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Published in J. Environ. Qual. 33:2357-2366 (2004).
© ASA, CSSA, SSSA
677 S. Segoe Rd., Madison, WI 53711 USA

TECHNICAL REPORTS

Wetlands and Aquatic Processes

Response of Biogeochemical Indicators to a Drawdown and Subsequent Reflood

R. Corstanje* and K. R. Reddy

Wetland Biogeochemistry Laboratory, Soil and Water Science Department, University of Florida Institute of Food and Agricultural Sciences, 106 Newell Hall, P.O. Box 110510, Gainesville, FL 32611

* Corresponding author (Corstanje{at}mail.ifas.ufl.edu)

Received for publication February 25, 2004.

   ABSTRACT
TOP
 ABSTRACT
INTRODUCTION
 MATERIALS AND METHODS
 RESULTS
 DISCUSSION
 REFERENCES
 
Temporal oscillations in hydrology are a common occurrence in wetlands and can result in alternating flooded and drained conditions in the surface soil. These oscillations in water levels can stimulate microbial activities and result in the mobilization and redistribution of significant amounts of carbon (C), nitrogen (N), and phosphorus (P). The goal of this study was to experimentally simulate a drawdown and reflood of marsh soil from a nutrient-enriched site and a reference site of a wetland (Blue Cypress Marsh Conservation Area, Florida). The goal was to better understand the changes in biogeochemistry and microbial activities present in these soils as a result of hydrological fluctuations. Measurements of dissolved reactive phosphorus (DRP), ammonia, and nitrate in the floodwater indicated significantly higher ({alpha} = 0.05) NH4+ and DRP fluxes from the nutrient-enriched site; floodwaters in the cores from both sites contained significant NO3– concentrations (9.6 mg N L–1), which was rapidly consumed over the core incubation period (30 d). Water level drawdown and reflooding initially stimulated the soil microbial biomass, methanogenic rates, and extracellular enzyme activities (acid phosphatase and ß-glucosidase). The anaerobic microbial metabolic activities (CO2) where initially significantly ({alpha} = 0.05) enhanced by the reflood, resulting in roughly equivalent rates as the aerobic respiratory activities (CO2), presumably as a function of the high water column NO3– levels. This study illustrates that the reflood event in the hydrological cycles in a wetland can significantly stimulate the activities of hydrolytic enzymes and microbiological communities in these soils.

Abbreviations: APA, acid phosphatase activity • ßGA, ß-glucosidase activity • DRP, dissolved reactive phosphorus • MBC, microbial biomass carbon • MUF, methyl-umbelliferone • TP, total phosphorus


   INTRODUCTION
TOP
 ABSTRACT
 INTRODUCTION
MATERIALS AND METHODS
 RESULTS
 DISCUSSION
 REFERENCES
 
HYDROLOGY, SPECIFICALLY THE PRESENCE, quantity, quality, and timing of water defines wetland ecosystems. The hydrological regime affects the vegetative communities, generates most of the hydric soil characteristics, and attracts wildlife characteristic to wetland ecosystems. Water-level changes at different magnitudes, frequencies, timing, and duration are common in natural wetlands (Olila et al., 1997). The effects on plant communities and microbial habitat differ in response to these fluctuations (Wilcox et al., 2001). Intermittent flooding and draining of wetland soils result in considerable temporal variability in the soil redox potentials, and the spatial variability is further compounded by rhizosphere oxygenation by wetland macrophytes (Lorenzen et al., 2001). Wetland soils therefore are relatively rich in functional microbial communities that are capable of utilizing a large range of electron acceptors, such as O2, NO3–, Fe(III), SO42–, and CO2 (D'Angelo and Reddy, 1999; Wright and Reddy, 2001). The absence of O2 under saturated conditions and the presence of alternate electron acceptors can control the rates of microbially mediated organic matter mineralization (McLatchey and Reddy, 1998; D'Angelo and Reddy, 1999).

Controlled water level oscillations in a wetland resulted in both a suppression of the microbial respiratory activity (Freeman et al., 1995) and no apparent effect on these activities (Freeman et al., 1996). Extracellular enzyme activities (sulfatase, ß-glucosidase, and phosphatase) were significantly stimulated by an experimental drawdown of a peatland (Freeman et al., 1996). The increase in organic matter mineralization in this system was therefore seen as the result of higher hydrolytic enzyme activities versus an overall increase in metabolic activities (Freeman et al., 1996). The effect of experimentally controlled variation of the redox potential on organic matter mineralization has been documented in soil slurries (McLatchey and Reddy, 1998; D'Angelo and Reddy, 1999; Wright and Reddy, 2001). Nutrient release experiments in marsh and lake sediments have been conducted both on soil slurries (Moore et al., 1998) and with intact cores (Holdren and Armstrong, 1980; Olila et al., 1997), in which the depth of the overlying water is controlled. Drying and rewetting processes are expected to generate a sequence of alterations to the microbial communities, concurrent with and affecting the changing soil properties and water column nutrient concentration. We found in a parallel study (Corstanje, 2003) that a drawdown in a subtropical marsh system resulted in significant stimulation of microbial activities and ß-glucosidase activities.

In the present study, we hypothesized that in a marsh system, drought and subsequent reflood remobilize and redistribute significant amounts of labile P, N, and C. We conducted a controlled drawdown and subsequent reflood experiment using the type of intact cores currently used in nutrient flux studies (Olila and Reddy, 1995) to monitor the nutrient dynamics and microbial community responses to changing redox levels as a result of soil flooding.


   MATERIALS AND METHODS
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
RESULTS
 DISCUSSION
 REFERENCES
 
Site Description and Sampling Protocol
Blue Cypress Marsh Conservation Area (BCMCA) is located at the headwaters of the St. Johns River in south-central Florida (Corstanje, 2003). The marsh, an 8000-ha subtropical freshwater system, is surrounded to the east, south, and west by levies and delimited to the north by Blue Cypress Lake (Fig. 1) . The marsh received two significant nutrient influxes from the surrounding agricultural lands in the northeast and southwest. Most of the nutrient influxes were diverted from the marsh in the early 1990s. This study focused on two particular areas of the marsh, a nutrient-enriched area in the northeast and a reference area in the northwest. Both areas had previously been the subject of a seasonal study (Corstanje, 2003) and the cores used in this study were obtained in the same locations in September 2002. The nutrient-enriched site had documented high levels of P (D'Angelo and Reddy, 1999; Olila and Reddy, 1995) whereas the reference location was chosen close to a St. Johns River Water Management District water quality sampling site. The nutrient-enriched site was characterized by a relative monospecific stand of cattail (Typha spp.); cores from the reference site were taken in an area dominated primarily by switch grass (Panicum spp.). The cores were 60-cm Plexiglas tubes with 10-cm inner diameters. The tubes were placed on the soil surface including any detrital material present. An indenture of the tube diameter was made and the detrital material not in tube was cut away. The tube was reinserted and while exerting a slight pressure and rotating the tube, the peat surrounding the tube was cut first with a knife and subsequently with a peat cutter, attempting to minimize soil compaction. At least 20-cm-deep soil cores were obtained creating a total sample volume of at least about 1500 cm3. A total of 26 cores were collected. The cores were capped and placed in a temperature-controlled greenhouse (approximately 22°C).



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  		Fig. 1. Geographic location of the Blue Cypress Marsh Conservation Area and approximate position of the sampling sites.

 
Incubation and Water Column Analysis
The cores were left capped in the greenhouse for a period of 8 mo to stabilize with no water column present. During this time the soil surface was colonized by a mixture of swamp-loosestrife [Decodon verticillatus (L.) Elliott], nutsedge (Cyperus spp.), and spike-rush (Eleocharis spp.) in all cores, and with maidencane (Panicum hemitomon Schult.) in the cores obtained in the reference site. Most of these species were either indigenous to the sampling areas (e.g., switch grass) or were noted present on the soil surface during a preceding drought at the Blue Cypress Marsh Conservation Area (e.g., swamp-loosestrife). When water levels recede, less competitive species such as swamp-loosestrife are able to grow from dormant seeds and propagules, complete at least one life cycle, and replenish the seed bank before being replaced through competitive interactions (Wilcox et al., 2001).

Two cores, one from each site, were removed from the greenhouse and brought to the laboratory (approximately 20°C). The bottoms were sealed and the cores were placed in a water bath. Triplicate redox probes were placed at the soil surface and 5- and 10-cm depth intervals, and connected to a data logger (CR10; Campbell Scientific, Logan, UT). The two cores were reflooded and maintained with a constant 30-cm water column from the soil through the periodic addition of site water for 30 d (Fig. 2) . The tanks were covered with a light foil that excluded light. The redox profiles over time by depth obtained for both cores were used to establish the destructive core sampling regime at 10, 16, and 30 d after reflooding. Upon completion of this preliminary experiment, all cores were randomly distributed over two water baths and flooded with site water at a 30-cm depth. Flooding of these cores was executed by wrapping the bottom ends with a 5-mm fiberglass mesh, slowly immersing the cores in tubs containing site water and then sealing the bottom ends. This approach was chosen to approximate reflooding by raising water tables. Three cores were excluded from the nutrient-enriched and reference sites, respectively (total = 6), which were destructively sampled before the initiation of the reflood. The cores were fitted with redox probes at the soil surface and 5- and 10-cm soil depths; water column pH and dissolved oxygen were monitored regularly throughout the experimental period (30 d).



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  		Fig. 2. Experimental setup of a sample core. Reflood was only executed once; the arrow is for illustrative purposes only.

 
The water column was sampled for DRP, NH4–N, and NO3–N on Days 1, 2, 4, 10, 15, 20, and 30 and total Kjeldahl N and total phosphorus (TP) on Days 1, 10, 15, and 30. Determination of DRP, NH4–N, and NO3–N required the water samples to be immediately filtered through 0.45-µm membrane filter paper after extraction. For DRP, filtered water samples were directly analyzed through automated colorimetric analysis technique (Method 365.1; USEPA, 1993). Water column TP was determined using an 5.5 M H2SO4 autoclave water digestion of unfiltered water samples and ascorbic acid analysis (Method 365.1; USEPA, 1993). The total Kjeldahl N, NH4–N, and NO3–N were analyzed colorimetrically (Methods 351.2, 350.1 and 353.2, respectively; USEPA, 1993). Three cores from each site were sacrificed at t = 10, 16, and 30 d into the experiment, and were cut into 0- to 5- and 5- to 10-cm segments and the overlying vegetation and detritus was sampled.

Analytical Methods
After removal of any visible live plant material, a coffee grinder was used to homogenize soil and vegetation samples. Soil bulk density was determined on a oven-dried (70°C), dry-weight basis, and the soils were subsequently ground. Total C and total N concentrations were determined with a Carlo-Erba NA 1500 CNS Analyzer (Haak-Buchler Instruments, Saddlebrook NJ), while TP was determined using the ashing method (Andersen, 1976) and analyzed by the ascorbic acid colorimetric procedure described by Kuo (1996) (Autoanalyzer II; Technicon, Tarrytown, NY).

Anaerobic and aerobic incubations were conducted on soil slurries of approximately 4 g of sample in anaerobic tubes with 10 mL of deionized, distilled water. Anaerobic soil slurries were subsequently actively purged with O2–free N2 and incubated horizontally in a longitudinal shaker to determine the anaerobic metabolic activities. Similarly, aerobic metabolic activities were determined by incubating 4 g of sample with 10 mL of oxygen-saturated water (8.2 mg O2 L–1). Analysis of headspace CO2 and O2 was done with a thermal conductivity detector at 30°C (Model 8AIT GC; Shimadzu, Kyoto, Japan) and headspace CH4 was analyzed by flame induction detection at 110°C (8AIF GC; Shimadzu). In order not to change significantly the dominant microbial communities as a result of the incubations, the incubations were performed over relatively short time frames (30 h). Upon terminating the incubation period, analysis of the headspace O2 concentrations and slurry dissolved oxygen concentrations was performed to ensure that aerobic conditions were maintained throughout the incubation.

Microbial biomass carbon (MBC) was determined on the soil samples by the chloroform fumigation incubation procedure coupled to a 0.5 M K2SO4 extraction (Vance et al., 1987; White and Reddy, 2001); the resultant dissolved organic C was determined on a Shimadzu TOC-5050A total organic carbon analyzer. The nonfumigated K2SO4–extracted dissolved organic C was reported as labile organic carbon (LOC). The difference between the fumigated and nonfumigated soil, corrected with an extraction efficiency factor kEC = 0.37 (Sparling et al., 1990), resulted in the MBC. The extracellular enzyme activities of ß-1,4-glucosidase (EC 3.2.1.21) and acid phosphatase (EC 3.1.3.1) were assayed (Prenger and Reddy, 2004) using a fluorescent artificial substrate methyl-umbelliferone (MUF-phosphate and MUF-ß-D-glucoside, respectively).

Calculation of Nutrient Flux
The nutrient flux (N and P) was calculated on the water column DRP, NO3–N, and NH4–N concentrations of the cores. The flux rate was calculated for the first four days over the two sites as this was the most significant period of P and N release. Therefore, the calculated flux rates are the maximum P and N release. The nutrient flux was calculated by determining the change in concentration vs. time, with linear regression, and then adjusting this by the floodwater volume and topsoil surface area ratio of the soil core:

[1]
where Ji is the flux of the nutrient constituent i (mg m–2 d–1), Ci is the concentration of component i in the floodwater (mg L–1), V is floodwater volume (2.3 L), A is area of the top of the soil core (78 cm2), and t is the time interval (days). As the nitrate flux was immediate (1 d), the flux was calculated as a single event versus a rate (i.e., dt was not included in the above formula) (Olila and Reddy, 1995; Fisher and Reddy, 2001).

Data Analysis
The rates of CO2 and CH4 production were analyzed as zero-order kinetic reactions and estimated as the coefficient of simple linear regression using Excel (Microsoft Corporation, 2003). Contrasts and comparisons were executed in JMP Version 4.0.2 (SAS Institute, 2001a) and SAS Version 8.2 (SAS Institute, 2001b), using a general linear model and, where appropriate, using repeated measures analysis of variance. Simple contrasts were executed as t tests, and the Tukey–Kramer adjustment (Kramer, 1956) was used for multiple comparison of means (all at {alpha} = 0.05 unless stated otherwise). As all the above-mentioned procedures carry the normality assumption, the data were examined for normality and homoscedacity of variance. Outliers were identified as observations that fell beyond the 1.5 ± interquartile range.


   RESULTS
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS
DISCUSSION
 REFERENCES
 
Soil Physicochemical Properties
The total P, C, and N soil (0–10 cm) concentrations between cores from the nutrient-enriched and reference sites did not differ significantly (Table 1; {alpha} = 0.05), with only slightly higher TP levels in the nutrient-enriched core. The final soil TP content was on average 784 (±78) mg P kg–1 in the nutrient-enriched soils and 686 (±126) mg P kg–1 in the reference soils on termination of the reflood; these values do not differ significantly from the original contents. Final soil total C and total N contents in the nutrient-enriched soil were 474 (±2) g C kg–1 and 33 (±5) g N kg–1 and in the reference soil were 471 (±10) g C kg–1 and 32 (±1.6) g N kg–1, respectively. These were not affected by the reflood experiment. The soil pH increased an average of 0.5 pH unit over the course of the experiment [4.98 (±0.06) and 4.90 (±0.17) at the start and 5.48 (±0.28) and 5.29 (±0.21) after 30 d for the reference and nutrient-enriched soils, respectively]. Soil bulk density and ash content decreased significantly (Table 2) as result of the reflood.


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  		Table 1. Selected soil physicochemical properties for the Blue Cypress Marsh Conservation Area nutrient-enriched northeast site and reference northwest site (n = 3; values in parentheses are one standard deviation).{dagger}

 

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  		Table 2. Changes in soil physicochemical properties as a result of the reflood of the Blue Cypress Marsh Conservation Area nutrient-enriched northeast site and reference northwest site (n = 3; values in parentheses are one standard deviation).

 
As the soils were flooded, the redox levels at the three soil depths (surface, 5 cm, and 10 cm) initially decreased (Fig. 3) at similar rates for the cores from both sites. In the deeper soils (10 cm), the rates of decrease of the redox potential occurred significantly faster compared with the shallower soil depths (5 cm) and at the soil–water interface. The soil–water interface stabilized within 15 d for the reference site at approximately 200 mV and at 20 d for the nutrient-enriched site at approximately 100 mV. The redox levels in the soil matrix at a 10-cm depth continued to drop continually throughout the experiment to about –100 mV for the both sites.



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  		Fig. 3. Soil redox (Eh) profiles over the experimental period (days) as a result of the reflood for the nutrient-enriched northeast site and reference northwest site. The symbols depict the redox potential at three different depths (diamonds represent the soil surface, boxes represent a 5-cm depth, and triangles represent a 10-cm depth).

 
Nutrient Flux and Dynamics
The water column dissolved oxygen levels oscillated between 0.5 and 1.5 mg L–1 throughout the course of the experiment with no significant difference between the cores from the two sites. Likewise, the water column pH remained relatively constant throughout the experiment with an average pH of 5.8 and 5.9 for the cores from both sites, respectively, with an average oscillation of one pH unit for both sites over the time period.

Nutrient release in this study is the result of the flooding approach, the presence of plants in the water column, and the manner in which the soil was dried. Typically, soil columns receded from the sides of the Plexiglas tubes when left to dry. The amount of nutrients (N and P) released from the soil into the water column as a result of a marsh reflood results in significant releases in DRP and NH4–N (Fig. 4) with respect to the original reflood water nutrient contents (Table 3).



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  		Fig. 4. Water column dissolved reactive phosphorus (DRP), ammonia (NH4–N), and nitrate (NO3–N) profiles over the experimental period in response to a reflood from below. The symbols depict the nutrient water column concentration for cores obtained at the two sites (boxes represent the nutrient-enriched northeast site and triangles represent the reference northwest site; error bars represent one standard deviation).

 

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  		Table 3. Chemistry of water sampled at the nutrient-enriched northeast site and reference northwest site in the Blue Cypress Marsh Conservation Area used as reflood water.

 
The soils sampled at the nutrient-enriched site released significantly higher levels of DRP: 109 (±56) mg P m–2 d–1 versus 6.5 (±3) mg–2 d–1 for the reference cores. Comparable NH4–N fluxes from the soil columns were estimated at 460 (±178) mg N m–2 d–1 and 109 (±56) mg N m–2 d–1 for the nutrient-enriched and reference cores, respectively. These calculations assume that the flux occurred over the top of the soil surface only. As there was some soil recession from the tube sides, the soil–water interaction during the reflood was presumably more than just over the top of the soil column. Correcting the above values for the total soil surface in interaction with the water column resulted in estimated DRP flux rates of 12 (±6) and 0.7 (± 0.3) mg P m–2 d–1 and in NH4–N fluxes of 51 (±23) and 7 (±3) mg N m–2 d–1 for the nutrient-enriched and reference cores, respectively. The latter probably better reflect the total release of the P from this soil as these soils were reflooded from below. Therefore, significantly more of the reflood water is in contact with the sides as would be the case when reflooded from the top, which is typically done in lake core studies (Olila and Reddy, 1995). After the initial increases in water column DRP and NH4–N concentrations, these generally showed gradual decreases. The initial reflood resulted in dramatic increases in NO3–N (from 0.02 to 9.5 mg L–1 in the cores from the reference site and 0.03 to 10.3 mg L–1 in the cores from the nutrient-enriched site) in the floodwater, which was estimated to be an equivalent of 134 and 144 mg m–2 for the reference and nutrient-enriched cores, respectively, over a single event. After this initial spike in the levels of NO3–N, they decreased exponentially in a similar fashion for both sites (Fig. 4).

Water column TP and total Kjeldahl N levels mirrored the dynamics presented by DRP and NH4–N and NO3–N. Initial (Day 1) total Kjeldahl N levels were 6.4 (±0.1) and 17.0 (±1.0) mg N L–1 for the reference and nutrient-enriched cores, respectively, decreasing to 4.3 (±0.2) and 12 (±4.0) mg N L–1, respectively, at the end of the experiment. Likewise, water column TP levels decreased from 0.23 (±0.01) and 2.1 (±0.1) mg P L–1 to 0.05 (±0.005) and 1.6 (0.9) mg P L–1 for the reference and nutrient-enriched cores, respectively, yet unlike the N dynamics, DRP remained a sizable fraction of the water column TP (60–80%) throughout the experiment, generally increasing over the duration of the experiment.

The variability in total nutrient contents precluded comparison across sites. The total amount of P released into the water column was estimated as a product of the maximum released and the total volume of water (3.5 mg P), which was well within the error margin associated with the TP values. Similarly, the total amount of N was estimated to 40 mg N, which is a fraction (>1%) of the margin of error associated with the total N values.

Microbial Community Response to Flooding Soils
The levels of aerobic respiratory activity did not change appreciably as a result of flooding or between the cores from the two sites (Fig. 5a) , averaging 0.46 (±0.04) µmol CO2 g–1 h–1 for the nutrient-enriched cores and 0.46 (±0.05) µmol CO2 g–1 h–1 for the reference cores. The surface soil (0–5 cm) strata had significantly higher activities than the deeper soils (5–10 cm) [i.e., 0.5 (±0.04) and 0.39 (±0.03) µmol CO2 g–1 h–1, respectively]. The highest anaerobic microbial activities were noted at the beginning (Day 1) of the flooding experiment and were roughly equivalent to the aerobic activities [i.e., 0.46 (±0.09) and 0.53 (±0.07) µmol CO2 g–1 h–1 for the soils from nutrient-enriched and reference cores, respectively] (Fig. 5b). By the end of the experiment, the levels of anaerobic respiratory activity had dropped to about 30 to 40% of the aerobic respiration levels, 0.18 (±0.03) µmol CO2 g–1 h–1 in the nutrient-enriched cores and 0.17 (±0.02) µmol CO2 g–1 h–1 in the reference cores. There were no significant differences in anaerobic respiratory activities by depth or by site, resulting in the changes in time (P = 0.0108) as the only significant factor.



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  		Fig. 5. Changes in (a) aerobic and (b) anaerobic microbial respiratory activity and (c) methanogenic activity over the course of the reflood study averaged over the entire soil core. The symbols depict the respective microbial activity for cores obtained at the two sites (boxes represent the nutrient-enriched northeast site and triangles represent the reference northwest site; error bars represent one standard deviation).

 
Whereas the overall levels of anaerobic respiration level dropped as the experiment proceeded, methanogenic activities generally increased over the course of the experimental period (Fig. 5c). Little to no CH4 production was detected in the soils from the cores destructively sampled at the initiation of the experiment (Day 1). Toward the end of the experiment, significant levels of methanogenesis were detected, with surface soils producing 0.56 (±0.28) and 0.21 (±0.05) µmol CH4 g–1 h–1 for the nutrient-enriched and reference cores, respectively. The deeper soils generated significantly less CH4, 0.06 (±0.03) and 0.04 (±0.008) µmol CH4 g–1 h–1 for the nutrient-enriched and reference cores, respectively. The methanogenic rates did not vary significantly by site, but showed a slightly (P = 0.104) significant interaction of depth by time.

Microbial biomass carbon did not differ significantly by depth (Fig. 6) ; however, the highest levels were noted in the surface layers in the nutrient-enriched cores, 9.0 (±3.4) g kg–1 on Day 10 and in the reference cores 8.2 (±3.9) g kg–1 on Day 16. This increase in MBC was muted in the deeper soils, with the MBC levels in nutrient-enriched cores peaking at 6.7 (±4.2) g kg–1 and in the reference cores at 5.3 (±2.3) g kg–1 on Day 16. The initial microbial biomass carbon in these soils was low (mean = 0.6 ± 0.2 g kg–1) and after peaking on Day 16, by Day 30 had dropped to an average of 2 to 4 g kg–1. On average, microbial biomass levels in the cores from the nutrient-enriched and reference sites did not show significant differences; however, the timing of the MBC peak in the different cores did result in a slight (P = 0.06) time by site interaction. Soil labile organic carbon content increased significantly as a result of the reflood (Fig. 6), after which it leveled off at about 5 g kg–1 in surface soils and 3.5 g kg–1 in deeper soils for most of the experimental period. The changes in the labile organic C levels over the experimental period differed for cores coming from different sites (reflood time by site; P = 0.015), a response primarily driven by the oscillation in labile organic C levels in the surface soils between nutrient-enriched and reference sites.



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  		Fig. 6. Soil labile organic carbon content (LOC) and microbial biomass carbon (MBC) content two depth intervals over the course of the reflood study. The symbols depict LOC and MBC content for cores obtained at the two sites (boxes represent the nutrient-enriched northeast site and triangles represent the reference northwest site; error bars represent one standard deviation).

 
The extracellular enzymatic activities associated with the decaying plant material increased over the course of the experiment (Fig. 7) , in which ß-glucosidase activity (ßGA) increased similarly over all cores irrespective of site. The initial levels of ßGA in plant material was relatively low, 0.54 (±0.25) and 3.3 (±2.9) µg MUF g–1 h–1 for the nutrient-enriched and reference sites, respectively. The final levels of ßGA were considerably higher, 40 (±2.7) and 38 (±17) µg MUF g–1 h–1 for the nutrient-enriched and reference cores, respectively. In contrast to the soil ßGA levels (Fig. 8) , the activity levels were significantly lower for the decaying plant material (µsoil = 80 ± 3; µplant = 27 ± 4 µg MUF g–1 h–1). The increases in acid phosphatase activities (APA) in the nutrient-enriched cores were slightly lower than at the reference site (Fig. 7). These differences were not significantly different when taken over the entire experimental period. The initial APA levels associated with the plant material were 4 (±2) and 7 (±4) µg MUF g–1 h–1 for the nutrient-enriched and reference sites, respectively, while the final APA levels were 123 (±22) and 121 (±19) µg MUF g–1 h–1, denoting significant increases in APA as a consequence of the flooding. The biggest difference in APA levels between the nutrient-enriched and reference sites was seen on Day 16 (66 ± 33 vs. 99 ± 23 µg MUF g–1 h–1). In comparing the phosphatase enzyme activities associated with plant material to that present in the soils, the levels present with the plant material were significantly lower than the soil enzyme levels (µsoil = 90 ± 39; µplant = 63 ± 10 µg MUF g–1 h–1).



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  		Fig. 7. Extracellular enzyme activities associated with the decaying plant material. The symbols depict the enzymatic activity for cores obtained at the two sites (boxes represent the nutrient-enriched northeast site and triangles represent the reference northwest site; error bars represent one standard deviation).

 


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  		Fig. 8. Changes in levels of extracellular enzyme activities by depth over the reflood period. The symbols depict the enzymatic activity for cores obtained at the two sites (boxes represent the nutrient-enriched northeast site and triangles represent the reference northwest site; error bars represent one standard deviation).

 
The levels of extracellular enzyme activities showed a parabolic-type response similar to the microbial biomass (Fig. 8), with the highest activities noted in the 10- to 16-d interval. The levels of ßGA showed no significant differences across depths, with 86 (±12) µg MUF g–1 h–1 at the 0- to 5-cm interval and 82 (±16) µg MUF g–1 h–1 at the 5- to 10-cm interval for the cores from the nutrient-enriched site, and 82 (±14) µg MUF g–1 h–1 in the surface layer (0–5 cm) versus 81 (±20) µg MUF g–1 h–1 in the subsurface layer (5–10 cm) for the cores from the reference site. The overall mean ßGA levels by site did not differ significantly (84 ± 14 and 82 ± 17 µg MUF g–1 h–1), the main dynamics of interest were the changes in time of ßGA in the cores from the two sites (P = 0.07; Fig. 8). The changes in APA levels over the experimental period seem to depict a similar parabolic response curve (Fig. 8), with the highest activities on Day 16. No significant differences were noted in APA dynamics over time between the cores from the two sites. The mean APA activity over all cores was significantly lower in the deeper soils, with 117 (±20) µg MUF g–1 h–1 in the surface layer (0–5 cm) and 71 (±40) µg MUF g–1 h–1 in the deeper soil layer (5–10 cm).


   DISCUSSION
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS
 DISCUSSION
REFERENCES
 
Cycles of drawdown–reflood occur naturally in wetlands, introducing oscillations of soil oxygenation. A sustained drawdown in a marsh can result in sediment mineralization and subsequent nutrient releases from the oxidized soils and sediments (De Groot and Van Wijck, 1993; Qiu and McComb, 1994; Baldwin, 1996; Olila et al., 1997; Michell and Baldwin, 1998; Fisher and Reddy, 2001). The largest portion of P flux is due to solubilization of accumulated nutrients, which are products of the enhanced mineralization under aerobic conditions (Ogwada et al., 1984).

Subsurface versus surface reflooding resulted in nutrient flux rates that were significantly higher than those reported in the literature. Studies with fluxes from sediments have ranged between <1 and 100 mg P m–2 (Watts, 2000; Pant and Reddy, 2001). Comparable rates in this study were 48 and 2.8 mg P m–2 for the nutrient-enriched and reference cores, respectively, when taken over the entire core surface. They are significantly higher when computed only over the top core surface (436 and 26 mg P m–2, respectively). These flux rates are similar to those found for wetland soils and sediments, which ranged from 1.5 to 334 mg P m–2 d–1 (Olila et al., 1997; Moore et al., 1998; Fisher and Reddy, 2001). Both flux rates have significance for the actual reflooding from the surface in the wetland system from which they obtained the cores, as organic soils tend to form cracks when dry.

The effect of decomposing macrophytes on increasing the water column nutrient content has been noted before (Moore et al., 1998) for a limited number of littoral zone cores where the decomposing plant material released significant levels of P. In contrast to the flux rates obtained in this study, Bostic and White (personal communication, 2004) obtained P-flux rates of 3.6 (±1.5) and 1.1 (±0.9) mg P m–2 d–1 in the nutrient-enriched and reference sites, respectively, in cores that included no plants. When standing plant material was included in the study, the overall DRP release rates increased to 26 (±31) mg P m–2 d–1 for both sites. In our experimental approach, we were unable to distinguish between the reflood approach and the presence of vegetation. However, the P-release rates in this study were relatively instantaneous, whereas it took 10 d to attain similar concentrations in a similar study at the Blue Cypress Marsh Conservation Area (Bostic and White, personal communication, 2004). This study indicates that a substantially faster pulse of nutrients is released when the reflood occurs through the subsurface versus a surface reflood.

Flooded soils typically release NH4+ in the soil porewater (Patrick and Mahapatra, 1968) as a result of the diffusion gradient between the soil and floodwater. In the case of a subsurface reflood, the mass transport of NH4+ into the floodwaters results in a relatively instantaneous spike in NH4+ (Fig. 4). In the presence of oxygen, the NH4 is then oxidized to NO3–. Given that NO3– is a significant electron acceptor in the absence of oxygen, the decreasing NO3– levels in the overlying water column can be either a result of microbial denitrifying activities in the water column or in the soil column with an ensuing flux of NO3– into the soil resulting in an exponential decrease in NO3– levels (Fig. 4). Fisher and Reddy (2001) experimentally dosed flooded cores with NO3– and found it was consumed immediately. Given the relatively high levels of NO3– initially found in the cores in the present study when contrasted to comparable studies (Reddy and Rao, 1983), it can be expected that the ensuing microbial community metabolic activities are at least initially dominated by NO3– reduction.

In wetlands, organic matter decay and nutrient regeneration have been summarized as a function of (i) substrate quality, (ii) hydrological regime and the supply of electron acceptors (e.g., O2, NO3–, SO42–), and (iii) environmental factors such as pH and temperature (Godshalk and Wetzel, 1978; Fenchel and Jorgensen, 1977; DeBusk and Reddy, 1998; McLatchey and Reddy, 1998; D'Angelo and Reddy, 1994; Wright and Reddy, 2001). Freeman et al. (1996) suggested that the enhanced decomposition following drawdowns may also be due to the reactivation of the extracellular enzyme activities that are responsible for organic matter decomposition. In the cores in this study, the levels of ßGA were slightly but significantly higher in this study when compared with enzyme activity levels in flooded soils from a previous study in the Blue Cypress Marsh Conservation Area (in 1999; Corstanje, 2003) and similar to those after a drought event (in 2000; Corstanje, 2003). Alkaline phosphatase is primarily regulated by presence of bioavailable phosphate (Newman and Reddy, 1993; Sinsabaugh and Moorhead, 1994; Wright and Reddy, 2001); the release of SRP into the water column might repress APA levels. All extracellular enzyme activities associated with plant material in the water column increased over the full experimental period, irrespective of the water column chemistry. Likewise, soil APA activities seemed to decrease only toward the end, indicating that if the enhanced levels of DRP were inhibiting, it was only toward the end of the experimental period. This lag in extracellular enzyme activity response to the changing environmental conditions might be significant as they are generally viewed as sensitive indicators of nutrient status (Sinsabaugh and Moorhead, 1994).

The reflood appeared to initially stimulate MBC levels, which were associated with a slight increase in labile organic carbon levels and ßGA. Soil microbial biomass has been shown to regulate transformation and storage of nutrients (Martens, 1995). The dynamics associated with MBC seem to indicate that the microbial community is initially stimulated by the reflood.

Drawdown in wetlands introduces oxygen into the sediment and decomposition rates under aerobic conditions have been shown to be significantly higher than under anaerobic conditions (McLatchey and Reddy, 1998; Wright and Reddy, 2001). DeBusk and Reddy (1998) found that anaerobic mineralization of wetland soils was one-third of the rate of aerobic soils. The initial respiratory activities in this study were not significantly different. Toward the end of the experimental period the anaerobic respiratory levels were one-third of the aerobic decomposition rates, presumably as a result of high NO3– levels initially present in the floodwaters. There was a slight increase in aerobic microbial activity over the course of the incubation, but the relative activities did not change substantially despite the fall in redox levels in the soils. In this particular system, overall anaerobic conditions were not attained until after Day 20, whereas methane production was detected by Day 10. Within the soil core matrix the conditions are presumably favorable for methanogenesis irrespective of the overall core and water column environmental conditions.

In summary, this core study resulted in significant nutrient releases from the cores taken from the nutrient-enriched areas and significantly lower release rates from the cores taken from the reference areas. The combination of the presence of decaying plants and reflooding from below resulted in relatively high flux rates when compared with other similar studies. However, we were not able to differentiate between the two factors. Our results show that subsurface reflooding will result in significant nutrient flux rates from these soils. We found that a drawdown–reflood event did stimulate extracellular enzyme activities and soil microbial biomass and activities. Drawdown and reflood events could remobilize and redistribute N and P from the localized nutrient-enriched areas over a wider area of the marsh.


   ACKNOWLEDGMENTS
 
Florida Agricultural Experimental Station Journal Series no. R-10143. This research was supported in part by the St. Johns River Water Management District. The authors would like to express their thanks to the analytical assistance provided by Ms. Yu Wang, and the two anonymous reviewers as well as the associate editor for their insightful comments.


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