oxygen, mg L-1 Surface 0.30 2.15 1.1 10.6 Mid 0.25 0.55 1.1 3.1 Bottom 0.15 0.60 0.60 0.65 pH Surface 7.50 7.66 7.38 8.29 Mid 7.33 7.34 7.26 7.34 Bottom 7.47 7.15 7.22 7.26 Temperature, C Surface 31.0 31.5 30.3 31.8 Mid 29.7 28.9 29.3 29.0 Bottom 29.7 29.1 28.3 28.8 Water Depth, m 1.10 0.74 1.45 0.70

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47 0 10 20 30 40 50 60 70 051015202530354045Fraction Incident Light Remaining (%)Water Depth (cm) SAV Typha Open Water SAV shade Typha shade 0 10 20 30 40 50 60 70 051015202530354045Fraction Incident Light Remaining (%)Water Depth (cm) SAV Typha Open Water SAV shade Typha shade 0 10 20 30 40 50 60 70 051015202530354045Fraction Incident Light Remaining (%)Water Depth (cm) SAV Typha Open Water SAV shade Typha shade Figure 2-11. Incident light (measured as photon flux) remaining at water depths of 0+, 35 and 65 cm in emergent and submerged macrophyte communities, as well as within open water reaches, of STA 1W, Cell 1, in September 2002. The shaded region s represent the contribution of SAV and Typha canopy shading to light reductions, beyond the attenuation attributed to the water column alone. Even in shallow (<1m) treatment wetla nds, the sediment-water interface and a portion of the water column ma y be below the euphotic zone. Light availability controls aquatic photosynthes is and alters surface water chemistry. The influence of light and aquatic photos ynthesis on water column P dynamics, then, is greater near the air-water interface when the area is vegetated, while the remainder of the water column is influenced by respiration processes.

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48 Porewater Chemistry Significant rainfall within the S-5 basin during the 17-day equilibration required pumping of ADW into STA-1W. Water levels increased approximately 0.9 m between deployment June 10, and retrieval June 27, 2002 (Figure 2-12). At the time of retrieval, recent inflow waters had likely influenced surface water chemistry (Table 2-3). Figure 2-12. Stage level recorded during the seventeen day equilibration period for porewater samplers deployed in the outflow region of STA-1W Cell 1, during June 2002. The bottom dept hs for the SAV and Emergent vegetation stations are shown for reference. Profiles of porewater pH levels associated with emergent and submerged vegetation were circumneutral (6.77 7.57) in the both sediment types (Figure 2-13) and below the pH levels in the overlying water column. While pH values 2.5 3.0 3.5 4.0 4.5 5.0 6/106/176/24Date (2002)NGVD Stage (m) Cell 1 Stage SAV Bottom Elevation EAV Bottom Elevation

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49 Figure 2-13. Vertical profiles of porewate r pH values from soil collected from emergent and submerged vegetati on. Porewater equilibrators were deployed for 17 days in June 2002, in the outflow region of STA-1W Cell 1. were measured immediately after peeper retrieval, the disparity between replicates may be an artifact of oxygen reintroduction into, and/or CO2 outgassing from, porewater samples. However, the disparity is more likely due -25 -20 -15 -10 -5 0 5 10 15 20 25 6.06.57.07.58.0pHDepth (cm) EAV 6.06.57.07.58.0 SAV

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50 to the natural spatial variability of the soils. All three replicate peepers were retrieved at once at each site, and the first station’s samples were processed before the second station peepers were withdrawn. Those replicates processed first had lower pH values than replicates sampled last. All pH determinations were made within 4 hours of peeper retrieval. Profiles of SRP concentrations were different between emergent and submerged community sediments (Figure 2-14). Emergent aquatic vegetation (EAV) sediments exhibited slightly high er mean porewater SRP concentrations (0-6 cm depth, 996 131 g L-1) and lower overlying water column concentrations (0+6 cm depth, 402 293 g L-1) than the submerged plant communities (0-6 cm depth, 728 279 g L-1; 0+6 cm depth, 616 256 g L-1). The resulting diffusive P flux was greater from the emergent sediments (0.39 0.11 mg P m-2 d-1) than from SAV sediments (0.07 0.05 mg P m-2 d-1). Porewater DOP concentrations in EAV soils ranged from 0 to 350 g L-1, with one measurement at 658 g L-1, 18-20 cm below the soil surface (Figure 2-15). The SAV soil DOP profile ranged from 0-830 g L-1, with maximum concentrations 2-6 cm below the soil surface in all three replicates. DOP may diffuse upward into the water column, or downward into deeper soils, where further transformations may increase P recalcitrance, or conversely make it bioavailable. Surface soils also contained greater fulvicand humic acidassociated P in both soil types than th e muck soils lower in the soil profile (Figure 2-5). Microbial biomass and activity have been reported higher in WCA-

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51 2A surface soils than in deeper soils (White and Reddy, 2001), and may influence both the porewater DOP pool and th e soil fulvicand humic acid-P pools through release of metabolic by-products or mineralization of organic matter. Water columns at both EAV and SAV stations were likely anoxic during the 17-day peeper equilibration period. High flows (Figure 2-6) likely Figure 2-14. Vertical profiles of soluble reactive phosphorus concentrations in soil porewater collected from emergent and submerged vegetation. Porewater equilibrators were deployed for 17 days in June 2002, in the outflow region of STA-1W Cell 1. -25 -20 -15 -10 -5 0 5 10 15 20 25 0.00.51.01.52.0SRP (mg L-1)Depth (cm) Water column at time of deployment Water column at time of retrieval EAV 0.00.51.01.52.0 SAV

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52 Figure 2-15. Vertical profiles of dissolv ed organic P concentrations in soil porewater collected from emerge nt and submerged vegetation. Porewater equilibrators were deployed for 17 days in June 2002, in the outflow region of STA-1W Cell 1. resuspended surface sediments, increasing light attenuation and decreasing photosynthesis. Dissolved oxygen conc entrations on June 27, 2002 (retrieval) were low even in surface waters (0.30 and 1.1 mg L-1) within the SAV and EAV communities, respectively (Table 2-3). -25 -20 -15 -10 -5 0 5 10 15 20 25 0.00.20.40.60.81.0DOP (mg L-1)Depth (cm) EAV 0.00.20.40.60.81.0 SAV

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53 Under low-oxygen conditions, Fe-oxides can be reduced to soluble Fe2+. This was observed, as dissolved Fe concentrations increased with depth through the emergent macrophyte porewater (Figure 2-16). The profile in submerged community was more uniform with depth. All concentrations were low compared to those typical (50-100 mg L-1) of reduced mineral soil environments (Patrick and Khalid, 1974). Newman and Pietro (2001) reported similarly low total Fe concentrations (0.15-0.25 mg L-1) in STA-1W Cell 4 in 1993-1994 surface water samples taken just after field flooding. Even under oxygenated conditions, such low Fe concentrations may not provide the P sorption capacity characteristic of acidic mineral wetland soils (Richardson, 1985). Calcium, specific conductance and especially alkalinity concentration profiles all increased with depth in the EAV profiles, but were constant or decreasing with depth in the submerged profiles (Figure 2-17 to Figure 2-19). Calcium carbonate precipitation in su rface waters within the SAV community may have led to elevated values, while dissolution of the underlying limerock may have influenced the porewater concentrations in both soil types. The EAV community porewater exhibited a clear gradient of alkalinity concentrations, suggesting upward diffusive flux of thes e components through the soil profile, and potentially into the water column. Calcium carbonate precipitation is concomitant with aquatic photosynthesis in hard waters. Everglades ADW was often saturated with CaCO3 before entering STA 1W, as observed in mesocosm inflow waters.

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54 Figure 2-16. Vertical profiles of dissolved iron concentrations in sediment porewater collected from emerge nt and submerged vegetation. Porewater equilibrators were deployed for 17 days in June 2002, in the outflow region of STA-1W Cell 1. Daytime aquatic photosynthesis elevated pH levels in the mesocosms, and the water column became supersaturated. The formation of CaCO3 precipitates is therefore an important mechanism for P so rption capacity of newly accrued soils. -25 -20 -15 -10 -5 0 5 10 15 20 25 0.000.050.100.15Dissolved Fe (mg L-1)Depth (cm) EAV 0.000.050.100.15 SAV

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55 Saturation index values indicated CaCO3 precipitation was favored in the porewater of both community types and at all depths (Figure 2-20). Over the long term, water column CaCO3 precipitation may be less important to soil sorption capacity than previously though t. Differences in the water column Figure 2-17. Vertical profiles of dissolved calcium concentrations in sediment porewater collected from soils below emergent and submerged vegetation. Porewater equilibrators were deployed for 17 days in June 2002, in the outflow region of STA-1W Cell 1. -25 -20 -15 -10 -5 0 5 10 15 20 25 0100200300Calcium (mg L-1)Depth (cm) EAV 0100200300 SAV

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56 -25 -20 -15 -10 -5 0 5 10 15 20 25 02004006008001000Alkalinity (mg CaCO3 L-1)Depth (cm) EAV 02004006008001000 SAV Figure 2-18. Vertical profiles of alkalinity concentrations (as CaCO3) in sediment porewater collected from soils below emergent and submerged vegetation. Porewater equilibrators were deployed for 17 days in June 2002, in the outflow region of STA-1W Cell 1. chemistry observed for submerged and emergent macrophyte stands were not as apparent in the porewater. Influence of the native calcareous substrata on emergent community porewater may account for P storage in the Ca-bound, HCl-extractable P pool in those soils, despite the lack of evidence for calcium

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57 -25 -20 -15 -10 -5 0 5 10 15 20 25 0100020003000Specific Conductivity (S cm-1)Depth (cm) EAV 0100020003000 SAV Figure 2-19. Vertical profiles of specif ic conductance values in sediment porewater collected from soils below emergent and submerged vegetation. Porewater equilibrators were deployed for 17 days in June 2002, in the outflow region of STA-1W Cell 1. carbonate precipitation in the EAV water column. Water column CaCO3 precipitation in the SAV beds probably explains the higher HCl-bound P pool found in those soils (271 80 mg P kg-1) than those of the Typha -region soils (119 45 mg P kg-1).

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58 0.1110100CaCO3 SI SAV -25 -20 -15 -10 -5 0 5 10 15 20 25 0.1110100Depth (cm) EAV Figure 2-20. Vertical profiles of calcium c arbonate saturation index (SI) values in porewater collected from soils beneath emergent and submerged vegetation. All index values were greater than one, indicating supersaturated conditions with respect to CaCO3 Note log scale on xaxis. Intact Core P Flux Study Soils collected from beneath the Typha and SAV communities released P into overlying water under dark, oxic conditions (Figure 2-21). Initial floodwater was particle free, and contained 13 g DOP and 3 g SRP L-1. After 28 days of

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59 incubation, SRP concentrations were highes t in two of the three replicate waters overlying SAV soils (98 60 g L-1 for all three SAV cores). The cattail soils also released P into the water column at 49 49 g L-1. 0 20 40 60 80 100 120 140 160 180 051015202530 Elapsed Time (days)Water Column SRP (g L-1) SAV soils EAV soils Water Control Figure 2-21. Soluble reactive phosphorus (SRP) concentrations in the water columns of SAVand Typharegion soils and control columns (no soil) during 28-day dark laboratory in cubation. Error bars for each treatment indicate one standard deviation between triplicate cores. Note: Typha soil treatment values are fr om duplicate cores, while control and SAV soil values were from triplicate cores. Of the three un-amended cattail cores, one replicate had visible fragments of leaf litter within the surface soils. The results observed from this core were excluded from flux calculations, due to the possible effects of the litter fragments on the results. Phosphorus movement from the soil to the water during the 7-20 day period of linear SRP concentration increase was greatest from SAV soils.

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60 Corresponding flux rates from SAV soil, adjusted for change in control column concentrations, was 1.84 1.04 mg P m-2 day-1. Emergent soils showed slightly less P release (Table 2-4), though the di fference was not significant (p > 0.05). Near-zero (-0.02 to 0.03 mg P m-2 day-1) flux was observed in the control cores. Water column pH and temperature valu es stayed consistent throughout the 28-day incubation. Initial flood water pH (8.06) increased slightly to 8.36 0.06 before the 8-hour sampling, and remained similar in all treatments (8.41 0.09) during the remaining 28 days. Temperature was moderated by the incubation water bath, and averaged 24 3 C during the incubation period. Table 2-4. Flux estimates from intact SAV and Emergent soil cores kept under dark conditions for 35 days. The 30 cm water column was continuously aerated. A 100 g L-1 phosphorus spike was added after 30 days to measure short-term uptake rates. All values are means ( st. dev.) of triplicate cores minus mean flux in control cores, in mg P m-2 day-1. Period of flux estimate SAV soil EAV soil Release Day 7 to Day 20 1.84 1.04 1.20 0.27 Uptake 0 8 hours after spike 4.97 12.6 0.17 2.7 Similarly, dissolved calcium (Ca) concentrations were stable after a short initial equilibration period (Figure 2-22). Alkalinity concentrations increased marginally in the Typha soil treatment over the 28 days (Figure 2-22). SAV soil and water control treatments were constant with respect to dissolved calcium and alkalinity concentrations.

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61 Dissolved organic P concentrations were slightly elevated in SAV-soil incubations, relative to control treatments or treatments with Typha soils (Figure 2-23). During the incubation period, me an DOP concentrations for both soil Figure 2-22. Mean dissolved calcium and alkalinity concentrations in the water columns of three treatments (see te xt for details) during 28-day dark laboratory incubation. Error bars indi cate 1 standard deviation from the mean of three replicates. 200 220 240 260 280 300 320051015202530 Elapsed Time (days)Alkalinity (mg CaCO 3 L-1) 60 80 100 120Dissolved Ca (mg L-1) EAV soil SAV soil Water control

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62 treatments decreased slightly over the first 24-hours, then increased for the next 27 days. After 28 days, water column DOP concentrations above SAV soils (28 3 g L-1) were higher than in waters above cattail soil (22 3). These concentrations are very low relative to those observed in the porewater (e.g. as high as 830 g L-1 in the surface 2-6 cm SAV soils), which were also higher in the SAV soils (Figure 2-15). Concentrations of dissolved organic carbon (DOC) were similar during the period of linear P flux (monitored only between 4 and 14 days of incubation) (Figure 2-24). Figure 2-23. Mean dissolved organic ph osphorus concentrations in the water columns above SAV and Typha -region soils and in control (water only) columns during 28-day dark laborator y incubation. Error bars indicate 1 standard deviation from the mean of three replicates. 0 10 20 30 40 0.1122028 Elapsed Time (Days)Water Column DOP (g L-1) SAV EAV Water Control

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63 Greater P flux from SAV soils compared to the Typha -region soils, though insignificant (p> 0.05), was supported by gr eater total and exchangeable P in the SAV soils (Figure 2-5). The potential diffusive P flux of 0.39 mg P m-2 day-1 in the Typha region, as calculated from porewater gradients, was a small contributor to overall flux (1.20 mg P m-2 day-1). Figure 2-24. Dissolved organic carbon co ncentrations in water columns over Typha soils, SAV soils and control colu mns (no soil) after 4, 7, 10 and 14 days of dark laboratory incubation. Error bars indicate 1 standard deviation from the mean of three replicates. Rapid biological P uptake occurred during the 8 hours following P amendment (100 g P) (Figure 2-25). Control columns showed a small decrease in SRP concentration, then leveled off near 60 g L-1. Phosphorus in the control 36 37 38 39 40 41 42 43 44 45 46 471014Elapsed Time (days)DOC (mg L-1) EAV soil SAV Soil Water Control

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64 waters may have been absorbed by bacteria in the water column or attached to core walls, aerator, etc., and converted to non-reactive phosphorus. 0 30 60 90 120 061218243036424854Time after PO4 Amendment (hr)Water column SRP (g L-1) SAV EAV Control Figure 2-25. Water column SRP conc entrations above a “background” concentration as determined before a 100 g spike was added to each column. Background concentration was assumed stable through the 53-hour period. Triplicate water columns were assembled with or without soils from cattail or SA V communities and aerobically incubated for 30 days prior to the P-spike amendment. The change in the control water concentration was assumed to be an experimental artifact and was subtracted from the observed concentration for each treatment. The effects of soil type on P flux could then be evaluated directly. In the eight hours following P additions, SAV soils showed greater potential for P uptake than the EAV soils (Table 2-4). The EAV soil cores provided almost no uptake capacity above that observed for the control cores. This may be indicative

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65 of higher P sorption capacity and the influence of CaCO3 in the SAV surfaces soils, compared to EAV soils. Implications for STA and WCA Management In both STA-1W and WCA-2A, recently accreted soils were enriched in P relative to deeper soils (Craft and Richardson, 1993; Reddy et al., 1993; Reddy et al., 1998; Koch-Rose et al. 1994; Newman and Pietro, 2001). In this study, SAV soils were enriched relative to soils from EAV areas within STA-1W Cell 1. Surface-soil P enrichment creates a potent ial for diffusive flux both upward into the water column, as well as downward into underlying soil material. Upward flux can increase outflow water P concentrations and reduce wetland removal efficiency. Downward flux, on the other hand, may be a beneficial process for long-term storage. SAV soils (0-2 cm layer) contained more P in stable pools (705 68 mg P kg-1) than EAV soils (555 111 mg P kg-1) after eight years of flow-through STA operations. Labile pools of exchangeable and Aland Fe-bound P were similar in both surface-soil types (84 8 mg P kg-1 for SAV; 83 15 mg P kg-1 for Typha ). Because of similar labile-P pools, the two soil types each released P to an oxic water column at similar flux rates, with slightly greater flux rates coming from the SAV soils. SAV soils also reduced P concentrations following water column P amendments, whereas Typha soils did not. Within the submerged community, uptake mechanisms for water column P reduction are mostly associated with so il sorption and leaf surfaces. Leaves are

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66 concentrated towards the top of the canopy to intercept light, and senesce from older, lower portions of the plant. A Najas canopy submerged under increasing water depths would isolate water beneath the submerged canopy and close to the soil surface from the overlying water column. This lack of water exchange above the soil-water interface could allow bottom water to reach P concentrations similar to those of the porewater through diffusion. Removal mechanisms were not effectively maintaining water column concentrations lower than the porewater in SAV beds during the high flow event of June 2002. Net removal of SRP did not occur in the water column within 5 cm of the soil-water interface, which in e ffect increased the diffusion distance and reduced the overall flux rate. Had the peeper incubation period occurred under conditions of a shallow water column, biological P uptake near the soil-water interface may have resulted in a higher P flux. Potential diffusion rates appeared to account for only a portion (4-33%) of the flux measured in intact cores. Othe r studies estimating P flux in wetlands have found diffusive flux rates lower than intact core flux or in situ benthic chamber measurements (Fisher and Reddy, 2001). Burrowing macroinvertebrate activity may cause advective water exchange from soils into overlying water. Additionally, uptake of water column P near the soil surface by macrophytes and their epiphytes, phytoplankton and benthic microorganisms is necessary to maintain strong concentration gradients over short diffusion distances. Inorganic

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67 retention of P through adsorption and Fe precipitate formation also influence the distribution of dissolved P along the water column-porewater gradient. In Everglades soils, soil respiration rate (as CO2 production) and decomposition of litter/detritus was correlated to soil P content along a nutrient enrichment gradient in the northern Everglades (Davis, 1991; DeBusk and Reddy, 1998; Qualls and Richardson, 2000). Thus increased P mobilization may be expected in P-enriched areas, relative to areas of lower P levels. The rate of organic matter mineralization for soils along a eutrophication gradient in WCA2A was controlled primarily, however, by the availability of electron acceptors (White and Reddy, 2001), of which oxygen is the most efficient. The nutrientimpacted region of WCA-2A is frequent ly anoxic due to high sediment oxygen demand and shade-induced limitations of aquatic photosynthesis. Therefore, organic matter mineralization rates may be higher in P-enriched areas dominated by submerged macrophytes that are capable of supplying O2 to the water column, than in those areas dominated by dense Typha stands. The nutrient-impacted areas of WCA-2A and STA-1W are near monospecific stands of Typha While SAV communities seem to be outcompeted in these inflow regions, they do ex ist in STA-1W further downstream. Data collected from other porewater studies in STA-1W (DBE, unpublished data) along with the results presented here, su ggest diffusive flux rates decrease with distance through the wetland for the SAV communities (Figure 2-26). Such a relationship is useful for predicting inte rnal P loading at various locations along

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68 the treatment gradient. However, the diffusive flux is likely only a fraction of the total mobilization of P from soils to overlying water. Figure 2-26. Phosphorus diffusive flux estimates based on 22 porewater equilibrators deployed in STA-1W Cell 1 (June-July 2002; closed markers) and Cell 4 (November-De cember 2001; open markers), as a function of distance through the entire wetland. Diamond markers indicate SAV; square markers indi cate EAV. The regression line was calculated using only SAV diffusive flux rates. Emergent vegetation diffusive P flux did not show similar dependence on fractional distance. This difference may ha ve been due to the influence of root uptake on soil porewater SRP concentrations, which was greater for soils in the outflow Typha -dominated region than for soils from dense Typha stands near the inflow region. Porewater SRP concentrations in the upper 10 cm of emergent y = 2.25e-8.19xR2 = 0.74 0.0 0.2 0.4 0.6 0.8 0.00.20.40.60.81.0 Fractional Distance through STA-1WDiffusive P Flux (mg P m-2 d-1) SAV EAVCell 1 In Cell 1 Out Cell 4 Out Cell 4 In

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69 sediments were much lower in the inflow region (0.18 mg P L-1) than near the Cell 1 outflow (0.99 mg P L-1). The Cell 1 inflow emergent station may also have been isolated from the primary flow paths through the dense em ergent in the inflow region, whereas the outflow station was adjacent to open water. This difference in proximity to flow paths may have subjected the inflow-region Typha station to lower P loads than the outflow-region Typha station. These data suggest that the internal P load from porewater diffusion depends on vegetation, distance from the inflow, and the proximity to the main flow paths through the wetland. Conclusions Managers of treatment wetlands must consider options for reducing internal P loading, which can occur through diffusion and other processes (e.g. resuspension, dissolution), by creating stable soil-P pools. At a minimum, internal P loading should be minimized in outflow regions of the STAs to achieve low TP concentrations in surface water outflows. Options may include maximizing natural processes of new soil humification and Ca-P mineral formation, or enhancing stability through chemical additions. Areas dominated by SAV may provide a benefit during early stages of wetland creation and soil development through water column CaCO3 precipitation, which appeared to increase the Ca-bound P pool in SAV soils, relative to EAV soils. Soils beneath EAV were P-depleted relative to soil P concentrations shortly after flooding, which suggests that pools operationally defined as “more recalcitrant” are still

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70 subject to mobilization. Over time and especially in inflow regions, high rates of soil accrual may necessitate mechanical soil removal to reduce internal P loads, which occurred regardless of vegetation type.

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71 CHAPTER 3 BIOMASS PHOSPHORUS STORAGES AND DYNAMICS Introduction Tracts of the northern Everglades were transformed into three water conservation areas (WCAs) in the 1940-50’s by surrounding the vast marshes with earthen berms (Figure 3-1). These areas provided water storage and flood control for urban and agricultural development in South Florida. To the northwest of the WCAs lies the Everglades Agricultural Area (EAA), a 200,000 ha region of drained Everglades soils in sugarcane and winter vegetable production. WCA-2A has received agricultural drainage water (ADW) discharges since 1955 (Bartow et al., 1996). Excessive phosphorus (P) loading from these discharges has been identified as the primary cause for an observed eutrophication gradient. Changes observ ed near the discharge structures include increased water column, soil, and plant tissue P concentrations, and change in ecosystem function, relative to the interior marsh (Craft and Richardson, 1993; DeBusk et al., 1994; Reddy et al., 1993; Reddy et al., 1998). In order to reduce P loads to the Everglades, stormwater treatment areas (STAs) were constructed to intercept P in EAA drainage waters before it enters the WCA marsh. The STAs have performed well since construction, reducing TP concentrations to below 50 g L-1 as required by the Everglades Forever Act

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72 (EFA) of 1994. The EFA mandates further P load reductions, requiring TP concentrations of 10 g L-1 for discharges into the WCAs. Figure 3-1. The historic Everglades region of south Florida is now three distinct parcels, including the Everglades Agricultural Area (EAA), Water Conservation Areas (WCA) and Evergl ades National Park. Stormwater treatment areas (STA) are shown (in gray), including STA-1W. N E W S Everglades National Park WCA-1 WCA-3 WCA-2A FLORIDA, USA EAA Lake Okeechobee STA-1W

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73 Macrophytes influence P removal performance in treatment wetlands through direct P uptake as well as by influencing the physicochemical characteristics of its environment. Typha domingensis can grow in dense stands when nutrient and hydrologic conditions allow. Phosphorus enrichment in the northern Everglades has increased Typha above-ground biomass and shoot density, and slough communities of WCA-2A were replaced by Typha near inflow structures (Wu et al., 1997; Miao et al., 2000). Dense Typha stands and open water areas differ in rates of water exchange and community metabolism (Belanger et al., 1989; Brix, 1997; McCormick et al., 1997). In WCA-2A, Typha canopy shading of the water column has reduced aquatic photosynthesis and wind-induced mixing, leading to reduced oxygen supply relative to nearby open water areas (Belanger et al.,1989; Grimshaw et al., 1997). Dense stands of macrophytes can have large accumulations of organic materials, either settled from flowing water or produced on-site. These accumulations increase heterotrophic oxygen demand which, coupled with canopy shading, can result in anoxic conditions (Belanger et al., 1989; McCormick et al., 1997). Also important to P dynamics in wetlands are microbial processes in the soil and water column, which respond quickly to system inputs or environmental change. Qualls and Richardson (2000) reported Typha litter in WCA-2A flumes sequestered SRP up to 10 times the original litter content over a one-year period, with little net change to the macrophyte P content. Microbial biomass on the leaf litter increased in P content, yet the factors controlling

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74 microbial P uptake and release from wetland biomass remain unidentified. Specifically, the effect of community metabolism on biomass P uptake within Typha leaf litter is unknown. As an open water community transitions towards a dense macrophyte stand, such as occurred in WCA-2A, P cy cling and removal processes within the water column are likely affected. A similar increase in Typha stand density likely occurs within the STAs due to high rates of nutrient loading, yet the effects of water column shading by Typha on P cycling process remains unknown. Through field observations, mesocosm experiments, and controlled laboratory incubations, P storage, uptake and release were examined from Typha tissues as well as from microbial populations associated with Typha leaf litter. Specific objectives included: Estimating the relative partitioning of P between live and dead components of Typha stands (Mesocosm Phosphorus Storage) Investigating the role of leaf litter accumulations on P release from the soil (Flux Study Using Intact Soil Cores) Quantifying the rate of P uptake from the water column by microbial populations associated with Typha litter under oxygenated and anoxic conditions (Litter Incubations) Materials and Methods Bench-scale studies took place at the Wetland Biogeochemistry Laboratory at the University of Florida in Gainesville, FL. Outdoor mesocosms were located next to the STA-1W inflow canal on an experimental platform provided by DB Environmental. Field investigations took place in STA-1W Cell 1.

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75 STA-1W Site Description Agricultural runoff from the Everglades Agricultural Area (EAA) Basin S5 (Figure 3-1) is pumped in canals to STA-1W, a full-scale (2699 ha) treatment wetland operated by SFWMD to reduce P loadings to the Everglades (SFWMD, 2003). Everglades muck soils were drained for agricultural production decades ago, and were reflooded in 1994. The wetland contains three flowpaths of two cells each (Figure 3-2). Cell 1 is the first cell of the eastern flowpath, and contains emergent vegetation ( Typha spp.) in the inflow region and along the eastern edge (Figure 3-3). A mixture of emergent, floating, and submerged vegetation occupy the downstream reaches of the cell. This distribution of community types has remained relatively constant in Cell 1 since the wetland was first flooded (Newman and Pietro, 2001). During eight years of flow-through operations, the wetland has accrued new wetland soils. Mesocosm Phosphorus Storage Two outdoor mesocosms (2.2m L x 0.79m W x 1.0m D) located near STA1W received STA-inflow water for 2.7 years at a hydraulic loading rate of 10 cm day-1, as part of another study (DBEL, 2001). Muck soils in these systems were inoculated with Typha collected from within STA-1W. The vegetation and soils in the two cattail mesocosms were sampled destructively in late August and early September 2001, at the end of the water quality monitoring period (December 29, 1998 through August 14, 2001). All vegetation was collected from each meso cosm and segregated into inflow and

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76 outflow regions. Cattail shoots were cut at the soil surface, and separated into live (green) and dead shoots. Triplicate soil cores (38.5 cm2 each) were retrieved from the inflow and outflow regions. Th e underlying muck soil was discarded. Figure 3-2. Stormwater Treatment Area 1 W in Palm Beach county, Florida. Three flowpaths receive surface water drained from adjacent agricultural soils and reduce the phosphorus load to Water Conservation Area 2A. Cells 1, 2 and 3 are comprised of mixed emergent, submerged and floating vegetation, while Cells 4, 5A, and 5B are primarily submerged and floating vegetation. Arrows indicate general direction of flow. 4 2 5B 5A 1 3 Loxahatchee NWR Water Conservation Area 1 STA 1W (2699 ha) NorthernFlowpath EasternFlowpath WesternFlowpath 4 2 5B 5A 1 3 Loxahatchee NWR Water Conservation Area 1 STA 1W (2699 ha) NorthernFlowpath EasternFlowpath WesternFlowpath

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77 Figure 3-3. Aerial photograph of STA-1W Ce ll 1, first cell of the eastern flowpath, in November 2000 (courtesy SFWMD). Also shown is the location of the emergent vegetation field station. Belowground cattail biomass and submerged macrophyte tissues were collected after soil sampling was complete. Wet weights for vegetation samples were recorded in the field. Bulk density was determined for accrued soils. Vegetation and soil samples were oven dried (65C), weighed, and homogenized for analysis. TP content was determined by digesting 50 mg sample in concentrated nitric acid, followed by perchloric acid digestion at 210 C (COE 3227, Plumb 1981). Cell 1Outflow InflowWCA-1 Loxahatchee NWR Field Station

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78 Phosphorus mass removals were calculated for each mesocosm based on inflow and outflow TP concentrations of weekly or bi-weekly grab samples and a constant hydraulic loading rate of 10 cm day-1 (DBEL, 2001). The P mass recovered in accrued soil, together with the change in biomass P storage, was compared to mass removal based on water column concentration reductions. Plant Tissue Desiccation Study Biomass components within the water column of a treatment wetland are largely comprised of live SAV in submerged macrophyte beds, and both live and dead shoots in emergent stands. Samples of fresh Najas tissue and dead Typha shoots (leaf litter) were collected from the outflow region of Cell 1 on May 29, 2002. Triplicate vertical acrylic columns (7 cm i.d.) received 13 g tissue (wet wt.), along with 0.75 L of reflood water collected from the STA-1W Cell 4 outflow weir. A pair of control columns contained no plant tissue. All treatments and controls were placed in a water bath, co ntinuously-aerated, and covered with an opaque shroud. Water samples were withdrawn at times 0, 1 and 9 days, and filtered (0.45 m). SRP and acidified TSP samples were analyzed colorimetrically (potassium antimony tartrate, sulfuric acid, ammonium molybdate, ascorbic acid) on a Spectronics Genesys 5 spectrophotometer (EPA 365.2; EPA 1979). Analysis for TSP was preceded by persulfate acid digestion and neutralization. Flux Study Using Intact Soil Cores With and Without Typha Litter Six replicate intact soil cores were retrieved from a Typha stand (26.6292N, 80.4219 W) in the outflow region of STA-1W Cell 1 on July 4, 2002.

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79 Each core was retrieved by inserting an acrylic core tube (38.5 cm2) through the newly accrued soil layer into the underlying native farm muck to a minimum depth of 10 cm. The top end of the core was sealed with a #13 rubber stopper prior to extraction. Each of the six intact cores was completely filled with site water and sealed with a rubber stoppe r. An opaque shroud minimized solar heating and blocked light during transport to the lab. Typha leaf litter was collected from STA-1W Cell 1 at the time of soil core collection, and was kept in site water on ice for transport to the lab. Litter consisted of intact shoots that had little to no visible damage (from grazers, breakage, etc.), yet had become neutrallyor negatively buoyant. This phase of shoot decomposition was chosen because of three attributes: Fresh live shoots would likely leach tissue P rapidly after being collected from the source plant Ample time in the water column for colonization by aquatic microorganisms Negative buoyancy facilitates the experimental incubation design Each shoot was divided into uniform lengths of ~5 cm. Pieces from all shoots were mixed at random, then added to six cores, three with soil and three with reflood water only (litter controls). Overlying water was replaced with 1.15 L filtered (0.45 m) STA-1W Cell 4 outflow water (15 g TSP L-1). The 30 cm water column was aerated and incubated in a dark water bath at ~22C for 28 days.

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80 Water samples (30 mL) were withdrawn at t = 0.1, 0.5, 1, 1.5, 2, 4, 6, 10, 14, and 20 days, and analyzed for SRP and pH. Representative samples were also analyzed for TSP, dissolved calcium, total alkalinity, and dissolved organic carbon (DOC). Dissolved calcium was determined using flame atomic absoption spectroscopy (EPA 215.1; EPA 1979) on a Perkin-Elmer 3110. Alkalinity was titrated with 0.02N H2SO4 (EPA 310.1; EPA 1979). Dissolved organic carbon was on acidified, filtered (0.45 m) samples, and measured with a Shimadzu TOC5050A (Duisburg, Germany) TOC analyzer equipped with an ASI-5000A autosampler (5310-A; APHA, 1992). Sample pH was recorded immediately following collection, using a 3 in 1 gel filled combination pH electrode and Corning 313 pH meter. Water bath temperature was continuously recorded by a StowAway Tidbit logging probe (Onset Computer; Bourne, MA) as well as monitored periodically with a thermometer. After 30 days, an amendment of 100 g P L-1 (as KH2PO4) was added to each core. The water volumes above each co re differed slightly ( 5 mL) from the original water volume of 1.15 L added one month prior, likely due to different evaporation rates induced by the aerators. These differences were recorded but volumes were not adjusted at that time. Water samples (30 mL) were withdrawn t = 0, 4, 8, 24 and 53 hours after the amendment, and analyzed for SRP. Each core received 30 mL of P-unamended reflood water after sampling.

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81 Typha Litter Incubations Through a series of oxic and anoxic incubations, P uptake and release rates were determined for freshly submerged Typha litter and associated microbial biomass. Senescent submerged plant tissue from a stand of Typha ~30 m in diameter (25 50.275’ N; 80 42.991’ W) were collected in southern WCA-3A, and kept in site water on ice for transport to the lab. Directly south of WCA-2 and the EAA, west of urban Miami and north of Everglades National Park (Figure 3-1), WCA-3A is a mosaic of pristine Everglades ecotypes, from wet sloughs dominated by Nymphaea spp. and Utricularia spp. to sawgrass ( Cladium ) ridge and tree island communities. The slough surrounding the Typha stand was dominated by Nymphae oderata has historically low water TP and SRP levels of 9 2 g L-1 and <2 g L-1, respectively (Swift and Nicholas, 1987), and was the source of water for the incubations. Part I. P uptake by Typha litter Incubation water was filtered (0.45 m) and amended with KH2PO4 to concentrations of 0, 10, 30, 100 and 300 and 1000 g P L-1 above background SRP concentrations. This concentration range was representative of the P gradient to which litter is exposed: low-P surface water to high-P porewater. Short acrylic columns received 25 g litter (wet wt.) and 0.50 L P-amended water (Figure 3-4). Cores were sealed with rubber stoppers at each end. The top stopper was penetrated by two glass tubes which permitted sparging with air for oxic trials, or N2 + 0.03% CO2 for anoxic trials. Gas flow mixed the water column

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82 during sparge (0.5 3hrs), after which gas lines were clamped shut. Water samples were withdrawn after 0.3, 1, 2, 3, and 5 days of incubation, and analyzed for SRP, pH and DOC. Dissolved oxygen concentrations were monitored periodically during sampling, and measured for each replicate at the end of the incubation to ensure experimental conditions were achieved. Figure 3-4. Schematic of incubation design, containing Typha litter and Pamended surface waters from WCA 3A. Not to scale: Water volume was 0.5 L, head space was ~ 0.05L. Part II. P release from enriched litter After seven days of incubation, the incubation waters for each litter sample were replaced by un-amended site water. Water column SRP concentrations were measured immediately (t=0), and at 1, 2, 3, and 5 days after the water exchange. Water SRP concentrat ions were used to calculate oxic and Purge line out Sampling port Spargeline in Purge line out Sampling port Spargeline in

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83 anoxic P uptake and release rates for Typha litter across the range of initial water P concentrations. Part III. Microbial Biomass Phosphorus At the conclusion of the release phase, Typha litter subsamples of ~10 g wet weight were subjected to either 24-hour CHCl3 fumigation and 16 hour 0.5 M NaHCO3 (pH 8.5) extraction on a reciprocating shaker, followed by centrifugation (10 min, 6000 rpm), or direct extraction without fumigation. Additional P extracted after fumigation was considered the contribution of the microbial biomass phosphorus (MBP) pool that was present and susceptible to chloroform (Hedley and Stewart, 1982). Typha tissue samples were analyzed in their original condition (preserved field-moist in the dark at 3.5C) for comparison. Extracts were analyzed for SRP and TP using the ascorbic acidmolybdenum blue method (EPA 365.2; EPA 1979) on a Technicon Autoanalyzer. TP samples were prepared by persulfate digestion at 150C, increasing to 380C, prior to analysis. Statistical Methods Statistical analyses on experimental data were performed using MSExcel (v. 2000 Microsoft Corp.) ANOVA and t-test macros. Error around mean values is presented as one standard deviation of replicate measurements.

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84 Results and Discussion Mesocosm Phosphorus Storages The following narrative describes observ ed changes to vegetation in the mesocosms, though no quantitative measurements of biomass occurred prior to August 2001. In February of 1999, the surface of the water in rep 1 was covered in Lemna which likely reduced light penetration and gas exchange to the water column. Lemna persisted throughout 1999 in rep 1, with none noted in rep 2. Dense mats of floating Lemna minor were shown by Ngo (1987) to inhibit phytoplankton growth by shading the water column, causing algae to die and settle out. Water column chemistry can also be affected below dense mats of floating macrophytes. Low pH and oxygen levels can develop, and increased nutrient levels may result from internal loading (Gopal et al., 1984). Greater Najas biomass was observed in rep 2 than rep 1 in December 1999. Through the first year of operation, rep 2 appeared to outperform rep 1 with respect to water column P reduction (Figure 3-5). From December 29, 1998 through December 21, 1999, mean inflow TP concentration (106 g L-1) was reduced to mean outflow concentrations of 70 and 33 g L-1 for rep 1 and rep 2, respectively. Lemna may have created anoxic conditions and internal P loading from the initial muck substrate, while Najas in rep 2 may have provided oxygenated water column conditions which can minimize soil-P release (Gonsiorczyk et al., 2001). By September 2000, rep 1 appeared to have more Najas while rep 2 had

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85 Figure 3-5. Inflow and outflow TP conc entrations for mesocosms dominated by Typha and operated from December 1998 through August 2001. 0 20 40 60 80 100 120 140 160 180 200Jan 99 May 99 Sep 99 Jan 00 May 00 Sep 00 Jan 01 May 01 Sep 01MonthTotal P (g L-1) Cattail Rep 1 Cattail Rep 2Outflow (273) (240) 0 50 100 150 200 250 300 350Jan 99 May 99 Sep 99 Jan 00 May 00 Sep 00 Jan 01 May 01 Sep 01Total P (g L-1) Inflow

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86 developed an underwater periphyton mat. At the time of the whole-mesocosm biomass sampling in August 2001, no peri phyton was found in either mesocosm, and rep 1 contained more Najas (289 g dry wt.) than rep 2 (12.4 g dry wt.). Mean surface water TP concentration reductions were greater for the latter mesocosm, which had a more dense Typha canopy (rep 2), though variability in vegetation and P removal performance occurred throughout the period of record (Figure 3-5). It is not clear whether fluctuations in biomass and species composition of SAV, floating plants, and periphyton within and between the two cattail mesocosms contributed to variability in P removal performance. Based on mean inflow and outflow water TP concentrations, the two cattail mesocosms removed 1.2 (rep 1) and 2.0 (rep 2) g P m-2 yr-1 of an average annual loading rate of 3.5 g P m-2 yr-1. Phosphorus removal resulted in an increased biomass P pool and accrued soils. Of the total P mass recovered in biomass and new soils, 66 (rep 1) and 96% (rep 2) was attributed to TP removal from the water column. Unfortunately, the change in P in the initial muck substrate was not quantified for the pe riod of record. The difference between water column P mass removal and P recovered in biomass and new-soil P reflected experimental error inherent in sampling, but P mobilization from the initial muck substrate (Table 3-1) was al so likely. The substrate in one mesocosm, therefore, may have had only a minor net P contribution (0.2 g P m-2), while the other mesocosm substrate provided nearly a third (1.5 g P m-2) of the P mass accrued in biomass and new soil storages.

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87 Typha P concentrations for live tissues, dead leaves and root/rhizomes were comparable to those reported for the Everglades (e.g. Toth, 1988) (Table 3-2). Belowground cattail tissues (roots+rhizomes) comprised 39% and 65% of the total recovered biomass P in rep 1 and rep 2, respectively (Figure 3-6). Previous Everglades studies estimate 30-40% of total Typha biomass P was associated with belowground tissues in P-enriched areas, while 60% was belowground in low-P environments (Toth, 1988; Miao and Sklar, 1998). The disparity in the below-ground cattail P storage between the replicates in this study was echoed by the above-ground standing crop (Figure 3-6). Greater biomass P storage observed in replicate 2 agreed with greater reductions between inflow and outflow water TP concentrations over the 2.7-year study period. Table 3-1. Phosphorus removed from th e water column over the 2.7-year study period and recovered in the vegetation ( Typha + Najas ) and soils upon termination of the study on August 20, 2001. Rep 1 Rep 2 Average Biomass Storage, g P m-2 1.2 1.8 1.5 (29%) New Soil Storage, g P m-2 3.6 3.9 3.7 (71%) Total Recovered P, g P m-2 4.9 5.6 5.2 (100%) Inflow-Outflow Water Mass P Removed, g P m-2 3.2 5.4 4.3 Potential net P contribution from muck substrate into biomass and new soils, g P m-2 1.6 0.2 0.9

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88 Figure 3-6. A: Partitioning of recovered Najas and Typha tissue P within the inflow (In) and outflow (Out) halves of two mesocosms operated from December 1998 through August 2001. B: Total P mass recovered (solid line) in biomass and sediments rela tive to annual mass removal rates (dashed line) calculated from water column TP concentration reductions. 0.0 0.5 1.0 1.5 2.0 2.5Rep-1 InRep-1 OutRep-2 InRep-2 OutMass recovered (g P m-2 yr-1) New Soils Biomass 0.0 0.1 0.2 0.3 0.4 0.5 0.6 0.7 0.8 0.9 1.0Rep-1 InRep-1 OutRep-2 InRep-2 OutFraction of Biomass P Recovered Typha roots Typha live shoots Typha dead shoots Najas A B

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89 Table 3-2 Dry matter and P concentrations in accrued soil and tissue storages (live and dead Typha shoots, below-ground Typha (roots and rhizomes), and Najas tissues) retrieved from mesocosms on August 20, 2001. Typha tissue P concentrations are co mparable to values reported by Toth (1988). Live Typha shoots Dead Typha shoots Belowground Typha All Najas Tissues Newly accrued soil (g dry m-2) Rep 1 416 1163 745 41 3932 Rep 2 466 2141 1616 1.8 5426 (mg P kg-1) Rep 1 790 180 620 5300 924 Rep 2 710 190 630 3300 747 Toth 1988 580-1000 160-260 570-740 High biomass in rep 2 caused the belowground portions of the Typha stand to become space limited, and root structures extended upwards into the water column. These structures may have provided a pathway for direct P uptake from the water column, more so in rep 2 than in the less dense rep 1, since greater P reduction was observed over the 2.7-year period of record in the mesocosm with greater cattail density and less SAV biomass. Root and rhizome biomass in rep 2 (1.6 kg m-2) was more than double that of rep 1 (0.75 kg m-2), but root TP concentrations were similar (Table 3-2). Additionally, rep 2 had twice the dead Typha biomass (2.1 kg m-2) that was present in rep 1 (1.2 kg m-2). Dead biomass in rep 2 may have provided substrate for a larger microbial community than rep 1, and increased P removal from the water column in that mesocosm. Live above-ground Typha biomass in this study was similar in the two mesocosms (0.42 and 0.47 kg m-2, for rep 1 and rep 2

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90 respectively), and with similar TP contents (Table 3-2), the two had similar P mass stored in that compartment. Substantial fine-grained organic soil (not fibrous cattail detritus) was produced in the 2.7-year study period. Total P content of the accrued soils was more uniform within rep 1 (inflow, 922 mg kg-1; and outflow, 926 mg kg-1) than rep -2 (inflow, 640 mg kg-1; outflow, 854 mg kg-1). When P associated with the dead Typha leaves and accrued soil is compared to P in live plant biomass ( Najas + live Typha leaves, roots), nearly four-fold greater P mass was sequestered into “dead storage” (4.05 g P m-2) than into “live storage” (1.19 g P m-2). If it is assumed that P found in the soils which accrued during the period of record resulted from the live biomass (SAV tissues, epiphytes, and live Typha leaves), years) POR, ( P) g Biomass, (Live P) g (Soil, yr P, Biomass Live for rate er Net turnov1 then a turn-over rate of 4 yr-1 for the live macrophyte P storage is approximated. This calculation is subject to several errors (e.g. (it ignores direct sedimentation and P adsorption by the soil matrix), but approximates the rate of P movement from biomass to soil storages. In a full-scale wetland, mass balance of P into biomass would be more complex. Non-uniform soil accretion can occur when a high flow event scours the soil or introduces suspended solids. Macrophyte tissue senescence occurs inconsistently throughout the year. Cattails in temperate wetlands may senesce

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91 annually (Kadlec and Knight, 1996), but the mild subtropical climate of South Florida does not necessarily cause Typha die-back in the winter months. Tissue Desiccation Study Emergent plant tissue has high proportions of structural (lignin and cellulose) tissues in order to support the aboveground portions of the plant (Debusk and Reddy 1998). Submerged vegetation, in contrast, relies on buoyancy within the water column for vertical suppo rt. Lignin and cellulose components of macrophyte tissues are slowly decomposed in wetlands (Godshalk and Wetzel, 1978), which results in slower decomposition of Typha tissues than SAV tissues. Typha tissues persisted in litter bags for several years in WCA 2A (Davis, 1991), with approximately 50% weight remaining after two years. In contrast, Dierberg (1993) found only 9-14% of the original mass of the submerged macrophytes Hydrilla verticillata Royle and Vallisneria americana Michx. remained after three weeks in the Lake Okeechobee littoral zone following a decline in lake level. Tissues in that study, however, desiccated when exposed to the air, and decomposition in the water column may be less rapid or incomplete, allowing the SAV fragment to regrow. Upon senescence, macrophyte tissue (litter) undergoes a series of P exchanges before incorporated into the so il. Submerged surfaces become sites for epiphyte attachment. Microbial “attack” on the leached litterfall acts to enrich litter P content, as organisms sequester soluble P from the surrounding water to maintain suitable ratios of C, N, and P. As the available carbon and P sources are

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92 depleted, recalcitrant organic-P is partially mineralized by extracellular enzyme production in the low-P environment. Over time (several months to years) an apparent increase in litter P content may also be due to losses of other elements (e.g., carbon to CO2 through respiration, N through denitrification) relative to P within decomposing litter (Davis, 1991). In the case of Typha where senescent tissue was used for incubation, labile ph osphorus would have been translocated into the plant base during senescence or was rapidly leached into the surrounding water prior to collection (Christiansen et al., 1985, Davis, 1991). Initial floodwaters of the desiccation study were low in bioavailable SRP (3 g L-1), and contained 24 g TP L-1 (Table 3-3). Soluble reactive phosphorus in the floodwater in the Najas treatment increased markedly above initial levels over the nine-day incubation, but not in either the Typha incubations or the control waters (Figure 3-7). Because of the variability in Najas SRP release, however, differences in SRP between treatments were not significant (p > 0.05) after either 1 or 9 days. DOP concentrations in floodwater from the Najas Table 3-3. Mean ( 1 s.d.) water quality parameter values measured in the Cell 4 outflow water used in the preliminary desiccation experiment. Temperature pH TP SRP DOP PP C g L-1 g L-1 g L-1 g L-1 28.0 8.73 24 3 11 10 treatment increased significantly (p < 0.05) above control water concentrations after both 1 and 9 days, from 11 to 29 g L-1 over the nine days, while floodwater

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93 DOP from the Typha treatment (17 g L-1) was not significantly higher than control concentrations (14 g L-1) or significantly different from Najas treatment (p > 0.05). A B 0 20 40 60 80 100 0246810DOP (g L-1) Najas Typha Control 0 20 40 60 80 100 0246810 Days of Dark IncubationSRP (g L-1) Najas Typha Control Figure 3-7. A: Dissolved organic phosph orus (DOP) and B: soluble reactive phosphorus (SRP) concentrations in floodwaters containing fresh Najas tissues and Typha litter, compared to control waters containing no plant tissues, during a nine day dark incubation. Error bars indicate one standard deviation from trip licate (experimental) or duplicate (control) cores. Dierberg (1993) reported that the ma jority of tissue-P was rapidly (days) released from desiccating SAV exposed by lake level draw-down. The released P was largely bioavailable. In this study, the soluble P released from SAV over nine

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94 days of dark conditions was 77% SRP. Vascular aquatic plants translocate some nutrients, including P, from senescing older tissues to new leaves to avoid such losses (Christiansen et al., 1985). Najas tissue became chlorotic and fragmented during the nine-day incubation, while Typha litter remained intact. Light appeared to be important in maintaining P content in SAV, and the plant may be susceptible to soluble P release during short-duration turbid even ts which inhibit light penetration into the water column. Phosphorus stored in submerged dead Typha tissues was not sensitive to dark conditions, as the tissues did not release SRP or DOP to the water column. Since the Najas community produces little discernible “leaf litter,” and fresh Najas tissue rapidly senesced in dark lab conditions, only the effects of Typha litter on P flux were explored with intact soil cores. Flux Study using Intact Soil Cores Soils collected from beneath the Typha community in STA-1W Cell 1 outflow region released P into overlying water under dark, aerobic conditions (Figure 3-8). Initial floodwater was particle free (filtered 0.45 m), and contained 13 g DOP L-1 and 3 g SRP L-1. Of the three cattail soil cores without litter amendments, only one replicate had visibl e fragments of leaf litter within the surface soils. The results observed from this core were more similar to the litter amended treatments than the replicate treatments without litter, and therefore were excluded from this discussion.

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95 0 10 20 30 40 50 60 70 80 90 05101520 Elapsed Time (days)Water Column SRP (g L-1) Typha Soil Only Typha Soils with Litter Litter Only Water Control Figure 3-8. Soluble reactive phosphorus concentrations in the water columns of Typharegion intact soil cores with and without Typha litter amendments during 20-day dark laboratory incubation. Litter and water control cores contained no so il. Error bars for each treatment indicate one standard deviat ion between triplicate cores. Typha soils with litter, litter only, and water control treatment SRP concentrations were each near the detection limit of 2 g L-1 for 28 days. Typha soil treatment values are from duplicate cores. Between 7 and 20 days of incubation, SRP concentrations were significantly higher (p < 0.05) in the remaining two replicate waters overlying Typha soils without litter amendments (mean at 20 days= 63 g L-1; n=2), than in litter only and water control treatments (mean at 20 days = 2 g L-1; n=6). After 10 days, SRP concentrations in the water columns above soil-only treatments were significantly higher (15 g L-1) than those with litter plus soil (4 g L-1) (p< 0.05). Typha litter amendments (25g wet wt.) had a marked effect on SRP

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96 concentration in the floodwater during the 20-day incubation. Concentrations never exceeded 4 g SRP L-1 in any replicate without soils (litter only and water control), indicating that there was no net P flux out of the litter amendments. When leaf litter was added to cattail soils that alone had shown a positive flux into the water column, the flux was abse nt. Instead, SRP concentrations in the overlying water were maintained throughout the 20-day incubation (4 1 g L-1), or as low as the control treatments. Initial flood water pH (8.06) increased slightly to 8.36 0.06 before the 8hour sampling, and was similar among all treatments (8.41 0.09) during the remaining 20 days. Water bath temperature averaged 24 3 C during the incubation period. Water column DOP concentrations for the two Typha soil treatments after 20 days (20 and 35 g L-1) were higher than the Typha soil + litter, litter only and incubation water control, which after 20 days were 15 1, 15 2, and 14 3 g L1, respectively, though differences were no t significant (p > 0.05) (Figure 3-9). All treatment and control water DOP concentrations were not significantly different from one another or different over time during the 20-day incubation (p > 0.05).These results agree with the results of the 9-day tissue desiccation experiment: Typha litter does not release DOP under dark aerobic conditions.

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97 Figure 3-9. Mean dissolved organic phos phorus concentrations in the water columns of four treatments (see text for details) during 28-day dark laboratory incubation. Error bars indi cate 1 standard deviation from the mean of three replicates. Mean calcium values for the soil + litter and litter only treatments were significantly higher (123 mg L-1) than those of the treatments without litter (104 mg L-1) after 0.1 days of incubation (p < 0.05). After 0.5 days, however, no significant differences were observed in dissolved calcium concentrations between any treatment. Calcium concentrations were stable after that initial equilibration period (Figure 3-10). From 7 to 20 days of incubation, water column Ca concentrations in Typha soil + litter treatments (85 5 mg L-1) were 0 10 20 30 40 0.11220Elapsed Time (Days)Water Column DOP (g L-1) Typha soil Typha soil + Litter Litter Control Water Control

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98 lower than litter only and water control treatments, but not significantly different. Figure 3-10. Mean dissolved calcium and alkalinity concentrations in the water columns of four treatments (see text for details) during a 28-day dark laboratory incubation. Error bars indicate 1 standard deviation of the mean from three replicates. 200 220 240 260 280 300 32005101520 Elapsed Time (days)Alkalinity (mg CaCO 3 L-1) 60 80 100 120 140 160Dissolved Ca (mg L-1) Typha soil Typha soil + Litter Litter Control Water control

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99 Alkalinity concentrations decreased significantly (p < 0.05) in litter control treatments from initial floodwater concentrations of 271 6 mg L-1 to 258 6 mg CaCO3 L-1 after 14 days (Figure 3-10). After 14 days, the litter only treatment was lower in alkalinity (258 6 mg CaCO3 L-1) than the Typha soil-only treatment (286 3 mg CaCO3 L-1). All other treatments were constant with respect to alkalinity concentrations. Dissolved organic carbon (DOC) concentrations were significantly higher after 4, 7, 10, and 14 days in treatments containing litter amendments (mean s.d. for all time intervals, 46 1.9 mg L-1), relative to soil-only and water-only control treatments (42 1.3 mg L-1) (p > 0.05) (Figure 3-11). Release of DOC from Typha litter may explain the lack of increased water column P, as microbial P uptake in the soil-only treatment could have been carbon-limited. The method of DOC quantification used here did not describe the quality of carbon for microbial utilization, however there may be additional differences in the carbon quality between treatments that influenced P removal from the water column (Koshmanesh et al., 1999). Rapid biological P uptake during the half-hour following the P amendment of 100 g L-1 may explain an observed increase in SRP concentrations (94 g L-1 for all columns, 88 g L-1 for the control columns) of less than 100 g L-1 (Figure 3-12). The water-only control columns showed a small decrease in concentration, then leveled off near 60 g L-1. Phosphorus in the control waters may have been absorbed by bacteria in the water column or

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100 attached to core walls, aerator, etc., and converted to non-reactive phosphorus. The change in the control water P co ncentration was assumed to be an experimental effect and was subtracted from the observed concentration for each treatment. Figure 3-11. Dissolved organic carbon concentrations in water columns of four treatments (see text for details) af ter 4, 7, 10 and 14 days of a dark laboratory incubation. Error bars indi cate 1 standard deviation from the mean of three replicates. The Typha soil cores provided no significant P uptake capacity when compared with the water-only control cores (p > 0.05), while litter-amended treatments decreased the SRP concentrations to 3-11 g L-1 in 53 hours. Eight hours after the amendment, concentratio ns in the litter-only treatment were 36 38 40 42 44 46 48 50 52 Typha soilTypha soil + LitterLitter ControlWater ControlDOC (mg L-1) Day 4 Day 7 Day 10 Day 14

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101 significantly lower than the other treatments, including the Typha soil + litter treatments (Figure 3-12). In the eight hours following the P-amendment, litteronly columns exhibited the greatest uptake rate of any of the controls or treatments (27.3 2.3 mg m-2 day-1), followed by the litter-amended cattail soils (22.5 2.4 mg m-2 day-1) (Table 3-4). The significant difference (p < 0.05) between P uptake rates of either litter treatment and the soil-only treatment is likely due to continued P flux out of the soil into the water column, as observed during the initial 20-day release period. Soil N release into the water column increased the TN content of the litter over the 32-day experiment, from 0.71 0.04 to 1.22 0.11%, while the TN content of litter without soil decreased slightly to 0.66 0.11%. Total C content of the litter was similar before (45.5 1.1 %) and after incubation (44.5 to 46.6 %) in both litter treatments. There was likely some conversion of Typha tissue biomass into microbial biomass carbon. In decomposition studies on Typha biomass, half of the live tissue carbon may remain after a two year period of decay (Davis, 1991). Thus the litter carbon pool may be stable relative to the nutrient pools. Phosphorus Uptake by Litter Microbial Populations After sampling events at T = 1, 2 and 3 days, three anoxic flasks and three oxic flasks were monitored for D.O. concentration. Short daily sparge periods (0.5 to 3 hr) during the 5-day P uptake monitoring period with N2 + 0.03 % CO2 (anoxic) or air (oxic) followed by stagnant incubation conditions resulted in low D.O. concentrations in anoxic treatments (0.59 0.13 mg O2 L-1) relative to

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102 oxic treatments (4.63 0.86 mg O2 L-1). After 5 days, D.O. measured in all flasks was still significantly higher (p < 0.05) for oxic treatments (2.92 1.00 mg O2 L-1) than for anoxic treatments (0.21 0.13 mg O2 L-1), even though no flask was continuously purged. 0 30 60 90 120 061218243036424854Time (hr)Water column SRP (g L-1) Typha Soil Only Water Control Typha Soil and Litter Litter Only Figure 3-12. Water column SRP conc entrations above a “background” concentration as determined before a 100 g spike was added to each column. Background concentration was assumed stable through the 24-hour period. Triplicate water columns were assembled with or without soils from a cattail stand in STA-1W Cell 1, and aerobically incubated for 30 days prior to the P-spike amendment.

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103 Table 3-4. Flux estimates from intact Typha -region soil cores and submerged Typha litter kept under dark conditio ns for 35 days. The 30 cm water column was continuously aerated. A 100 g L-1 phosphorus spike was added after 30 days to measure shortterm uptake rates. All values are means ( st. dev.) of triplicate cores minus mean flux in control cores, in mg P m-2 day-1. Typha soil Typha soil Typha litter Typha litter Period of flux estimate (mg P m-2 day-1) Release 7 – 20 days incubation 1.20 0.27 0.00 0.05 -0.01 0.01 Uptake 0 – 8 hours after spike 0.17 2.7 22.5 2.4 27.3 2.3 Rapid P uptake occurred under oxic conditions, reducing water SRP concentrations from 1000 g L-1 initial concentration to 7 3 g SRP L-1 (Figure 3-13). The maximum uptake rate (56 g P g dry matter -1 d-1) in the 1000 g L-1 oxic treatment was observed prior to the first 8 hour sampling. Based on field estimates of litter accumulations in STA-1W Cell 1 of 0.93 kg dry matter m-2, the equivalent areal uptake rate would be 52 mg m-2 d-1, nearly twice the rate observed in the intact core incubation P uptake period (Table 3-4). This potential rate may not be sustainable over the long term, however short-term uptake at that rate appears possible if another element is not limiting, and if adequate P was loaded to the system. For treatments that began the 5-day P uptake incubation at lower concentrations (100, 30, 10 and 0 g P L-1 amendments), SRP concentrations were reduced to below the analytical detection limit of 2 g L-1

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104 (Figure 3-13) in as little as 8 hours. Because of lower P mass added, uptake rates were slower in those treatments. 1 10 100 0246 Elasped Time (Days)SRP (g L-1) 100 Anaerobic 100 Aerobic 30 Anaerobic 30 AerobicUptake 1 10 100 1000 0246 Elasped Time (Days)SRP (g L-1) 1000 Anaerobic 1000 Aerobic 300 Anaerobic 300 AerobicUptake Figure 3-13. Mean soluble reactive phosph orus (SRP) concentrations in triplicate 0.5 L flasks, containing Typha litter incubated in the dark under oxic or anoxic conditions for 5 days. In itial concentrations of 1000, 300, 100 and 30 g L-1 above background SRP level (< 2 g L-1) were achieved with KH2PO4 amendments at Time = 0 da ys. Note SRP concentrations are shown on a logarithmic scale. When waters with initial P concentrations of 1000 or 300 g L-1 above background were sparged with N2 + CO2 gas, P uptake was slower and less

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105 complete than in the corresponding oxic treatments. Final SRP concentrations for the 1000 and 300 anoxic treatments were 580 and 63 g P L-1, representing reductions of 42 and 79%, respectively. Anoxic P uptake by litter in initial P concentrations lower than 300 g L-1 was no different than oxic P uptake, and SRP concentrations were maintained below 2 g L-1 after the 8 hour sampling. Phosphorus-amended waters were replaced with fresh unamended water (<2 g SRP L-1) after seven days of incubation. Concentrations of SRP measured immediately after the exchange were below detection for all treatments, with the exception of the 1000 g L-1 anoxic treatments (3-4 g L-1) (Figure 3-14). The minor amount of P detected was possibly residual from the wetted flask and litter surfaces, which were drained of Pamended water, but not rinsed before new low-P water was added. In the same treatment (1000 g L-1 anoxic), SRP concentrations reached 23 6 g L-1 24 hours after water exchange. Concentrations remained similar over the next five days (19 6 g L-1 ) in each of the replicate samples. Phosphorus concentrations also increased in the 300 g L-1 anoxic treatment, rising steadily over the 5-day release monitoring period. Anoxic treatments initially incubated with 100 or less g P L-1 showed no detectable P release over five days of incubation in un-amended waters. While the oxic litter treatments exhibited greater P uptake than the comparable anoxic treatments, no P release was observed from any oxic treatment. Even the 1000 g L-1 oxic Typha

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106 litter maintained water column SRP concentrations at 2 g L-1 for five days in low-P reflood waters (Figure 3-14). At the end of the entire 12-day incubation period (uptake followed by release), Typha litter showed an increase in microbial biomass P (MBP) in the 1000 and 300 g L-1 oxic treatments (104.4 29.2 and 65.8 11.1 mg P kg dry litter-1, respectively), relative to control (40.0 5.9 mg P kg-1) tissues preserved in the dark at 3.5 C during the study (Figure 3-15). 0 5 10 15 20 25 30 35 0123456 Elasped Time (Days)SRP (g/L) 1000 Anaerobic 300 Anaerobic 1000 Aerobic 300 AerobicRelease Figure 3-14. Mean soluble reactive phosph orus (SRP) concentrations in triplicate 0.5 L flasks, containing Typha litter incubated in the dark under oxic or anoxic conditions. After incubation under initial concentrations of 1000 or 300 g L-1 above background SRP level (< 2 g L-1), water was replaced with fresh low-P water (< 2 g SRP L-1) at Time = 0 days. Typha litter exposed to lower initial water P concentrations under aerobic conditions showed similar MBP levels to control litter. Microbial biomass P levels for the anoxic Typha incubations were at or below levels measured in the control samples. While small decreases in MBP were observed in the 30 and 1000

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107 anoxic litter treatments (27.4 7.8 and 27.6 4.5 mg P kg-1 dry litter) relative to control tissues, differences in MBP for all anoxic tissue incubations did not appear to depend on initial water P concentration (Figure 3-15). y = 0.06x + 9.56 R2 = 0.89 -20 0 20 40 60 80 100 1101001000 Initial Water P Conc. (g P L-1) MBP (mg P kg-1) Oxic -20 0 20 40 60 80 100 1101001000 Initial Water P Conc. (g P L-1) MBP (mg P kg-1) Anoxic Figure 3-15. Change in microbia l biomass phosphorus (MBP) of Typha litter incubated for seven days at initia l water concentrations of 0, 10, 30, 100, 300 and 1000 g L-1 SRP, under oxic (top) and anoxic (above) conditions.

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108 Koshmanesh and others (1999) showed the importance of carbon substrate quality in the uptake and retention of P by soil microbes in an Australian constructed wetland receiving ADW. Bioavailable carbon additions (glucose) to the soil increased P uptake by 35-45%, and was attributed to increased microbial uptake and growth. In that study, sequestered P was retained under anoxic conditions. Acetate amendments increased P uptake to rates comparable to those from glucose addition, but the acetate-induced “luxury P uptake” as polyphosphate was subject to release again during anoxic conditions. Utilization of acetate as a carbon source was likely less efficient than glucose, and therefore P incorporation into cell tissue was lower for the same incubation period (Koshmanesh et al., 1999). Implications for STA management Reduced efficiency of microbial P uptake and retention under anoxic conditions has important implications for both STA management and remediation of the nutrient-impacted WC A-2A marsh. Several studies implicate microbial P acquisition as the primary mechanism of P immobilization in wetlands (Brix, 1997; Grimshaw et al., 1997), especially those with soils of low sorption capacity (Qualls and Richardson, 2000). Typha domingensis in outdoor enclosures retained P within the litter layer through microbial uptake processes (Richardson and Marshall, 1986; Richardson, 1999), and P content increased up to ten-fold above live tissue concentration during long-term (2 yr) nutrient enrichment studies.

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109 Long-term nutrient loading has resulted in the continued expansion of a P-enriched, low-oxygen zone extending southward into WCA-2A from ADW discharges (Belanger et al., 1989; Reddy et al., 1993; Reddy et al., 1998). Similar conditions may develop within Typha -dominated regions of STA-1W Cell 1 inflow. Surface water TP reductions in WCA-2A are gradual with distance from the inflow, until a point where a mixtur e of sawgrass and open-water sloughs compete with encroaching Typha (Reddy et al., 1998). High soil-P levels in the WCA-2A impacted area adjacent to inflow structures may represent deposition that occurred before Typha colonized to the density observed in recent surveys. Phosphorus remains mobile in the impa cted region because the water column lacks sufficient oxygen supply to support rapid microbial sequestration. In contrast, oxygen supply from slough areas adjacent to the low-oxygen region allows rapid microbial P uptake along th e transition zone between slough and cattail-dominated areas. Total P concentration reduction rates with distance increase dramatically through the transi tion region of WCA-2A (Reddy et al., 1998). An analysis of the soil sorption capacity along the WCA-2A eutrophication gradient has shown that soil sorption capacity was not saturated even in the enriched areas, yet elevated SRP concentrations persist through that region (Richardson and Vaithiyanathan, 1995). This suggests that the water column may not be in contact with the sorption sites, possibly due to the litter accumulations or macrophyte P uptake depleting exchangeable soil-P storages.

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110 More important are the uptake mechanisms of microbial biomass within the litter layer and in contact with the water column. Due to O2 limitation, however, this mechanism is inefficient in dense stan ds. One possibility for controlling oxygen levels may be periodic drawdown. The drawdown would consolidate and oxidize newly accreted soil material. A subsequent reflood may allow SAV and algal communities to colonize and oxygenate the water column, and microbial communities to rapidly sequester released nutrients. Oxygen supply to the water column is not expected to significantly increase mobilization of soil-P through decomposition as long as sufficient dissolved carbon is available as an energy source for metabolism. Conclusions The majority of P removed from the water column by Typha mesocosm communities was incorporated into newl y accreted soils (78%) during 2.7 years, though the new soil did not appear to be of cattail origin. Typha incorporated 39 and 65% of total mesocosm biomass P to belowground tissue. Healthy Najas tissues became chlorotic after nine days in the dark, and may be a source of SRP or DOP to the water column during events of high turbidity. In contrast, Typha litter material was relatively stable in dark conditions. While Typha biomass appears to persist as leaf litter and detrit us to a greater extent than SAV tissues, the extensive root system has the potential to hinder long-term storage by mobilizing P from enriched soils.

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111 Typha leaf litter at the soil surface may limit diffusive flux of soil-P into the water column if Typha stand density is low and the water column is oxygenated by aquatic autotrophs. Similarly in treatment wetlands, the litter layer can become P-enriched from extern al loading to the water column through microbial and algal uptake mechanisms. However, at higher stand densities, P removal from surface water is limited to soil sorption and microbial uptake mechanisms because primary productivi ty and oxygen supply is light-limited. When oxygen is depleted in the high ly enriched regions near ADW discharge structures, P uptake by litter-associated microbial biomass is likely oxygen limited. Treatment wetland managers may need to monitor oxygen levels in inflow regions with dense Typha stands, and provide dissolved oxygen through SAV aquatic photosynthesis or other means to ensure low outflow water TP concentrations. Further research will likely identify the long-term sustainability of P uptake processes by leaf litter, and conditions under which the sequestered P is re-mobilized into surface waters.

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112 CHAPTER 4 SYNTHESIS Constructed treatment wetlands are useful tools for removing phosphorus (P) from agricultural drainage waters (ADW). Besides being a cost-effective treatment alternative with secondary benefits (e.g., wildlife values, aesthetics), wetland communities typically reduce P concentrations in surface waters between inlet and outflow. However, characteristics of submerged aquatic vegetation (SAV) and emergent wetland communities, including water column shading, oxygen supply, and community metabolism, can influence treatment wetland P removal performance. Emergent and submerged vegetation communities each remove P from the water column, though the dominant processes are different. There is significant debate in the literature over the uptake of P through SAV roots vs. shoots (Denny, 1972; Bole and Allen, 1978; Hill, 1979; Carignan and Kalff 1980; Barko and Smart, 1980,1981; Carignan, 1982; Carpenter, 1983; Moeller et al., 1988; Rattray et al., 1991; Gumbricht, 1993; Stephen et al., 1997; Kufel and Kufel, 2002). Many of these studies, including those using radiotracers, maintained SAV tissue cultures in P-free overlying water and concluded sediments can act as a source for nearly all P taken up by rooted macrophytes. However, fragments of rooting macrophytes, as well as rootless species and macroalgae (e.g., charophytes) can

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113 grow well in soil-free cultures of agricultural drainage waters (Pietro, 1998, Kufel and Kufel, 2002, and personal observation). It is generally agreed that the source of P to Typha spp. is through root uptake (Davis, 1991; Newman et al., 1996; Lorenzen et al., 2001), and below-ground Typha biomass can potentially deplete soil P. In a review on the effects of SAV on nutrient dynamics during sediment deposition and resuspension, Barko and James (1998) concluded that SAV beds efficiently trap suspended matter from the water column by lowering water velocities and reducing resuspension. The SAV community itself also can quickly produce appreciable sediment. Mesocosms vegetated with SAV accumulated deep organic sediment layers over three years of operations, at rates of up to 3.1 cm yr-1 under high hydraulic loads (DBEL 2001). It is possible that deposition of P-laden organic matter resulted in P-en richment of soils beneath submerged vegetation observed in this study. The Typha community likely relies on slow flows and rapid microbial uptake mechanisms to retain P “on-site.” Emergent biomass-P is subject to rapid leaching from senescent tissues and can be transported downstream. To counteract such losses, emergent litter accumulation appears to provide some capacity for assimilating and retaining P. Typha leaf litter can provide substrata for microbial growth and P uptake, where incorporation of inorganic P into organic P pools during soil formation delays recycling into a bioavailable inorganic form. Without continuous additions of new leaf litter, bacterial films

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114 and algae become space-limited on subst rata, and overextended colonies slough particulates into the water column under high flows. Phosphorus uptake by the litter layer may also function as a driver of diffusive flux from the soil. Litter/microbial P uptake mechanisms were able to maintain low water column SRP concentrations, which led to strong diffusion gradients and higher potential diffusive P flux in the emergent macrophyte stand, relative to SAV. This flux may not impact water column P concentrations if the nutrient is incorporated into the litter layer at the soil surface, but would mobilize P from long-term storage in the accrued soils. Water column P removal is thus aided by the biological P cycling that occurs primarily in association with submerged surfaces. Humification of organic detritus and leaf litter results in lower P bioavailability than in the original tissue and is an important proce ss in soil development. It is within the soil matrix that physicochemically stable P forms must be sequestered over the long term for wetland water treatment to be successful. Irradiance and Water Column Shading Solar UV irradiance of the water column is moderated by turbidity (inorganic particles, planktonic algal cells, etc.), color from dissolved organic matter (DOM), and canopy shading (Wetzel et al. 1995). In the absence of these controlling factors, UV–B irradiance (280-320 nm) can reach depths of 5 m or more. During periods of high flows, low-density organic matter can become suspended in the water column, increasing turbidity and decreasing light

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115 penetration. After short periods under aphotic conditions, SAV tissues can atrophy and release soluble reactive P (SRP) and dissolved organic P (DOP). Meanwhile, dead EAV tissues can persist in the dark without releasing P into surrounding waters. Dissolved organic compounds are protected from photolysis by the emergent canopy. Rose and Crumpton (1996) found standing live and dead Typha shoots, leaf litter, and ephemeral floating plant populations combined to reduce light levels, measured at 5 cm below the water surface, to 2% of ambient levels. In contrast, less dense emergent macrophyte stands allow UV penetration into the water column for photolysis of organic compounds into bioavailable SRP. Submerged macrophytes can similarly reduce light penetration, photosynthesis and dissolved oxygen generation in waters below the leaf canopy. Depending on current water depth and recent depth changes, the proportion of the water column thus isolated from the euphotic zone can be variable. Wetzel et al. (1995) examined with laboratory experiments the UV-B photolysis of DOM compounds derived from Typha latifolia and Juncus effusus. Total DOM concentration was found to change little under natural irradiance levels. However, a conversion of recalcitrant humic substances to labile compounds did occur, which increased the bioavailability of those organic compounds to further microbial degradation.

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116 In the absence of dense vegetation, shallow wetlands (<2m deep) are sometimes entirely euphotic, with bent hic algae capturing incident light and reflecting the excess. Shading of incident radiation by an emergent Typha canopy was observed to reduce the net primary productivity of periphyton within the stand by 80%, relative to adjacent unshaded communities (Grimshaw et al., 1997). Irradiance can control the distribution of algal species because of differences in optimum light levels between species. In oligotrophic waters, planktonic bacteria and algae are subject to short-term negative effects of high ultra-violet (UV) irradiance (Karentz et al., 1994), but are otherwise stimulated by increased light availability. Bunn et al. (1999) estimate that diatoms are in optimal light regime where shading by ri parian canopy exceeds 80% (< 9 mol m-2 day-1), while the filamentous alga Cladophora requires less than 50% shading (26 mol m-2 day-1) (Graham et al., 1995). Submerged macrophytes are typically able to grow in full sunlight. Some researchers have reported photoinhibition in marine macrophytes within intertidal areas (King and Schramm, 1976), when falling water levels expose macrophytes to increasing light. The SAV canopy can also preclude light penetration to the soil surface if dense be ds develop under static flow conditions and static or decreasing water depths. Under dynamic water levels and flow conditions, however, the submerged macrophyte bed is continuously reshaped and increases light penetration.

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117 In the northern Everglades mixed-marsh community of sawgrass prairies and wet sloughs, periphyton dominates productivity in the sloughs, while the sawgrass community dominates marginally higher land elevations. Emergent vegetation is typically found in more shallow waters than SAV sloughs, and as a result, soil and detritus may be exposed more frequently. In a shallow water column the oxygen demands of the soil, detritus and biota are concentrated in a smaller water volume than under deep-water conditions, and the water column can become anoxic. Oxygen Supply In the continuously flooded environment of a treatment wetland, oxygen supply comes primarily from aquatic photosynthesis and secondarily from pressure-induced oxygen transport through macrophyte aerenchyma tissues, or physical (e.g. wind-induced) mixing. Oxygen demand is high in the flooded wetland environment. Biological oxygen demand by heterotrophic organisms and respiring plant cells, and chemical oxygen demand of reduced soil compounds, all deplete dissolved oxygen (DO) levels. Water circulation is important at the site of oxygen consumption. Boundary layers develop along soil and detritus surfaces which limit oxygen supply to consumers and decomposers even in aerated water. Jorgensen and Revsbech (1985) described boundary layers adjacent to coastal sediments and detritus as 0.2 to > 1 mm in thickness, depending on flow velocity and surface roughness. Diffusion times for O2 passing through the layer were 1.2 to 9 min.

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118 Aquatic photosynthesis, meanwhile, increases DO concentrations and potential for metabolic activity (Karjalainen et al., 2001). Atmospheric oxygen diffuses into still waters at slow rates (0.72 g O2 m-2 day-1) relative to potential oxygen production rates via aquatic photosynthesis (Odum, 1956). Windinduced mixing of surface waters can in crease diffusion rates, though emergent macrophyte stands reduce wind velocities near the water surface. Aquatic photosynthesis by algae, cyanobacteria, and macrophytes can be extremely high. Submerged macrophytes in Florida spring runs can generate as much as 64 g O2 m-2 day-1. This rate is among the highest reported for any system (Duarte and Canfield, 1990). The high rates of productivity observed in Everglades cyanobacterial mats (wet season, 3 to 23 g O2 m-2 day-1) (McCormick and Laing, 2003) are possible only through intense recycle pathways of nutrients, carbon and gases (Wetzel, 1993). Rooted emergent macrophytes (Jespersen et al., 1998) and submerged aquatic plants (Karjalainen et al., 2001) both can transfer oxygen from shoots to roots through aerenchyma tissues, at rates which sometimes exceed that required for root respiration. Excess oxygen is lost to the surrounding soil environment. The microenvironment around root hairs, or the rhizosphere, can experience increased redox potential and dissolved oxygen levels, relative to the bulk soil (Wium-Anderson and Anderson, 1972; Flessa 1994; Wigand et al. 1997). The effects of an oxidized rhizosphere on soil processes include coupled nitrification/denitrification reactions, increased microbial metabolism (including

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119 the activation of aerobic extracellular enzymes), and concentration of Fe along the oxic-anoxic interface. The importance of oxic surface sedi ments in maintaining oligotrophic conditions was described for Lake Stechlin (Germany) by Gonsiorczyk et al. (2001). In that study, oxidized Fe hy droxides within the surface sediments provided P sorption sites and microbial activity acted to minimize P release from sediment porewater. Similarly in this st udy, leaf litter and detritus surfaces above wetland soils minimized P loading to the water column, but was sensitive to oxygen supply. Due to the low Fe concentrations in Everglades soils, Fe hydroxide formation may not be as important to maintaining low water column P concentrations as for lakes with mineral sediments. Community Metabolism Wetland community metabolism has a fundamental influence on P cycling and stability in treatment wetlands. Light penetration into the water column provides energy for aquatic photosynthesis and organic matter provides energy to heterotrophic organisms (Odum 1956). In shaded areas, heterotrophic respiration exceeds autotrophic photosynthesis, and the community metabolism is said to be negative. Negative metabolism results in an absence of diurnal increases in dissolved oxygen and pH levels, as well as reduced rates of CaCO3 precipitation. Limited available P in the freshwater environment drives intense competition for this essential nutrient between all life forms. Through biological

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120 uptake, chemical sorption and precipitation, and physical settling processes, P is concentrated on solid surfaces, and dilute d in the ambient water column. In beds of the submerged macrophyte, Najas guadalupensis surfaces consist of live photosynthetic leaf and shoot tissues, inor ganic precipitates and settled particles. The emergent macrophyte, Typha spp., has stands of live shoots, along with dead non-photosynthetic shoots, decomposing leaf litter, and settled particles in contact with the water column. Epiphytic and microbial colonies utilize submerged surfaces, especially those which provide necessary nutrients and carbon. Nutrients absent in the substrate must be acquired from the ambient water solution. Due to low P availability relative to carbon, oxygen, nitrogen, and other essential elements, P uptake can be rapid. In some systems, uptake of P from the water column by epiphytes and microbes could equal or ex ceed direct uptake by the macrophytes which are used to describe the community. However, microbes and metaphyton are well-known to internally cycle nutrients many times faster than macrophytes because of their faster turnover rates. Due to rapid desiccation and inadequate sampling methodology, microbial and epiphytic biomass is often co nsidered small if not insignificant to the P mass balance of the larger community. Nevertheless, P sequestration to surfaces by epiphytes and microbes may be an important link between the ambient water column and the limitations of direct uptake mechanisms of macrophytes. In this way, the aquatic epiphytes, bacteria and fungi are

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121 analogous to mycorrhizae which extend the P depletion zone in soils around terrestrial plant roots. In fact, aquatic my corrhizae associated with the freshwater submerged macrophyte, Vallisneria americana were observed to enhance P uptake by 85% compared to plants without active mycorrhizae (Wigand and Stevenson, 1997). Microbial activity in treatment wetlands is controlled by numerous factors including the abundance of available oxygen and alternative electron acceptors. In this study, microbial P uptake was dependent on DO concentrations. At low oxygen concentrations, soluble P concentration reductions by Typha litter were slower and less complete than in simil ar treatments incubated under higher DO levels. As an emergent stand increases in density, shading of the water column may reduce photosynthetic oxygen re-s upply to detritusand leaf litterassociated heterotrophs. At the same time, accumulations of litter in dense stands increase the biological oxygen demand. Less dense stands of emergent macrophytes that allow light penetration to aquatic autotrophs are likely to host microbial colonies capable of reducing P concentrations to very low levels (SRP < 2 g L-1 ). Phosphorus sequestered into microbial biomass under low oxygen conditions is subject to release into th e water column. This suggests that under stagnant conditions, water beneath the canopy of emergent vegetation where respiration exceeds photosynthesis, dissolved oxygen levels are reduced and water column P concentrations can become elevated.

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122 The differences in morphology between the emergent and submerged macrophytes therefore may be generalized in terms of metabolism. Dense emergent stands contain many surfaces that consume oxygen, while SAV communities provide more surfaces that are oxygenated through primary productivity. Implications for STA Management The uptake potential of the Typha litter may account for the water column P reduction capacity observed in Typha -dominated wetlands. Cell 1 of a Stormwater Treatment Area (STA-1W) wetland removed an average 0.860 g P m2 yr-1 during 2002 (SFWMD, 2003), which equates to 2.4 mg m-2 day-1. In comparison, SAV-dominated Cell 4 is a smaller cell that has consistently provided higher P reductions than Cell 1. In water year 2002, Cell 4 P reductions averaged 2.0 g P m-2 yr-1, or 5.4 mg P m-2 day-1. The P removal rate for either cell is much lower than the 27.3 mg P m-2 day-1 taken up by fresh Typha litter during short-term (8 hours) intact soil core incubations, and the 56 mg P m-2 day-1 estimated from soil-less Typha litter incubations. Compared to Typha soils without fresh Typha litter, the addition of litter increased the rate of uptake by two-fold. With litter accumulating available P from the water column as well as any mobilized soil-P from below, a P-enriched litter layer will result. Uptake rates measured in the cores containing Typha litter cannot be extrapolated to the entire cell. The infl ow P load to an STA is not entirely bioavailable, nor is Cell 1 a uniform Typha stand with bare soils or accumulated

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123 fresh litter. Instead, Cell 1 is a patchwor k of floating, submerged and emergent vegetation. Sparse and dense stands of ea ch are found throughout the cell. Water column conditions at the center of a Typha stand can be anoxic, and litterassociated microbial P uptake is slower under low oxygen conditions. The region of a wetland that receives point discharge of ADW can be densely colonized by cattails. It is not clear, however, if cattail vegetation colonize in response to elevated water column P conditions, or elevated soil P conditions, or a combination of both. After a dense stand of the emergent macrophyte has formed however, little P removal from external loads is expected, due to anoxic conditions typical within such a stand (Figure 4-1). The slough communities of the Everglades are shallow (< 1m) systems which contain considerable DOM released from vegeta tion. High light levels in the slough communities probably helps drive primary production by autotrophs (e.g., algae, SAV) that assimilate the photolysate P. Calcium enrichment of soils through water column CaCO3 precipitation may have increased the soil Ca-bound P pool in SAV soils, but not to the extent that P mobilization from the soil was negated. Reddy et al. (1998) found that along a eutrophication gradient in Water Conservation Area (WCA)-2A, calcium enrichment in surface soils was thir ty times greater than P enrichment. When ADW containing high concentrations of Ca2+ passes through a photosynthetically-active wetland, CaCO3 precipitation occurs. Suspended particles in ADW that contain calcium may also settle within the wetland

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124 Figure 4-1. Conceptual diagrams of phos phorus, calcium and oxygen levels in the water column in relation to vegetation type. Dense cattail stands are found near point discharges of nutrient and Ca-enriched agricultural drainage waters, while sawgrass slough communities are un-enriched (with respect to Ca and P) areas of the wetland. environment to enrich soil calcium co ntents. The sorption capacity of CaCO3 is finite, however, and lower than that of Fe and aluminum hydroxides P P Ca Cattail Sawgrass Sawgrass Slough Dense Cattail O2Water Flow P P Ca Cattail Sawgrass Sawgrass Slough Dense Cattail O2Water Flow

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125 (Richardson, 1985). Continuous P loadings to treatment wetland soils may exhaust existing or newly-formed calcareous sorption sites. Soils below cattail stands may also become Ca-enriched through interactions with groundwater. Since CaCO3 precipitation is not likely to occur in dense cattail stands such as those near the inflow of WCA-2A, the aforementioned Ca-enrichment observed at those locations either occurred prior to cattail invasion, during periods of lower stand density, or through interactions with the groundwater. Calcium deposition would likely have occurred immediately upon commencement of ADW discharges to the WCA, because the slough environment present at that time would have been capable of precipitating CaCO3. Over time, the calcium precipitate may have increased soil sorption capacity, and led to P enrichment of WCA-2A inflow-region soils. Cattails that colonized the enriched soils may have altered the P retention process at work in that region. Cattail roots mobilize P from the soils (White et al., in press), canopy shade hinders the regeneration of soil sorption capacity by reducing CaCO3 precipitation, and oxygen demand slows microbial activities. An STA inflow region colonized by cattails through a similar process may provide some initial P removal, but as stand density increases, removal efficiency is likely to decline. Open water areas may therefore be essentia l to successful P removal from surface waters and long-term P retention in the newly accreted soils.

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126 Conclusions Availablility of light, carbon, oxygen an d other essential nutrients creates an intense pressure for organisms to acquire and conserve P to effectively maximize growth when favorable conditions arise. As a controlling nutrient in many freshwater systems, P is closely associated with macrophyte and microbial biomass in wetlands. Macrophytes incorp orate soil-P into biomass, then when tissues senesce, P is released into the water column in both organic and inorganic water-soluble forms. Macrophyte vegetation also plays an important role in structuring the physicochemical underwater environment. Adsorptive surfaces of amorphous CaCO3 precipitates can effectively retard the movement of P through the landscape, and may enhance Ca-bound P pools in newly accrued soils. SAV beds oxygenate the water column which allows increased microbial metabolism and P uptake. Oxygen demand maintains anaerobic conditions in deep soils and dense macrophyte stands. The greatest consequence of vegetation type on P removal performance of a wetland may be the effect the macrophyte has on the microbial biomass which it supports. The synthesis of these P cycling processes can aid wetland managers in achieving water quality treatment objectives and ultimately aid in the restoration of P-impacted marshes.

PAGE 138

127 LIST OF REFERENCES American Public Health Association (APH A). Standard methods for examination of water and wastewater. 18th Ed. 1992. Barko, J.W., and W. James. 1998. Effects of submerged aquatic macrophytes on nutrient dynamics, sedimentation, and resuspension. In : The structuring role of submerged macrophytes in lakes. E. Jeppesen et al. (eds.) Springer, NY. Barko, J.W., and R.M. Smart. 1980. Mobilization of sediment phosphorus by submersed freshwater macrophytes. Freshwater Biology 10: 229-238. Barko, J.W., and R.M. Smart. 1981. Sediment-based nutrition of submersed macrophytes. Aquatic Botany 10: 339-352. Barko, J.W., and R.M. Smart. 1983. Effects of organic matter additions to sediment on the growth of aquatic plants. Journal of Ecology 71: 161-175. Bartow, S. M., C. Craft, and C. J. Richardson. 1996. Reconstructing historical changes in Everglades plant community composition using pollen distributions in peat. Journal of Lake and Reservoir Management 12 (3): 313-322. Belanger, T. V., D. J. Scheidt, and J. R. Platko II. 1989. Effects of nutrient enrichment on the Florida Everglades. Lake and Reservoir Management 5 (1): 101-111. Bloesch, J. 1995. Mechansims, measurement and importance of sediment resuspension in lakes. Marine and Freshwater Research 46: 295-304. Boers, P. C. M., W. V. Raaphorst, and D. T. Van der Molen. 1998. Phosphorus retention in sediments. Water Science and Technology 37 (3): 31-39. Bole, J. B. and J. R. Allan. 1978. Uptake of phosphorus from sediment by aquatic plants, Myriophyllum spicatum and Hydrilla verticillata Water Research 12: 353358. Brix, H. 1997. Do macrophytes play a role in constructed treatment wetlands? Water Science and Technology 35 (5): 11-17.

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131 Khoshmanesh, A., B. Hart, A. Duncan and R. Beckett. 1999. Biotic uptake and release of phosphorus by a wetland sediment. Environmental Technology 20: 8591. King, R. J., and W. Schramm 1976. Photosynthetic rates of benthic marine algae in relation to light intensity and seasonal variation. Marine Biology 37: 215-222. Koch, M. S. and K. R. Reddy 1992. Distribution of soil and plant nutrients along a trophic gradient in the Florida Everglades. Soil Science Society of America Journal 56:1492-1499. Koch-Rose, M. S., K. R. Reddy, and J. P. Chanton. 1994. Factors controlling seasonal nutrient profiles in a subtropical peatland of the Florida Everglades. Journal of Environmental Quality 23: 526-533. Krause-Jensen, D., and K. Sand-Jensen. 1998. Light attenuation and photosynthesis of aquatic plant communities. Limnology and Oceanography 43: 396-407. Kufel, L. and Kufel, I. 2002. Chara beds acting as nutrient sinks in shallow lakes a review. Aquatic Botany 72: 249-260. Lorenzen, B., H. Brix, I. Mendelson, K. McKee and S. L. Miao. 2001. Growth, biomass allocation and nutrient use efficiency in Cladium jamaicense and Typha domingensis as affected by phosphorus and oxygen availability. Aquatic Botany 70: 117-133. Madsen, J. D., P. A. Chambers, W. F. James, E. W. Koch, and D. F. Westlake. 2001. The interaction between water moveme nt, sediment dynamics and submersed macrophytes. Hydrobiologia 444:71-84. McCormick, P. V., M. J. Chimney, D. R. Swift. 1997. Diel oxygen profiles and water column community metabolism in the Florida Everglades, U.S.A. Archiv fur Hydrobiologie 140: 117-129. McCormick, P. V., and J. A. Laing. 2003. Effects of increased phosphorus loading on dissolved oxygen in a subtropical wetland, the Florida Everglades. Wetlands Ecology and Management 11: 199-216. Miao, S.L. and W.F. DeBusk. 1999. Effects of phosphorus enrichment on structure and function of sawgrass and cattail communities in the Everglades. In : K.R. Reddy, G.A. O’Conner, and C.L. Schelske. (Eds.) Phosphorus biogeochemistry in subtropical ecosystems. Lewis Publ. Boca Raton.

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132 Miao, S. L., S. Newman, and F. Sklar. 2000. Effects of habitat nutrients and seed sources on growth and expansion of Typha domingensis Aquatic Botany 68: 297311. Miao, S.L. and F.H. Sklar. 1998. Biomass and nutrient allocation of sawgrass and cattail along a nutrient gradient in the Florida Everglades. Wetlands Ecology and Management 5: 245-263. Moeller, R. E., J. M. Burkholder, R. G. Wetzel. 1988. Significance of sedimentary phosphorus to a rooted submersed macrophyte ( Najas flexis (Willd.) Rostk. and Schmidt) and its algal epiphytes. Aquatic Botany 32: 261-281. Moore, P.A., K.R. Reddy, and D.A. Graetz. 1991. Phosphorus geochemistry in the sediment-water column of a hypereutrophic lake. Journal of Environmental Quality 20: 869-875. Newman, S., J. B. Grace and J. W. Koebel. 1996. Effects of nutrients and hydroperiod on Typha Cladium and Eleocharis : Implications for Everglades restoration. Ecological Applications 6: 774-783. Newman, S. and K. Pietro. 2001. Phosphorus storage and release in response to flooding: implications for Everglades stormwater treatment areas. Ecological Engineering 18: 23-38. Ngo, V. 1987. Boosting pond performance with aquaculture. Operations Forum 4: 20-23. Nurnberg, G.K. 1996. Trophic state of clear and colored, softand hardwater lakes with special consideration of nutrients, anoxia, phytoplankton and fish. Lake and Reservoir Management 12: 432-447. Odum, H. T. 1956. Primary production in flowing waters. Limnology and Oceanography 1: 102-117. Otsuki, A., and R. G. Wetzel. 1972. Coprecipitation of phosphate with carbonate in a marl lake. Limnology and Oceanography 17: 763-767. Patrick, W. H. Jr., and R. A. Khalid 1974. Phosphate release and sorption by soils and sediments: Effect of aerobic and anaerobic conditions Pietro, K. 1998. Phosphorus uptake rates of a Ceratophyllum /periphyton community in a Southern Florida freshwater marsh. Thesis. Florida Atlantic Univ.

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134 Rose, C., and W. Crumpton. 1996. Effects of emergent macrophytes on dissolved oxygen dynamics in a prairie pothole wetland. Wetlands 16: 495-502. South Florida Water Management District (SFWMD). 2003. Everglades Consolidated Report. West Palm Beach, FL. Stephen, D., B. Moss and G. Phillips. 1997. Do rooted marcophytes increase sediment phosphorus release? Hydrobiologia 342/343:27-34. Sweerts, J. R. A., C. A. Kelly, J. W. M. Rudd, R. Hesslein, and T. E. Cappenberg. 1991. Similarity of whole-sediment diffusion coefficients in freshwater sediments of low and high porosity. Limnology and Oceanography 36 (2): 335-342. Swift, D. R., and R. B. Nicholas. 1987. Periphyton and water quality relationships in the Everglades Water Conservation Areas. Tech. Publ. 87-2. South Florida Water Management District. West Palm Beach, FL. Toth, L. A. 1988. Effects of Hydrologic regimes on lifetime production and nutrient dynamics of cattail. SFWMD Tech Publ. 88-6. West Palm FL. Wetzel, R.G.1993. Microcommunities and micronutrients: Linking nutrient regeneration, microbial mutualism, and high sustained aquatic primary production. Netherland Journal of Aquatic Ecology 27: 3-9 Wetzel, R.G., P.G. Hatcher and T.S. Bianchi. 1995. Natural photolysis by ultraviolet irradiance of recalcitrant dissolved organic matter to simple substrates for rapid bacterial metabolism. Limnology and Oceanography 40: 1369-1380. White, J.R., and K.R. Reddy. 2001. Influence of selected inorganic electron acceptors on organic nitrogen mineralization in Everglades soils. Soil Science Society of America Journal 65: 941-948. White, J.R., K.R. Reddy, and M. Z. Moustafa. In press. Influence of hydrology and vegetation on phosphorus retention in Everglades Stormwater Treatment Area wetlands. Hydrological Processes. Wigand, C., and J. C. Stevenson. 1997. Facilitation of phosphorus assimilation by aquatic mycorrhizae of Vallisneria americana Michx. Hydrobiologia 342/343: 3541. Wigand, C., J. C. Stevenson, J. C. Cornwell. 1997. Effects of different submersed macrophytes on sediment biogeochemistry. Aquatic Botany 56: 233-244.

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135 Wium-Anderson, S., and J. A. Anderson. 1972. The influence of vegetation on the redox profile of the sediment of Grane Langso, a Danish lake. Limnology and Oceanography 17: 948-952. Wu, Y., F. Sklar, and K. Rutchey. 1997. Analysis and simulations of fragmentation patterns in the Everglades. Ecological Applications 7 (1): 268-276.

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136 BIOGRAPHICAL SKETCH Kevin Grace was born in Philadelphia, PA in 1976, and grew up in Silver Spring, MD. At the age of 16, Kevin left Maryland to attend the Florida Institute of Technology, in Melbourne, FL, where he received a BS degree in environmental science. After graduation, Kevin remained in Melbourne, a fair city by any measure, and began working for DB Environmental. In 2001, he moved to Gainesville, FL to pursue his interests both at the university and the greater north-central Florida region, where hiking, camping and canoeing opportunities abound. With luck, Kevin will continue to interact with good people through interesting work.


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Title: Phosphorus removal and soil stability within emergent and submerged vegetation communities in treatment wetlands
Physical Description: Mixed Material
Creator: Grace, Kevin ( Author, Primary )
Publication Date: 2003
Copyright Date: 2003

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PHOSPHORUS REMOVAL AND SOIL STABILITY WITHIN EMERGENT AND
SUBMERGED VEGETATION COMMUNITIES IN TREATMENT WETLANDS














By
KEVIN GRACE















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


2003













ACKNOWLEDGMENTS

This research was made possible through moral, technical, and financial

support from DB Environmental, Inc. of Rockledge FL, its staff, and founders

Thomas DeBusk and Dr. Forrest Dierberg, whose encouragement has been

invaluable. Additional support was provided by the Everglades Agricultural

Area, Environmental Protection District and the Soil and Water Science

Department staff and faculty. I extend thanks to Dr. John White and Dr. Ramesh

Reddy for their guidance; and for the improvements they made to this thesis. I

also appreciate the contributions of friends and fellow students. Thanks go to

Patrick Owens and Scott Jackson for their expert assistance in the field. Finally,

my father Don, sister Becca and brother Geoff are also an important part of

anything I do, and we fondly remember my mother Nancy, always.















TABLE OF CONTENTS

page

ACKNOWLEDGMENTS..................................................... ..... ....................... ii

LIST O F TA BLES ........................................................................................................ v

LIST O F FIG U R ES..................................................................................................... vi

A BST R A C T ............................................................................................................... xii

CHAPTER

1 PHOSPHORUS DYNAMICS IN TREATMENT WETLANDS: A REVIEW...... 1

Introd u action .......................................................................................................... 1
Phosphorus in the Water Column................................................................. 5
Phosphorus Removal Processes............................. ......................................... 6
Everglades Research Site Description............................................. ........... 13
N eed for Research .............................................................................................. 14

2 PHOSPHORUS STABILITY IN ACCRETED WETLAND SOILS................... 15

Introduction ........................................................................................................ 15
Materials and Methods ....................................................... ........................... 18
Results and D iscussion...................................................................................... 31
C onclu sion s......................................................................................................... 69

3 BIOMASS PHOSPHORUS STORAGE AND DYNAMICS........................... 71

Introduction ........................................................................................................ 71
Materials and Methods ...................................................... ............................ 74
Results and Discussion...................................................................................... 84
C onclu sions ............................................................................................................. 110

4 SY N T H ESIS....................................................................................................... 112


iii









Irradiance and W ater C olum n Shading......................................................... 114
O xygen Supply ....................................................................................................... 117
C om m unity M etabolism ....................................................................................... 119
Im plications for STA M anagem ent ..................................................................... 122
C conclusions ............................................................................................................. 126

LIST O F REFEREN C ES ............................................................................................... 127

BIO G RA PH IC A L SK ETC H ........................................................................................ 136















LIST OF TABLES


Table page

2-1 Sequential extraction of phosphorus from wetland soils using a method
adapted from Hieltjes and Lijklema (1980). ...............................................24

2-2 Water column characteristics at 3 cm and 20 cm depths in duplicate
mesocosms (2.2m L x 0.79m W x 0.4m D) dominated by emergent and
submerged aquatic vegetation (EAV and SAV, respectively) on May 30,
2001.................................................................................................................. 4 0

2-3 Water column D.O., pH, and temperature profiles in the STA-1W Cell 1
surface waters at the time of peeper retrieval on June 27, and on
Septem ber 29, 2002........................................................................................... 46

2-4 Flux estimates from intact SAV and EAV soil cores kept under dark
conditions for 35 days......................................................................................60

3-1 Phosphorus removed from the water column over the 2.7-year study
period and recovered in the vegetation (Typha + Najas) and soils upon
termination of the study on August 20, 2001........................................ ....87

3-2 Dry matter and P concentrations in accrued soil and tissue storage (live
and dead Typha shoots, below-ground Typha (roots and rhizomes), and
Najas tissues) retrieved from mesocosms on August 20, 2001 .......................89

3-3 Mean ( 1 s.d.) water quality parameter values measured in the Cell 4
outflow water used in the preliminary desiccation experiment....................92

3-4 Flux estimates from intact Typha-region soil cores and submerged Typha
litter kept under dark conditions for 35 days ...................................... ...103















LIST OF FIGURES


Figure page

1-1 The historic Everglades region of south Florida is now three distinct
parcels, including the Everglades Agricultural Area (EAA), Water
Conservation Areas (WCA) and Everglades National Park. Stormwater
treatment areas (STA) are shown (in gray), including STA-1W, where
field investigations took place........................................................ ............... 3

1-2 General model of phosphorus storage in wetlands. Arrows indicate
potential P exchange pathways between storages.............................................4

2-1 Stormwater Treatment Area 1W in Palm Beach county, Florida. Three
flowpaths receive surface water drained from adjacent agricultural soils.....19

2-2 Aerial photograph of STA-1W Cell 1, first cell of the eastern flowpath, in
N ovem ber 2000.................................................................................................21

2-3 Schematic of porewater equilibrator used in estimates of P flux from
STA-1W Cell 1 soils into the overlying water column....................................... 29

2-4 Average bulk density values for accrued sediments at depth intervals
below the sediment water interface. .................................................... ..........32

2-5 Soil phosphorus pools determined through a sequential extraction
procedure using 1.0 M NH4C1, 0.1 M NaOH, and 0.5 M HC1. .........................34

2-6 Inflow rates and stage level of water in the outflow region of Cell 1
compared to bottom elevations at the emergent and submerged aquatic
vegetation (SAV) stations, prior to and during peeper and Hydrolab
deploym ents. .....................................................................................................42

2-7 Water column temperature within Cell 1 emergent and SAV
communities during the Hydrolab monitoring period (July 3 17, 2002).....43

2-8 Water column dissolved oxygen saturation levels within Cell 1 emergent
and SAV communities during the hydrolab monitoring period (July 3 -
16, 2002). ............................................................................................................. 43









2-9 Water column specific conductance levels within Cell 1 emergent and
SAV communities during the hydrolab monitoring period (July 3 16,
2002). ............................................................................................................... . 44

2-10 Water column pH levels within Cell 1 emergent and SAV communities
during the hydrolab monitoring period (July 3 16, 2002). ...........................44

2-11 Incident light (measured as photon flux) remaining at water depths of 0+,
35 and 65 cm in emergent and submerged macrophyte communities, as
well as within open water reaches, of STA 1W, Cell 1, in September 2002....47

2-12 Stage level recorded during the seventeen day equilibration period for
porewater samplers deployed in the outflow region of STA-1W Cell 1,
during June 2002. ..............................................................................................48

2-13 Vertical profiles of porewater pH values from soil collected from
emergent and submerged vegetation.................... .......................................49

2-14 Vertical profiles of soluble reactive phosphorus concentrations in soil
porewater collected from emergent and submerged vegetation .................. 51

2-15 Vertical profiles of dissolved organic P concentrations in soil porewater
collected from emergent and submerged vegetation ......................................52

2-16 Vertical profiles of dissolved iron concentrations in sediment porewater
collected from emergent and submerged vegetation ......................................54

2-17Vertical profiles of alkalinity concentrations (as CaCO3) in sediment
porewater collected from soils below emergent and submerged
vegetation. ..........................................................................................................55

2-18 Vertical profiles of specific conductance values in sediment porewater
collected from soils below emergent and submerged vegetation ..................56

2-19Vertical profiles of dissolved calcium concentrations in sediment
porewater collected from soils below emergent and submerged
vegetation. ..........................................................................................................57

2-20 Vertical profiles of calcium carbonate saturation index (SI) values in
porewater collected from soils beneath emergent and submerged
vegetation. ..........................................................................................................58

2-21 Soluble reactive phosphorus (SRP) concentrations in the water columns
of SAV- and Typha- region soils and control columns (no soil) during 28-
day dark laboratory incubation. .......................................................................59









2-22 Mean dissolved calcium and alkalinity concentrations in the water
columns of three treatments (see text for details) during 28-day dark
laboratory incubation. .................................................... .............................. 61

2-23 Mean dissolved organic phosphorus concentrations in the water columns
above SAV and Typha-region soils and in control (water only) columns
during 28-day dark laboratory incubation................................................62

2-24 Dissolved organic carbon concentrations in water columns over Typha
soils, SAV soils and control columns (no soil) after 4, 7, 10 and 14 days of
dark laboratory incubation. ................................... ............................................63

2-25 Water column SRP concentrations above a "background" concentration
as determined before a 100 gg spike was added to each column ..................64

2-26 Phosphorus diffusive flux estimates based on 22 porewater equilibrators
deployed in STA-1W Cell 1 (June-July 2002; closed markers) and Cell 4
(November-December 2001; open markers), as a function of distance
through the entire w etland.................................... .............................................68

3-1 The historic Everglades region of south Florida is now three distinct
parcels, including the Everglades Agricultural Area (EAA), Water
Conservation Areas (WCA) and Everglades National Park. Stormwater
treatment areas (STA) are shown (in gray), including STA-1W ....................72

3-2 Stormwater Treatment Area 1 W in Palm Beach county, Florida. Three
flowpaths receive surface water drained from adjacent agricultural soils
and reduce the phosphorus load to Water Conservation Area 2A. ................76

3-3 Aerial photograph of STA-1W Cell 1, first cell of the eastern flowpath, in
November 2000 (courtesy SFW M D)................................................................77

3-4 Schematic of incubation design, containing Typha litter and P-amended
surface w aters from W CA 3A ..........................................................................82

3-5 Inflow and outflow TP concentrations for mesocosms dominated by
Typha and operated from December 1998 through August 2001 ...................85

3-6 Partitioning of recovered Najas and Typha tissue P within the inflow (In)
and outflow (Out) halves of two mesocosms operated from December
1998 through August 2001................................................ ........................... 88

3-7 Dissolved organic phosphorus (DOP) and soluble reactive phosphorus
(SRP) concentrations in floodwaters containing fresh Najas tissues and
Typha litter......................................................................................................... 93









3-8 Soluble reactive phosphorus concentrations in the water columns of
Typha-region intact soil cores with and without Typha litter amendments
during 20-day dark laboratory incubation........................................................95

3-9 Mean dissolved organic phosphorus concentrations in the water columns
of four treatments (see text for details) during 28-day dark laboratory
in cu b action ........................................................................................................... 97

3-10 Mean dissolved calcium and alkalinity concentrations in the water
columns of four treatments (see text for details) during a 28-day dark
laboratory incubation................................................................ .....................98

3-11 Dissolved organic carbon concentrations in water columns of four
treatments (see text for details) after 4, 7, 10 and 14 days of a dark
laboratory incubation...................................................... ............................. 100

3-12 Water column SRP concentrations above a "background" concentration
as determined before a 100 gg spike was added to each column.................102

3-13 Mean soluble reactive phosphorus (SRP) concentrations in triplicate 0.5 L
flasks, containing Typha litter incubated in the dark under oxic or anoxic
conditions for 5 days.......................................................................................104

3-14 Mean soluble reactive phosphorus (SRP) concentrations in triplicate 0.5 L
flasks, containing Typha litter incubated in the dark under oxic or anoxic
con edition s. ......................................................................................................... 106

3-15 Change in microbial biomass phosphorus (MBP) of Typha litter incubated
for seven days ..................................................................................................107

4-1 Conceptual diagrams of phosphorus, calcium and oxygen levels in the
water column in relation to vegetation type................................................ 124














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

PHOSPHORUS REMOVAL AND SOIL STABILITY WITHIN EMERGENT AND
SUBMERGED VEGETATION COMMUNITIES IN TREATMENT WETLANDS

By

Kevin Grace

August 2003

Chair: John R. White
Major Department: Soil and Water Science

Phosphorus (P) removal by treatment wetlands is an integral part of

Everglades protection and restoration. The effects of water column shading on P

cycling and retention were explored in emergent (Typha spp.) and submerged

(Najas guadalupensis) vegetation communities within a Stormwater Treatment

Area (STA-1W) wetland in south Florida. The physicochemical aquatic

environments within these two vegetation communities were hypothesized to

differentially affect community metabolism, which in turn would affect

biological P uptake rates and P stability in accrued soils.

Muck soils beneath emergent aquatic vegetation (EAV) were P-depleted

over 8 years of operation in STA-1W, while the muck soils beneath submerged

aquatic vegetation (SAV) beds were P-enriched. The stability of P within newly









accrued soils was dependent on macrophyte community type, and was likely

increased through water column CaCO3 precipitation. Soils (0-4 cm layer) in SAV

communities contained significantly more P in residual, Ca/Mg-bound and

fulvic/humic acid-bound pools than soils in EAV. Because of similar pools of

exchangeable and Al/Fe-bound P, the two soil types each released P to an

oxygenated water column at similar flux rates.

Typha litter and associated microbial biomass retained P mineralized from

soils under oxic water column conditions, but retention was lower under anoxic

conditions. Dense EAV stands accumulate oxygen demand, reduce light

penetration and may have little microbial P uptake and retention capacity due to

anoxic conditions. While Typha biomass persists as leaf litter and detritus to a

greater extent than Najas tissues, the extensive Typha root system has the

potential to hinder long-term storage by mobilizing P from enriched soils.

Community metabolism was influenced by water column shading, which

reduced CaCO3 precipitation and Ca-bound P pools; and reduced oxygen supply

to microorganisms in EAV communities. Managing for SAV and eliminating

dense EAV stands from treatment wetlands may reduce surface water TP

concentrations. Phosphorus-enriched areas within the northern Everglades may

also be contained or restored by increasing light penetration to the water column,

in order to enhance soil P sorption capacity through CaCO3 precipitation and

increase photosynthetic oxygen supply to the aquatic microbial communities.













CHAPTER 1
PHOSPHORUS DYNAMICS IN TREATMENT WETLANDS A REVIEW
Introduction

Phosphorus (P) availability limits primary productivity in many

freshwater systems (Boers et al., 1998; Reddy et al., 1999). Cyanobacterial mats

are periphyton communities that can maintain high productivity in low P

environments due to rapid internal nutrient recycling (Wetzel, 1993).

Cyanobacteria (Schizothrix, Scytonema) within the mats are competitive with

emergent macrophytes (i.e., cattails (Typha spp.)) in waters with low P

concentrations, however they are sensitive to increases in P levels. In contrast,

Typha grows well in P-enriched soils, and can produce a large biomass annually

(Toth, 1988; Davis, 1991; Davis, 1994). A shift from open water areas dominated

by calcareous, cyanobacteria periphyton to monoculture stands of Typha, in areas

downstream of water control structures, has raised concerns among resource

managers about the loss of ecological integrity of the oligotrophic Everglades

marsh (Davis 1994).

Constructed wetlands have become a practical treatment technology for

removing P from surface waters. In south Florida, farmland in the Everglades

Agricultural Area (EAA) has been taken out of cultivation and flooded to create

wetlands for surface water treatment (Figure 1-1). The created wetlands, or









stormwater treatment areas (STAs), were subsequently colonized by a mixture of

emergent, submerged and floating wetland vegetation.

Submerged aquatic vegetation (SAV) including Najas guadalupensis

Spreng., and emergent cattail (Typha latifolia L., and T. domingensis Pers.),

communities coexist throughout the STAs. Floating macrophytes (Eichhornia

crassipes, Pistia stratiotes) and periphyton appear transient in the upstream and

downstream reaches of the wetland flow path, respectively.

Current water management objectives in south Florida for treating

agricultural drainage waters (ADW) with treatment wetlands include achieving

low effluent total phosphorus (TP) concentrations, maximizing phosphorus

removal from the water column, and producing stable soils for long-term P

storage. Essential to achieving these objectives is an understanding of P exchange

rates between three major storage: soil, biomass and the water column (Figure

1-2). The water column is a compartment of variable P storage that must be

minimized for effective P removal. The biomass of macrophytes and

microorganisms influences several biogeochemical cycles that control P exchange

between compartments. Soils are viewed as the long-term storage compartment,

though surface soils and associated P interact with the biomass and water

column compartments.









FLORIDA, USA


Park


N

W E


Figure 1-1. The historic Everglades region of south Florida is now three distinct
parcels, including the Everglades Agricultural Area (EAA), Water
Conservation Areas (WCA) and Everglades National Park. Stormwater
treatment areas (STA) are shown (in gray), including STA-1W, where
field investigations took place.


*" .































Figure 1-2. General model of phosphorus storage in wetlands. Arrows indicate
potential P exchange pathways between storage. Deep soils represent
a long-term P storage compartment.

Water column P removal and soil stability (P storage potential) can be

understood through examination of biological processes (e.g., plant P uptake,

tissue senescence, and microbial decomposition), chemical processes (e.g.,

adsorption/desorption from surfaces and precipitation/dissolution of P

compounds), and physical processes (e.g., particulate settling and resuspension,

soil burial). Each of these processes within the treatment system can influence the

P retention capacity of the system.









Phosphorus in the Water Column

Phosphorus Species

Phosphorus exists in a variety of chemical forms within natural waters.

Compounds can range from dissolved ion species (e.g. H2P04-, HP042-) to large

complex molecules. Because of the range in molecular size, it is important to

describe two classes of phosphorus compounds, dissolved and particulate

species. An operational definition of dissolved P is that which passes through a

0.45 pm membrane filter, whereas the particulate P is retained. Particulates > 0.45

tpm in diameter are generally too large for direct uptake by plants, algae and

microorganisms. However, particles can be transported by surface waters, and

subsequently mineralized downstream.

Phosphorus compounds also exist as organic or inorganic forms, though

there can be considerable exchange between the two. Organic P is either derived

from cell components or is soluble P that has been completed by organic matter,

and can be resistant to mineralization or dissociation. Inorganic P can be either

dissolved phosphate ions or sorbed onto (or incorporated within) a mineral

particle.

Soluble reactive phosphorus (SRP) is the dissolved inorganic fraction that

is readily bioavailable. Dissolved organic phosphorus (DOP) is less reactive and

must be hydrolyzed prior to biological assimilation. The remaining fraction is

particulate phosphorus (PP), which is retained by a 0.45 tpm membrane.









Phosphorus Sources

Phosphorus in any form can be imported into a wetland through surface

inflow waters, precipitation, or groundwater seepage (allochthonous). It can also

be transformed within the wetland from one form to another autochthonouss).

Internal loading represents P released to the water column from biomass and soil

storage. The allochthonous load can be sequestered during the growing season

by plant uptake and then released as an internal load during periods of plant

senescence. Substantial internal loads can also be attributed to soil P release. Both

allochthonous (external) P loads to the Everglades system and autochthonous

(internal) loads from nutrient-impacted areas must be reduced in order to

achieve management and restoration objectives.

Phosphorus Removal Processes

Biological Processes

Assimilation

Biological P uptake is almost exclusively in the soluble reactive form

(SRP). Carignan (1982) correlated 97% of P uptake by macrophytes with changes

in soil porewater and water column SRP concentrations. Bacteria and other

microorganisms assimilate P and rapidly recycle it to the water column upon

death (Currie and Kalff, 1984). Rates of SRP assimilation by macrophytes are

slower and highly variable across species. Richardson and Marshall (1986) found

that P04 additions to emergent macrophyte enclosures were initially assimilated

into the microbial biomass and detritus, rather than into the emergent biomass.









Emergent macrophytes and some submerged plants obtain P from the soil

porewater, while other submerged macrophytes obtain the nutrient from water

directly through stem and leaf tissues. Phytoplankton assimilate P directly from

the surrounding water column, and are dependent on mixing to circulate

sufficient nutrients for growth. Epiphytes and benthic populations of algae,

bacteria and fungi depend on water flow to carry necessary nutrients across the

substrate surface, and reduce boundary layers that can otherwise limit nutrient

supply to slow rates of molecular diffusion.

Phosphorus uptake rates depend on the microbial, algal or higher plant

biomass. Macrophyte biomass P storage increases throughout the growing

season as biomass increases, though tissue P concentrations may be variable

(Hill, 1979). Cattails can allocate 60% of the total biomass to below-ground tissue

in low P environments, but only 40% in P-enriched environments (Miao and

Sklar, 1998; Miao and DeBusk, 1999). Biomass P is then subject to recycle into the

water column, or incorporation into the soil detritus/microorganism

compartment where it is slowly buried into the deep sediments (Richardson and

Marshall, 1986).

Nutrient availability in the water column dictates the importance of soil

nutrients. Rattray et al. (1991) suggested that in eutrophic, aquatic environments

macrophyte growth was not necessarily limited by soil nutrient deficiencies.

Using radioisotopes, however, Carignan and Kalff (1980) have shown that SAV

rooted in eutrophic waters (167 gg SRP L-1) would still obtain nearly all tissue P









from the soils, and they implicate macrophytes as potential nutrient "pumps"

that return P from the soils to the water column. In oligotrophic water, soil

nutrition (especially P) can limit growth of rooted, submerged macrophytes

(Rattray et al., 1991). Organic matter accumulation on mineral soils was shown to

influence SAV species succession in lakes (Barko and Smart, 1983) with

maximum growth rates in soils containing 5-20% organic matter. If either SAV

species or growth rate affects P removal by SAV beds, organic matter

accumulations would influence P removal capacity of the community.

Despite the potential importance of soil P chemistry in rooted macrophyte

nutrition, other aquatic plants such as phytoplankton, periphyton, floating

plants, unrooted macrophyte species (e.g. Ceratophyllum demersum), and SAV

fragments depend entirely on nutrient uptake from the water column.

Furthermore, water column P uptake beyond the requirements for maximum

growth, referred to as "luxury uptake", has been reported for a

Ceratophyllum/periphyton complex growing in South Florida ADW (Pietro 1998).

Currie and Kalff (1984) suggested similar maximum P uptake kinetics

exist for bacteria and phytoplankton. However, they also observed a five-fold

greater increase in bacteria intracellular P concentrations over phytoplankton

during "P-limited" lab incubations. Algae are therefore capable of greater P

uptake per mass in the water column, because its biomass is usually greater than

that of bacteria (Currie and Kalff, 1984). However, under the same P-limited









conditions, bacteria are able to incorporate P more readily into cellular tissue,

which may affect the rate at which P becomes available again.

Chemical Processes

Adsorption

Phosphorus adsorption onto mineral surfaces is often related to the

presence of calcium carbonates and Fe and Al hydroxides in soils (Patrick and

Khalid, 1974; Richardson, 1985; Porter and Sanchez, 1992). Patrick and Khalid

(1974) demonstrated the influence of iron chemistry on P sorption and release

from flooded soils. In the oxidized water column, dissolved iron is low because

Fe3+ dominates and is precipitated as ferric oxyhydroxides. Under reducing

conditions after oxygen is consumed by chemical and biological oxidation, iron is

solvent as the Fe2+ ion.

In acid soils, the phosphorus adsorption potential can be estimated

accurately from the extractable aluminum and iron contents in the soil

(Richardson, 1985). In soils of higher pH, carbonates control P sorption capacity

and availability. Due to relatively low Fe and Al contents of EAA histosols, P

sorption isotherms were unrelated to Al levels and weakly correlated to Fe, but

were significantly correlated with total Ca and free carbonates (Porter and

Sanchez, 1992).

Richardson and Vaithiyanathan (1995) examined P sorption along a

gradient of soil-P enrichment in the northern Everglades. They found the P

adsorption coefficient (a measure of P sorption capacity) was lower in Everglades









histosols than for mineral wetland soils where iron regulates P sorption

characteristics. In Everglades soils where anoxic, reducing conditions are

common, P sorption is controlled by CaCO3 (Koch and Reddy, 1992) rather than

iron hydroxides (Reddy et al., 1993; Richardson and Vaithiyanathan, 1995).

Precipitation

In addition to biological uptake and chemical adsorption processes,

CaCO3 precipitation is an important P removal mechanism in SAV systems,

either through CaCO3 sorption of P or direct coprecipitation (Gumbricht, 1993).

CaCO3 compounds are not subject to dissolution and subsequent P release when

reducing conditions develop, but are sensitive to change in CO32- equilibria, pH

and temperature.

CaCOs(s) Ca2+ + CO32- Ksp = 10-8.4 (at 25C) [1]

In Fe and Al-dominated systems, Fe and Al-oxide precipitates play a

similar role by incorporating phosphate from the water column into insoluble

precipitates (Richardson, 1985). Insoluble calcium phosphate minerals such as

apatite can only form under high concentrations of P, well above the levels

observed in most surface waters (Golterman, 1998). Porewater P concentrations

are frequently higher than those in the water column, however, and precipitation

of metastable Ca-P minerals (e.g., tricalcium phosphate) may occur within the

soil environment.









Physical Processes

Settling

Newly accrued wetland soil is a particulate matrix comprised of microbial

biomass, macrophyte detritus, and inorganic solids settled from the water

column. Settling of suspended particles occurs faster in emergent and submerged

macrophyte communities than in open water (Madsen et al., 2001), due to

reduced flow velocities and mixing. Wetland substrata including standing

shoots and litter in Typha stands, SAV biomass, and the soil surface collect

flocculent material as it settles onto surfaces which expands the "active" surface

into three dimensions. Recently settled particles are in dynamic exchange with

the water column via diffusion gradients, decomposition and leaching,

bioturbation and resuspension. They can also be buried deeper into the soil

profile. Surface soil P is still potentially bioavailable to rooted macrophytes,

benthic algae, and soil microbial populations.

Surface soil is subject to resuspension by turbulent, high-velocity flow, as

well as through bioturbation. Soil resuspension increases turbidity and decreases

light penetration through the water column (Bloesch, 1995), which affects

community metabolism and P uptake. Since P concentrations are generally lower

in the water column than in the soil porewater, resuspension of sediment

particles increases desorption from solid phases into the bulk solution.

Unvegetated reaches of a wetland flow path may dramatically alter

physical soil stability. In addition to faster flow velocities and shorter hydraulic









retention times, unvegetated areas lack roots to stabilize the soil. High flow

conditions may scour the bottom, suspending these soils in the water column for

transport downstream. This process maintains deep-water conditions relative to

nearby vegetated areas. Emergent and submerged macrophyte communities

stabilize the water column and reduce soil resuspension (Gumbricht, 1993;

Madsen et al. 2001), compared to open water sites.

Burial

Whether initial removal processes are biological, chemical, or physical,

long-term storage of P in wetlands requires burial into deep sediments. Burial by

organic matter accumulation occurs as the net result of a productive community

metabolism or whenever primary production exceeds respiration. Excessive

external P loading to the Everglades marsh has resulted in long-term

accumulation of P in the sediments and biomass storage (Reddy et al. 1993,

Reddy et al. 1998). Using 137Cs dating techniques, Reddy et al. (1993) estimated P

accumulation rates in Everglades soils from 0.11 to 1.14 g m-2 yr-1. They also

identified higher nutrient retention by Typha than Cladium (sawgrass)

communities.

Long-term burial rates are also influenced by temperature, hydrology, and

fire regime (Reddy et al. 1993), as each of these factors affects the rate at which

accumulating organic matter is oxidized to CO2 and water. Decomposition of

organic matter increases with temperature due to increased microbial

metabolism. The warm temperatures in south Florida allow rapid









decomposition, yet flooded conditions reduce the supply of oxygen to

decomposers. Alternate electron acceptors (other than 02) are used in oxidation

reactions under flooded conditions, and may control the overall rates of organic

matter decomposition/ accumulation and associated P burial in wetlands (White

and Reddy 2001). Periodic fires, often the result of lightning strikes, can rapidly

oxidize organic matter and affect long-term rates of organic matter accumulation.

Everglades Research Site Description

Located between the EAA and Water Conservation Area (WCA) 1, STA-

1W was the first of six STAs to begin flow-through operations in 1994. Wetland

environments in STA-1W that were dominated by SAV and cattail communities

were examined in this study with respect to phosphorus removal and retention

into new soils. Currently managed as the Arthur R. Marshall Loxahatchee

National Wildlife Refuge, WCA-1 represents the northern extent of the largely

unaltered Everglades land.

To the south, WCA-2 has been the focal point of much research on the

impacts of P on wetland processes. A clear and well-documented transition from

Cladium jamaicense prairies to Typha spp.-dominated wetlands exists south and

west from canal discharge points. Data from this study are compared to the

many relevant studies that were conducted within WCA-2A along the northern

eutrophication gradient.

Directly south of WCA-2, and between urban Miami and the Everglades

National Park, is WCA-3. This region is a mosaic of pristine Everglades ecotypes,









from wet sloughs dominated by Nymphaea spp. and Utricularia spp. to sawgrass

(Cladium) ridge and tree island communities. Isolated stands of Typha exist across

the landscape, and are associated with alligator holes.

Need for Research

Emergent and submerged vegetation communities differ in physical and

chemical structure, yet the effects of these differences on P removal and soil

stability characteristics within the treatment wetland context are largely

unexplored. Until recently, the two vegetation types were each considered a part

of the naturally-recruited wetland community, and little effort was made to

separate the effects of one independent of the other. For example, the biomass

storage potential of cattails has been viewed as an asset to wetland treatment,

and many Typha wetlands function well as wastewater polishing areas (Kadlec

and Knight 1996). However the contribution towards P removal and retention of

the submerged macrophytes, epiphyton and microbial community, is unknown.

In eutrophic, aquatic environments continuously loaded with P, cattails

are capable of creating large monoculture stands (Davis 1991). Since, they are

ubiquitous, aggressive colonizers, the differences between SAV and cattail

communities in providing different P removal and effluent TP concentrations are

important. My research, therefore, compared emergent and submerged

macrophyte communities with respect to P interactions within the water column

and the development of new soils.
















CHAPTER 2
PHOSPHORUS STABILITY IN ACCRETED WETLAND SOILS
Introduction

Long-term retention of phosphorus (P) in wetlands can occur through the

accretion of new soil (Reddy et al., 1993). Treatment wetlands have been

constructed for P removal from surface waters, and accrete soils to store P and

protect downstream waters from eutrophication (Kadlec and Knight, 1996).

Stability of newly accrued soil-P is an important issue in such wetlands, as

internal P cycling can elevate water column P above target outflow

concentrations. Newly accreted surface soils are in contact with both the water

column and macrophyte roots, thus soil-P can be released into the water column.

Soil P retention depends on the characteristics of the wetland vegetation as well

as the chemical environment within the water column and surface soil layer.

Aquatic photosynthesis elevates water column pH levels during daylight

hours due to consumption of dissolved inorganic carbon species (i.e.CO2, HCO3-)

that equilibrate with solid carbonates. When water column calcium and

alkalinity levels are sufficiently high, pH elevations lead to calcium carbonate

supersaturation and precipitation (Otsuki and Wetzel, 1972). The precipitate, in

turn, provides sorption sites for dissolved inorganic P, and results in a calcium-









and phosphorus-enriched soil. At 25 'C, the following equilibrium reactions

govern CaCO3 solubility in water:

CaCO3(s) + C02(g) + H20 Ca2+ + 2HC03- K = 10-6.03 [2]

2HC03- 2H+ + 2C32- Ksp = 10-1033 [3]

CaCO3(s) + C02(g) + H20 Ca2+ + 2H+ + 2CO32- K = 10-26.63 [4]

Calcium carbonate chemistry may play an important role in wetland

treatment of drainage water from the Everglades Agricultural Area (EAA), a

region productive in sugarcane and winter vegetables. The EAA has muck soils

underlain with calcareous limestone (Gleason and Stone, 1994). Shallow surface

waters coupled with a high water table increases surface water and groundwater

interactions, which elevate Ca and carbonate (alkalinity) concentrations in

irrigation waters, agricultural drainage water (ADW), and in the soil itself.

Excessive phosphorus (P) loading from EAA ADW discharges has been

identified as the primary cause for an observed eutrophication gradient in the

northern Everglades. Changes observed near the discharge structures include

increased water column, soil, and plant tissue P concentrations, and change in

ecosystem function, relative to the interior marsh (Craft and Richardson, 1993;

DeBusk et al., 1994; Reddy et al., 1993; Reddy et al., 1998). Phosphorus

enrichment has led to increased Typha spp. above-ground biomass and shoot

density (Grimshaw et al., 1997; Wu et al., 1997; Miao et al., 2000), and reduced

PAR levels at the air-water interface, relative to levels at nearby open slough









sites. Reduced PAR may limit aquatic photosynthesis, which could reduce

CaCO3 precipitation and soil P retention.

Rates of P diffusion from soils to overlying water can be controlled by the

strength of the ion activity gradient, temperature and soil porosity. Unequal

distribution of phosphate, due to biological uptake mechanisms or sorption onto

solid minerals such as CaCO3 and FeOOH (Golterman, 1998), can establish an ion

activity (related to ion concentration) gradient. Moore and others (1991) found a

diffusive P flux rate of 1.69 mg m-2 day-' for sediments in a hypereutrophic

freshwater lake (Lake Apopka, FL). The magnitude of internal nutrient flux can

equal external loads to such a system, and maintain water column P above 30 Rg

L-1, concentrations typical of eutrophic systems (Nurnberg, 1996).

Internal loading can potentially impair treatment wetland performance.

As P-enriched soils accumulate, the potential for diffusive flux to mobilize soil-P

into the water column may increase. Additionally, emergent macrophyte shade

may limit water column photosynthesis and CaCO3 precipitation, relative to

submerged macrophyte areas, thereby producing soils of different P retention

capacity and internal P flux rates. The influence of macrophyte vegetation

shading of the water column on CaCO3 precipitation and accrued soil stability is

currently unknown.

The objectives of this study were to:

S Characterize the stability of P within accreted soils of emergent and
submerged macrophyte stands through sequential P extraction,









* Determine the potential for CaCO3 formation in emergent and submerged
macrophyte communities, both in the water column and the soil
porewater,

* Estimate the potential internal P flux to the water column from newly
accrued soils formed within emergent and submerged macrophyte stands.

These research objectives will address the stability of P in soils formed in

emergent and submerged macrophyte-dominated wetlands, which is essential to

the long-term management of STAs as well as the nutrient-impacted northern

Everglades.

Materials and Methods

This investigation was pursued through bench-scale incubations, in field-

operated mesocosms, and in situ within the full-scale wetland environment.

Bench-scale studies took place at DB Environmental, Inc., in Rockledge FL; and

at the Wetland Biogeochemistry Laboratory at the University of Florida in

Gainesville, FL. Outdoor mesocosms were located next to the STA-1W inflow

canal on an experimental platform provided by DB Environmental.

STA-1W Site Description

Agricultural runoff from the Everglades Agricultural Area (EAA) Basin S-

5 (Figure 2-1) is pumped via canals to STA-1W, a full-scale (2699 ha) treatment

wetland operated by the South Florida Water Management District (SFWMD) to

reduce P loadings to the Everglades (SFWMD, 2003). Everglades muck soils

were drained for agricultural production decades ago. Nearly 50, 000 acres of

former ag land have recently been reflooded to create STAs, of which the








Everglades Nutrient Removal project, now part of STA-1W, was the prototype. A

detailed spatial analysis of phosphorus and other constituents of the pre-existing

farm soils in the STA 1W footprint was conducted prior to flooding (Reddy and

Graetz, 1991).


STA 1W (2699 ha)


5B


Loxahatchee NWR
Water Conservation Area 1



Northern Flowpath


i
U-


Eastern Flowpath
Western Flowpath


Figure 2-1. Stormwater Treatment Area 1W in Palm Beach county, Florida. Three
flowpaths receive surface water drained from adjacent agricultural
soils. The wetland functions to reduce water phosphorus
concentrations prior to discharge into Water Conservation Area 2A.
Cells 1, 2 and 3 are comprised of mixed emergent, submerged and
floating vegetation, while Cells 4, 5A, and 5B are primarily submerged
and floating vegetation. Arrows indicate general direction of flow.

Cell 1 is the first cell of the eastern flowpath, and is comprised of emergent

aquatic vegetation, or EAV, (Typha spp.) in the inflow region and along the









eastern edge (Figure 2-2). A mixture of emergent, floating, and submerged

aquatic vegetation (SAV) occupy the downstream reaches of the cell. This

distribution of community types has remained relatively constant in Cell 1 since

the wetland was first flooded in 1994 (Newman and Pietro, 2001). The

juxtaposition of areas dominated by submerged and emergent vegetation in the

Cell 1 outflow region suggests that the two community types have developed

under similar hydraulic and nutrient loadings. Eight years of flow-through

operations in this wetland have been sufficient to allow accretion of soils in both

community types, and to establish a record of long-term effectiveness of P

sequestration from the water column.

Two sampling sites were selected in the outflow region of Cell 1. One

station (26.6292N, 80.4219 W) represented EAV, Typha domingensis, while SAV

species Najas guadalupensis and Ceratophyllum demersum occupied the second

station (26.6278N, 80.4352 W). The longevity of the contrasting vegetation

communities observed at the two locations was verified with aerial photos

provided by SFWMD, as well as through personal communication with field

personnel.

Stability of recently accreted wetland soil-P was characterized through

bulk density analysis and sequential extractions. Vertical profiles of water

column chemistry were constructed for mesocosms dominated by emergent and

submerged macrophytes and for emergent and submerged macrophyte

communities in southeastern Cell 1 of STA-1W. Profiles of soil porewater









constituents were also constructed for the STA soils. For each profile, the CaCO3

saturation index was calculated to determine whether conditions for

precipitation were present.


Figure 2-2. Aerial photograph of STA-1W Cell 1, first cell of the eastern flowpath,
in November 2000 (courtesy SFWMD). Also shown is the location of
the field station in emergent (pink) and submerged (grey) vegetation
stands.

Phosphorus diffusion flux rates across the sediment-water interface were

calculated based on porewater concentration gradients. Potential P flux from

these soils was then determined experimentally using intact cores incubated

under lab conditions.









Soil Collection and Analysis

Soil cores were retrieved from both field sites on July 3 and 4, 2002. Each

core was retrieved by pushing acrylic core tubes (7 cm dia.) through the accrued

soil layer into the underlying native farm muck to a minimum depth of 10 cm.

The top end of the core was sealed with a #1312 rubber stopper prior to soil core

extraction. The horizon dividing new wetland soils from the native muck soil

was determined by differences in color and texture, and the depth of accrued

wetland soil was recorded.

Cores were sectioned in the field at 2 cm depth intervals. The 0-2 cm, 2-4

cm, and 4-6 cm intervals were retained for analysis, along with the underlying

muck. Like depth increments from five replicate cores were composite and

homogenized before analysis to account for field variability. This procedure was

repeated three times at both sites, and resulted in triplicate samples for each of

the four soil layers. A fresh 90 cc subsample of composite soil samples was dried

to constant weight (65C) for bulk density. Flocculent surface (0-2 cm) samples

were allowed to settle in graduated cylinders overnight (dark, 4 C) and the

excess water was discarded prior to analyses.

Sequential Extractions for Inorganic Phosphorus Pools

Inorganic P pools in the soil surface intervals of 0-2 cm and 2-4 cm were

characterized by sequential extraction (modified from Hieltjes and Lijklema,

1980) and contrasted to the underlying muck soils. Fresh soil samples were

weighed (5g wet) into 50 mL centrifuge tubes. To each centrifuge tube, 40 mL of









IM NH4C1 was added. The slurry was then shaken for two hours. Samples were

then centrifuged for 15 minutes at 2800 rpm before the supernatant was decanted

and filtered (0.45 pm) for SRP analysis. The NH4Cl extraction was repeated, and

the supernatant was added to that of the first extraction before analysis. The tube

and sample residue were then weighed, and 40 mL of the next extractant was

added (Table 2-1).

Sodium hydroxide (NaOH) and HC1 extractions were shaken for 17 and

24 hr, respectively. Both SRP and TP analysis were performed on NaOH

extractions, while only SRP analysis was performed on the HC1 extracts. Soil

residue remaining after HC1 extraction was subjected to TP analysis. TP samples

were filtered through Whatman 41 qualitative filters. Extractants were

refrigerated at 4C until analysis. Soluble reactive phosphorus colorimetric

analysis (potassium antimony tartrate, sulfuric acid, ammonium molybdate,

ascorbic acid) was performed on a Spectronics Genesys 5 spectrophotometer.

Total P analysis included a persulfate digestion and neutralization prior to SRP

analysis.

Using the residue weights recorded after each extraction, an estimation of

P carry-over (the extractant volume and P mass carried over between extractions)

was calculated and used to adjust each P pool.

Flux Study Using Intact Soil Cores

Three additional replicate intact soil cores were retrieved from the SAV

and EAV communities for intact core incubations. Each soil core was topped off









with site water, then sealed with rubber stoppers. An opaque shroud minimized

solar heating and blocked light during transport to the lab.

Table 2-1. Sequential extraction of phosphorus from wetland soils using a
method adapted from Hieltjes and Lijklema (1980).
Sediment P pool Extractant Shake Time Analysis Stability

Exchangeable 1.0 M NH4C1 Two 2 hr SRP Readily
at pH 7 periods bioavailable

Fe- and Al-bound SRP

Humic and Fulvic 0.1 N NaOH 17 hr TP-SRP
Acid-bound

Ca-bound 0.5 N HCI 24 hr SRP

Residue Remainder digested for TP Recalcitrant


Overlying water was replaced with 1.15 L filtered (0.45 pm) STA-1W Cell

4 outflow water. The 30 cm water column was aerated and incubated in a dark

water bath at -22C for 28 days. Water samples (30 mL) were withdrawn at At =

0, 0.5, 1, 1.5, 2, 4, 6, 10, 14, 20, and 28 days, and analyzed for SRP and pH.

Representative samples were also analyzed for TSP, dissolved calcium, total

alkalinity, and DOC. Reflood water was added (30 mL) after each sampling to

maintain water volume.

Samples for SRP and TSP were filtered through a 0.45 tpm polyether

sulfone filter immediately after sample collection. Persulfate digestion followed

by neutralization was performed on TSP samples. SRP and TSP water

determinations followed the ascorbic acid-molybdenum blue method (EPA 365.2;









EPA 1979) using a Spectronics Genesys 5 spectrophotometer. Dissolved calcium

was determined using flame atomic absoption spectroscopy (EPA 215.1; EPA

1979) on a Perkin-Elmer 3110. Alkalinity was titrated with 0.02N H2SO4 (EPA

310.1; EPA 1979). Dissolved organic carbon analysis was on acidified, filtered

(0.45 pm) samples, and measured with a Shimadzu TOC-5050A (Duisburg,

Germany) TOC analyzer equipped with an ASI-5000A autosampler (5310-A;

APHA 1992).

Sample pH was recorded immediately following collection, using a 3 in 1

gel filled combination pH electrode and Corning 313 pH meter. Water bath

temperature was continuously recorded by a StowAway Tidbit logging probe

(Onset Computer; Bourne, MA) as well as monitored periodically with a

thermometer.

At the conclusion of the P Flux study, the water column of each core was

spiked with 100 tg P L-1 (as KH2PO4). The water volumes above each core

differed slightly ( 5 mL) from the original water volume of 1.15 L added one

month prior, likely due to different evaporation rates induced by the aerators.

These differences were recorded but volumes were not adjusted at that time.

Water samples were withdrawn At = 0, 4, 8, 24 and 53 hours after the

amendment, and analyzed for SRP. Each core received 30 mL of unamended

reflood water after sampling to maintain water volume, and was kept under an

opaque shroud.









Mesocosm Design and Background

One SAV-dominated and two EAV-dominated mesocosms (4.2m L x

0.79m W x 1.0m D) were maintained for 2.7 years at a hydraulic loading rate of

10 cm day-1, as part of another study (DBEL, 2001). These systems were

inoculated with Najas guadalupensis and Typha spp. plants, respectively, which

were collected from within STA-1W Cell 4. Weekly monitoring of inflow and

outflow TP concentrations, temperature and pH, and periodic monitoring of

other constituents (TSP, SRP, dissolved Ca (dCa), total alkalinity (TA), specific

conductance) was performed from December 29, 1998 through August 8, 2001.

Shade Effects within EAV and SAV Mesocosm Communities

In order to investigate the effects of shade on P removal in the EAV

mesocosms, I examined the ambient light regime and calcium carbonate

saturation index (SI). Duplicate SAV mesocosms with similar operational history

were sampled for comparison.

Incident light was measured at 2m above the water surface (above cattail

canopy), and at the water surface. Two 47r spherical quantum sensors recorded

available PAR during peak daylight hours (1000-1400), using a Li-COR LI-1000

data logger (Lincoln, Nebraska). Eight replicate one-second measurements of

photon flux were averaged for each datum value, and recorded when the value

had stabilized to roughly within 1%. All comparisons were made to the

simultaneous "ambient" light levels, to adjust for short-term temporal changes in

incident radiation (e.g. change in cloud cover).









Chemical profiles through mesocosm water columns were constructed

from in situ measurements of dissolved oxygen (DO) concentrations, pH, specific

conductance and temperature taken on May 30, 2001. Concurrently, surface (3

cm) and bottom (20 cm) water samples were collected and analyzed for SRP,

dCa, TA concentrations. Water samples were withdrawn from the inflow and

outflow regions of each tank (approximately 10 cm from end wall), as well as

from each inflow stream.

The SI of CaCO3 was calculated from water column and porewater

chemistry profiles of submerged and emergent macrophyte communities.

Specific conductance values provided an approximation for ionic strength.

Hydrogen ion concentration was calculated from pH values. Calcium and

alkalinity concentrations were used to calculate activity products of the Ca2+

CO32- and HCO3- ions. Temperature values were used to adjust all solubility

constants. The CaCO3 SI was then computed for water at the surface and at 20 cm

depth according the following relationship:


S[Ca HCO, HCO, K
c 2+ ICa L HCO L[HC3 J
SI= -I[5]


where: y indicates the activity coefficient of Ca2+ and HCO3
[ ] indicates the concentration of Ca2+, HCO3- and H+
K is the acidity constant of HCO3-, and
Kso is the solubility constant of CaCO3 at equilibrium.









Diel monitoring of STA-1W Typha and Najas communities

The mesocosm platforms provided a controlled environment for

examining the influence of macrophytes on the water column chemistry.

However, the water column environment in a full-scale STA during operations

(e.g. high flow events) may be different than that observed at the small-scale. To

investigate STA water column chemistry, diurnal monitoring of emergent and

SAV communities took place in the southeastern region of STA-1W Cell 1 (Figure

2-2) in July 2002. Two Datasonde Hydrolab multiprobes were deployed to record

pH, temperature, D.O. and specific conductance at 15 min intervals during a

two-week period of high flows (up to 1200 cfs through STA 1W). Each device

was suspended from a tripod to an initial probe depth of 10 cm below the water

surface.

Dissolved oxygen, temperature and pH profiles through the water column

were recorded for the same SAV and emergent stations on July 27 and September

29, 2002, using Yellow Springs Instrument dissolved oxygen meter and Corning

pH meter. In September, the water column light regime was also assessed in SAV

and emergent macrophyte stands, as well as in open water reaches of STA 1W

Cell 1. Incident light was recorded simultaneously above the water column

(ambient level), in surface waters (3 cm), at mid-depth and 10 cm above the

substrate, using the methodology described for mesocosm light measurements.









Flux Study Using Porewater Equilibrators

Porewater was collected using porewater equilibrators, or peepers,

modeled after Hesslein (1976) (Figure 2-3). Triplicate peepers were deployed in

Cell 1 (- 1 m apart) within emergent and submerged vegetation on June 10, 2002,

and retrieved 17 days later. Three peepers at each station were inserted vertically

through the accrued soil into the underlying muck soils to a depth of -25 cm.

Cover plate Coarse filter

0.2 unm
Membrane







Water

Soil surface



8 mL cell


Figure 2-3. Schematic of porewater equilibrator used in estimates of P flux from
STA-1W Cell 1 soils into the overlying water column. A 0.2 pm
polyether sulfone membrane (Supor 200) was inserted between a
coarse particle filter and the 8 mL sampling cells. Some sample cells
are above the soil-water interface, while others equilibrate with
porewater below the soil surface.

The soil-water interface bisected the peeper such that some sampling

chambers would equilibrate with soil porewater, and others would equilibrate

with the water column above the interface. Specific conductance, SRP, TSP,









dissolved calcium and iron, and alkalinity concentrations and pH levels were

measured above and below the soil-water interface. Dissolved iron

determinations were made by a bathophenanthroline method, modified from

APHA 3500-Fe D (APHA, 1992) for small sample volume and using a Spectronics

Genesys 5 spectrophotometer.

Diffusive fluxes of P (i.e. H2PO4- and HP042-) at each coring location were

calculated based on changes in SRP concentration with depth. The concentration

gradient of soil porewater constituents with distance can drive diffusive flux

according to Fick's first law, where:

dC/
J = -D dz [6]
0.470 +1.91

where J = diffusive flux, mg m-2 S-1
0 = porosity
Ds = the sediment diffusion coefficient, cm2 sec-1, and
dC/dz is concentration change, gg L-1, per depth interval cm.

The final term accounts for sediment porewater tortuosity, a parameter

that is linearly related to porosity (Sweerts et al., 1991). The concentration

gradient dC/dz, was determined using the slope of a linear regression through the

+5 to -5 cm depth interval.

Temperature was recorded during the equilibration period with max/min

thermometers deployed at the soil-water interface adjacent to one peeper at each

station. Alkalinity concentrations and pH levels were recorded in the field

immediately after peeper retrieval. The thermodynamic potential for calcium









carbonate and calcium phosphate saturation and mineral formation was then

calculated for the water column-porewater continuum at each site.

Statistical analyses on experimental data were performed using MSExcel

(v. 2000 Microsoft Corp.) ANOVA and t-test macros. Error around mean values

is presented as one standard deviation for replicate samples.

Results and Discussion

Soil Characterization

STA-1W was previously farmland, with irrigation canals transecting the

acreage. Since the land was flooded in 1993, cattail preference for shallow waters

has encouraged growth along old canal banks. Dredged spoil from pre-flooding

canal maintenance resulted in raised soil surfaces, and shallow water depths.

This pattern is evident in aerial photographs taken of the wetland in November

2000, after six years of flow-through operation (Figure 2-2).

Measured from water surface to the top of the litter (or accumulated soil)

layer, free-water depths on June 10, 2002, ranged spatially in the outflow region

of STA-1W Cell 1 from 0 to 56 cm. Additional water column was occupied by

Typha leaf litter, with up to 77 cm from air to soil surface. Based on probing-rod

sounding measurements, litter accrued to depths ranging from 0 to 30 cm.

Wetland soils had accrued above the native muck soils to variable depths

beneath the emergent (6-14 cm) and submerged (4-20 cm) macrophyte

communities. The accrued material was flocculent organic matter of lower bulk

density than the underlying muck soils (Figure 2-4). Accrued soil bulk density in








the two community types was not significantly different for any given depth (p >

0.05). Muck soils were near the surface in some cores, while other cores were

characteristic of deeper (> 6 cm) soil accrual, which may explain wide variability

in the 4-6 cm layer bulk density measurements at both the SAV station (0.169

0.049 g cm-3) and Typha station (0.223 0.093 g cm-3).



E 0-2 El D Submerged
| Emergent
% 2-4


O 4-6


Muck


0.00 0.05 0.10 0.15 0.20 0.25 0.30
Bulk Density (g cm3)

Figure 2-4. Average bulk density values for accrued sediments at depth intervals
below the sediment water interface. The underlying farm muck was
collected from below the horizon of the accrued sediment. Error bars
indicate 1 s.d. for triplicate samples, and each sample is a composite
of five discrete soil samples.

Soil Phosphorus Pools

A sequential extraction was used to characterize the relative

bioavailability of P within the surface soils and the underlying muck soils. Reddy

and Graetz (1991) used a similar sequential extraction, except 1M KC1 was

substituted for NH4LC to characterize the readily available P pool. Soil TP









decreased significantly from surface soils (0-2 cm and 2-4 cm layers) in the

emergent stand to the underlying muck soils(p < 0.05), while the SAV 2-4 cm soil

layer showed a slight increase above the surface layer in each of three replicate

composites (Figure 2-5). In the upper 4 cm, soils from the Typha community were

P-enriched (584 mg P kg-1) relative to the native muck soils (334 mg P kg-1, p <

0.05), but enriched less than 0-4 cm SAV soils (813 mg P kg-1; p < 0.01).

There was a significant difference in muck soil P levels beneath the SAV

and the emergent communities. SAV-region muck soils were 2-3 times higher in

P than the emergent muck soils (500 212 and 168 94 mg P kg-1, respectively).

In comparison to the native farm (Knight's Farm) soils (335 31 mg P kg-1) and

soils 10 months after flooding (358 35 mg P kg-1), the muck in the emergent

region has become depleted in P over the 7 years of operation, while submerged

macrophyte muck has been P-enriched (Figure 2-5). The residual P pool was

increased 10 months after flooding, and in recent samples from all depths.

Variation in soil TP on Knight's farm prior to flooding was greater than

variation for other soil parameters (Reddy and Graetz, 1991), yet no soil-P value

was reported lower than those observed below the emergent stand in this study.

Local variation in P distribution was accounted for with composite soil samples

from multiple cores, taken several meters apart. The SAV and emergent stations

were several 100 meters apart.

Compositing vertically, however, may complicate the interpretation of

observed differences between this study and that of Reddy and Graetz (1991).








Soil Phosphorus (mg P kg1)
0 200 400 600 800 1000


0-2 cm

< 2-4 cm

Muck


0-2 cm


~1





[~4


< 2-4 cm

Muck



E Drained
Hu 0-30cm
C/)

P0 Flooded
5 0-5cm


D Residue U Ca/Mg Bound

O Fulvic/Humic Acid D Fe/AI bound

D Readily Available

Figure 2-5. Soil phosphorus pools determined through a sequential extraction
procedure using 1.0 M NH4C1, 0.1 M NaOH, and 0.5 M HC1. The
residual P pool remained associated with the soil through all three
extractions. Knight's Farm data from Reddy and Graetz, 1991.









Those investigators homogenized soils at 30 cm increments before flooding, and

5 cm increments after flooding. During agricultural production surface soils are

mixed through tillage. However, wetland soils lower in the profile may have

lower P concentrations than surface soils, and large depth intervals do not

adequately describe surface soil chemistry. Thus, small depth increments were

used in this study.

Accrued soil-P occurred primarily (34-60%) as residual P, a highly

recalcitrant form. Calcium-bound HCl-extractable P was also a major pool in new

wetland soils (15-45%), especially in the SAV surface 0-4 cm soils (mean 33%, or

271 mg HCl-extractable P kg-1). This is a substantial increase from 8 mg HCl-

extractable P kg-1 found in the drained surface (0-30 cm) soils (Reddy and Graetz,

1991).

Significantly lower fractions of total soil P were associated with fulvic and

humic acids in accrued soils (5-26%), relative to drained farm muck (58%) (p <

0.05). This pool of organic, moderately available P characterized by 0.1 M NaOH

extraction was highest in the Typha surface 0-2 cm soils (19%, or 128 mg NaOH-

extractable organic P kg-1), and decreased with depth at both stations.

The Fe- and Al-bound P pool was relatively small in accrued wetland

soils, representing 2-8% of soil TP. Nevertheless, emergent surface 0-2 cm soils

contained 45 5 mg NaOH-extractable inorganic P kg-1, nearly twice that of SAV

surface soils (26 11 mg kg-1). If such a pool was primarily P associated with Fe-

hydroxides, it would be subject to mobilization during prolonged anoxic









conditions. Release of the whole 0-2 cm pool into a 1-m deep water column,

assuming equal soil bulk density of 0.1 g cm-3, would elevate TP concentrations

by 52 pg L-1 in SAV communities, and 90 pg L-1 in the emergent region. Muck

soils below emergent vegetation were three-fold lower in Fe- and Al-bound P (4

3 mg kg-1) than muck below SAV (12 5 mg kg-1).

The NH4-extractable P pool (readily available fraction) was higher in the

SAV accrued soils (56 8 mg kg-1) than in emergent accrued soils (32 10 mg kg-

1). Muck soils were low in NH4Cl-extractable P, with 5 2 mg kg-1 readily

bioavailable. These values are within the range of values reported by Reddy and

Graetz (1991) for farm soils prior to flooding (0-30 cm, 42 mg kg-1), and muck

soils after eight years of submergence (0-5 cm, 2 mg kg-1).

Accumulation of organic matter, P enrichment of surface soils and

detritus, and subsequent humification of the organic material, have been viewed

as positive wetland attributes for P removal treatment applications. Phosphorus

can be removed from the ambient water and become concentrated in recalcitrant

organic soil components. When the availability of soil-P was characterized by

sequential extraction, however, it was shown that the "bioavailable" and

"recalcitrant" P pools alike were reduced in the soil beneath emergent

vegetation, relative to either pre-flooded conditions or a contrasting community

type, namely submersed macrophytes (Najas guadalupensis) (Figure 2-5). It seems

likely that recalcitrant P compounds may be susceptible to mobilization through

biotic mechanisms such as organic acid mineralization or enzymatic hydrolysis.









In contrast to the Najas community soils, the accrued soil beneath Typha was

likely in close contact with, and under the influence of, extensive below ground

biomass.


Water Column Profile and CaC03 Saturation Index

Each mesocosm received a mean inflow P loading of 3.5 g P m-2 yr-1

(average inflow concentration = 96 pg TP L-1) over the 2.7-year period of

operation. Total P concentration reduction was significantly greater (ac = 0.05) in

the SAV mesocosm, with an average outflow concentration of 26 tg L-1, than in

the cattail mesocosms, whose outflows averaged 42 and 62 tg L-1. Differences

between cattail communities were not significant with respect to P removal

performance.

Calcium and alkalinity concentration reductions were also observed in

SAV-mesocosm surface waters, but not in the cattail stands. Between February

15, 2000 and August 8, 2001, Ca levels in inflow waters (72 mg L-1) common to

both SAV and cattail mesocosms were reduced to 47 mg L-1 at the outflow of the

SAV mesocosm. Cattail mesocosm outflows averaged 69 and 70 mg Ca L-1.

Likewise, alkalinity was reduced from 206 to 135 mg CaCO3 L-1 by SAV while

cattail outflows averaged 202 and 204 mg CaCO3 L-1.

On May 30, 2001, SAV mesocosm surface water outflow SRP

concentrations were low (5 tg L-1 ), relative to the inflow region (20-43 tg L-1),

and Post-BMP inflow waters (49-58 tg L-1) (Table 2-2). In the Typha-dominated

mesocosms, SRP concentration reductions were smaller, with surface outflow









waters of 18 and 37 gg L-1. Water at 20 cm depth had higher SRP concentration

than surface waters in all mesocosms, perhaps a result of internal P loading from

the soil to the bottom waters.

Similar trends were seen in the Ca and alkalinity concentrations (Table 2-

2). For example, surface outflow Ca levels (43 and 45 mg L-1) from SAV

mesocosms were appreciably lower than those of the ADW inflow water (61 mg

L-1). Water at 20 cm depth in both community types was nearly equal in Ca and

alkalinity concentration, and similar to the inflow waters (Table 2-2). In the SAV

systems, constituent concentrations declined between inflow and outflow, but

reductions were more apparent in surface water than at 20 cm depth. Calcium,

alkalinity, pH and temperature levels in the Typha mesocosms were uniform

internally, and showed little change with depth or with distance, compared to

the SAV mesocosms (Table 2-2). Surface waters within the SAV mesocosms had

elevated pH levels and temperatures, compared to the surface waters shaded by

Typha.

Calcium carbonate precipitation was thermodynamically favored (SI>1)

only in SAV surface waters (Table 2-2). Typha surface waters and waters at 20 cm

depth in either community were unsaturated with respect to CaCO3, and

dissolution of the mineral was favored (SI < 1). While CaCO3 formation may

have occurred in the surface waters, the cooler temperatures and lower pH levels

of waters at 20 cm depth created an unsaturated environment. Calcium and

alkalinity concentrations were higher at depth than in surface waters, which









suggests that bottom waters may be in equilibrium with settled precipitates in

the surface soils. Calcium carbonate precipitates may slowly dissolved in the

CaCO3 -undersaturated bottom water environment.

The influence of shade on aquatic photosynthesis was apparent in the D.O.

concentrations of the two community types (Table 2-2). One replicate cattail

mesocosm contained more Najas than the other, and had a more open Typha

canopy. Increased light penetration was observed for this mesocosm, with an

878% reduction in incident irradiance at the water surface compared to 972%

reduction within the more dense Typha community. SAV surface water D.O.

concentrations were 16.2 mg 02 L-1 or greater, whereas Typha surface waters

were 0.5 3.5 mg 02 L-1. Waters at depth were lower in D.O. than surface waters

in both SAV and Typha, due to lower aquatic productivity and greater distance

from the air-water interface.


Diel Water Quality Monitoring of Emergent and Submerged Communities

Since it was created in 1994, surface water in Cell 1 has been maintained at

an average stage of 3.64 0.16 m NGVD. During the 2002 deployments of the

porewater equilibrators (June 10- 27) and hydrolabs (July 3 17), the cell stage

was only slightly higher than average (3.89 m) (Figure 2-6). Prior months had

declining water levels and low (<10 cfs) hydraulic loads, typical flows during the

spring dry season.












Table 2-2. Water column characteristics at 3 cm and 20 cm depths in duplicate mesocosms (2.2m L x 0.79m W x 0.4m D)
dominated by emergent and submerged aquatic vegetation (EAV and SAV, respectively) on May 30, 2001.
Water from STA-1W inflow canal was delivered to each mesocosm at 10 cm day-1 between December 1998 and
August 2001.


Time DO CaCO3 SI
mg L-1


Alkalinity Diss. Ca


mg CaCO3 L1


mg L-1


EAV 1 13:22 148 59.6 58 7.36 31.0
Inflow 2 13:22 150 60.5 56 7.59 33.3
SAV 1 13:40 142 61.1 52 7.37 33.5
2 13:40 148 62.7 49 7.56 34.3
EAV 1 10:00 3.5 0.7 134 60.1 14 7.38 25.7
InfloRen 2 11:20 1.5 0.4 146 58.2 36 7.18 27.3
Inflow Region
SV 1 11:40 16.2 15 144 59.7 29 8.60 31.5
SAV
3 cm 2 12:00 20+ 32 108 41.6 11 9.26 31.9
EAV 1 10:24 2.3 0.8 146 58.5 18 7.42 25.4
Ouw R n 2 10:45 0.3 0.3 147 59 37 6.99 25.1
Outflow Region
SV 1 12:20 20+ 103 122 42.7 5 9.90 33.2
SAV
2 12:35 20+ 105 111 42.3 5 9.91 33.9
EAV 1 10:15 0.5 0.3 146 58.4 20 7.09 24.5
Inion 2 11:30 0.2 0.2 164 59.5 49 6.89 24.7
Inflow Region
SV 1 11:50 0.8 0.6 172 67.7 56 7.22 24.8
SAV
S____2 12:10 0.1 0.6 154 59.1 30 7.33 25.0
20 cm
EAV 1 10:28 0.5 0.3 150 58.5 25 7.00 24.6
on 2 10:55 0.2 0.2 140 59.4 40 6.94 24.2
Outflow Region
SV 1 12:28 0.1 0.7 176 60.5 7 7.30 25.2
12SAV0.5 0.7 162 60.7 5 7.34 25.4
2 12:55 0.5 0.7 162 60.7 5 7.34 25.4


Rep


SRP
g L-1


Temp









Fluctuating water levels exposed the Hydrolab probes to air twice during

the two-week deployment, and only a subset of the data (July 9 July 17, 2002)

was used in this discussion. Differences in bottom elevation between stations

meant that water depth was on average 0.3 m deeper at the SAV station than at

the nearby emergent station (Figure 2-6).

Water quality monitoring in the Cell 1 outflow region revealed similar

temperature, dissolved oxygen saturation and specific conductance levels

between emergent and SAV communities (Figure 2-7, Figure 2-8, and Figure 2-9).

The emergent community, however, had pH levels roughly 1 pH unit lower than

the submerged community (Figure 2-10). Average ( one s.d.) pH level from July

9 16, 2002 was 8.05 0.17 in the SAV community, as compared to 7.10 0.34 in

the emergent community. Such a difference in pH may result from inorganic

carbon uptake by SAV for aquatic photosynthesis.

Conversion of HCO3- and CO2 into organic cell components likely

depleted these constituents despite the evidence of well-mixed conditions

provided by other parameters. Calcium carbonate equilibria can buffer water

column pH around 8.3. In the SAV community, sustained underwater

photosynthesis would tend to increase local pH levels, drive CaCO3

precipitation, and create a CaCO3 -buffered environment.

The lack of a diel pattern in pH levels suggests that during the high flow

event in June, little photosynthesis was occurring at either station. The high flows

may have suspended bottom sediment, and water depths were substantially








5.0


4.5


4.0


3.5


3.0


2.5


4/1 5/1 6/1 7/1


,1000
. 800
600
400
200


4/1 5/1 6/1 7/1 8/1
Date (2002)


Figure 2-6. Inflow rates and stage level of water in the outflow region of Cell 1
compared to bottom elevations at the emergent and submerged
aquatic vegetation (SAV) stations, prior to and during peeper and
Hydrolab deployments.

increased. The dissolved organic carbon in wetland surface waters can attenuate

light in a few meters depth (Krause-Jensen and Sand-Jensen, 1998). While D.O.


8/1 9/1 10/1 11/1


9/1 10/1 11/1










-SAV -EAV









I/> I I II


7/10


7/11


7/12


7/13 7/14 7/15


7/16


Date

Figure 2-7. Water column temperature within Cell 1 emergent and SAV
communities during the Hydrolab monitoring period (July 3 17,
2002).


100


80


60


40


20


0


7/10 7/11


7/12 7/13 7/14 7/15


Date


Figure 2-8. Water column dissolved oxygen saturation levels within Cell 1
emergent and SAV communities during the hydrolab monitoring
period (July 3 16, 2002).


-SAV -EAV












I I I I A I


7/16










1300


1200


1100


1000


900


800


7/10 7/11


7/12 7/13 7/14 7/15 7/16


Date


Figure 2-9. Water column specific conductance levels within Cell 1 emergent and
SAV communities during the hydrolab monitoring period (July 3 16,
2002).


9.0

8.5

8.0


.r 7.5

7.0

6.5

6.0


7/10 7/11 7/12 7/13 7/14 7/15


7/16


Date



Figure 2-10. Water column pH levels within Cell 1 emergent and SAV
communities during the hydrolab monitoring period (July 3 16,
2002).


SAV -EAV




i i-ii- i-


SAV -EAV







i* i-- i i**









concentrations did increase from 0 to 20% saturation during daylight hours at

each station, this is lower than the levels observed in SAV communities under

more typical, quiescent conditions (DBEL, 2001) or as observed in the mesocosms

(Table 2-2). Oxygen concentrations in June were lower than those observed on

September 29, 2002 at the same Cell 1 stations (Table 2-3).

In September, D.O. concentration profiles at the SAV station revealed

supersaturated conditions in the SAV surface waters (10.6 mg L-1, 31.8C) and

lower levels (0.65 mg L-1) near the bottom of the water column (0.65 m from

water surface and 0.1 m above the sediment surface). Emergent-stand surface

waters were less saturated than the SAV station with respect to D.O. (2.15 mg L-1,

31.5C), but waters at depth were similar (0.60 mg D.O. L-1). Surface water pH

levels at that time were 8.29 and 7.66 for SAV and emergent stations,

respectively.

Light penetration into the water column is essential for aquatic

photosynthesis. On September 29, 2002, below-surface light levels were reduced

in emergent stands relative to open water and SAV beds (Figure 2-11). In the

SAV and open water areas, light available just below the surface was

substantially reduced (-65%) from ambient light 2 m above the surface. Light

reflecting off the water surface may have increased the "ambient" measurements

and decreased the submerged surface measurements. Further light reduction

within the open water column may have resulted from attenuation and

scattering by dissolved organic matter or suspended particulate matter.









Table 2-3. Water column dissolved oxygen (D.O.) concentrations, pH, and
temperature profiles in the STA-1W Cell 1 surface waters at the time of
peeper retrieval on June 27, and on September 29, 2002.


EAV SAV


Parameter


Depth


June


Sept.


June


Sept.


Dissolved oxygen, mg L-1 Surface 0.30 2.15 1.1 10.6

Mid 0.25 0.55 1.1 3.1

Bottom 0.15 0.60 0.60 0.65

pH Surface 7.50 7.66 7.38 8.29

Mid 7.33 7.34 7.26 7.34

Bottom 7.47 7.15 7.22 7.26

Temperature, C Surface 31.0 31.5 30.3 31.8

Mid 29.7 28.9 29.3 29.0

Bottom 29.7 29.1 28.3 28.8


Water Depth, m


1.10


0.74


1.45


0.70


The Typha canopy drastically reduced the amount of light available just

below the water surface. SAV tissues reduced available light by shading the

waters below the leaf canopy. Interesting to note, however, is the near total light

extinction (>99% attenuated) at bottom depths (62 65cm) in both SAV and

emergent communities, regardless of the shallow water column (75 cm). The

phototrophic benthos at the open water stations, in contrast, had 7.7 0.3 % (as

mean 1 s.d.) of the incident photon flux available as light energy.










Fraction Incident Light Remaining (%)
0 5 10 15 20 25 30 35 40 45
0

10 -

S20 /

30 -
/ --SAV

40 Typha

A 5- Open Water
50
SAV shade
60 / Typha shade
T A
70


Figure 2-11. Incident light (measured as photon flux) remaining at water depths
of 0+, 35 and 65 cm in emergent and submerged macrophyte
communities, as well as within open water reaches, of STA 1W, Cell 1,
in September 2002. The shaded regions represent the contribution of
SAV and Typha canopy shading to light reductions, beyond the
attenuation attributed to the water column alone.

Even in shallow (
and a portion of the water column may be below the euphotic zone. Light

availability controls aquatic photosynthesis and alters surface water chemistry.

The influence of light and aquatic photosynthesis on water column P dynamics,

then, is greater near the air-water interface when the area is vegetated, while the

remainder of the water column is influenced by respiration processes.






48


Porewater Chemistry

Significant rainfall within the S-5 basin during the 17-day equilibration

required pumping of ADW into STA-1W. Water levels increased approximately

0.9 m between deployment June 10, and retrieval June 27, 2002 (Figure 2-12). At

the time of retrieval, recent inflow waters had likely influenced surface water

chemistry (Table 2-3).


5.0


4.5 Cell 1 Stage


S4.0 -


0 3.5
0 EAV Bottom Elevation
z
3.0 SAV Bottom Elevation


2.5 -
6/10 6/17 6/24

Date (2002)

Figure 2-12. Stage level recorded during the seventeen day equilibration period
for porewater samplers deployed in the outflow region of STA-1W Cell
1, during June 2002. The bottom depths for the SAV and Emergent
vegetation stations are shown for reference.

Profiles of porewater pH levels associated with emergent and submerged

vegetation were circumneutral (6.77 -7.57) in the both sediment types (Figure

2-13) and below the pH levels in the overlying water column. While pH values










pH
8.0 6.0


6.0


-25 I I

Figure 2-13. Vertical profiles of porewater pH values from soil collected from
emergent and submerged vegetation. Porewater equilibrators were
deployed for 17 days in June 2002, in the outflow region of STA-1W
Cell 1.

were measured immediately after peeper retrieval, the disparity between

replicates may be an artifact of oxygen reintroduction into, and/or CO2

outgassing from, porewater samples. However, the disparity is more likely due









to the natural spatial variability of the soils. All three replicate peepers were

retrieved at once at each site, and the first station's samples were processed

before the second station peepers were withdrawn. Those replicates processed

first had lower pH values than replicates sampled last. All pH determinations

were made within 4 hours of peeper retrieval.

Profiles of SRP concentrations were different between emergent and

submerged community sediments (Figure 2-14). Emergent aquatic vegetation

(EAV) sediments exhibited slightly higher mean porewater SRP concentrations

(0-6 cm depth, 996 131 gg L-1) and lower overlying water column

concentrations (0+6 cm depth, 402 293 pg L-1) than the submerged plant

communities (0-6 cm depth, 728 279 tg L-1; 0+6 cm depth, 616 256 tg L-1). The

resulting diffusive P flux was greater from the emergent sediments (0.39 0.11

mg P m-2 d-1) than from SAV sediments (0.07 0.05 mg P m-2 d-1).

Porewater DOP concentrations in EAV soils ranged from 0 to 350 tg L-1,

with one measurement at 658 ptg L-1, 18-20 cm below the soil surface (Figure

2-15). The SAV soil DOP profile ranged from 0-830 ptg L-1, with maximum

concentrations 2-6 cm below the soil surface in all three replicates. DOP may

diffuse upward into the water column, or downward into deeper soils, where

further transformations may increase P recalcitrance, or conversely make it

bioavailable. Surface soils also contained greater fulvic- and humic acid-

associated P in both soil types than the muck soils lower in the soil profile

(Figure 2-5). Microbial biomass and activity have been reported higher in WCA-










2A surface soils than in deeper soils (White and Reddy, 2001), and may influence

both the porewater DOP pool and the soil fulvic- and humic acid-P pools

through release of metabolic by-products or mineralization of organic matter.

Water columns at both EAV and SAV stations were likely anoxic during

the 17-day peeper equilibration period. High flows (Figure 2-6) likely


SRP (mg L1)
0.0 0.5 1.0 1.5 2.0 0.0 0.5 1.0 1.5 2.0
25 0' 1
EAV SAV
20
*Water column at time of
deployment
15
OWater column at time of
retrieval
10

E

** 0
*" /

S-5

-10


-15

-20

-25

Figure 2-14. Vertical profiles of soluble reactive phosphorus concentrations in
soil porewater collected from emergent and submerged vegetation.
Porewater equilibrators were deployed for 17 days in June 2002, in the
outflow region of STA-1W Cell 1.











DOP (mg L1)
0.0 0.2 0.4 0.6 0.8 1.0 0.0
25 ---
EAV
20


15 -


10 -


5 -

0


-5


0.2 0.4 0.6 0.8 1.0


Figure 2-15. Vertical profiles of dissolved organic P concentrations in soil
porewater collected from emergent and submerged vegetation.
Porewater equilibrators were deployed for 17 days in June 2002, in the
outflow region of STA-1W Cell 1.

resuspended surface sediments, increasing light attenuation and decreasing

photosynthesis. Dissolved oxygen concentrations on June 27, 2002 (retrieval)

were low even in surface waters (0.30 and 1.1 mg L-1) within the SAV and EAV

communities, respectively (Table 2-3).









Under low-oxygen conditions, Fe-oxides can be reduced to soluble Fe2+

This was observed, as dissolved Fe concentrations increased with depth through

the emergent macrophyte porewater (Figure 2-16). The profile in submerged

community was more uniform with depth. All concentrations were low

compared to those typical (50-100 mg L-1) of reduced mineral soil environments

(Patrick and Khalid, 1974). Newman and Pietro (2001) reported similarly low

total Fe concentrations (0.15-0.25 mg L-1) in STA-1W Cell 4 in 1993-1994 surface

water samples taken just after field flooding. Even under oxygenated conditions,

such low Fe concentrations may not provide the P sorption capacity

characteristic of acidic mineral wetland soils (Richardson, 1985).

Calcium, specific conductance and especially alkalinity concentration

profiles all increased with depth in the EAV profiles, but were constant or

decreasing with depth in the submerged profiles (Figure 2-17 to Figure 2-19).

Calcium carbonate precipitation in surface waters within the SAV community

may have led to elevated values, while dissolution of the underlying limerock

may have influenced the porewater concentrations in both soil types. The EAV

community porewater exhibited a clear gradient of alkalinity concentrations,

suggesting upward diffusive flux of these components through the soil profile,

and potentially into the water column.

Calcium carbonate precipitation is concomitant with aquatic

photosynthesis in hard waters. Everglades ADW was often saturated with

CaCO3 before entering STA 1W, as observed in mesocosm inflow waters.






54


Dissolved Fe (mg L1)


0.00
25 -


0.05


0.10


0.15 0.00


0.05


0.10


0.15


Figure 2-16. Vertical profiles of dissolved iron concentrations in sediment
porewater collected from emergent and submerged vegetation.
Porewater equilibrators were deployed for 17 days in June 2002, in the
outflow region of STA-1W Cell 1.

Daytime aquatic photosynthesis elevated pH levels in the mesocosms, and the

water column became supersaturated. The formation of CaCO3 precipitates is

therefore an important mechanism for P sorption capacity of newly accrued soils.









Saturation index values indicated CaCO3 precipitation was favored in the

porewater of both community types and at all depths (Figure 2-20). Over the

long term, water column CaCO3 precipitation may be less important to soil

sorption capacity than previously thought. Differences in the water column


Calcium (mg L1)

0 100 200 300 0 100 200 300
25 -

EAV SAV
20


15


10






Q -5



-10-


-15


-20


-25

Figure 2-17. Vertical profiles of dissolved calcium concentrations in sediment
porewater collected from soils below emergent and submerged
vegetation. Porewater equilibrators were deployed for 17 days in June
2002, in the outflow region of STA-1W Cell 1.









Alkalinity (mg CaCO3 L1)
0 200 400 600 800 1000 0 200 400 600 800 1000


Figure 2-18. Vertical profiles of alkalinity concentrations (as CaCO3) in sediment
porewater collected from soils below emergent and submerged
vegetation. Porewater equilibrators were deployed for 17 days in June
2002, in the outflow region of STA-1W Cell 1.

chemistry observed for submerged and emergent macrophyte stands were not as

apparent in the porewater. Influence of the native calcareous substrata on

emergent community porewater may account for P storage in the Ca-bound,

HCl-extractable P pool in those soils, despite the lack of evidence for calcium










Specific Conductivity (pS cm1)

0 1000 2000 3000 0 1000


2000


3000


Figure 2-19. Vertical profiles of specific conductance values in sediment
porewater collected from soils below emergent and submerged
vegetation. Porewater equilibrators were deployed for 17 days in June
2002, in the outflow region of STA-1W Cell 1.

carbonate precipitation in the EAV water column. Water column CaCO3

precipitation in the SAV beds probably explains the higher HCl-bound P pool

found in those soils (271 80 mg P kg-1) than those of the Typha-region soils (119

45 mg P kg-1).











CaC03 SI
100 0.1


0.1
25

20

15


-5

-10

-15

-20

-25


10 100


Figure 2-20. Vertical profiles of calcium carbonate saturation index (SI) values in
porewater collected from soils beneath emergent and submerged
vegetation. All index values were greater than one, indicating
supersaturated conditions with respect to CaCO3 Note log scale on x-
axis.
Intact Core P Flux Study

Soils collected from beneath the Typha and SAV communities released P

into overlying water under dark, oxic conditions (Figure 2-21). Initial floodwater

was particle free, and contained 13 pg DOP and 3 gg SRP L-1. After 28 days of









incubation, SRP concentrations were highest in two of the three replicate waters

overlying SAV soils (98 60 gig L-1 for all three SAV cores). The cattail soils also

released P into the water column at 49 49 itg L-1.


0)
.-
C'
Cn
c
E
03
o
I-
C)
(0


180

160

140

120

100

80

60

40

20

0


0 5 10 15 20 25 30

Elapsed Time (days)


Figure 2-21. Soluble reactive phosphorus (SRP) concentrations in the water
columns of SAV- and Typha- region soils and control columns (no soil)
during 28-day dark laboratory incubation. Error bars for each
treatment indicate one standard deviation between triplicate cores.
Note: Typha soil treatment values are from duplicate cores, while
control and SAV soil values were from triplicate cores.

Of the three un-amended cattail cores, one replicate had visible fragments

of leaf litter within the surface soils. The results observed from this core were

excluded from flux calculations, due to the possible effects of the litter fragments

on the results. Phosphorus movement from the soil to the water during the 7-20

day period of linear SRP concentration increase was greatest from SAV soils.









Corresponding flux rates from SAV soil, adjusted for change in control column

concentrations, was 1.84 1.04 mg P m-2 day-1. Emergent soils showed slightly

less P release (Table 2-4), though the difference was not significant (p > 0.05).

Near-zero (-0.02 to -0.03 mg P m-2 day-1) flux was observed in the control cores.

Water column pH and temperature values stayed consistent throughout

the 28-day incubation. Initial flood water pH (8.06) increased slightly to 8.36

0.06 before the 8-hour sampling, and remained similar in all treatments (8.41

0.09) during the remaining 28 days. Temperature was moderated by the

incubation water bath, and averaged 24 3 o C during the incubation period.

Table 2-4. Flux estimates from intact SAV and Emergent soil cores kept under
dark conditions for 35 days. The 30 cm water column was
continuously aerated. A 100 gg L-1 phosphorus spike was added after
30 days to measure short-term uptake rates. All values are means ( st.
dev.) of triplicate cores minus mean flux in control cores, in mg P m-2
day1.
Period of flux estimate SAV soil EAV soil
Release Day 7 to Day 20 1.84 1.04 1.20 + 0.27
Uptake 0 8 hours after spike 4.97 12.6 0.17 2.7


Similarly, dissolved calcium (Ca) concentrations were stable after a short

initial equilibration period (Figure 2-22). Alkalinity concentrations increased

marginally in the Typha soil treatment over the 28 days (Figure 2-22). SAV soil

and water control treatments were constant with respect to dissolved calcium

and alkalinity concentrations.






61


Dissolved organic P concentrations were slightly elevated in SAV-soil

incubations, relative to control treatments or treatments with Typha soils (Figure

2-23). During the incubation period, mean DOP concentrations for both soil


-4- EAV soil


-e- SAV soil


--- Water control


0 5


10 15 20
Elapsed Time (days)


Figure 2-22. Mean dissolved calcium and alkalinity concentrations in the water
columns of three treatments (see text for details) during 28-day dark
laboratory incubation. Error bars indicate 1 standard deviation from
the mean of three replicates.


>
E

0

mo



0
0






Ej
CO

(0r


120



100



80



60
320

300

280

260

240

220

200









treatments decreased slightly over the first 24-hours, then increased for the next

27 days. After 28 days, water column DOP concentrations above SAV soils (28 +

3 gg L-1) were higher than in waters above cattail soil (22 3). These

concentrations are very low relative to those observed in the porewater (e.g. as

high as 830 pg L-1 in the surface 2-6 cm SAV soils), which were also higher in the

SAV soils (Figure 2-15). Concentrations of dissolved organic carbon (DOC) were

similar during the period of linear P flux (monitored only between 4 and 14 days

of incubation) (Figure 2-24).


40
E O SAV

, 30 EAV

O O Water Control

| 20 -
o




0


0.1 1 2 20 28

Elapsed Time (Days)

Figure 2-23. Mean dissolved organic phosphorus concentrations in the water
columns above SAV and Typha-region soils and in control (water only)
columns during 28-day dark laboratory incubation. Error bars indicate
1 standard deviation from the mean of three replicates.










Greater P flux from SAV soils compared to the Typha-region soils, though

insignificant (p> 0.05), was supported by greater total and exchangeable P in the

SAV soils (Figure 2-5). The potential diffusive P flux of 0.39 mg P m-2 day-1 in the

Typha region, as calculated from porewater gradients, was a small contributor to

overall flux (1.20 mg P m-2 day-l).


46
45 EAV soil
m SAV Soil
44
4O E Water Control
S43
42
E 41
0 40
O
S39 -
38
37
36
4 7 10 14

Elapsed Time (days)

Figure 2-24. Dissolved organic carbon concentrations in water columns over
Typha soils, SAV soils and control columns (no soil) after 4, 7, 10 and 14
days of dark laboratory incubation. Error bars indicate 1 standard
deviation from the mean of three replicates.

Rapid biological P uptake occurred during the 8 hours following P

amendment (100 gg P) (Figure 2-25). Control columns showed a small decrease

in SRP concentration, then leveled off near 60 ptg L-1. Phosphorus in the control






64

waters may have been absorbed by bacteria in the water column or attached to

core walls, aerator, etc., and converted to non-reactive phosphorus.


E
0
0
I-


120



90



60



30



0


0 6 12 18 24 30 36 42

Time after P04 Amendment (hr)


48 54


Figure 2-25. Water column SRP concentrations above a "background"
concentration as determined before a 100 gg spike was added to each
column. Background concentration was assumed stable through the
53-hour period. Triplicate water columns were assembled with or
without soils from cattail or SAV communities and aerobically
incubated for 30 days prior to the P-spike amendment.

The change in the control water concentration was assumed to be an

experimental artifact and was subtracted from the observed concentration for

each treatment. The effects of soil type on P flux could then be evaluated directly.

In the eight hours following P additions, SAV soils showed greater potential for

P uptake than the EAV soils (Table 2-4). The EAV soil cores provided almost no

uptake capacity above that observed for the control cores. This may be indicative









of higher P sorption capacity and the influence of CaCO3 in the SAV surfaces

soils, compared to EAV soils.

Implications for STA and WCA Management

In both STA-1W and WCA-2A, recently accreted soils were enriched in P

relative to deeper soils (Craft and Richardson, 1993; Reddy et al., 1993; Reddy et

al., 1998; Koch-Rose et al. 1994; Newman and Pietro, 2001). In this study, SAV

soils were enriched relative to soils from EAV areas within STA-1W Cell 1.

Surface-soil P enrichment creates a potential for diffusive flux both upward into

the water column, as well as downward into underlying soil material. Upward

flux can increase outflow water P concentrations and reduce wetland removal

efficiency. Downward flux, on the other hand, may be a beneficial process for

long-term storage.

SAV soils (0-2 cm layer) contained more P in stable pools (705 68 mg P

kg-1) than EAV soils (555 111 mg P kg-1) after eight years of flow-through STA

operations. Labile pools of exchangeable and Al- and Fe-bound P were similar in

both surface-soil types (84 8 mg P kg-1 for SAV; 83 15 mg P kg- for Typha).

Because of similar labile-P pools, the two soil types each released P to an oxic

water column at similar flux rates, with slightly greater flux rates coming from

the SAV soils. SAV soils also reduced P concentrations following water column P

amendments, whereas Typha soils did not.

Within the submerged community, uptake mechanisms for water column

P reduction are mostly associated with soil sorption and leaf surfaces. Leaves are









concentrated towards the top of the canopy to intercept light, and senesce from

older, lower portions of the plant. A Najas canopy submerged under increasing

water depths would isolate water beneath the submerged canopy and close to

the soil surface from the overlying water column. This lack of water exchange

above the soil-water interface could allow bottom water to reach P

concentrations similar to those of the porewater through diffusion.

Removal mechanisms were not effectively maintaining water column

concentrations lower than the porewater in SAV beds during the high flow event

of June 2002. Net removal of SRP did not occur in the water column within 5 cm

of the soil-water interface, which in effect increased the diffusion distance and

reduced the overall flux rate. Had the peeper incubation period occurred under

conditions of a shallow water column, biological P uptake near the soil-water

interface may have resulted in a higher P flux.

Potential diffusion rates appeared to account for only a portion (4-33%) of

the flux measured in intact cores. Other studies estimating P flux in wetlands

have found diffusive flux rates lower than intact core flux or in situ benthic

chamber measurements (Fisher and Reddy, 2001). Burrowing macroinvertebrate

activity may cause advective water exchange from soils into overlying water.

Additionally, uptake of water column P near the soil surface by macrophytes and

their epiphytes, phytoplankton and benthic microorganisms is necessary to

maintain strong concentration gradients over short diffusion distances. Inorganic









retention of P through adsorption and Fe precipitate formation also influence the

distribution of dissolved P along the water column-porewater gradient.

In Everglades soils, soil respiration rate (as CO2 production) and

decomposition of litter/detritus was correlated to soil P content along a nutrient

enrichment gradient in the northern Everglades (Davis, 1991; DeBusk and

Reddy, 1998; Qualls and Richardson, 2000). Thus increased P mobilization may

be expected in P-enriched areas, relative to areas of lower P levels. The rate of

organic matter mineralization for soils along a eutrophication gradient in WCA-

2A was controlled primarily, however, by the availability of electron acceptors

(White and Reddy, 2001), of which oxygen is the most efficient. The nutrient-

impacted region of WCA-2A is frequently anoxic due to high sediment oxygen

demand and shade-induced limitations of aquatic photosynthesis. Therefore,

organic matter mineralization rates may be higher in P-enriched areas dominated

by submerged macrophytes that are capable of supplying 02 to the water

column, than in those areas dominated by dense Typha stands.

The nutrient-impacted areas of WCA-2A and STA-1W are near

monospecific stands of Typha. While SAV communities seem to be outcompeted

in these inflow regions, they do exist in STA-1W further downstream. Data

collected from other porewater studies in STA-1W (DBE, unpublished data)

along with the results presented here, suggest diffusive flux rates decrease with

distance through the wetland for the SAV communities (Figure 2-26). Such a

relationship is useful for predicting internal P loading at various locations along








the treatment gradient. However, the diffusive flux is likely only a fraction of the

total mobilization of P from soils to overlying water.


0.8
08 SAV EAV
'E y = 2.25e-8.19x
S0.6 Cell 1 Out 2= 0.74

E '
0.4 -
Cell 1 In Cell 4 In

i 0.2 -
0. Cell 4 Out

S0.0o
0.0 0.2 0.4 0.6 0.8 1.0
Fractional Distance through STA-1W

Figure 2-26. Phosphorus diffusive flux estimates based on 22 porewater
equilibrators deployed in STA-1W Cell 1 (June-July 2002; closed
markers) and Cell 4 (November-December 2001; open markers), as a
function of distance through the entire wetland. Diamond markers
indicate SAV; square markers indicate EAV. The regression line was
calculated using only SAV diffusive flux rates.

Emergent vegetation diffusive P flux did not show similar dependence on

fractional distance. This difference may have been due to the influence of root

uptake on soil porewater SRP concentrations, which was greater for soils in the

outflow Typha-dominated region than for soils from dense Typha stands near the

inflow region. Porewater SRP concentrations in the upper 10 cm of emergent









sediments were much lower in the inflow region (0.18 mg P L-1) than near the

Cell 1 outflow (0.99 mg P L-1).

The Cell 1 inflow emergent station may also have been isolated from the

primary flow paths through the dense emergent in the inflow region, whereas

the outflow station was adjacent to open water. This difference in proximity to

flow paths may have subjected the inflow-region Typha station to lower P loads

than the outflow-region Typha station. These data suggest that the internal P load

from porewater diffusion depends on vegetation, distance from the inflow, and

the proximity to the main flow paths through the wetland.

Conclusions

Managers of treatment wetlands must consider options for reducing

internal P loading, which can occur through diffusion and other processes (e.g.

resuspension, dissolution), by creating stable soil-P pools. At a minimum,

internal P loading should be minimized in outflow regions of the STAs to

achieve low TP concentrations in surface water outflows. Options may include

maximizing natural processes of new soil humification and Ca-P mineral

formation, or enhancing stability through chemical additions. Areas dominated

by SAV may provide a benefit during early stages of wetland creation and soil

development through water column CaCO3 precipitation, which appeared to

increase the Ca-bound P pool in SAV soils, relative to EAV soils. Soils beneath

EAV were P-depleted relative to soil P concentrations shortly after flooding,

which suggests that pools operationally defined as "more recalcitrant" are still






70


subject to mobilization. Over time and especially in inflow regions, high rates of

soil accrual may necessitate mechanical soil removal to reduce internal P loads,

which occurred regardless of vegetation type.













CHAPTER 3
BIOMASS PHOSPHORUS STORAGE AND DYNAMICS
Introduction

Tracts of the northern Everglades were transformed into three water

conservation areas (WCAs) in the 1940-50's by surrounding the vast marshes

with earthen berms (Figure 3-1). These areas provided water storage and flood

control for urban and agricultural development in South Florida. To the

northwest of the WCAs lies the Everglades Agricultural Area (EAA), a 200,000 ha

region of drained Everglades soils in sugarcane and winter vegetable production.

WCA-2A has received agricultural drainage water (ADW) discharges

since 1955 (Bartow et al., 1996). Excessive phosphorus (P) loading from these

discharges has been identified as the primary cause for an observed

eutrophication gradient. Changes observed near the discharge structures include

increased water column, soil, and plant tissue P concentrations, and change in

ecosystem function, relative to the interior marsh (Craft and Richardson, 1993;

DeBusk et al., 1994; Reddy et al., 1993; Reddy et al., 1998).

In order to reduce P loads to the Everglades, stormwater treatment areas

(STAs) were constructed to intercept P in EAA drainage waters before it enters

the WCA marsh. The STAs have performed well since construction, reducing TP

concentrations to below 50 gg L-l as required by the Everglades Forever Act










(EFA) of 1994. The EFA mandates further P load reductions, requiring TP

concentrations of 10 gg L-1 for discharges into the WCAs.

FLORIDA, USA


N

W*E

S


Figure 3-1. The historic Everglades region of south Florida is now three distinct
parcels, including the Everglades Agricultural Area (EAA), Water
Conservation Areas (WCA) and Everglades National Park. Stormwater
treatment areas (STA) are shown (in gray), including STA-1W.


I

i









Macrophytes influence P removal performance in treatment wetlands

through direct P uptake as well as by influencing the physicochemical

characteristics of its environment. Typha domingensis can grow in dense stands

when nutrient and hydrologic conditions allow. Phosphorus enrichment in the

northern Everglades has increased Typha above-ground biomass and shoot

density, and slough communities of WCA-2A were replaced by Typha near

inflow structures (Wu et al., 1997; Miao et al., 2000). Dense Typha stands and

open water areas differ in rates of water exchange and community metabolism

(Belanger et al., 1989; Brix, 1997; McCormick et al., 1997). In WCA-2A, Typha

canopy shading of the water column has reduced aquatic photosynthesis and

wind-induced mixing, leading to reduced oxygen supply relative to nearby open

water areas (Belanger et al.,1989; Grimshaw et al., 1997). Dense stands of

macrophytes can have large accumulations of organic materials, either settled

from flowing water or produced on-site. These accumulations increase

heterotrophic oxygen demand which, coupled with canopy shading, can result in

anoxic conditions (Belanger et al., 1989; McCormick et al., 1997).

Also important to P dynamics in wetlands are microbial processes in the

soil and water column, which respond quickly to system inputs or

environmental change. Qualls and Richardson (2000) reported Typha litter in

WCA-2A flumes sequestered SRP up to 10 times the original litter content over a

one-year period, with little net change to the macrophyte P content. Microbial

biomass on the leaf litter increased in P content, yet the factors controlling









microbial P uptake and release from wetland biomass remain unidentified.

Specifically, the effect of community metabolism on biomass P uptake within

Typha leaf litter is unknown.

As an open water community transitions towards a dense macrophyte

stand, such as occurred in WCA-2A, P cycling and removal processes within the

water column are likely affected. A similar increase in Typha stand density likely

occurs within the STAs due to high rates of nutrient loading, yet the effects of

water column shading by Typha on P cycling process remains unknown.

Through field observations, mesocosm experiments, and controlled laboratory

incubations, P storage, uptake and release were examined from Typha tissues as

well as from microbial populations associated with Typha leaf litter. Specific

objectives included:

* Estimating the relative partitioning of P between live and dead
components of Typha stands (Mesocosm Phosphorus Storage)

* Investigating the role of leaf litter accumulations on P release from the soil
(Flux Study Using Intact Soil Cores)

* Quantifying the rate of P uptake from the water column by microbial
populations associated with Typha litter under oxygenated and anoxic
conditions (Litter Incubations)

Materials and Methods

Bench-scale studies took place at the Wetland Biogeochemistry Laboratory

at the University of Florida in Gainesville, FL. Outdoor mesocosms were located

next to the STA-1W inflow canal on an experimental platform provided by DB

Environmental. Field investigations took place in STA-1W Cell 1.









STA-1W Site Description

Agricultural runoff from the Everglades Agricultural Area (EAA) Basin S-

5 (Figure 3-1) is pumped in canals to STA-1W, a full-scale (2699 ha) treatment

wetland operated by SFWMD to reduce P loadings to the Everglades (SFWMD,

2003). Everglades muck soils were drained for agricultural production decades

ago, and were reflooded in 1994. The wetland contains three flowpaths of two

cells each (Figure 3-2). Cell 1 is the first cell of the eastern flowpath, and contains

emergent vegetation (Typha spp.) in the inflow region and along the eastern edge

(Figure 3-3). A mixture of emergent, floating, and submerged vegetation occupy

the downstream reaches of the cell. This distribution of community types has

remained relatively constant in Cell 1 since the wetland was first flooded

(Newman and Pietro, 2001). During eight years of flow-through operations, the

wetland has accrued new wetland soils.

Mesocosm Phosphorus Storage

Two outdoor mesocosms (2.2m L x 0.79m W x 1.0m D) located near STA-

1W received STA-inflow water for 2.7 years at a hydraulic loading rate of 10 cm

day-1, as part of another study (DBEL, 2001). Muck soils in these systems were

inoculated with Typha collected from within STA-1W.

The vegetation and soils in the two cattail mesocosms were sampled

destructively in late August and early September 2001, at the end of the water

quality monitoring period (December 29, 1998 through August 14, 2001). All

vegetation was collected from each mesocosm and segregated into inflow and








outflow regions. Cattail shoots were cut at the soil surface, and separated into

live (green) and dead shoots. Triplicate soil cores (38.5 cm2 each) were retrieved

from the inflow and outflow regions. The underlying muck soil was discarded.




STA 1W (2699 ha)


5B


Loxahatchee NWR
Water Conservation Area 1



D Northern Flowpath
D Eastern Flowpath
f Western Flowpath


Figure 3-2. Stormwater Treatment Area 1 W in Palm Beach county, Florida. Three
flowpaths receive surface water drained from adjacent agricultural
soils and reduce the phosphorus load to Water Conservation Area 2A.
Cells 1, 2 and 3 are comprised of mixed emergent, submerged and
floating vegetation, while Cells 4, 5A, and 5B are primarily submerged
and floating vegetation. Arrows indicate general direction of flow.



































Figure 3-3. Aerial photograph of STA-1W Cell 1, first cell of the eastern flowpath,
in November 2000 (courtesy SFWMD). Also shown is the location of
the emergent vegetation field station.

Belowground cattail biomass and submerged macrophyte tissues were

collected after soil sampling was complete. Wet weights for vegetation samples

were recorded in the field. Bulk density was determined for accrued soils.

Vegetation and soil samples were oven dried (65C), weighed, and homogenized

for analysis. TP content was determined by digesting 50 mg sample in

concentrated nitric acid, followed by perchloric acid digestion at 210 C (COE 3-

227, Plumb 1981).









Phosphorus mass removals were calculated for each mesocosm based on

inflow and outflow TP concentrations of weekly or bi-weekly grab samples and a

constant hydraulic loading rate of 10 cm day-' (DBEL, 2001). The P mass

recovered in accrued soil, together with the change in biomass P storage, was

compared to mass removal based on water column concentration reductions.

Plant Tissue Desiccation Study

Biomass components within the water column of a treatment wetland are

largely comprised of live SAV in submerged macrophyte beds, and both live and

dead shoots in emergent stands. Samples of fresh Najas tissue and dead Typha

shoots (leaf litter) were collected from the outflow region of Cell 1 on May 29,

2002. Triplicate vertical acrylic columns (7 cm i.d.) received 13 g tissue (wet wt.),

along with 0.75 L of reflood water collected from the STA-1W Cell 4 outflow

weir. A pair of control columns contained no plant tissue. All treatments and

controls were placed in a water bath, continuously-aerated, and covered with an

opaque shroud. Water samples were withdrawn at times 0, 1 and 9 days, and

filtered (0.45 pm). SRP and acidified TSP samples were analyzed colorimetrically

(potassium antimony tartrate, sulfuric acid, ammonium molybdate, ascorbic

acid) on a Spectronics Genesys 5 spectrophotometer (EPA 365.2; EPA 1979).

Analysis for TSP was preceded by persulfate acid digestion and neutralization.

Flux Study Using Intact Soil Cores With and Without Typha Litter

Six replicate intact soil cores were retrieved from a Typha stand

(26.6292N, 80.4219 W) in the outflow region of STA-1W Cell 1 on July 4, 2002.









Each core was retrieved by inserting an acrylic core tube (38.5 cm2) through the

newly accrued soil layer into the underlying native farm muck to a minimum

depth of 10 cm. The top end of the core was sealed with a #1312 rubber stopper

prior to extraction. Each of the six intact cores was completely filled with site

water and sealed with a rubber stopper. An opaque shroud minimized solar

heating and blocked light during transport to the lab.

Typha leaf litter was collected from STA-1W Cell 1 at the time of soil core

collection, and was kept in site water on ice for transport to the lab. Litter

consisted of intact shoots that had little to no visible damage (from grazers,

breakage, etc.), yet had become neutrally- or negatively buoyant. This phase of

shoot decomposition was chosen because of three attributes:

* Fresh live shoots would likely leach tissue P rapidly after being collected
from the source plant

* Ample time in the water column for colonization by aquatic
microorganisms

* Negative buoyancy facilitates the experimental incubation design

Each shoot was divided into uniform lengths of -5 cm. Pieces from all

shoots were mixed at random, then added to six cores, three with soil and three

with reflood water only (litter controls). Overlying water was replaced with 1.15

L filtered (0.45 pm) STA-1W Cell 4 outflow water (15 gg TSP L-1). The 30 cm

water column was aerated and incubated in a dark water bath at -22C for 28

days.









Water samples (30 mL) were withdrawn at At = 0.1, 0.5, 1, 1.5, 2, 4, 6, 10,

14, and 20 days, and analyzed for SRP and pH. Representative samples were also

analyzed for TSP, dissolved calcium, total alkalinity, and dissolved organic

carbon (DOC).

Dissolved calcium was determined using flame atomic absoption

spectroscopy (EPA 215.1; EPA 1979) on a Perkin-Elmer 3110. Alkalinity was

titrated with 0.02N H2SO4 (EPA 310.1; EPA 1979). Dissolved organic carbon was

on acidified, filtered (0.45 pm) samples, and measured with a Shimadzu TOC-

5050A (Duisburg, Germany) TOC analyzer equipped with an ASI-5000A

autosampler (5310-A; APHA, 1992).

Sample pH was recorded immediately following collection, using a 3 in 1

gel filled combination pH electrode and Corning 313 pH meter. Water bath

temperature was continuously recorded by a StowAway Tidbit logging probe

(Onset Computer; Bourne, MA) as well as monitored periodically with a

thermometer.

After 30 days, an amendment of 100 tg P L-1 (as KH2PO4) was added to

each core. The water volumes above each core differed slightly ( 5 mL) from the

original water volume of 1.15 L added one month prior, likely due to different

evaporation rates induced by the aerators. These differences were recorded but

volumes were not adjusted at that time. Water samples (30 mL) were withdrawn

At = 0, 4, 8, 24 and 53 hours after the amendment, and analyzed for SRP. Each

core received 30 mL of P-unamended reflood water after sampling.









Typha Litter Incubations

Through a series of oxic and anoxic incubations, P uptake and release

rates were determined for freshly submerged Typha litter and associated

microbial biomass. Senescent submerged plant tissue from a stand of Typha -30

m in diameter (25 50.275' N; 80 42.991' W) were collected in southern WCA-3A,

and kept in site water on ice for transport to the lab. Directly south of WCA-2

and the EAA, west of urban Miami and north of Everglades National Park

(Figure 3-1), WCA-3A is a mosaic of pristine Everglades ecotypes, from wet

sloughs dominated by Nymphaea spp. and Utricularia spp. to sawgrass (Cladium)

ridge and tree island communities. The slough surrounding the Typha stand was

dominated by Nymphae oderata, has historically low water TP and SRP levels of 9

2 pg L-1 and <2 pg L-1, respectively (Swift and Nicholas, 1987), and was the

source of water for the incubations.

Part I. P uptake by Typha litter

Incubation water was filtered (0.45 tpm) and amended with KH2PO4 to

concentrations of 0, 10, 30, 100 and 300 and 1000 tg P L-1 above background SRP

concentrations. This concentration range was representative of the P gradient to

which litter is exposed: low-P surface water to high-P porewater.

Short acrylic columns received 25 g litter (wet wt.) and 0.50 L P-amended

water (Figure 3-4). Cores were sealed with rubber stoppers at each end. The top

stopper was penetrated by two glass tubes which permitted sparging with air for

oxic trials, or N2 + 0.03% CO2 for anoxic trials. Gas flow mixed the water column









during sparge (0.5 3hrs), after which gas lines were clamped shut. Water

samples were withdrawn after 0.3, 1, 2, 3, and 5 days of incubation, and analyzed

for SRP, pH and DOC.

Dissolved oxygen concentrations were monitored periodically during

sampling, and measured for each replicate at the end of the incubation to ensure

experimental conditions were achieved.

Sampling port E
Sparglinine in Purge line out

















Figure 3-4. Schematic of incubation design, containing Typha litter and P-
amended surface waters from WCA 3A. Not to scale: Water volume
was 0.5 L, head space was 0.05L.

Part II. P release from enriched litter

After seven days of incubation, the incubation waters for each litter

sample were replaced by un-amended site water. Water column SRP

concentrations were measured immediately (At=0), and at 1, 2, 3, and 5 days after

the water exchange. Water SRP concentrations were used to calculate oxic and









anoxic P uptake and release rates for Typha litter across the range of initial water

P concentrations.

Part III. Microbial Biomass Phosphorus

At the conclusion of the release phase, Typha litter subsamples of -10 g

wet weight were subjected to either 24-hour CHCb fumigation and 16 hour 0.5 M

NaHCO3 (pH 8.5) extraction on a reciprocating shaker, followed by

centrifugation (10 min, 6000 rpm), or direct extraction without fumigation.

Additional P extracted after fumigation was considered the contribution of the

microbial biomass phosphorus (MBP) pool that was present and susceptible to

chloroform (Hedley and Stewart, 1982). Typha tissue samples were analyzed in

their original condition (preserved field-moist in the dark at 3.5C) for

comparison.

Extracts were analyzed for SRP and TP using the ascorbic acid-

molybdenum blue method (EPA 365.2; EPA 1979) on a Technicon Autoanalyzer.

TP samples were prepared by persulfate digestion at 150C, increasing to 380C,

prior to analysis.

Statistical Methods

Statistical analyses on experimental data were performed using MSExcel

(v. 2000 Microsoft Corp.) ANOVA and t-test macros. Error around mean values

is presented as one standard deviation of replicate measurements.









Results and Discussion

Mesocosm Phosphorus Storages

The following narrative describes observed changes to vegetation in the

mesocosms, though no quantitative measurements of biomass occurred prior to

August 2001. In February of 1999, the surface of the water in rep 1 was covered in

Lemna, which likely reduced light penetration and gas exchange to the water

column. Lemna persisted throughout 1999 in rep 1, with none noted in rep 2.

Dense mats of floating Lemna minor were shown by Ngo (1987) to inhibit

phytoplankton growth by shading the water column, causing algae to die and

settle out. Water column chemistry can also be affected below dense mats of

floating macrophytes. Low pH and oxygen levels can develop, and increased

nutrient levels may result from internal loading (Gopal et al., 1984).

Greater Najas biomass was observed in rep 2 than rep 1 in December 1999.

Through the first year of operation, rep 2 appeared to outperform rep 1 with

respect to water column P reduction (Figure 3-5). From December 29, 1998

through December 21, 1999, mean inflow TP concentration (106 Rg L-1) was

reduced to mean outflow concentrations of 70 and 33 pg L-1 for rep 1 and rep 2,

respectively.

Lemna may have created anoxic conditions and internal P loading from the

initial muck substrate, while Najas in rep 2 may have provided oxygenated water

column conditions which can minimize soil-P release (Gonsiorczyk et al., 2001).

By September 2000, rep 1 appeared to have more Najas, while rep 2 had








350
' 300 -o- Inflow
) 250
3 200
0.. 150
C 100
0 50
I-
S0

Jan May Sep Jan May Sep Jan May Se
99 99 99 00 00 00 01 01 01

200 -(273) (240)
180 Outflow
160 -+ Cattail Rep 1
SCattail Rep 2
140
) 120
100
80
o 60
40
20
0
Jan May Sep Jan May Sep Jan May Se
99 99 99 00 00 00 01 01 01

Month

Figure 3-5. Inflow and outflow TP concentrations for mesocosms dominated by
Typha and operated from December 1998 through August 2001.


p


p









developed an underwater periphyton mat. At the time of the whole-mesocosm

biomass sampling in August 2001, no periphyton was found in either mesocosm,

and rep 1 contained more Najas (289 g dry wt.) than rep 2 (12.4 g dry wt.). Mean

surface water TP concentration reductions were greater for the latter mesocosm,

which had a more dense Typha canopy (rep 2), though variability in vegetation

and P removal performance occurred throughout the period of record (Figure

3-5). It is not clear whether fluctuations in biomass and species composition of

SAV, floating plants, and periphyton within and between the two cattail

mesocosms contributed to variability in P removal performance.

Based on mean inflow and outflow water TP concentrations, the two

cattail mesocosms removed 1.2 (rep 1) and 2.0 (rep 2) g P m-2 yr-1 of an average

annual loading rate of 3.5 g P m-2 yr-1. Phosphorus removal resulted in an

increased biomass P pool and accrued soils. Of the total P mass recovered in

biomass and new soils, 66 (rep 1) and 96% (rep 2) was attributed to TP removal

from the water column. Unfortunately, the change in P in the initial muck

substrate was not quantified for the period of record. The difference between

water column P mass removal and P recovered in biomass and new-soil P

reflected experimental error inherent in sampling, but P mobilization from the

initial muck substrate (Table 3-1) was also likely. The substrate in one mesocosm,

therefore, may have had only a minor net P contribution (0.2 g P m-2), while the

other mesocosm substrate provided nearly a third (1.5 g P m-2) of the P mass

accrued in biomass and new soil storage.









Typha P concentrations for live tissues, dead leaves and root/rhizomes

were comparable to those reported for the Everglades (e.g. Toth, 1988) (Table

3-2). Belowground cattail tissues (roots+rhizomes) comprised 39% and 65% of

the total recovered biomass P in rep 1 and rep 2, respectively (Figure 3-6).

Previous Everglades studies estimate 30-40% of total Typha biomass P was

associated with belowground tissues in P-enriched areas, while 60% was

belowground in low-P environments (Toth, 1988; Miao and Sklar, 1998).

The disparity in the below-ground cattail P storage between the replicates

in this study was echoed by the above-ground standing crop (Figure 3-6).

Greater biomass P storage observed in replicate 2 agreed with greater reductions

between inflow and outflow water TP concentrations over the 2.7-year study

period.


Table 3-1. Phosphorus removed from the water column over the 2.7-year study
period and recovered in the vegetation (Typha + Najas) and soils upon
termination of the study on August 20, 2001.
Rep 1 Rep 2 Average
Biomass Storage, g P m-2 1.2 1.8 1.5 (29%)

New Soil Storage, g P m-2 3.6 3.9 3.7 (71%)

Total Recovered P, g P m-2 4.9 5.6 5.2 (100%)
Inflow-Outflow Water
3.2 5.4 4.3


Mass P Removed, g P m-
Potential net P contribution
from muck substrate into
biomass and new soils, g P m-2


1.6 0.2 0.9


1








O Typha roots U Typha live shoots U Typha dead shoots E Najas
1.0
0.9 -
0.8
0.7 -
0.6
0.5
0.4
0.3
0.2
0.1
0.0 R- I-
Rep-1 In Rep-1 Out Rep-2 In Rep-2 Out


* New Soils


D Biomass


2.0


1.5


1.0


0.5

nn


- i =


Rep-1 In


Rep-1 Out


Rep-2 In Rep-2 Out


Figure 3-6. A: Partitioning of recovered Najas and Typha tissue P within the
inflow (In) and outflow (Out) halves of two mesocosms operated from
December 1998 through August 2001. B: Total P mass recovered (solid
line) in biomass and sediments relative to annual mass removal rates
(dashed line) calculated from water column TP concentration
reductions.


A









Table 3-2. Dry matter and P concentrations in accrued soil and tissue storage
(live and dead Typha shoots, below-ground Typha (roots and
rhizomes), and Najas tissues) retrieved from mesocosms on August 20,
2001. Typha tissue P concentrations are comparable to values reported
by Toth (1988).

Below- A N Newly
Live Typha Dead Typha o All Najas e
ground A. N accrued
shoots shoots Tissues s
Typha soil
(g dry m-2)
Rep 1 416 1163 745 41 3932
Rep 2 466 2141 1616 1.8 5426
(mg P kg-1)
Rep 1 790 180 620 5300 924
Rep 2 710 190 630 3300 747
Toth 1988 580-1000 160-260 570-740

High biomass in rep 2 caused the belowground portions of the Typha

stand to become space limited, and root structures extended upwards into the

water column. These structures may have provided a pathway for direct P

uptake from the water column, more so in rep 2 than in the less dense rep 1,

since greater P reduction was observed over the 2.7-year period of record in the

mesocosm with greater cattail density and less SAV biomass. Root and rhizome

biomass in rep 2 (1.6 kg m-2) was more than double that of rep 1 (0.75 kg m-2), but

root TP concentrations were similar (Table 3-2).

Additionally, rep 2 had twice the dead Typha biomass (2.1 kg m-2) that was

present in rep 1 (1.2 kg m-2). Dead biomass in rep 2 may have provided substrate

for a larger microbial community than rep 1, and increased P removal from the

water column in that mesocosm. Live above-ground Typha biomass in this study

was similar in the two mesocosms (0.42 and 0.47 kg m-2, for rep 1 and rep 2