Distributions and Budgets of Carbon, Phosphorus, Iron and Manganese in a Floodplain Swamp Ecosystem

Edward J. Kuenzler Professor of Environmental Biology

Patrick 3. Mu1 hol land* Research Assistant

Laura Anne Yarbro** Research Assistant

Leonard A. Smock*** Research Assistant

Department of Environmental Sciences and Engineering University of North Carolina at Chapel Hill Chapel Hi 11 , North Carol ina 27514

* Present Address: Division of Environmental Sciences Oak Ridge National Laboratory Oak Ridge, Tennessee 37830

** Present Address: Horn Point Environmental Laboratory Cambridge, Mary1and 21 61 3

*** Present Address: Department of Biology Virginia Commonwealth University Richmond, Virginia 23284

The work upon which this publication is based was supported in part by funds provided by the Office of Water Research and Technology, U.S. Department of the Interior, b!ashington, D.C., through the Water Resources Research Institute of The University of North Carolina as authorized by the Water Research and Development Act of 1978; in part by a Predoctoral , Fellowship from the National Science Foundation to L. A. Yarbro.

Project No. B-110-NC I Agreement No. 7 4-34-0001-8105

May 1980 ACKNOWLEDGEMENTS

This research was carried out with the generous assistance of many associates and colleagues. Robert Sniffen and Jeffery Koenings worked closely with us on many field trips and provided critical suggestions on numerous aspects of the study. Carol Parker, Shirley Wasson, Charles Page, and Thomas Smith assisted with laboratory analyses; George McRae provided assistance in the field. Sharon Shramm was very helpful with computer programming and Frank Malcolm with construction of field gear. Willow Baker of the N.C. Forest Service kindly provided daily precipitation data and Art Belk, also of the Forest Service, provided the opportunity for an aerial survey of Creeping Swamp. The Extension Service of Pitt County provided data on phosphorus fertilizer useage and crop acreage in the Creeping Swamp watershed. Special appreciation is due Margaret Schimert for her care in preparing the final typescript, Finally, several faculty members, Mark M. Brinson, Christopher S. Martens, Charles R. O'Melia, Frederic K. Pfaender, Seth R. Reice, and Richard A. Yarnell provided many he1 pful suggestions as we1 1 as critical judgement and unflagging support and encouragement.

DISCLAIMER STATEMENT

Contents of this pub1 ication do not necessarily reflect the views and policies of the Office of Water Research and Technology, U. S. Depart- ment of the Interior, nor does mention of trade names or commercial products constitute their endorsement or recommendation for use by the U. S. Government. ABSTRACT

A freshwater floodplain swamp ecosystem, Creeping Swamp, on the 1 ower Coastal Plain of North Carolina was the site of intensive, coordinated studies of system structure and functioning focused on organic carbon and phosphorus. The swamp, usually flooded in winter and early spring by darkly-colored water low in conductivity, pH, and plant nutrients, supports a typical deciduous hardwood bottomland forest. The investigations emphasized field measurements of the major compartments and of the dominant fluxes between these compartments in the natural swamp for two years. In addition the fluxes of carbon and phosphorus into and out of the swamp ecosystem were determined and material budgets were constructed. The abundances of iron and manganese in several physical and chemical forms in water and soils were also measured. Most of the organic C in the swamp ecosystem (22 kgem'2) was present in living trees and sap1 ings (63%), and almost all of the remainder occurred as swamp floor litter, logs, and soil organic matter (to $5 cm depth). Annual inputs to the swamp in 1977 amounted to 902 g Gem- ayr-l of which 21% was organic C in stream flows into the swamp and 77% was net primary productivity of trees, saplings, shrubs and herbs. Internal cycling of organic carbon, amounting to 352 g cem-?eyr-l, was comprised of litterfall (77%), macro1 i tter (I 7%), and throughfall and sternf low (6%). Organic C outputs from the swamp (541 g corn-2ayr-l)were partly provided by stream flow (40%) but mostly by litter respiration when the swamp was not flooded (45%), by aquatic benthic respiration (8%), water column respiration (5%), and anaerobic respiration (2%). The swamp exported organic C (about 20 g ~.m-~a~r-l),mostly as dissolved and colloidal humic material s.

Most of the phosphorus in the swamp ecosystem (about 39 g pornm2) was in the upper 25 cm of soil (86%), with the remainder mostly in aboveground vegetation. Annual P input to the ecosystem (1 050-1 2%) mg em-2.yr-1) was carried predominantly by surface waters (94%), with bulk precipitation delivering the rest. Internal P cycling consisted for the most part in filterable reactive P uptake from the floodwaters to algae and to the forest floor, and sedimentation; a substantial portion of this uptake supported the vegetation cycle from tree roots to canopy to litterfall and throughfall back to the swamp floor. Much of the particulate P in the floodwaters sedimented out on the swamp floor. The swamp floor also returned filterable unreactive P back to the water. There was marked retention of phosphorus by the swamp resulting in very low P concentrations in the water downstream. Iron and manganese exist in swamp waters and soils in a large variety of physical and chemical forms which change in concentration through the year. Largest amounts were associated with soil particles, next highest concentrations were in soil pore water, and lowest amounts were in the swamp floodwaters. Significant positive correlations existed between particulate P concentrations and unreactive particulate Fe concentrations in floodwaters.

TABLE OF CONTENTS

Page

ACKNOWLEDGEMENTS ...... ii

ABSTRACT ...... iii

LIST OF FIGURES ...... xi

LISTOFTABLES ...... xv

SUMMARY. CONCLUSIONS AND RECOMMENDATIONS ...... xix 1 . INTRODUCTION General ...... 1

Background ...... 1

Wetland Hydrology ...... 2

Organic Carbon Fluxes ...... 4

Phosphorus ...... 5 Organic Carbon. Phosphorus and Iron Interactions .... 7 Description of the Study Area ...... 9 Structureofthe Report ...... 13

ORGANIC CARBON CYCLING AND EXPORT

INTRODUCTION ...... 15

METHODS

Study Area ...... 18

Inundation Patterns ...... 18 Biomass ...... 19 Precipitation and Streamflow ...... 19 Organic Carbon Analysis ...... 20 Biologic Inputs ...... 20 TABLE OF CONTENTS (Continued) Page Biologic Outputs ...... 23 Particulate Organic Carbon Formation ...... 25 Dissolved Organic Carbon Leaching From Ground Litter ...... 25 Additional Methods ...... 26 RESULTS Inundation Patterns ...... 26 Living Biomass ...... 27 Swamp Floor Detritus ...... 31 Hydrologic Inputs and Outputs ...... 33 Biologic Inputs ...... 39 Biologic Outputs ...... 47 Particulate Organic Carbon Formation ...... 54 Leaching of DOC from Ground Litter ...... 57 Natural and Channelized Streams in Eastern North Carolina ...... 61 Effect of Spates on Organic Carbon Concentrations in Creeping and Clayroot Swamps ...... 64 DISCUSSION Organic Carbon Concentrations ...... 64 Productivity and Biomass ...... 68 Respiration ...... 71 Production/Respiration Ratios for the Aquatic Ecosystem ...... 76 Particulate Organic Carbon Formation ...... 76 TABLE OF CONTENTS (Continued)

Page Annual Organic Carbon Budget for Creeping Swamp . . 77 Organic Carbon Export from Watersheds ...... 83 Fate of Exported Organic Carbon ...... 88

Comparison Between Natural and Channelized Streams ...... 89 3 . PHOSPHORUS CYCLING IN THE FLOODPLAIN ECOSYSTEM AND EXPORTS FROM THE WATERSHED

INTRODUCTION

General ...... 91

Phosphorus Cycling in Wetlands ...... 91

Factors controlling phosphorus cycling in wetlands ...... 92 Phosphorus cycling in Creeping Swamp ..... 92

METHODS

Phosphorus Budget for Creeping Swamp ...... 94 Hydrologic measurements ...... 94 Water chemistry ...... 95 Hydrologic fluxes of phosphorus ...... 97 Phosphorus Cycling in Creeping Swamp Floodplain . . 97 Study area ...... 97

Floodpl ain hydro1ogy ...... 98

Phosphorus in herbs. shrubs. vines. and bryophytes ...... 98 Phosphorus in ground litter ...... 98 Soil phosphorus ...... 99 TABLE OF CONTENTS (Continued)

Page Throughfall ...... 99 Stemflow ...... 99 Litterfall ...... 100 Sedimentation ...... 100 Forest Floor-Water Exchanges of Phosphorus . . . , . 101 Flowing-water chambers -- 1977 ...... 101 Still -water chambers -- 1978 and 1979 . . . . . 103 RESULTS Water Chemistry and Hydrologic Fluxes Hydrology and inundation patterns ...... 109 Phosphorus in bul k precipitation ...... 113 Surface water chemistry ...... 114 Surface water exports of phosphorus from the Creeping Swamp watershed ...... 120 Ground water phosphorus concentrations and losses from the watershed ...... 124 Phosphorus Cycling in the Swamp Floodplain Ecosystem ...... 124 Annual surface water imports and exports of phosphorus ...... , ...... 126 Standing stocks of phosphorus in the floodplain ecosystem ...... 130 Fluxes of phosphorus within the swamp floodplain ecosystem ...... 131 Forest Floor - Floodwater Exchanges of Phosphorus . 141 FRP exchanges ...... 141 TABLE OF CONTENTS (Continued)

Page FUP exchanges ...... 144

Fate of 32~removed from the water . , . . . . 144

Biotic and abiotic components of phosphorus fluxes ...... 147 Annual estimates of floor-water exchanges of filterable phosphorus ...... 151 Algal uptake of FRP ...... 154 DISCUSSION

Phosphorus Budget for the Creeping Swamp Watershed Annual inputs and outputs ...... 154 Phosphorus Cycling in the Floodplain Swamp Ecosystem ...... 159 Phosphorus concentrations in swamp waters . . . 159 Standing stocks of phosphorus in the Creeping Swamp ecosystem ...... 161 Intrasystem transfers of phosphorus ...... 162 Forest floor-water exchanges of phosphorus . . 163 A budget of phosphorus cycling in Creeping Swamp ...... 168 Comparison to other wetland and forested ecosystems ...... 177 4. SEASONAL CHANGES IN FORMS AND SPECIES OF IRON AND MANGANESE IN SWAMP WATER AND SOILS INTRODUCTION ...... 181 METHODS Sampling Sites and Schedule ...... 183 TABLE OF CONTENTS (Continued)

Page Water Samples Collection and particle size separation ....183 Iron analysis ...... 184 Manganese analysis ...... 186 Forms of iron and manganese associated with suspended particulate matter ...... 186 Phosphorus and organic carbon analysis ....188 Field Measurements ...... 188 Soil Samples and Pore Water Soil samples ...... 189 Pore water ...... 189 RESULTS Water Quality Characteristics ...... 190 Iron ...... 190 Manganese ...... 194 Forms of Fe and Mn Associated with Suspended Particulate Matter ...... 196 Phosphorus and Organic Carbon Concentrations ....198 Pore Water ...... 198 Forms of Fe and Mn Associated with Swamp Soils ...... 202 DISCUSSION ...... 204 LITERATURE CITED ...... 209 LIST OF FIGURES Page

1. Map of the Study Region Showing Location of Streams and Sampling Stations ......

2. Mapof Creeping Swamp Study Area...... 3. Conceptual Model of Organic Carbon Cycling in a Swamp-Stream Ecosystem ...... 4. Seasonal Patterns of Mean Daily Fraction of Area Inundated for the Period 1973-1978 ...... Seasonal Record of Ground Litter Standing Crop from September 1976 to December 1977 ...... Annual Record of Streamflow and Floodplain Inundation at CP-10...... Seasonal Patterns of Fine Total Organic Carbon (FTOC) Concentration at Three Stations in Creeping Swamp and Streamflow at CP-10 from January 1975 to May 1978 . . Aboveground Net Annual Wood Increment in 1977 as a Function of Diameter at Breast Height (DBH) for Seven Common Species in Creeping Swamp ...... Seasonal Patterns of Average Daily Litterfall in Creeping Swamp for 1976 and 1977 ...... Seasonal Patterns of Macro-litterfall in Creeping Swamp from September 1976 to October 1977 ...... Weighted Mean Organic Carbon Concentration in Throughfall as a Function of Amount of Throughfall for the Dormant Season (November to March) and Growing Season (April to October) in Creeping Swamp ...... Seasonal Patterns of Gross Primary Productivity (GPP) of Aquatic Plants, Mostly Algae ...... Effect of Temperature on Rates of Terrestrial Litter Respiration in Creeping Swamp ...... Effect of Temperature on Rates of Total Soil Respiration in Creeping Swamp ...... LIST OF FIGURES (Continued)

Page

Seasonal Patterns of Water Column Respiration in Creeping Swamp from November 1976 to April 1978 .... 52 Effect of Temperature on Rates of Aquatic Benthic Respiration in Creeping Swamp for the (a) Cold Season (November to February), and (b) Warm Season (March to October ...... 53 Methane Evolution in Creeping Swamp from March 1978 to February 1979: (a) Seasonal Patterns, and (b) as a Function of Water Temperature ...... 55 Seasonal Patterns of Monthly Terrestrial Litter Respiration (Rtl) and Aquatic Benthic Respiration (Ra) Rates in Creeping Swamp for 1976 and 1977 .... 56 Time Course of Weight Loss from Newly Fallen Leaves during the Initial Seven Days of Leaching in Distilled Water...... 59 Effect of Water Temperature on Rates of Slow Leaching of Ground Litter during Four Experiments with Creeping Swamp Floor Cores ...... 60 Seasonal Patterns of Fine Total Organic Carbon (FTOC) Concentration and Streamflow at (a) Chicod Creek (CH-20), (b) Palmetto Swamp (PM-10) , (c) Clayroot Swamp (CY-10) , and (d) Tracey Swamp (TR-10) in 1976 ...... 63 Patterns of Fine Total Organic Carbon (FTOC) Concentration and Streamflow at Two Stations in Creeping Swamp (CP-10 and CP-20) during Spates in (a) January, (b) June, and (c) December 1976 ...... 65 Patterns of Fine Total Organic Carbon (FTOC) Concentration and Streamflow at Two Stations in Clayroot Swamp (CY-10 and CY-20) during Spates in (a) January, (b) June, and (c) December 1976 ...... 66 Model of Organic Carbon Flow in Creeping Swamp for 1977 ...... 79 Seasonal Patterns of Monthly Organic Carbon Export from the Watershed Drained by Creeping Swamp in 1976 and 1977 ...... 82 LIST OF FIGURES (Continued)

Page

Annual Export of Organic Carbon as a Function of Annual Runoff for Various Up1 and,Swamp-draining , and the Nanaimo River Watersheds ...... 86

A Descriptive Model of Phosphorus Cycling in Creeping Swamp...... 93

Design of Chamber for in situ Measurements of Floor-Water Exchanges of Phosphorus ...... 102

Precipitation, Inundation and Discharge for Creeping Swamp during Water Year 1377 ...... 110

Precipitation, Inundation and Discharge for Creeping Swamp during Water Year 1978 ...... 111 Seasonal Patterns of Phosphorus Concentrations at CP-14 . . 121 Seasonal Patterns of Phosphorus and Biomass in Ground Litter ...... 132

Soil Phosphorus and Organic Matter along a Transect of the Floodplain ...... 133

Seasonal Patterns of Throughfall Volumes and Phosphorus Concentrations ...... 135 Seasonal Patterns of Phosphorus and Biomass in Litterfall . 138 Sedimentation of Particulate Phosphorus onto the Creeping Swamp Floodpl ain ...... 139 FRP Uptake by the Swamp Forest Floor as a Function of Initial FRP Concentration ...... 145 The Interactive Effect of Initial FRP Concentrations and Ambient Water Temperature on FRP Uptake ...... 146 The Effect of Bacterial Antibiotics, Formalin and Sodium Arsenate on FRP Uptake at Enhanced FRP Concentrations . 152 A Phosphorus Budget for the Creeping Swamp Floodplain ... 173

Concentrations of BPN-reactive and Unreactive Ferric Fe in Particulate, Colloidal, and Dialyzable Forms ..... 192 LIST OF FIGURES (Continued)

Page

42. Concentrations of BPN-reactive and Unreactive Ferrous Fe in Particulate, Colloidal, and Dialyzable Forms ..... 193 43. Concentrations of Manganese in Particulate, Colloidal, and Dialyzable Forms ...... 44. Concentrations of the Forms of Fe Associated with Suspended Particulate Matter ...... 45. Concentrations of the Forms of Mn Associated with Suspended Particulate Matter ...... 199 46. Concentrations of the Forms of Phosphorus in the Water at CP-14 ...... 200 47. Concentrations of the Forms of Organic Carbon in the Water at CP-14 ...... 201 LIST OF TABLES

Page

1. Mean Monthly Inundated Fraction of Creeping Swamp .... 27

2. Estimated Biomass (Dry Weight) and Density for Trees and Saplings in Creeping Swamp ...... 29 3. Relative Density and Basal Area for Species of Trees in the Low and High Areas in Creeping Swamp as Computed from Point-quarter Data Collected in August 1974 ..... Weighted Mean Annual Organic Carbon Concentrations in Creeping Swamp and Tributary Streams ...... Flux of Organic Carbon -via Streamflow and Groundwater . . Total Aboveground Plant Litterfall in Creeping Swamp . . Dry Weight Loss as a Percentage of Original Dry Weight during a 7-day Leaching Study of Newly Fallen Leaves Collected in Autumn from Several Species of Trees in Creeping Swamp ......

Utilization of Readily Leachable Material from Newly Fa1 len Red Maple Leaves in Swamp Floor Cores Taken in November 1978 ......

Weighted Mean Annual DOC and FPOC Concentrations in Natural and Channelized Streams in Eastern North Carol ina ...... Concentration of Organic Carbon in Natural Waters .... Annual Productivity and Standing Crop of Biomass and Detritus in Swamps and Upland Forests ...... Comparison of Field and Laboratory Measurements of Aquatic Benthic Respiration Rates ...... Annual Total Soil Respiration in Various Swamps and Upland Forests ...... Monthly P/R Ratios for the Aquatic Ecosystem in 1977 . .

Inputs and Outputs of Organic Carbon in Creeping Swamp for1976and1977 ...... LIST OF TABLES (Continued)

Page

Characteristics of Three Stream Segment Ecosystems and Indices Describing Organic Matter Dynamics ..... 81 Annual Organic Carbon Export and Runoff from Various Upland and Swamp-draining Watersheds ...... 84,85 Net Hydrologic Export of TOC from Creepin Swamp for 1976 and 1977 in g corn-2 of S~amp.~r-Y ...... 87 Inundation Patterns of the Creeping Swamp Floodplain . . 113 Annual Weighted Mean Concentrations and Inputs of Phosphorus in Bulk Precipitation ...... 113 Annual Mean Conductivity, Turbidity, Color, pH and Phosphorus Concentrations at CP-10 and CP-20 .... 115 Annual Mean Conductivity, Turbidity, Color, pH and Phosphorus Concentrations at TB-02 and TB-04 .... 116 Annual Mean Conductivity, Turbidity, Color, pH and Phosphorus Concentrations at TB-01 and TB-03 .... 117 24. Annual Mean Conductivity, Turbidity, Color, pH and Phosphorus Concentrations at TB-07, TB-09 and TB-10...... 118 25. Fractionation of Filterable Phosphorus into Dialyzable and Colloidal Forms ...... 122,123 26. Annual Exports of Phosphorus from the Upstream Watersheds at CP-20 and CP-10 ...... 123 27. Average Conductivities and Phosphorus Concentrations in Wells in the Creeping Swamp Study Area and Annual Losses of Phosphorus in Deep Ground Water Outflow ..... 125 28. A Water Budget in the Creeping Swamp Floodplain ..... 127 29. Annual Surface Water Imports and Exports of Phosphorus from the Swamp Floodplain ...... 129 30. Biomass and Phosphorus Content of Above-Ground Vegetation in Creeping Swamp ...... 131 LIST OF TABLES (Continued)

Page

31. Phosphorus in Ground Litter and Soil ...... 134

32. Volumes, Annual Weighted Mean Concentrations and Fluxes of Phosphorus in Throughfall and Stemflow ...... 137

33. Sedimentation of Particulate Phosphorus from Floodwaters ...... 140

34. Average Rates of Exchange of FRP and FUP, Turnover Times and Correlation Coefficients of Flux with Independent Variables ...... 142

35. Transformation of FR~~Pinto FU~~Pand Distribution of the Fraction of 32~Removed from the Water Column . . 147 36. Effects of Formalin Solution of FRP and FUP Floor-Water Exchanges ...... 149

37. Regeneration of FRP and FUP in Chambers Following Experi- ments when FRP Concentrations were Increased over Ambient ...... 150

38. Comparison of the Ratios of Initial As:FRP Concentrations with the Ratios of the Uptake of As and FRP .....153 39. Estimates of the Annual Fluxes of FRP and FUP at the Forest Floor-Water Interface during Water Years 1977 and 1978 ...... 153

40. Phosphorus Budget for the Creeping Swamp Watershed ...155 /

41. Comparison of Total Phosphorus Concentrations and Inputs in Bulk Precipitation for Several Locations in the Eastern United States ...... 157 42. Rates of Phosphorus Exchange at the Sediment-Water Interface of Various Aquatic and Semi-Aquatic Ecosystems ...169

43. A Budget for Phosphorus Cycling in the Creeping Swamp Floodplain Ecosystem ...... 171,172 44. Comparison of Ecosystem and Watershed Inputs and Outputs of Phosphorus in Wetlands, Upland Forests and Agriculatural Areas ...... 178

xvi i LIST OF TABLES (Continued)

Page 45. Physical and Chemical Characteristics of Swamp Water . . 191 46. Concentrations of Dialyzable Fe and P in Soil Pore Water from the Low Floodplain ...... 202 47. Concentrations of the Forms of Iron and Manganese Associated with Swarnp Soils, and Percent Organic Matter of Swamp Soils ...... 203 SUMMARY, CONCLUSIONS, AND RECOMMENDATIONS

Floodplain swamps are closely 1inked to most Coastal Plain streams. These swamp ecosystems include not only the structural elements such as the stream, the swamp forest, the atmosphere, the soil, and the fauna, but a1 so the functional attributes such as primary productivity and nutrient cycl ing. One natural floodplain swamp ecosystem, Creeping Swamp, on the lower Coastal Plain of North Carolina, was selected for intensive field studies of phosphorus and organic carbon as they exist in the structural matrix and as they cycle into, out of, and within the ecosystem. The study area, about 3.2 km2, supports a deciduous bottom1 and hardwood forest which was last cut about 40 years ago, Although flooding may occur after heavy rains in any season, continuous flooding of the swamp usually occurs in winter and spring. The flood waters are typically clear, dark-colored, acid, and low in conductivity and plant nutrients. The studies reported here show details of swamp structure and functioning through the storages and processing of organic carbon and of phosphorus. Investigation of the physical and chemical forms of carbon, phosphorus, iron, and manganese give insight into the many interactions between the water, the soils, and the swamp vegetation.

The largest compartments of organic C were the living trees and saplings (13.8 kg Corn-2) and the litter, logs, and soil organic materials (to 25 cm) (8.2 kg corn-2) of the swamp floor. Shrubs, herbs, and algae amounted to less than 25 g Csm-2. Important processes of organic carbon cycling of the swamp ecosystem were measured and an annual organic carbon budget was constructed for calendar year 1977. Major inputs included the organic C in hydrologic flows from upstream and tributaries (192 cj ~-m-2*~r-l),from rainfall (2.4 g ~.m-2*~r-l), and from net aboveground primary productivity (706 g ~-m-2e~r-l).Net aboveground primary pro ucti ity was comprised of tree and sapling net productivity (676 g Ca '8. yr-v ) and shrub and herbaceous plant net productivity (16.7 g C*m-y-yr-l) during the warm season, and algal net productivity (14.1 g corn-2.yr-1) during winter and early spring. Tree and sap1 ing net aboveground ~roductivitywas further frac- tioned into net wood increment (324 g Come eyr-I), litterfall (272 g ~*m-2-~r-l),macro-1 i tterfall 58.9 g C*m-2Wyr-l), and net throughfall and stemflow leaching (20.7 g C*m-kyr-1). Major outputs were organic C in hydrologic flows of surface water (214 g ~-m-2-yr-l)and heterotrophic respiration (327 g em-2-yr-1) . Respiration was strongly dependent on temperature and on extent of flooding. It was partitioned into respiration in the water column (25.5 g ~-m-2.~r-l)on the swamp floor when flooded (45.9 g ~*m-~e~r-l)and dry (246 g ~*m-leyr-l),and metabolism under anaerobic conditions (9.8 g C-m-2eyr-1). Total inputs of organic C to the ecos stem exceeded total outputs by 361 g C*m-2eyr-lY with a1 1 but 38 g C-m-5.yr-I of this in net wood increment. Creeping Swamp was in approximate balance with regard to detrital flows in 1977 and no peat or large- scal e detritus accumulations were observed.

xix Many swamp ecosystems are very productive. Those with water flow and hydrologic inputs of nutrients from upstream areas, such as small swamp-streams and larger riverine swamps, may be among the most productive forested ecosystems, particularly in terms of annual plant litterfall. Creeping Swamp had intermediate levels of plant biomass and high annual net primary productivity, compared to other swamps. Compared to upland forests, however, Creeping Swamp had high biomass and annual net primary productivity. Swamps are often areas of detritus and peat accumulation due to inefficient decomposition under oxygen stress (Given 1975). Oxygen stress in swamp-streams, such as Creeping Swamp, may not be important due to water flow and alternating flooded and dry conditions. Leaching and respiratory outputs balanced inputs of detrital organic carbon to the swamp floor. There was no apparent accumulation of detrital organic carbon in Creeping Swamp. Swamp waters have higher DOC concentrations than do those in upland watersheds. This is probably the result of increased leaching of plant detritus due to longer contact times between ground litter and surface water and increased evapotranspiration in swamps. There was no clear difference in weighted mean annual DOC concentrations between seven natural and channelized streams in eastern North Carolina. The important factor controlling mean annual DOC concentration in these streams may be the amount of swamp drainage relative to the entire watershed. Streams draining watersheds with similar fractions of swampland usually have similar mean annual DOC concentrations. Watersheds draining swamps export more organic carbon per unit of runoff than do upland watersheds, apparently due to increased export from the swamp portion of the watershed. Creeping Swamp itself exported more organic carbon per unit area than did the watershed drained by it. Swamps with hydrologic regimes involving surface water flow-through are likely sites of relatively large organic carbon leaching and export. Most of the export of organic carbon from Creeping Swamp was in the dissolved form due to leaching processes and relatively low water velocities. While dissolved organic C (DOC) is not readily utilizable by consumer organisms, transformation to the particulate state may greatly enhance its availability. DOC precipitation, while not important in most freshwaters (Lock and Hynes 1976), has been shown to increase dramatically as freshwater and seawater mix in estuaries (Shol kovitz 1976; Gardner and Menzel 1974). DOC export from swamps in coastal areas may be extremely important in maintaining high estuarine secondary productivity as a result of DOC precipitation and retention in estuaries. The two most important compartments with regard to organic C also contained most of the phosphorus. Most of the P was in the upper 25 cm of soil (32-35 g porn-')* and in above-ground vegetation (5.5 g Phos horus was much less abundant in the forest litter layer (370-540 mg Porn-!) and in floodwaters (0-10 mg porn-2).

The greatest source of phosphorus to the ecosystem was in surface water inflow (980-1 200 mg ~.m-2*yr-l). Precipitation inputs were re1atively small (60-79 mg ~*m-2-~r-l).A hog farm drained by a tributary to the swamp greatly increased imports of fi1 terable reactive phosphorus (FRP) and articulate phosphorus. During the two years of study, 22-29 mg P-m-gmyr-1 was exported by surface waters, giving 30-57% retention of phosphorus inputs to the swamp. FRP was retained during both years, whereas small net amounts of f il terabl e unreactive phosphorus (FUP) were exported from the ecosystem. Imports and exports of particulate phosphorus were nearly equivalent one year; imports increased by 2.5 times the next year without significant change in exports, resulting in a large net retention.

Transfers of phosphorus within the ecosystem occurred primarily between canopy vegetation and the forest floor, and floodwaters and the forest floor. Annual return of phosphorus from the canopy to the forest floor (390-480 mg ~-m-2eyr-l), in l itterfall (313-329 mg ~am-z=yr-l), branch- fall (22 mg ~-m-2*~r-l),and throughfall and stemflow (59-131 mg ~-m-2ayr-l), was similar to average standing stocks of phosphorus in litter on the forest floor, indicating a turnover time of about one year for this component. On the average, particulate P and FRP moved from floodwaters to the forest floor (-1 72 and -71 0 to -1 100 mg ~em-~*~r-l,respectively), whereas FUP was released from the forest floor to floodwaters (340 mg ~-m-2*yr-l).

Imports and exports due to fertilization (250 mg ~-m-~*yr-')and harvesting of crops (100 mg ~.m-2.~r-l)in the entire watershed overwhelmed natural imports in precipitation and exports in surface waters, even though crop lands covered only 25-30% of the watershed. By its location at the base of the watershed, the swamp floodplain ecosystem regulated the amount and form of phosphorus leaving the watershed in surface waters. Since fluxes of phosphorus to and from the ecosystem were predominantly hydrologic, they were significantly greater during storm flow periods. Phosphorus cycling in the swamp floodplain ecosystem was dominated by large exchanges between floodwaters and the forest floor. The standing stocks of phosphorus in floodwaters were extremely small compared to annual fluxes to and from this compartment. Measured exchange rates of both FRP and FUP between the forest floor and floodwaters were much greater than net fluxes estimated from budget calculations for the whole ecosystem, implying rapid recycling between the two components. FRP fluxes at the forest floor-water interface were under considerable biological influence, especially a filamentous algae bloom during the flooded season. However, the low levels of FRP in the floodwaters and the strong response of flux rates to FRP increments suggested that the magnitude of these exchanges were controlled by ambient FRP concentrations.

*Ranges shown in parentheses are values for Water Years 1977 and 1978. The forest floor thus appeared to have a high affinity for, and was not saturated with respect to, phosphorus. Of the phosphorus retained by the swamp ecosystem, some was stored in above-ground plant biomass. The remainder may have accumulated in roots or in the soil.

Senescence, death, and decomposition of the algae probably released the P which had been accumulated from floodwaters to the forest floor and roots of vascular plants when their nutrient demand was greatest. Canopy return of phosphorus was roughly equivalent to the average standing stock of phosphorus in the ground litter, implying a one year turnover time of phosphorus in litter. This short turnover time, the lack of a distinct fermentation or humus layer in the forest floor, and observed growth of surface roots into ground litter pointed to tight recycling of vegetation phosphorus.

Phosphorus cycling in both the Creeping Swamp watershed and in the floodplain ecosystem was characterized by the net retention of phosphorus even under the pressure of increased inputs due to hog farm releases. Movement of phosphorus from the watershed as a whole and from the floodplain itself appeared to be hydrologically control led. Surface water flow in the watershed and swamp floodplain varied in response to rain storms, seasonally, and from year to year. Phosphorus transport during storm flow periods was a significant component of total annual fluxes (Kuenzler, --et a1 . 1977; Yarbro 1979). The particulate phosphorus fraction increased in concentration during storm flow periods, especially downstream of agricultrual areas or in channelized tributaries. In the absence of known pollution inputs, FRP and FUP fractions showed little change in concentration during storm flow due to buffering processes in the ecosystem. There- fore, exports of these fractions were closely related to the total transport of water. Greatest water exports from the swamp occurred during the winter and early spring when vegetation was dormant and heterotrophic processes dominated in the ecosystem. Despite the large surface area of the swamp forest floor and the large volume of water passing through the swamp ecosystem, exports were very low under undisturbed conditions.

Runoff from the Creeping Swamp watershed differed by a factor of three between the two years studied. In spite of greater runoff in the second year, exports of particulate P remained quite similar to the year before. Because of its location at the base of the watershed, the swamp ecosystem may have reduced the effects of erosion (surface water sediment transport) occurring in agricul trual areas and in channel ized streams. FRP exports dropped dramatically during the year of greater runoff, due to a sharp decline in inputs from the hog farm. However, FUP exports increased nearly three times between the two years, suggesting that FUP originated from leaching processes which were water-vol ume dependent.

Phosphorus in the swamp waters was quite low in concentration and predominantly filterable in size, when pollution sources were absent. Large qortions of FRP and FUP were colloidal in size rather than truly dissolved.

xxi i Phosphorus cycling in Creeping Swamp was similar to that of other wetlands in two basic respects: the greatest storage of phosphorus was in the sediments, and fluxes and exports were closely linked with water movement. The accumulation of phosphorus by Creeping Swamp floodplain ecosystem was similar to values measured in other wetlands with the exception of brackish and saline marshes. Net accumulation of phosphorus by the swamp suggests that the ecosystem is accreting in biomass and thus is not in steady state and the youth of the swamp forest (40-50 years) 1ends support to this conclusion. More precise assessment of phosphorus accumulation by the ecosystem awaits measurements with less experimental error and better understanding of effects of annual variations in the hydrological budget . When the pathways of phosphorus and organic carbon cycling in Creeping Swamp are compared, several striking differences are evident. In stream waters, organic carbon was primarily in dissolved form (85-98% of total organic carbon, TOC) (Mu1 hol land 1979). Particulate phosphorus concentrations, on the other hand, tended to be the largest fractions in stream waters. T0C:TP ratios (by atoms) in stream waters averaged 7400: 1 and 1140:l in 1976 and 1977, respectively, with maximum ratios in TB-03 and TB-07 waters (1830-3OgO: 1) and minimum ratios in TB-02 waters (41 -52: 1). These large ratios, with the exception of those found at TB-02, suggest that the swamp ecosystem is severely phosphorus limited, especially with respect to utilization of hydrologic organic carbon. Inputs of carbon to the swamp were in atmospheric (78%)and hydrologic (22%) forms, whereas hydrologic inputs were the sole source of phosphorus to the ecosystem. When hydrologic inputs and outputs are compared, the swamp ecosystem exported net amounts of organic carbon (Mu1 hol land 1979). This was reflected in the increase in annual weighted mean concentrations of organic carbon from CP-20 to CP-10 and in the change in the C:P ratios in hydrologic inputs (500:l) to hydrologic outputs (740:l). When atmospheric inputs of carbon are included in the ecosystem carbon budget, the uptake of carbon in net primary productivity resulted in a large net increase in carbon in the ecosystem. Within the ecosystem, the ratio of carbon to phosphorus in litterfall (2240:1), branch-fall (6920:1), throughfall and stemflow (881 : 1 ) , reflects the phosphorus economy of the swamp biota. Iron and manganese in natural swamp water and soil were separated analytically into a number of chemical and physical forms. Not only did the total concentrations in the water change before and during the annual period of flooding but the proportions of the various forms also changed. High levels of Fe and Mn were found in stagnant pools in the stream channel in October prior to fall flooding. Particulate and colloidal Fe and Mn were usually dominant in these pools, with crystalline mineral and reducible forms being most important. The first flcod in November brought still higher levels of particulate Fe and Mn strongly dominated by the crystalline mineral forms. During the period of nearly constant swamp flooding (December-April) particulate Fe and Mn concentrations were lower

xxi i i and variable, but were predominantly crystal1 ine in form. Colloidal and dialyzable ferric iron were of similar importance, but colloidal unreactive ferrous iron and dialyzable Mn were more important than dialyzable ferrous iron and colloidal Mn, respectively, during the flood season. The colloidal and dialyzable iron in swamp water which reacted with batho henanthroline was entirely ferric and seldom exceeded 0.10 mg Fe.1- e . Iron in soil pore water, however, had 2-15 times as much Fez' as ~e3+, reaching concentrations as high as 2.0 m Feel-1. Even the BPN- reactive Fe3' was always at least 0.1 mg Fe.1-7, and thus usually at a higher concentration than occurred in the overlying water. The iron associated with swamp soils was mostly in crystalline mineral form; this form usually increased with depth in the soil. Next in importance were reduci bl e iron oxides which, however, general ly decreased i n concentration with depth. Highest concentrations of soil Mn were found in exchangeable and reducible forms in the leaf litter layer. At greater depths in the soil crystalline Mn usually was most important. The presence of both oxidized and reduced forms of iron in the soil is probably a consequence of non-equi 1i bri um conditions in the soi 1 whereby redox potential varies markedly over short distances. Particulate P concentrations were positively correlated with unreactive particulate Fezt and ~e3+concentrations in floodwaters suggesting co- precipitation or complexation of P with Fe on particulates. Other correlations between reactive forms of P and Fe, Mn, and organic C were not significant, suggesting that experimental studies are necessary to elucidate these interactions. Recommendations Based upon our investigations and upon studies of other wetland ecosystems we make the following recommendations: The State of North Carolina should commit itself to protection of floodplain swamp ecosystems along its rivers and streams. The natural functioning of these floodplain systems provides real values in terms of hydrology, sediment control, reduction of excessive plant nutrients, production of particular timber species, nursery areas for fish, and outdoor recreation. A previous study (Kuenzler, et al. 1977) demonstrated the controls which swamp ecosystems have sonwater qua1 ity. The high level of productivity demonstrated in the present study, and in particular the magnitude of organic carbon flux through the leaf- litter-detritus pathway, suggests a very important food base for swamp fishes and other wildlife. The remarkable ability of the flooded swamp ecosystem to transfer phosphate from the water to the soil has significant implications for water quality not only in the stream but also in our estuaries. Increased loadings of phosphate and nitrate into the Chowan River over the past one or two decades has resulted in massive algal blooms. Increased inputs from industry, from municipal

xxi v sewage, and from intensified agriculture and livestock operations have contributed phosphate and nitrate at the same time that stream channelization and swamp forest destruction have markedly weakened the ability of the watershed to remove these plant nutrients. Protection of the floodplain swamp ecosystem is compatable with numerous non-des- tructive uses of the resource. For example, timber harvest can continue; in fact, selective cutting would probably improve the quality of the timber stand in many places. Fishing and hunting are also compatable uses of this wetland system. Small floodplain swamps would serve for educational field trips for public school children to demonstrate many facets of Coastal Plain geology, stream hydrology, natural history, water quality, and land use management.

The recent slow-down in stream channelization projects in North Carolina is consistent with protection of floodplain swamps. However, reduction in the rate of swamp drainage does not conserve these valuable wetlands; it only postpones the date at which all are destroyed. The swamps which border first-, second-, and third-order streams in the headwaters of each river system are particularly worthy of preservation because they set the initial water quality and because they are so extensive relative to the volume flow. Destruction of these swamps by drainage should not be permitted to continue. In addition, very large tracts of pocosin which feed many Lower Coastal Plain streams should be left in their natural state; management practices should be instituted for all drained pocosins that insure continued high water quality.

Although the natural swamp system has the ability to reduce markedly levels of suspended matter, phosphate, and nitrate, there are limits to the loadings which may be accommodated without impairing water qua1 ity or damaging the natural system. It would be appropriate to determine experimentally the loadings which the swamp system can assimilate. Until such a study is performed, however, nutrients and suspended matter from runoff and from point sources should be monitored and, if necessary, controlled. Such a study would bear directly upon the feasibility of using North Carolina swamplands for tertiary treatment of sewage or other wastes. Two additional research programs seem appropriate at this time. The . first would be preliminary assessment of the functioning of natural pocosin ecosystems in eastern North Carolina. These pocosins constituted the largest tracts of wetlands in North Carolina in 1962 (Wilson 1962) and certainly are still an important factor in water quality and the direction of land development in that part of the state. The functional aspects to be studied should include hydrology, productivity, and nutrient cycling. Desirable experimental studies of the system include manipulation of drainage, of fire, and of nutrient (i.e,, waste) loadings.

xxv A second research program would link land use, swamp functioning, and estuarine water qua1 ity. It appears that intensive agricultural activity, swamp drainage and conversion, and eutrophication of tidal rivers and estuaries in recent years are related. Such an investigation would be of major value to state and federal resource management agencies.

xxv i 1. INTRODUCTION

General

Scientists and resource managers recognize the importance of freshwater wetlands because of their high levels of primary productivity, their contributions to water quality, their recreational potential, their refuge values for endangered species, and their indispensibility as nursery grounds for certain fishes (e.9. , Wharton 1970; Anonymous 1976; Brinson 1976). Unfortunately the dafa base on ecological functioning of swamp systems is limited and consequently understanding is often insufficient to permit accurate prediction of the environmental impacts resulting from modification of existing swamps. For example, Rulison and Martin (1972) evaluated twelve flood control and drainage projects in hardwood bottomlands in coastal North Carolina and South Carolina, including analysis of the benefit/cost procedures employed, They found that the deleterious effects of stream channelization upon water quality, such as increased concentrations of nitrogen, phosphorus, pesticides, other agricultural chemicals, and silt, were not included in the pre-channelization benefitlcost projec- tions. The Soil Conservation Service and the Corps of Engineers have in the past not sufficiently considered the value of swamps as determiners of water quality because the ecological and chemical data base was insufficient for accurate assessment.

Floodplain swamps occur in lowlands that border streams or rivers which seasonally inundate them. In 1960 they covered about 1,860 km2 in North Carolina, mostly within 160 km of the coast (Wilson 1962). The dominant trees are mostly broadleaved deciduous species such as red maple, river birch, tupelo, sycamore, willow, ironwood, sweet gum, black gum, oaks, and hickories (We1 1s 1928; Oosting 1942) ; cypress may a1 so be found. Pocosins are even more extensive, covering about 9,160 km2 in North Carolina in 1960. Although often considered wastelands because of their thin stands of merchantable timber, their impenetrable brush, and their boggy soils, pocosins comprise the headwater drainages of a very large portion of North Carolina's Coastal Plain streams and therefore are prime determinants of the natural flow regime and of several important water quality parameters, such as pH, color, and plant nutrient concentrations. Floodplain swamp and pocosin area has decreased markedly since Wilson (1962) published his report through modification, damage, and outright conversion, mostly because of their timber and agricultural potential. Clearing, drainage, stream channelization, grazing of livestock, and introduction of pollutants have adversely affected the flora and fauna of large tracts throughout the Southeast.

Background

The research reported here was initiated because of the scarcity of information about the functioning of small floodplain swamp ecosystems. It built upon studies of water quality of Creeping Swamp, two other natural streams, and four channelized streams (Kuenzler, et a1 . 1977). That study showed substantial differences between natural and--channel ized streams in amount of color, turbidity, conductivity, and nitrate-nitrogen. Color was about three times higher in natural streams than in channelized streams, with the exception of Tracey Swamp which received substantial pocosin drainage. The higher turbidities in channelized streams were attributed to scouring of the stream bed, and the higher conductivities to a larger proportion of ground water relative to surface runoff. Part of the increase in nitrate and nitrite may be attributable to runoff from fertilized agricultural fields, but part of it results from lack of swamp forest along the channelized streams as a sink for nitrogen. There were very low levels of dissolved oxygen and high levels of color, organic-N, and ammonium-N during summer in natural streams. Nitrate and soluble-P levels did not show marked seasonal changes, although particulate-P exhibited pulses during summer periods of stagnation. Iron concentrations sometimes exceeded 3 mg-1-1 and manganese exceeded 0.3 mg- 1-1 in summer and fa1 1 in Creeping Swamp; levels of Fe and Mn were about one-tenth as high in winter. Positive correlations indicate a possible coupling between iron, manganese, and organic carbon. Even in summer, however, iron concentrations in swamp water were lower than in groundwater of the surface sediments (unpublished data). Present research focused upon the abundance and cycling of carbon and phosphorus and to a lesser extent iron and manganese in a defined segment of Creeping Swamp. The availability of continuous stream-flow data for Creeping Swamp from the U.S. Geological Survey permitted calculation of hydrologic inputs to and outputs from the swamp system.

Wet1 and hydro1ogy The hydrologic regime of wetlands can be separated into four components (Gosselink and Turner 1978): 1 ) the amount and source of water introduced to the wetland ecosystem; 2) the periodicity and timing of water inputs; 3) the flushing rate or rate of flow-through sf water in the wetland and; 4) the effects of the biota. Water entering a wetland may come from a number of sources: precipitation, groundwater influx, stream flow,and tidal fluctuations. Water may leave a wetland as vapor, streamwater, groundwater, or on ebbing tides.

Wetlands with the most restricted hydrologic regimes are ombrotrophic (meaning "rain-fed") bogs which, by their massive peat accumulations and/or the presence of an impervious lens jn the underlying sediments, are isolated from groundwaters and which receive no upland surface drainage. Typical examples include the Sphagnum bogs of the northern latitudes (Heinselman 1970; Moor and Bellamy l974), the pocosins (meaning "bog on a hi11 ") of the southeastern United States and more broadly the Okefenokee Swamp of Georgia (Schlesinger 1978). In bogs, the timing of precipitation and thus nutrient inputs is dependent on seasonal variations. The flushing rate of water in these ecosystems is very low, most water movement being vertical in response to precipitation inputs and evapotranspiration processes. The restricted drainage of these systems maintains anaerobic conditions in the sediments which slows decomposition processes. Low rates of decomposition coupled with restricted nutrient inputs result in greater peat accumulation which further isolates the ecosystem from groundwaters (He inselman 1970). Some bogs maintain contact with mineral-bearing groundwaters; these ecosystems are called fens and are characterized by a more diverse biota which have significantly higher concentrations of nitrogen and phosphorus in their tissues (Richardson, --et a1. 1978). The relative inputs of water and minerals in precipitation and groundwater~are variable but groundwater inputs are significant with respect to nutrient fluxes. The periodicity and timing of aqueous fluxes and hydrologic residence times are similar to those of bogs (Heinselman 1978). Swamps and marshes adjacent to and drained by surface waters such as streams, rivers, and estuaries are next in the series of wetland types having more open hydrologic regimes. These two ecosystem types are easily differentiated by vegetation: swamps are wooded ecosystems while marshes are covered with non-woody vegetation such as grasses, sedges or other aquatic vascular plants. In these systems, precipitation and groundwater inputs of water and nutrients may still be significant; however, these ecosystems are periodical ly flooded and then drained by surface waters. The flux of nutrients as a result of inundation plays a significant role in nutrient cycling in swamps and marshes. The timing of inundation in relation to the growing season, the energy associated with flood and tidal waters (i.e. , the erosion potential ), the duration of inundation and the rate of flow through swamps and marshes affect the input, retention and export of nutrients. In addition, the sediments may be an important nutrient reservoir for these ecosystems. Swamps adjacent to large sediment-bearing rivers are often separated from these rivers by levees and are subjected to sporadic.flooding by riverwaters and to long periods of stagnation (Butler 1975; Day, --et al. 1976; Brinson 1977; Holmes 1977; Mitsch, --et al. 1977; Kemp 1978). Swamps adjacent to rivers and streams draining the Atlantic Coastal Plain are bounded by small levees, or none at a1 1, and are typically flowing-water systems when inundated (Kitchens, et a1 . 1975; Boyt 1976; Boyt, et a1 . 1977; Kuenzler, et al. 1977). ~reepiGSwamp is typical of this-Tind of . swamp. and has an annual cycle of cool season inundation followed by drawdown in the growing season which is due to enhanced evapotranspiration. Finally, there is a group of Southeastern swamp ecosystems which have certain characteristics of northern wooded bogs. These are the cypress domes of Florida (Odum and Ewe1 1976, l978), the Okefenokee Swamp of Georgia (Rykiel 1977; Schlesinger 1976, 1978) and the Great Dismal Swamp in North Carolina and Virginia (Whitehead 1972; Day 1978). These are chiefly rain-fed ecosystems having little contact with mineral substrata and with either restricted drainage (cypress domes) or slow-flowing streams draining the edges of the swamp (Okefenokee and Great Dismal Swamps). The flushing rate of these systems is low with relatively long residence times for surface waters. Organic carbon fluxes

Carbon flow through an ecosystem is basic to its functioning, and measurements of this f 1ow are fundamental to understanding the ecosystem. To date there have been few complete energy or organic material budgets for whole ecosystems. Budgets of organic carbon and energy have been constructed for a small lake (Rich and Wetzel 1972; Wetzel and Otsuki l974), a woodland stream (Fisher and Likens l973), salt marshes (Teal 1962; Day, --et al. 1973; Nixon and Oviatt 1973), and a mangrove forest (Golley, --et al. 1962), but until now no budget of organic carbon or energy has been made for a freshwater swamp ecosystem. The role of allochthonous organic material as a major energy input to woodland streams has been confirmed (Nelson and Scott 1962; Minshall 1967; Fisher and Likens 1973). Because of shading by trees and the amount of organic matter introduced by the forest, forest streams tend to be heterotrophic, as found by Hoskin (1959) and Hall (1971 ). Recent papers have demonstrated the importance of decomposing leaf litter in the energy budgets of temperate streams (Gosz, et a.1972, 1973; Cummins, --et al. 1973; Fisher and Likens 1973) and in tropical mangrove swamps (Heald 1969; Odum 1970). Investiaations concerning the biological processing of this organic input have been undertaken (Kaushik and Hynes 1971; Cummins, --et al. 1973; Suberkropp, --et -al. 1976; Triska and Sedell 1976).

Swamps are among the most productive of ecosystems (Rodin, et a1. 1975) ; this productivity has been attributed to abundant water aanutrients and relatively warm temperatures. The abundance of water and the warm temperatures of the Southeast are evident. That nutrients are plentiful has not been demonstrated; the mechanisms by which they are accumulated and conserved are not understood. In view of the large soil surface in contact with the water when Creeping Swamp is flooded, and the many possible chemical and biological transformations which nutrients may undergo, this study emphasizes the interactions between the water and the benthic systems. Bottomland hardwood forests adjacent to small streams in the Southeast are seasonally flooded by darkly colored, clear waters. Greatest stream flow and consequent inundation of the floodplain occur during the cool season when forest trees are dormant and when low water temperatures may restrict metabolic activities of microorganisms living on the forest floor (~uenzler,et a1 . 1977). Inundation of the floodplain greatly enhances the ratio ofsurface area to water volume in the swamp; the presence of a leaf-litter layer on the floodplain probably increases contact between water and potentially reactive sites. The above conditions suggest that materials may be leached or exported from the swamp during this period when stream discharge is greatest. However, an earlier study demonstrated that Creeping Swamp was retaining relatively large proportions of total nitrogen and phosphorus introduced by atmospheric precipitation and surface runoff (Kuenzler, --et a1 . 1977). Moreover, despite extensive agricultural and timbering activities on land bordering the swamp, only low levels of minerals were brought into the swamp by streamflow. These findings suggested that the swamp had an effective conservation mechanism for nutrients vital to plant and animal growth. Considering the large surface area of the swamp floodplain which is exposed to floodwaters during the cool season, exchange processes at the swamp forest floor-water interface may be responsible for the removal and retention of phosphorus from incoming f 1oodwa ters . Phosphorus Studies on phosphorus cycling in upland forested watersheds have been many and varied. The most comprehensive include studies on the Hubbard Brook Experimental Forest (Hobbie and Likens 1973; GOSZ, --et al. 1976) and on a temperate forest in Be1 gium (Duvi gneaud and Denaeyer-de-Smet 1970). Fewer reports on phosphorus cycling in freshwater swamps are available (Odum and Ewel 1 1976, 1978; Brinson 1977). Phosphorus dynamics were studied in and around the forest floor of two temperate forests (Reiners and Reiners 1970; Gosz, --et a1 . 1976). The sources of phosphorus to wetland ecosystems may be in atmospheric precipitation, ground water inflow, surface water inputs, and tidal fluctuations. Large reservoirs of phosphorus may be located in the sediments of the ecosystems. The amount of phosphorus received by a wetland depends on the hydrologic regime, the geology of the underlying sediments or of the upstream drainage system and the activities of man. With respect to phosphorus cycling, freshwater swamps can be divided into three types : 1 ) bog-1 i ke, primari ly rainfed ecosystems; 2) ecosystems flooded and drained by clear, often darkly-colored stream waters and; 3) swamps exposed to turbid, sediment-carrying floodwaters of rivers. Bog-like ecosystems include the Okefenokee Swamp, the Great Dismal Swanip and the Florida cypress domes. Little quantitative work has been done on phosphorus fluxes in the Okefenokee and Great Dismal Swamps although some intrasystem processes have been quantified in both systems (Schl esinger 1976, 1978; Day 1978). These ecosystems are relatively mature and probably function in a manner similar to ombrotrophic bogs with respect to phosphorus cycling. Odum and Ewel (1 976, 1978) are conducting a long term study on potential use of cypress domes for water and nutrient management. Pre- liminary results indicate that, of the phosphorus added to these ecosystems in sewage effluent, 60% was lost from the ecosystem by overflow or ground water infiltration. Initially, most of the phosphorus was taken up by duckweed; this was particularly true during late summer. In a natural cypress dome, the ecosystem appeared to be in steady state with respect to phosphorus fluxes. The long term phosphorus assimilative capacity of these ecosystems has not yet been assessed. Swamps adjacent to clear or darkly-colored stream waters have received little attention until recently. Kitchens, --et al. (1975) found a considerable reduction of phosphorus concentrations in river water flowing through the Santee River swamp over a three month period in the winter. Observations over an annual cycle were not made; the data, therefore, were not conclusive. Kuenzler, et a:. (1977) found that natural swamp stream ecosystems, including Creeping swamp, in the North Carolina Coastal Plain, retained relatively large amounts of phosphorus even when enhanced inputs occurred. Nessei (1978) found that the Waldo cypress strand in Florida retained up to 76% of the phosphorus brought in by polluted stream waters. Contributions by sediments were not assessed. Like other Florida wetland ecosystems, concentrations of phosphorus remained high in surface waters. Boyt (1 976) found that a hardwood swamp in Florida exposed to sewage effluent reduced stream water phosphorus levels to lower values (0.13 mg*l-l) than concentrations found in stream waters (0.27 mgel-1) leaving a nearby unpolluted swamp. The fate of the assimilated phosphorus was not known; there was no evidence of a greater buildup of phosphorus in the sediments of the polluted swamp (Boyt, --et al. 1977). A swamp in the Northwest Territories of Canad reduced phosphorus concentrations in sewage effluent from 11 mg*l-P to 0.25 mg1-1 , a 97% decrease (Hartland-Rowe and Wright 1975). It was suggested that most of the phosphorus removed by this ecosystem occurred as a result of increased primary production,

These floodplain swamps are typically not separated from stream channels by levees and are lotic systems when inundated, The rivers that feed these swamps carry little or no sediment and have very low concentrations of dissolved phosphorus under natural conditions. Therefore, fluxes of phosphorus in and out of these ecosystems are typically low. The significance of ground water inputs to these ecosystems has not been evaluated. These prel iminary studies suggest that lotic, floodplain swanlps may have relatively large assimilative capacities for phosphorus. The lotic character of these ecosystems when inundated and seasonal drying of the floodplain may contribute to phosphorus retention.

Although climatic effects on ecosystem biogeochemistry cannot be separated from hydrologic or biologic controls, a distinction is made here so that the relationship among these factors can be discussed. In temperate to subtropical marshes and swamps, such as Creeping Swamp, primary production occurs predominately during the warm season, but heterotrophic activities continue year-round and are probably strongly related to temperature fluctuations. These ecosystems typically remove phosphorus from incoming waters during winter and early spring and may release phosphorus in the growing season if hydrologic exports occur (2.9., Bender and Correll 1934; Kitchens, -et --al . 1975; Axel rad, --et -a1 . 1976).

In addition to seasonal changes in primary productivity and hetero- trophy, the maturity of the biotic community of a swamp may also influence the biogeochemi stry of the ecosystem. Vitousek (1977) and Vitousek and Reiners (1975) have shown that communities which are increasing in biomass tend to retain nutrients within the ecosystem whereas mature communities which are at steady-state with respect to biomass also tend to be at equilibrium with respect to nutrient inputs and outputs. Creeping Swamp was cut-over 40-50 years ago and the forest probably remains in a successional state. The swamp forest, then, may be contributing to the phosphorus economy of the ecosystem by storing phosphorus in incrementi ng biomass. In summary, swamp phosphorus biogeochemistry is the manifestation of complex interactions among the biota, climate, levels and sources of inputs and storages of phosphorus, the hydrology and, more recently, the influence of humankind. Few wetlands exist today whose nutrient and hydrologic regimes have not been drastically modified by human activities. For the most part, these changes were made with no understanding of the processes affected. To preserve or utilize wetlands most wisely, we must understand how they function. Mortimer (1941, 1942) found that redox potential of the sediment- water interface of l akes determined the phosphorus re1 ease from sediments. Under anaerobic conditions phosphorus is generally released from sediments to the water. The actual mechanism is not understood but hydrodynamics, molecular diffusion, the oxidation state of iron, adsorption to clays, and complexation reactions ar& believed to be important (Lee 1970; Stumm and Morgan 1970; Syers, --et al. 1973). The relationship between iron and phosphorus mobility, in conjunction with components of sediments, suggests that a complex, consisting of hydrated iron oxides, organic compounds, aluminum, and silicates, may be the phosphate adsorbent (Mortimer 1977; Shukla, --et al. 1971; Williams, --et al. 1977). The adsorbing character of this complex is postulated to change under anaerobic conditions because of changes in the iron oxidation state, thus reducing the effectiveness of the complex in holding phosphorus in the sediments. Brinson (1976) has shown that freshly-fallen leaves in an alluvial swamp accumulate P and Fe during the first winter but begin to release it after 24 weeks. Under stagnant, low dissolved-oxygen conditions phosphorus concentrations increase in isolated pools on the swamp floodplain (Kuenzler, --et al. 1977). Plants take part in the phosphorus dynamics of the land-water interface in several ways. Shallow roots and adventitious roots growing from lenticels on the trunks of certain trees (Hook, --et al. 1970, 1977 ) take up phosphorus from the soil surface, including the litter layer, and from the water. Bryophytes and algae, particularly the early spring bloom of filamentous desmids and green algae, also contribute to phosphorus uptake from the water and soil surface. Organic carbon, phosphorus and iron interactions Muck of the dissolved organic matter in natural waters consists of complex, heterogenous, brown or yellow acidic polymers formed, at least in part, during the decomposition of lignin and simpler organ" molecules and collectively known as humic matter (Christman and Ghassemi 1966; Jackson 1975). Dissolved humic matter has traditional 1y been divided into fulvic acid and humic acid fractions. Fulvic acid, the humic fraction remaining in solution at a pH of about 2.2, consists of smaller, less complex molecules and may represent younger, lesr-aged humic matter. Most of the humic matter in flowing waters is fulvic acid (Reuter and Perdue 1972; Jackson 1975). Humic matter plays an important role in determining the concentrations of various elements in rivers and streams. Dissolved and colloidal humic matter has been shown to form complexes with a number of elements, particularly the divalent and trivalent metals, in soils and sediments (Hodgson, --et al. 1966; Schnitzer 1969) and in surface waters (Shapiro 1964; Ghassemi and Christman 1968; Reuter and Perdue 1972; Ramamoorthy and Kushner 1975; Koenings 1976; Benes, --et al. 1976; Wilson and Kinney 1977; Giesy and Briese 1978). Interactions between humics and iron have been particularly we1 1 documented. Schni tzer (1 969) reported ~e"3and ~1+3were more strongly complexed by soil humics than nine divalent metals studied, and Ghassemi and Christman (1 968) found humic-iron interactions to be pH dependent with minimum association in the pH range 7-8. Theis and Singer (1974) have shown that appreciable quantities of ferrous iron could be maintained in aerobic aquatic systems due to complexation by organic matter. Interactions between organic matter, iron , a1 uminum, and phosphorus have been reported in organic soils and sediments. Although an early investigation (Bradley and Siel ing 1953) indicated that organic matter inhibited phosphate adsorption and precipitation by iron and aluminum compounds in soils, more recent investigators have attributed to organic matter a positive effect on phosphorus adsorption. Levesque and Schni tzer (1 967) reported a close relationship between aluminum, iron, phosphorus and organic matter in soils, with phosphorus complexed to soil fulvic acid via the metal. Jackson and Schindler (1975) demonstrated a similar association for lake sediment systems. Humic-metal-phosphorus associations have been proposed for surface waters as well (Jackson 1975). However, Koenings and Hooper (1976) found that colloidal organic matter (0.0048 - 0.45 partially inhibited phosphate adsorption to colloidal ferric oxyhydroxides in a humic-rich acid bog lake. They attributed this to reduced formation of col 1 oidal ferric oxyhydroxides, the principal phosphate adsorbing compound, due to increased scavenging of ferric ions by colloidal organic matter. It is yet unclear whether organic matter in most soil, sediment, and surface water systems inhibits phosphorus adsorption by inhibiting formation of amorphous iron and aluminum compounds or whether it enhances metal- phosphorus association by increasing metal availability through formation of humic-metal complexes. Humic-metal interactions may be particularly important in controlling the chemistry of Coastal Plain rivers and streams in the southeastern United States. Surface waters here are typically rich in dissolved humic matter and poor in the inorganic i ns dominating the chemistry of most other waters (Nat, cat?, C1-, SQ4' S ). Relatively high concentrations of iron and aluminum measured in some rivers and streams in Georgia have been attributed to complexation by fulvic acids (Beck, --et al. 1974). The high organic/inorganic ratios, low ionic strength, and the low pH of these waters appear to be characteristics which may fundamentally differentiate water chemistry of the world's poorly drained lowland rivers and streams from those of the upland. Description of the study area

The Creeping Swamp watershed is located in the Coastal Plain province of North Carolina in Pitt, Beaufort, and Craven Counties (Fig. 1). The main stream channel of the swamp separates Pitt County from Beaufort and Craven Counties. The watershed covers 80 km2, of which app roximately 10% is floodplain swamp. Elevations in the watershed vary from 6-18 meters above sea level and slopes are slight (Winner and Simmons 1977). Within the study area (see below) elevations ranged from 6-1 0 meters and the slope of the stream channel averaqed 0.06% or 0.6 m per km (Kuenzler, --et a1 . 1977). The el imate is subtFopical and moist with cool winters, warm summers and moderate precipitation which is relatively constant throughout the year (Sumsion 1970). Rainfall averages 122 ~m-~r-lwith a slight seasonal maximum during summer. The annual mean temperature is 17 C with a summer maximum of 39 C and winter minimum of -9 C. The average range is 4.5 to 24 C (Sumsion 1970). Relative humidities are high, seldom dropping be1 ow 50%.

The watershed is drained by a small stream, Creeping Swamp, which runs through a forested floodplain of the same name. A portion of the wetland- stream system about 8.2 km long and 3.2 km2 in area between State Highways 102 and 43 was selected as the intensive study area (Fig. 2). Floodplain width varied from about 100 to 500 m.

The drainage of Creeping Swamp begins in perched upland pocosins, some of which have been cleared. Within the study area, the main stream is at least third order with several first or second order tributaries; it is seasonally intermittent as well. Highest flows in the stream and maximal inundation of the floodplain occur during winter andlearly spring; depths reach 1 m and velocities are usually 0.05-0.15 mesec' . Discharge from the swamp varies widely depending on rainfall during the flooded season and is dominated by storm flow periods (Kuenzler, et al. 1977); even then, however, velocities seldom exceed 0.25 m-sec'l. In the spring, water levels in the swamp gradually decrease due to increased plant transpiration in the watershed and within the swamp itself, By mid-May to June, water usually remains only in the stream channel and the floodplain is dry. As summer progresses, the stream usually becomes intermittent and often by mid-October the only surface water in the swamp is in widely spaced pools in the stream channel. With leaf-fall and cessation of transpiration, rains replenish the groundwater until stream flow resumes (late November - early December) and inundation of the floodplain occurs (mid to late December). At any time during the growing season, very heavy or prolonged rains may result in short-term inundation of the floodplain. TR- 20

-8 km -- -- _- -

Figure 1. Map of the study region showing location of the streams and sampling stations. \\ LEGEND

Figure 2. Map of the Creeping Swamp study area showing the main stream, tributaries and sampling stations. Land use as of January, 1978. Winner and Simmons (1 977) constructed a hydrologic budget for Creeping Swamp watershed. The sole source of water is precipitation; deep groundwater input is probably very small because aquifer boundaries coincide with watershed drainage boundaries. Of the incoming precipitation, 37% leaves the watershed as stream runoff and an estimated 2% as deep groundwater loss. The remaining 61%, estimated by difference, is attributable to loss as evapotranspiration from the watershed. In view of the relatively uniform precipitation throughout the year, the influence of evapotranspiration on watershed hydrology and thus on water-borne transport of materials is of major proportions and may be a controlling variable of the flux of materials out of the watershed.

Creeping Swamp watershed is underlain by unconsolidated clays, sands and calcareous sediments of the Quaternary, Tertiary and Cretaceous Systems (Sumsion 1970). Four geological units make up the uppermost sediments: 1) Quaternary deposits; 2) the Yorktown formation; 3) the Pungo River formation; and 4) Castle Hayne limestone (Winner and Simmons 1977). Quaternary deposits occur at elevations greater than 12 m above sea level and thus do not occur in the swamp floodplain. These deposits consist of sandy silts and clays laid down by streams. The Yorktown formation under- lies the Quaternary deposits at high elevations and is exposed at lower elevations, specifically within the swamp floodplain. It is composed of marine-deposited clays, silts, and sands, is about 12 m thick, and is the medium for exchange between ground and stream waters (Winner and Simmons 1977). The Pungo River formation is composed of phosphatic sands mixed with clays and silts and is about 3 m thick. Below the Yorktown and the Pungo River formation is Castle Hayne limestone, an important aquifer in the North Carolian Coastal Flain; water from this aquifer moves up into the swamp floodplain because of a hydrostatic head (Winner and Simmons 1977).

Soils in the Creeping Swamp floodplain are inorganic, highly weathered, poorly-drained clays and silts formed either on Coastal Plain sediments (Byars loam), on recent or local a1 luvium (Bibb complex) or on a1 luvial sediment (Roanoke silt loam) (Soil Survey, Pitt County 1974). These soils are classified as Ultisols in the most recent soil classification system.

The tree canopy of the swamp floodplain is composed of hardwoods dominated by Nyssa sylvatica var. biflora, Acer rubrum, Fraxinus caroliniana, Nyssa aquatic=- Liquidamber styraciflua mnzler, --et a1 . 1977). Along the stream channel, F. caroliniana and N. aquatica are most abundant. The forest was partial ly-1 ogged about 40 years ago. Sap1 ings are dominated by A. rubrum and -F. caroliniana. Herbaceous growth is sparse but includes ~auhruscernuus, Carex spp. and Hypericum virginica; vines (Smilax spp. ; Berchemia scandens, Rhus radicans, Mi tchel la repens, Decumaria barbara) and shrubs'(Leucothoei 11 ari s, Vaccinium corymbosum))'ewhatre abundant. At low elevations in the floodplain, several bryophytes-.- occur extensively: the aquatic moss, ~ontina~issppi and the l iverworts, Pore1 la pinnata and Pallavacinia lyelli metal. unpubl. ). In the late winter and early", sprinq, a bloom of filamentous alqae develops in the swamp floodwaters; typical siiecies include the filamentois desmids. Hyalotheca dissiliens and Desmidium aptogonum and others, filamentous diatoms such Eunotia pectinalis, solitary diatoms, and filamentous greens such as spp. , Oedogoni um spp. and Bul bochaete spp. (verified by P.

The Creeping Swamp watershed remains about 65% wooded (Kuenzler, et -al. 1977). Of the cleared areas, some are pine plantations and some a= croplands. Associated with agricultural fields are drainage ditches; in addition, some of the tributaries to the main stream have been channelized. Lumbering operations occur sporadically and have included, during the period of this study, floodplain as well as upland areas. Several animal husbandry operations are located in the watershed; two of them are near the floodplain study area. These operations markedly influence streamwater nutrient concentrations if adequate treatment measures are not taken. Structure of the report

The research reported here was conducted mostly by Dr. Patrick J. Mulholland and Dr. Laura Anne Yarbro and prepared as dissertations in partial fulfillment of the requirements for Doctor of Philosophy degrees. The two major chapters of this report are comprised of these dissertations, somewhat modified and condensed. The reader may wish to refer to these dissertations for more details. The final chapter represents research conducted by Dr. Leonard Smock during the final year of the grant.

2. ORGANIC CARBON CYCLING AND EXPORT by Patrick J. Mulholland INTRODUCTION The processing of materials and energy are among the principal functions of ecosystems. Energy flow through ecosystems drives material cycles . The cycl ing of most materials in ecosystems involves transfers from less concentrated, inorganic forms to more concentrated organic forms, and back to the inorganic state. Energy is temporarily stored at relatively high density when materials are concentrated in organic form. It is released and dispersed as organic materials decompose. Energy quality is determined in large part by the density at which it is stored. Ecosystem biota functions both as a temporary storage of relatively high quality energy (concentrated materials), and as a medium for energy degradation and materials dispersion; the latter is the cost of maintaining the storage. Natural ecosystems have important maintenance and support functions for human systems (Odum, E.P. 1969; Odum, H. T. 1971 ). Natural ecosystems modify climate, buffer against catastrophic meteroloyical events, assimilate and dissipate waste products of human activities, contribute to agricultrual productivity, and directly or indirectly provide man with important food and fiber products. As human activities demand modification or destruction of increasingly larger tracts of natural ecosystems, it is important to determine precisely their values in the natural condition, and those lost if modified. Understanding the natural flows of energy and materials in ecosystems is critical to this evaluation. Impor- tant material flows include those of critical plant nutrients, such as nitrogen and phosphorus, and of carbon, central in organic material processing and closely related to ecosystem energy flow. Small swamp-stream ecosystems are abundant in the southeastern United States. There is considerable pressure to channelize and drain these ecosystems. Between 1955 and 1973, the Soil Conservation Service completed 679 miles (1093 km) of channelization projects in North Carolina (Soil Conservation Service, 1973). Little is known about the natural flows of materials and energy in freshwater wetlands, and in particular,swamps, compared to most terrestrial and aquatic ecosystems. Past ecosystem 1 eve1 studies of wetlands have primarily focused on salt marshes, but freshwater wetlands have recently received attention by ecologists. While Howard-Mil 1 iams (1 972) found relatively high concentrations of major ions in a marsh bordering an African lake and Crisp (1966) found relatively large exports of nutrients from an eroding English peat bog, nutrient concentrations are usual ly relatively low in freshwater wet1 ands . Mi tsch, -et --a1 . (1 977) found general ly 1 ower concentrations for most ions in swamp waters than in the bordering river waters in a southern I1 1 inois riverine swamp. Brinson (personal communi - cation) has similar findings for an eastern North Carolina riverine swamp. Schlesinger (1 978) reported extremely low outputs of calcium, magnesium, sodium, and potassium from the Okefenokee Swamp in Georgia and Verry (1975) found low nutrient yields from a Minnesota bog. Evidence indicating freshwater wetlands may function as nutrient sinks, especially for nitrogen and phosphorus, has come from a variety of widely scattered areas, incl uding an African papyrus swamp (Gaudet l976), a Michigan marsh (Richardson, et al. 1976), a Wisconsin marsh (Lee, et al. 1975), a large South Carol ina riverine swamp (Kitchens, et a1 . 197g,and small North Carol ina swamp-streams (Kuenzler, --et a1 . 1977. Kuenzler, --et a1 . (1977) found that nutrient concentrations and exports were generally lower in natural swamp-streams than in channel ized streams in eastern North Carol i na .

Forested wet1 ands support large amounts of 1iving biomass (Schlesinger 1978; Smith, et al. 1975), but do not necessarily exhibit high primary productivi ty scmesinger 1978). The 1eve1 of primary productivity may be a function of hydrologic delivery of nutrient suppl ies (Gossel ink and Turner 1978). Re1 atively high productivi ties have been reported for swamps with substantial hydrologic inputs in North Carol ina (Bri nson l977), Louisiana (Conner and Day l976), and Florida (Burns 1978). How- ever, an Illinois riverine swamp was not highly productive (Mitsch, et a1. 1977). Swamp hydro1 ogic characteristics may a1 so enhance the export of net primary production. Although Brinson (1976) reported exports of organic carbon of from 0.3 to 5.2 g ~*m-~-~r-lfrom several upland watersheds, mostly forested, Day, et al. (1977) reported exports of 10.4 g ~*m-z*~r-lfrom a Louisiana swamp watershed. Wetlands have been reputed to have very inefficient recycling of carbon due to lack of abundant oxygen for respiration (Given 1975). Large accumulations of soil organic matter have been reported for a Minnesota bog (Reiners 1972), some North and South Carolina pocosins (Woodwell l958), and the Okefenokee Swamp in Georgia (Schlesinger 1978). However, a Louisiana swamp-forest, which dried periodically, did not possess large amounts of soil organic matter. This report concerns the cycling of organic carbon in a forested wetland little studied by ecologists -- a small swamp-stream (or floodplain swamp) ecosystem. A conceptual model of organic carbon flow in one swamp-stream, defining major inputs, outputs, and internal -ecosystem transfers,is shown in Fig. 3. The ecosystem here was defined to include 1iving vegetation and detrital storages both in the water column and within the top 25 cm of soil. Major inputs included net primary productivity and hydrologic inputs. Major outputs were respiration and hydrologic output. Transfer rates were primarily controlled by rainfall, temperature, season, and the sun. This model was the basis for development of the organic carbon budget for Creeping Swamp. An input-output, mass balance approach was used similar to that of Fisher (1977) in his Fort River study. Inputs, outputs and important internal transformations of organic carbon were measured. Inputs and outputs were grouped into hydrologic and biologic flows according to the process involved.

METHODS

Study area

Organic carbon concentration was monitored in three small natural swamp-streams and four channelized streams in eastern North Carolina beginning in 1975 (Fig. 1). Two or three sampling stations were establ ished on each stream. The natural swamp-streams were Creeping Swamp, Palmetto Swamp, and Chicod Creek. The channelized streams were Tracey Swamp, Clayroot Swamp, Grindle Creek, and Conetoe Creek. Sampling of Grindle and Conetoe Creeks was discontinued after 1975, and sampling of Clayroot Swamp did not begin until 1976. Streamflow was measured and water samples for organic carbon measurement were col 1 ected periodical ly, general ly at intervals of from two to four weeks. Daily samples were col lected at Clayroot Swamp during a few high streamflow periods in 1976.

One of the natural swamp-streams, Creeping Swamp, was selected for the intensive study. Stream sampling stations were set up at the upstream (CP-20) and downstream (CP-10) boundaries of the swamp, draining 32 and 80 km2, respectively. An internal stream sampl ing station (CP-14) was set up approximately 3 km downstream from CP-20. Sampling stations were a1 so establ ished on the five principle tributary streams entering the swamp between CP-20 and CP-10. They were set up at bridges or culverts within 2 km of the point at which they enter the swamp and were numbered TB-01 , TB-02, TB-03, TB-04, and TB-87 (Fig. 2). One area of the swamp was chosen as a location for much of the experimental and data collection activities. This area was chosen so as to be representative of the entire swamp and away from major roads or other human influences. The area selected was a 500 m length of swamp about 3 km downstream from CP-20 and encompassing CP-14. This area will hereafter be referred to as the study site.

Inundation patterns

In order to partition the swamp into terrestrial (unflooded) and aquatic habitats, inundation patterns were studied. Five transects (Nos. 1, 2, 3, 4, 5) were established across the swamp at intervals along its length (Fig. 2). Linear relationships between water level at CP-10 and fraction of the transect inundated were developed in one of two ways. First, three of the transects (1, 4, 5) were paced periodically and the number of dry and flooded paces determined at various water levels. Second, the remaining two transects (2, 3) were paced during one period of high water and the water depths at each pace were recorded. Using these depths, fractions of the transects inundated were computed for various water levels below that on that date. Inundation data from each transect were weighted by transect length and combined to develop whole-swamp inundation fractions at various water 1eve1 s at CP-10. Daily inundation fractions were then computed from the CP-10 daily water level record. Biomass

Living:

The trees and saplings of Creeping Swamp were studied in August 1974 using a point-quarter analysis (Cox 1967) along three transects (transects 1, 4, and 5; Fig. 2). Trees were defined as >10 cm diameter at breast height (DBH) and saplings were 2.5 to 10 cm DBH. The shr b and herbaceous plant stratum was sampled by harvesting 15 plots (1 m Y ) across each of two transects in September 1976 and June and September 1977. Aboveground parts of non-woody plants and Smilax spp. vines were harvested compl etely. Woody plants were harvested and separated into wood and 1eaf fractions. A1 1 col 1ected material was separated by species or general plant group, dried at 85 C, weighed, and subsamples combusted at 550 C (-+50 C) for 4 h to determine ash-free dry weight (AFDW). Soi 1 detritus :

Swamp floor litter was measured seven times between July 1976 and December 1977 by sampl ing 30 plots (0.02 m2) along a transect across the swamp in the mid-swamp study site according to a stratified random sampling procedure. Litter was placed in plastic bags and returned to the laboratory where it was separated into wood and non-wood fractions, dried, subsampled, and combusted to determine AFDW.

Soil organic carbon (below the litter layer) was measured to a depth of 0.25 m in April 1977 and May 1978 by L. A. Yarbro. Litter was removed and 5 cores were taken at each of 15 sites along a transect across the swamp in the mid-swamp study site. Each core (7 cm2) was divided into two sections according to depth (0 to 5 cm and 5 to 25 cm) and cores from each site were combined. Core sections were dried, subsampled, and combus ted to obtain organic matter content (AFDW) .

-Precipitation and s treamf 1ow Precipitation was measured and recorded daily at a site approximately 5 km southeast of Creeping Swamp. Precipitation was collected for chemical analysis using palyethyl ene funnel col 1ectors (0.02 m2) emptying into plastic bottles mounted 1 to 1.5 m above the ground. Nylon mesh (0.25 mm) was placed over the funnels to exclude debris. No appreciable 1eaching of organic carbon from the col 1ection bottles was detected during a 2-week incubation of distilled water. Two collectors were placed in a field near the study site and a third collector was placed in a clearing near CP-10. Coll ections for organic carbon analysis were made for 2-week periods monthly from February to August 1978.

Streamflow was continuously monitored and recorded by the United States Geological Survey (U.S.G.S.) at both CP-20 and CP-10. Daily water levels, daily streamflow, and annual runoff were provided. In August 1977 the U.S.G.S. discontinued monitoring CP-20 streamflow. Stream- flow was thereafter estimated from stage-discharge relationships provided and stages on sampling dates. Streamflows were computed periodically for the five major tributary streams by measuring velocities (Gurl ey Model 622-E current meter) and depth profiles. Water levels were read from staff-type indicators or were measured from bridge rail ings. Stage- streamflow relationships were constructed to determine streamflow on sampl ing dates when it was not measured directly. Water samples were collected at CP-14 and CP-20 approximately once every three weeks when there was streamflow during 1975. In 1976 and 1977 samples were collected biweekly during the flow period and daily during a number of spates. Samples were collected sporadically in 1978. Tributary streams were sampled biweekly when streamflow was appreciable during 1976 and 1977. Samples were also collected daily during a number of spates. A few samples were collected at low streamflows. Organic carbon analysis Total organic carbon (TOC) in water was partitioned into three size fractions, coarse particulate organic carbon (CPOC) , greater than 0.25 mm, fine particulate organic carbon (FPOC), between 0.45 vm and 0.25 mm, and dissolved organic carbon (DOC), less than 0.45 vm, FPOC was not measured directly. Fine total organic carbon (FTOC), less than 0.25 mm, was measured and FPOC was computed as the difference between FTOC and DOC. Coarse particulate organic matter (CPOM) transported in stream and tributary waters was measured periodically, particularly at high streamflows, using a large (0.75 by 0.30 m at mouth), flow-metered net (0.25 mm mesh). Material coll ected was returned to the laboratory, dried, subsampled, and combusted to determine AFDW as described earlier. CPOC was assumed to be 50% of AFDW (Jordan and Likens 1975). Samples passed through 0.25 mm nylon mesh for FTOC measurement and filtered samples (pre-washed, Gelman type GA-6) for DOC measurement were placed in acid- washed glass bottles and immediately frozen or put on ice to be frozen upon arrival at the laboratory (less than 24 h). Samples remained frozen until measurement (0.3 m in length), and net leaching by throughfall and s temf 1ow. Annual tree and sapling net wood increment: In order to estimate annual aboveground net wood increment of living trees and saplings, spring-connected metal bands were placed around 65 trees of various sizes (DBH) and species at about 1.5 m above the ground in March 1975. The trees were located across the swamp in the study site and were chosen to represent a sequence of sizes of seven of the most important species in Creeping Swamp. Increases in tree circumference were measured annually by determining band expansion. Increases in DBH were computed and translated into annual net stem and branch wood increments using allometric equations of Dabel and Day (1977) constructed for hardwoods in The Great Dismal Swamp, Virginia. Regressions relating net annual wood increment in 1977 and tree DBH were computed for each of the seven species banded and a combined wood increment-DBH re1 ationshi p was computed to allow computation of net annual wood increment of non-banded species of trees. Total swamp wood increment for 1977 was computed using the point-quarter data and the various wood increment-DBH re1 ationships.

Li tterfall and macro-1 i tterfal 1 : Litterfall of plant debris ~0.3m in length from sources more than 1.5 m in height was measured at 3- to 4-week intervals from March 1975 to March 1978. During peak 1 i tterfall in late October and early Novenlber the collection interv 1 was shortened to 1-2 weeks. In 1975 and 1976 ten collectors (0.2 n~I suspended muslin bags) were placed in a stratified random fashion across transects 1 and 4, and 9 collectors across transect 5. After March 1977 only transect 4 was used and 5 additional collectors were added bringing the number of collectors to 15. Litter was returned to the laboratory, dried, weighed, and subsamples ashed to determine both total dry weight and AFDW. Fall of branches and other plant parts greater than 0.3 m in length were measured by sampling six initially cleared quadrats (25 m2), three near CP-20 and three in the mid-swamp study site, at approximately 6 to 8-week intervals from September 1976 through October 1977. Data from 1 November 1976 to 31 October 1977 was taken to be 1977 input. Material coll ected was brought back to the laboratory, weighed fresh, subsampled, dried, reweighed, further subsampled, and combusted to determine both total dry weight and AFDW. Throughfall and s temf 1 ow: Fifteen throughfall coll ectorS, similar to those used for collecting precipitation, were placed along a transect near the mid-swamp study site in a stratified random design. Collections were made biweekly for volume between September 1976 and August 1978. Samples for organic carbon analysis were collected every other collection period, i.e., for one 2-week period every 4 weeks. Samples were preserved in the field during the collection period by the addition of 10 ml of 0.015 -N HgC12 to the collection bottle at the beginning of the period. This gavz a Hg++ concentration of >40 mg-1-1 after throughfall for most periods.

Ten trees in the mid-swamp study site were collared about 1.5 m above the swamp floor with a polyurethane sealer and the stemflow directed into polyethylene funnels emptying into covered large plastic garbage cans. Stemflow samples were preserved in the field with an initial addition of 50 to 100 ml of 0.1 N H C12 in each collector can, resulting in a HgC+ concentration of 40mg~l-q after stemflow for most periods. Collections were made at approximately bimonthly intervals during 1977 and the first half of 1978. Samples for throughfall and stemflow organic carbon analysis were placed in acid-washed glass bottles and frozen until analysis.

Shrub and herbaceous plant net productivity:

Annual net productivity of the shrub and herbaceous 1ayer (

A1 gal distribution and productivity:

Algal biomass distribution was measured monthly during March, April, and May in 1979. Benthic algae were sampled by taking cores (0.0016 m2) of the swamp floor where flooded at stations spaced uniformly across one or two transects approximately 100-200 m downstream from CP-14. Material collected from the swamp floor surface was placed in plastic bags, put on ice, and returned to the laboratory. Algae were carefully washed from 1i tter surfaces onto glass fiber f il ters and chl orophyl 1-a and phaeophytin measured according to the method of Lorenmen (1967).

Algal productivity was measured using a modification of the one- station die1 procedure (Odum 1956; Odum and Hoskins 1958). Dissolved oxygen was measured (Amer. Pub1 . Health Assoc. 1975) at CP-10 in the evening, early morning, and two or three times throughout the day. Gross primary productivity (GPP) was computed from dissolved oxygen increases during the day plus the oxygen consumed in respiration at night. Pro- duction and respiration rates were corrected for diffusion of oxygen into the water from the atmosphere. A surface diffusion coefficient was measured using the nitrogen-filled plastic dome method of Copeland and Duffer (1965). A second computation was made using physical and hydrau- 1ic properties of the stream (Tsivoglou and Wallace 1972). Algal productivity measurements were made periodically during the months of January through June in 1976, 1977, and 1978. Rates were converted from oxygen to carbon units assuming a productivity quotient of 1.2. Net primary productivity (NPP) was assumed to be 60% of GPP (see also Likens 1973). Values were averaged monthly and whole-swamp monthly algal NPP was computed by multiplying mean monthly NPP rates by mean monthly inundation fractions. Biologic outputs Biologic outputs include (1 ) respiration in the swamp floor when inundated, both a) aerobic, referred to as aquatic benthic respiration (Ra), and b) anaerobic; (2) respiration within the water column (Rwc); and (3) respiration in the 1 itter layer when dry, referred to as terrestrial litter respiration (Rtl). Total soil respiration (Rts) was also measured under terrestrial conditions; however, it could not be partitioned between root and heterotrophic components. Aquatic benthic and water column respiration: Aquatic benthic respiration was measured by collecting cores of the swamp floor in plexiglass cylinders (0.14 m in diameter, >10 cm in depth), transporting them to the laboratory, reflooding them with swamp water, and measuring changes in dissol ved oxygen (Mi nkl er method) during incubations at three temperatures (approximately 5, 15, and 25 C). Six cores were collected and incubated simultaneously at approximately bimonthly intervals from March 1978 to January 1979. Cores were initially incubated at a temperature near ambient water temperature, except for the cores collected in August, which were incubated at 5 C first. Cores were per- mi tted to accl imate for at least 24 h prior to measurement of respiration rates at all temperatures and at least 48 h prior to the initial measurement. An experiment performed on cores collected 8 August 1978 to determine if length of equilibration at a particular temperature affected respiration rates indicated that a 1 -day equilibration period was suff i- cient. Rates measured after 1 day of incubation at 23 C were within 6% of those measured after 3 days in five of the six cores and there was no pattern as to which incubation gave higher rates. Respiration in the water col urnn was measured by incubating rep1 icate bottles of swamp water in the swamp for 24 to 48 h and measuring changes in dissolved oxygen concentration. Measurements were made one or two times monthly from November 1976 to April 1978. Terrestrial 1 i tter respiration: Respiration in the swamp floor litter layer when unflooded was measured at five to ten sites periodically from 3 August 1977 to 19 July 1978 (Mu1 hol 1 and 1979). Respiration rates were measured by is01 ating litter and measuring carbon dioxide (C02) evolution using an alkali absorption technique (Lieth and Ouel lette 1962; Wi tkamp 1966). Litter collected from either a 0.02 m2 or 0.008 m2 area was placed into a metal cyl inder (0.02 m2) and the ends sealed by covering with Saran (R, plastic wrap held with rubber bands. A cup containing 15 to 25 ml of 1 N KOH was added to the chamber at the beginning of the incubation to ab-sorb evolved C02. An empty chamber containing only the KOH absorbent was incubated simultaneously as a blank. Chambers were incubated in the field for 6 to 15 h. Laboratory studies with field apparatus permitted computation of a C02 absorption coefficient (kabs. ) (by the KOH) of 2.3. Correction factors were computed to account for incomplete C02 absorption during field incubation (Mu1 holland 1979). For incubations >6 h and kabs. >1.0, the correction factor was relatively independent of C02 production rate and kabs.. Following field incubation the KOH was transferred to glass bottles and titrated within 24 h to determine C02 content. A few ml of 2 N BaCl2 was added to precipitate the absorbed carbon and the sample titratex with 1 N HC1, first to the phenolphthalein, and finally to the brom-cresol green endpoint (Maciol ek 1962). For measurements made 4 November 1977 and thereafter the incubated litter was returned to the laboratory, weighed, dried, reweighed, subsampled, and combusted to determine both moisture content and AFDW.

Total soi 1 respiration:

Total C02 evolution from the swamp floor when unflooded, referred to as total soil respiration (Rts), was measured at two to seven sites on 17 dates from 8 December 1976 to 19 July 1978. Respiration was measured by inserting a metal cylinder (0.02 m2) into the swamp floor, adding a cup of 1 N KOH (20 to 25 ml ) , covering with Saran wrap, and incubating for 6 to % h. Hand1 ing and analysis of the KOH after incubation to determine the amount of C02 absorbed was identical to that described for litter respiration. Both soil and air temperatures were taken using a mercury thermometer.

Anaerobic respiration:

Dissolved methane evolution from inundated areas was measured periodically from March 1978 to February 1979. Polyvinyl chloride cylinders (0.021 m2) were driven into the swamp floor, leaving a 2 to 5 cm lip exposed, and allowed to equilibrate for at least 1 week. Cylindrical plastic sheaths, extending above the water surface, were slipped over the imbedded cylinder lips and fixed with rubber bands. The water column was isolated from the atmosphere by floating a piece of styrofoam fit snugly into the plastic sheath. Stirring rods were mounted in the styrofoam tops to enable mixing of the water column before the final sample was drawn. Water samples were collected in 60 ml plastic syringes fitted with 3-way valves before and after the incubation period (3 to 6 h), placed in ice water, and analyzed for dissolved methane concentration within 24 h. Dissolved methane was measured by gas chromatography (Perkin-Elmer Model 900) after nitrogen stripping and subsampl ing of the nitrogen-methane mixture. Particulate organic carbon formation POC formation due to precipitation of DOC as well as bacterial growth was measured periodically from November 1977 to May 1978. Swamp water was collected and a portion incubated in glass bottles, wrapped in opaque plastic, and placed in the swamp for 30 to 80 days. Another portion was returned to the 1aboratory , f i1 tered through pre-combus ted Gel man Type A/E glass fiber filters, dried at 85 C, weighed, recombusted, and reweighed to determine particulate AFDW initially in the water. After incubation the first portion of water was returned to the laboratory and treated in the same manner to determine the concentration of particulate AFDW. Increases in AFDW during the incubation were assumed entirely to be precipitated dissolved organic matter, half of which was assumed to be DOC. This technique certainly overestimates DOC precipitation because it does not differentiate between precipitation and bacterial growth.

Dissolved organic carbon 1eaching from ground li tter

Rapid leaching:

The amount of readily leachable material from newly fallen leaves of various tree species was measured in the fall of 1978. Leaves were picked from trees just prior to abscission and grouped according to species. All were air dried and weighed, and subsamples were dried at 85 C 2nd reweighed to determine dry weight. Leaves were immersed in distilled water for 1 to 7 days at 23 C, dried again at 85 C, and reweighed to determine dry weight loss.

The readily leachable material in newly fa1 len 1eaves may be leached by surface water or throughfall as it percolates through the ground litter when the swamp is unflooded. In the fall of 1976 and winter of 1977 simul taneous experiments were performed at Creeping Swamp and at Chapel Hill, N.C. to determine the rate of throughfall leaching of newly fallen leaves. Newly fa1 len red maple leaves were collected in Creeping Swamp just prior to abscission and known amounts placed in nylon mesh bags (1 mm mesh). One group of bags was placed in a relatively high area of the swamp and a second set was placed in a wooded area near Chapel Hill. Two or three bags were collected periodically, dried at 85 C, and reweighed to determine dry weight loss. Throughfall was also measured during each collection period at the incubation sites with collectors similar to those described earl ier.

In November 1978 an experiment was performed to estimate the amount of readily leachable material that was biologically readily utilizable and quickly respired. Three swamp floor cores were collected, reflooded with swamp water, and incubated at 25 C. Measurements of dissolved organic carbon concentration were made before and after the incubations to determine background leaching rates. The water was then drained and each core was reflooded with leachate derived from newly fallen red maple leaves leached by distilled water for 12 and 24 h periods immediately preceding the incubatlon. Dissolved organic carbon cancentration of the leachate in each core was measured for a period of 7 days under aerobic conditions . Slow leaching: Rates of slow, long-term leaching of DOC from organic material on the swamp floor were measured during May, August, and November 1978, and January 1979. Swamp floor cores (0.14 m in diameter, >10 cm in depth) were taken, returned to the laboratory, reflooded with swamp water, and incubated at two or three temperatures with continuous aeration. DOC concentration was measured before and after 24 to 60 h incubation periods. Leaching rates per unit of swamp floor were computed from increases in DOC concentration. A total of six cores were collected and incubated in each experiment.

Additional methods AFDW was taken to be organic matter and organic carbon was assumed to be 50% of organic matter by weight (Westlake 1963; Reichle, et al. 1973). Statistical analysis of data was performed using Statistical Package for the Social Sciences (Nie, et al. 1970). Sample variability is throughout expressed in terms of one standard error (S. E.) from the mean value (2 + S.E.).

RESULTS

Inundation patterns Inundation patterns in Creeping Swamp were directly related to water level fluctuations. Mean monthly inundation fractions ranged from 0.64 in January 1976 and March -1977 to 0.0 in Octdber 1976 (Tab1 e 1) . Flooded and dry seasons were clearly evident. The swamp was, on average, 50% flooded during December, January, February, and March in 1976 and 54% flooded during these months in 1977. During June through October the swamp was flooded, on average, 13% in 1976 and 21% in 1977. A six-year mean daily inundation record for Creeping Swamp shows, even more clearly, the seasonal variation in inundation (Fig. 4). Salient features of seasonal inundation include: (1 ) re1 atively sharp increases in inundation during late fall, (2) relatively stable and high winter inundation, (3) gradual declines in inundation during spring, and (4) very low summer inundation. The dip in February, and peak in late May and early June in mean inundation evident in Fig. 4, are probably a result of particular weather patterns during the six-year period 1973-1978, and do not reflect long-term swamp inundation patterns. Precipi ation recorded at a site 25 km away averaged between 9 and 12.4 cmemo-r during 1973-1978, with the exception of February (7.95 cm) , October (7.37 cm) , and November (6.40 cm) (National Oceanographic and Atmospheric Administration 1973-1 978). Sea- sonal variation in inundation was due to the seasonal changes in evapotranspiration and not precipitation.

Table 1. Mean monthly inundation fractions of Creeping Swamp in 1976 and 1977. - - Month !976 1977

January February March Apri1 May June July August September October November December

Living biomass

Trees and saplings: Red map1 e (Acer- rubrum), ash, mostly water ash (Fraxinus carol iniana) , and black gum (Nyssa sylvatica var biflora) were the most abundant trees, and red maple and ash the most abundant saplings in Creeping Swamp (see also Mulholland 1979). In terms of basal area, black gum, swamp chestnut oak (Quercus michauxii), tupelo gum (Nyssa aquatica), and red maple qom- inated the tree stratum. Total swamp stem density was 705 treesoha' and 1206 sapl ings*ha-1. Average swamp tree diameter was 28.1 cm (DBH) and average sap1 ing diameter was 5.6 cm (DBH). Using the double logarithmic regressions relating tree biomass to DBH reported by Dabel and Day (1977) for Great Dismal Swam hardwoods, I computed a total tree biomass of 26.7 kg dry weightem-8 (13.4 k t ern-^) and a total sapling biomass of 0.9 kg dry weightem-2 (0.5 kg Gem-j) (Tab1e 2).

Table 2. Estimated biomass (dry weight) and density for trees and saplings in Creeping Swamp. Average density was computed from point- quarter data. Biomass was computed using diameter-weight regressions of Dabel and Day (1977)".

Component Trees Sap1 ings

Average Biomass: Leaves 10.1 kg*tree-' 0.3 kg*sap-' Branches 59.0 kg. tree-' 1 .6 kg- sapm1 Stem 31 0 kg* tree-' 5.7 kg.sapm1 To ta l 379 kg* tree-' 7.7 kg. sap-' Average Swamp Dens i ty 705 trees. ha- ' Average Swamp Biomass 26.7 kg*m-2 0.9 kg*m-2

*log (stem dry weight, kg) = 2.4064 log(DBH, cm) - 1.0665. log (branch dry weight, kg) = 2.1880 log(DBH, cm) - 1.4297. log (leaf dry weight, kg) = 2.1516 log(DBH, cm) - 2.1381.

Approximately 60% of the swamp was flooded on the date the point- quarter analysis was made. This flooded portion roughly corresponds to the lower elevations of Creeping Swamp which are inundated much of the winter and spring. For the years 1975-1977 this portion was flooded an average of 41% of the time during the period December through May. Dividing the point-quarter data into low and high swamp portions according to the presence or absence of flooded conditions at each point permitted analysis of the effects of inundation pattern on tree distribution and size. While some tree species were not numerous enough in either area to draw conclusions, changes in relative density and basal area among the more abundant species from low to high areas were evident (Table 3). Red maple, ash, tupelo gum, and bald cypress (Taxodium distichum) were decidedly more dense in the lower area and black gum, ironwood (Carpinus carol iniana), water oak (Quercus nigra), and swamp chestnut oak were a1 1 relatively more dense in the higher area. Total tree density was 58% greater and average tree diameter (DBH) was about 9% greater in the lower area. Average tree biomass for the lower are was 33.4 kg dry weightem-2, or 93% greater than the 17.3 kg dry weightam-q for the high area.

Table 3. Relative density and basal area for species of trees in the low and high areas in Creeping Swamp as computed from point-quarter data coll ected in August 1974. Average tree density for the low area was 917 treesoha-1 and that for the high area was 580 trees. ha-1 . Average tree diameter (DBH) was 29.4 cm and 27.0 cm for the low and high areas, respectively."

Relative Re1 ative Number Dens i ty Basal Area Species low high low high low high

-Acer rubrum L. 24 21 .250 .I81 .I14 .I13 Nyssa s lvatica var biflora"r Wal ter) Sargent Fraxi nus spp. Liquidambar styraciflua L. Carpinus caroliniana Walter Quercus nigra L. Quercus michauxi i Nu ttal 1 Nyssa aquatica L. Taxodium dis tichum (L. ) Richard I1 ex opaca Ai ton Ulmus americana L. Quercus laurifolia Michauxii Quercus phellos L. Liriodendron tul ipifera L. Cra taegus spp. Popul us heterophyl 1a L.

Total

*Species from Radford, --et a1 . 1968. Shrubs and herbaceous pl ants : Creeping Swamp has a shrub and herbaceous plant stratum varying widely in density and composed of various species of tree seedlings, vines milax ax spp., ~erchemiascandens (Hill) K. Koch, Rhus radicans L., and Decumaria barbsffernsoclea sensi bil is rand Osmunda regal is var. spectabil is (Wil ld. ) ~raylmes(Carex spp. ), virginia willow (Itea virginica L. ) , leucothoe (Leucothoe axillaris (Lam. ) D. Don), cane (fidinaria i antea (Wal ter) M'mzard' s tai 1 (Saururus cernuus L.), violets ?Viola spp. ), and other less common speci-also Kuenzler --et al. 1977). Total aboveground biomass ranged from 36.9 g AFDW*~-~on 13 September 1976 to 42.4 g AFDW~~-~on 22 June 7 977, with Smilax spp. and Leucothoe axil laris dominating during a1 1 harvests (Mu1 hol land 1979). Swamp floor detritus Ground 1 i tter: Standing crop of ground litter in Creeping Swamp, excluding boles, varied seasonally, exhibiting a sharp increase following autumn 1eaf-fall , and a decline thereafter (Fig. 5). Leaves and leaf fragments comprised from 55 to 80% of total ground 1 itter; this fraction was generally highest in inter and lowest in late summer. Feak standing crop was 1068 g AFDW*~-' of total litter and 856 g AFDWsm- of leaf 1itter on 17 February 1977. Minimum levels were 825 and 560 g AFDW~~'~of total and leaf litter on 20 September 1977. The standard errors for adjacent measurements frequently over1 ap; however, the trend was unmistakable. From 17 February 1977 to 20 September 1977 there was a net decomposition of approximately 250 g AFDW*~-2(125 g ~.m-2), most of which was leaf 1 i tter. The decomposition rate appeared to be faster from 17 February to 28 April than thereafter. There was a net inwt of 1 itter durina the ~eakleaf- fall period in 1977 (20 September to' 21 November) of 156 g AFDW~~-2. On 29 July 1979 an estimate of detrital boles on the swamp floor was made by measuring dimensions of all logs within two 0.25 ha quadrats and classifying them into one of three categories depending upon the degree of decomposition. Three logs in each decompositional classification were subsampled and wood density was measured. Biomass of dead boles on the swamp floor was 240 gem-2 and 520 g-m-2 in the two quadrats sampl ed.

Soil : The soil profile in Creeping Swamp below the 1 i tter layer consisted of a thin organic layer of about 1-5 cm, underlain by sand and clay. Plant roots were shallow and small roots and root hairs penetrated into the litter layer. There was no evidence of large scale sedimentation or humus formation. The top 5 cm of swamp soil, below the 1 itter layer, had a mean organic matter content of 17.3(+1.7)%in April 1977 and 16.5(+2.0)% in May 1978. The soil profile between 5 and 25 cm had a mean organic content of 7.4(+1.6)% in 1977 and 5.5(+0.8)% in 1978 (Yarbro 1979). Total soil organic matter to a depth of 25 cm, excluding 1itter, averaged 14,580 (+890) g ~~i3W.m-2(7,290 g corn-2) in April 1977 and 15,180 (+1000) g AFDW*~-2(7,690 g corn-2) in May 1978.

Hydrologic inputs and outputs

S treamf 1ow:

Mean annual runoff for the period 1972-1976 was 34.1 cm at CP-10 and 32.6 cm at CP-20 (as computed from U.S. Geological Survey Water Data, 1972-1977). In 1976, a drier than average year, runoff was 22.3 cm at CP-10 and 17.6 cm at CP-20 (U.S.G.S. 1978). In 1977 runoff totalled 41.1 cm at CP-10 and was estimated to be 0.96 (average fraction for 1972-1976) times that, or 38.7 cm, at CP-2P. Runoff entering the swamp from tributaries, ditches, and non-point sources between CP-20 and CP-10 was assumed to be the same as runoff at CP-20 since the drainages were similar in vegetation type and land use.

Streamflow in Creeping Swamp exhibited distinct seasonal variation (Fig. 6). December through April normal ly comprised the high-fl ow period. Approximately 1.5 cm of rainfall produced spates. This period usually consisted of a series of sharp increases and gradual declines in streamflow. During the low-flow period, which generally coincided with the growing season, the floodplain was usually unflooded and streamflow often ceased for extended periods. Approximately 2.5 to 3.5 cm of rainfall were required to produce significant spates. Storms similar in size to those during the high-flow period produced much smaller spates. Spates during 1976-1977 had recurrence intervals of < two years,with the exception of the spate in early November 1977, whicn was about a 7-year event (U.S.G.S. personal communication). Since March 1971, when CP-10 was initially gaged, a 10-year spate event occurred on 2 October 1971 and an 8-year event occurred on 2 April 1973. All other spates during that period have been 2-year events. Organic carbon concentration:

FTOC concentrations at CP-20, CP-14, and CP-10 varied seasonally (Fig. 7). Concentrations ranged from 5.8 mg*l-1 at CP-20 on 19 January 1977 to 73.0 mg.1-7 at CP-14 on 9 November 1976, under stagnant conditions. Lowest concentrations were found December through February. There were steady increases in concentrations from March to June and they remained high but fluctuated widely July through November. Highest concentrations occurred under stagnant conditions during the summer and autumn months and at relatively high streamflows in November foll owing peak 1 itterfall . DOC was by far the most abundant organic carbon fraction in streamflow (Table 4). It comprised greater than 85% of total organic carbon (TOC) at all stations during 1976 and greater than 93% at CP-20 and CP-10 Figure 6. Annual record of streamflow and floodplain inundation at CP-10 during (a) 1976 and (b) 1977. Heavy lines along baseline indicate no streamflow.

34 CP-I0 A CP-14 0 CP-20 CP-IOFLOW - STAGNANT CONDITIONS

Figure 7. Seasonal patterns of fine total organic carbon (FTOC) concentration at three stations in Creeping Swamp and streamflow at CP-10 from January 1975 to May 1978. during 1976 and 1977. Mean nnual DOC concentrations, weighted by streamflow, ranged from 10.0 mg*1-8 at CP-20 in 1976 to 23.9 mg.l-l at TB-07 in 1976. Weighted mean annual DOC concentrations in some of the tributary streams in 1977 1i kely surpassed 23.9 mg.1-1; however, FPOC and DOC were not separated on one high-flow sample date and computations could not be made. FPOC was generally the next most abundant organic carbon fraction, ranging from 0.1 mg-1-1 at CP-20 in 1977 to 3.1 mg-1-1 at TB-02 in 1976 (Table 4). Weighted mean annual FTOC concentrations, the sum of DOC and FPOC, ranged from 10.5 mg*l-1 at CP-20 in 1976 to 32.2 mg-1-1 at TB-07 in 1977. Weighted mean CPOC concentrations were <1 mgel-1 at all stations (Table 4).

Table 4. Weighted mean annual organic carbon concentrations in Creeping Swamp and tributary streams. Percent of TOC in parentheses.

DOC FPOC CPOC* Year Station (mg-1-' ) (rng.1-' ) (mg*l-' )

1976 CP-10 14.8 (96.7%) 0.4 (2.6%) 0.1 (0.7%) CP-20 10.0 (93.5%) 0.5 (4.7%) 0.2 (1.8%) TB-01 12.8 (93.4%) 0.2 (1.5%) 0.7 (5.1%) TB-02 18.8 (85.7%) 3.1 (14.0%) 0.2 (0.9%) TB-03 10.3 (91.2%) 0.7 (6.2%) 0.3 (2.6%) TB-04 15.5 (95.2%) 0.6 (3.6%) 0.2 (1.2%) TB-07 23.9 (86.9%) 3.0 (10.9%) 0.6 (2.2%)

1977 CP-10 19.7 (94.7%) 1.0 (4.8%) 0.1 (0.5%) CP- 20 17.4 (98.3%) 0.1 (0.6%) 0.2 (1.1%) TB-01 23.7 (DOC + FPOC) 0.7 (2.9%) TB-02 19.8 II 0.2 (1.0%) TB-03 18.9 I I 0.3 (1.6%) TB-04 20.9 It 0.2 (1 .O%) TB-07 32.2 I I 0.6 (1.8%)

*weighted mean for data collected from January 1976 to January 1978

DOC and FPOC concentrations at CP-20 and CP-10 were significantly correlated (P <0.01) with a number of physical parameters and chemical species, among these color (positive) , pH (positive) , temperature (posi - tive), dissolved oxygen (negative) and a number of metals (positive) (Mu1 hol land 1979). Negative correlations between DOC concentration and loglO (streamflow) were significant but weak at both CP-20 and CP-10 (r = -0.46 and -0.38, respectively). DOC concentration was positively correlated with loglo (streamflow) at TB-03, TB-04, and TB-07 (r = 0.70, 0.60, and 0.83, respectively). Similar correlations were not significant at TB-01 and TB-02. Correlations between FPOC concentration and log10 (streamflow) were significant at CP-10 only (r = 0.33). CPOC concentra- .tion was both highly and positively correlated with streamflow (r = 0.84) and loglO (strearnflow) (r = 0.83), but only at CP-10.

Streamflow and groundwater organic carbon flux:

The flux of organic carbon via streamflow at CP-20 and CP-10 was computed by multiplying weighted mean annual DOC, FPOC, and CPOC concentrations by total annual streamflow, and dividing by the area of swamp (3.2 km2). The flux of organic carbon into the swamp via runoff entering the swamp between CP-20 and CP-10 (tributary input) wasestimated from weighted mean annual FTOC and CPOC concentrations computed for each of the five major tributaries, averaged in proportion to annual flow in each, and mu1 tip1ied by total runoff entering the swamp between CP-20 and CP-10 (in volume units) . Overall tributary weighted mean DOC concentration was computed to be 13.5 mgel-l in 197 and 18.9 mgml-1 in 1977, weighted mean FPOC concentration was 1 .O mg*l-f in 1976 and 1.8 mg.1-1 in 1977, and weighted mean CPOC concentration was 0.3 mg*l-l in both years.

In a study of the hydrology of the Creeping Swamp watershed upstream of CP-10, Winner and Simmons (1977) reported that the segment of swamp between CP-20 and CP-10 received inputs of groundwater from the uppermost major confined aquifer, the Castle Hayne Limestone. The Castle Hayne originates near the western boundary of the Creeping Swamp watershed,dips toward the east,and receives recharge from high elevation areas within the watershed. Upward movement of water from the Castle Hayne to the low- lying Creeping Swamp floodplain is thought to be mostly derived from groundwater recharge at higher elevations in the watershed. DOC concentration in the Castle Hayne aquifer measured at a site approximately 60 km from Creeping Swamp was 2.5 mg*l-l (Leenheer, --et a1 . 1974). Annual streamflow and groundwater flux of organic carbon in Creeping Swamp is presented for 1976 and 1977 (Table 5). The flux of carbon was mostly DOC, and, while total outputs of TOC and DOC exceeded total inputs in both years, POC input exceeded output in 1976. Both inputs and outputs of TOC were substantially greater in 1977 than in 1976, reflecting 85% more streamflow at CP-10 and 120% more streamflow at CP-20 in 1977 as compared to 1976. Table 5. Flux of organic carbon via streamflow and groundwater in Creeping Swamp (g c*m-2eyr-1). Positive net flux indicates net movement into the Creeping Swamp study segment and negative values net movement out .

DOC POC TOC DOC POC TOC

Inputs Main Stream (CP-20) Tributary Groundwater Total Inputs Outputs Main Stream (CP-10) Net Flux

*Tributary weighted mean annual DOC and FPOC concentrations assumed to be the same fractions of FTOC as in 1976 (i, e., DOC = 91 .3%of FTOC and FPOC = 8.7% of FTOC) .

Rainfall organic carbon flux: I measured organic carbon concentrations in rainfall of from 5.0 to 15.7 mg*l-l, with a mean of 10.3 mg-1-1, from February to August 1978. True mean annual concentration was 1 i kely much 1ess than 10.3 mg*l-l . Contamination of some form must have occurred. Rainfall coll ector bottles were imbedded in the soil and groundwater DOC may have diffused through the plastic. I have chosen to discard this data. Jordan and Likens (1975) reported an average rainfall organic carbon concentration of 2.4 mg*l-l at Hubbard Brook, New Hampshire, and Brinson (personal communication)measured values of about 1.8 mg*l-1 in an area within 25 km of Creeping Swamp. Assuming a mean annual rainfall organic carbon concentration of 1.8 mg-1 "1 for Creeping Swamp, precipitation flux prior to contact with vegetation was 1.8 g in 1976 and 2.4 g C-m-2 in 1977. Biologic inputs Annual tree and sapling wood increment: Annual aboveground net stem and branch wood increment in 1977 ranged from 0 to 100.4 kg dry weightstree-1 for selected trees in Creeping Swamp. Annual net wood increment was positively related to tree diameter (DBH) for each of seven species chosen (Fig. 8). Net wood increment-DBH relationships for red maple (A. rubrum), sweet gum (I--. st raciflua), water oak (Q. nigra), and swamp chestnut oak (Q. michauxiir"- were steeper and increments were higher than for black gum (N. s lvatica var. biflora), tupelo gum (1.aquatica), and ash (Fraxinus spp.r"- The relatively low productivity of black gum and tupelo gum is striking since both are common swamp species and black gum is especially abundant in Creeping Swamp. The high productivity of sweet gum and swamp chestnut oak, and their lower density compared to red maple, ash, and black gum, may reflect very favorable conditions for these species at the higher elevations in the swamp where they are commonly found (Table 3). Perhaps ash, black gum, and tupelo gum are abundant at the lower elevations due to exclusion of species such as the oaks and sweet gum because of flooding frequency rather than to their higher productivities there. Red maple may be particularly abundant due to its relatively high productivity, compared to ash, black gum, and tupelo gum, and tolerance to flooding. Annual aboveground net wood increment for 1977 was 648 g dry ~ei~htem-~(324 g ~-rn-2)with red maple, chestnut oak, and sweet gum dominating. Trees comprised 96% of this figure since small diameter stems of all species grew very slowly.

Li tterfal 1 : Litterfall in Creeping Swamp, measured by 0.2 m2 basket-type collectors, was distinctly seasonal (Fig. 9). Approximately 60% of annual 1 itterfall in 1975-1977 occurred in October and November, with peak 1 itterfall usually the last week of October and first week of November. Mean 1itterfall was 604(+43) g dry weightam-2 in 1975, 638(+43) g dry weightem-2 in 1976, and 572(+24) g dry weight-m-2 in 1977 (Mulholland 1979). In terms of organic carbon $1i tterfall in Creepin Swamp was 285 g corn-2 in 1975, 305 g cam-2 in 1976, and 272 g Corn-? in 1977. Leaves comprised 80.7% of litterfall in 1976 and 75.1% in 1977, and wood comprised 11.2% and 15.9% of 1 itterfall in 1976 and 1977. Fruits and flower parts were 8.1% of 1itterfall in 1976 and 9.0% in 1977, and exhibited an April peak comprising 70% and 56% of April litterfall those years. Litter fractions were periodically combusted to determine AFDW. Leaves ranged from 92.2% to 95.9% ash-free, with a mean of 94.7% over an annual period. Woody tissue ranged from 95.9% to 98.6% ash-free with a mean of 97.0%, and fruit and flower parts ranged from 95.0% to 99.1% ash-free with a mean of 97.0%.

Macro-1 i tterfall : Macro-litterfall (material >0.3 m in length which was excluded from the basket-type collectors) was most 1i kely a function of both season and storm frequency. The highest rates measured were for late fall-early log Y=3.10 log X-3.38 log Y444 log X- 1.50

Wa 0 Z-

log Y=2.10 log X-2.10 log Y =3.05 log X-3.94

0 10 20 30 40 50 0 10 20 30 40 50 60 TREE DBH (cm)

Figure 8, Aboveground net annual wood increment in 1977 as a function of diameter at breast height (DBH) for seven common species in Creeping Swamp. ( f ) Chestnut Oak

@ log Y = 1.55 log X- 1.15

0 10 20 30 40 50 60 70 80 TREE DBH (cm)

Figure 8. (continued). J FMAMJJASONDJFMAMJJASOND 1976 1977

Figure 9. Seasonal patterns of average daily litterfall in Creeping Swamp for 1976 and 1977. winter and late winter-early spring (Fig. 10). A large spate occurred in late May 1977 which precluded collection for the period 13 April 1977 to 2 June 1977. Macro-litterfall during this period was estimated to be the average of the collection periods immediately preceding and following, or 0.358 g dry weightem-2-day-1. Annual macro-1 i tterfall for 1977 (1 November 1976 to 31 October 1977) was 121 (516) g dry weightem-2 (58.9 g em-3). In a Florida cypress swamp Burns (1978) measured macro-litterfall of 66 g C-m-2. Fall of tree stems is not represented by macro-litterfall here due to the relatively small size of the collection plots (25 m2). At Hubbard Brook tree stems comprised 14.1% of total annual aboveground 1 i tterfall , about 40 g cam-2 (Gosz, et al. 1976). Shrub and herbaceous plant net productivity: A nual shrub and herbaceous plant net productivity was 14.7(+5.8) g in 1976 and 16.7(k5.9) g C-m-2 in 1977. These are probably under- estimates of true annual net productivity due to the omission of losses during the growing season not measured by the harvest technique used here; however, it is doubtful they are low by more than 50%. Total aboveground pl ant 1i tterf a1 1 : Assuming macro-litterfall was constant from year to year, total annual aboveground plant litterfall in Creeping Swamp was about 379 g in 1976 and 348 g corn-2 in 1977 (Table 6). Leaves made up 65% of 1976 and 58% of 1977 total 1i tterfall , and wood comprised 26% of 1976 and 30% of 1977 totals. Shrub and herbaceous plant 1i tterfall was computed from harvest data and is probably low due to losses during the growing season.

Table 6. Total above round plant litterfall in Creeping Swamp in g C.m'2.yr-q (rS. E. ) .

Litterfall (<0.3 m in length) 305 .0 (520.4) 272.2 (211.2) Macro-1 i tterfall (>0.3 m in length)" 58.9 (58.2) 58.9 (k8.2) Shrub and herbaceous plant 1i tterfall 14.7 (55.8) 16.7 (k5.9) To ta 1 378.6 347.8

*Macro-litterfall assumed to be constant from year to year.

Throughfall and stemflow net 1 eaching: A portion of annual net primary productivity is leached from living vegetation by throughfall and stemflow. In 1977 throughfall in Creeping Swamp total led 90% of precipitation recorded at a site about 5 km away. Helvey and Patric (l965), in their review of throughfall and stemflow in eastern United States hardwood forests, reported annual throughfall general ly averaged 80 to 85% of precipitation. Organic carbon concentration, as well as many other constituents leached from foliage by throughfall and stemflow, has been shown to be a function of season, volume of precipitation in any one event, and time since last rainfall (Eaton, et al. 1973). In Creeping Swamp throughfal l organic carbon concentration (vwas significantly correlated (P <0.01) with volume of throughfall coll ected during the 2-week period of measurement (X) for the dormant season (November through March) and, with April excluded, for the growing season (Fig. 11). In April 1977 and 1978 large throughfall organic carbon concentrations (51.1 and 61.7 mg-1-1) were measured, due probably to the 1arge amounts of pollen and flower debris found in the samples at these times. In 1977 throughfall was 1.18 m and weighted mean annual organic carbon concentration was 18.0 mg.l11. Total flux of or anic carbon into Creeping Swamp via throughfall in 1977 was 21.2 g Cam- 9 . Stemfl ow organic carbon concentration varied widely from period to period and from tree to tree, but tended to be inversely related to volume. Concentrations ranged from 7.0 mg.l -1 to 120 mgel-l, with a weighted mean annual concentration of 29.2 mg*l-1. Assuming stemflow volume to be equal to 5% of annual precipitation (Helvey and Patric l965), total stemflow organic carbon flux in 1977 was 1.9 g C-m-2 or 9% of total throughfall flux. Eaton, et a1 . (1973) reported an organic matter stemflow flux of 11% of that for throughfall at Hubbard Brook in the summer of 1969. Net input of organic carbon via leaching of living vegetation by throughfall and stemflow during 1977 was computee as the difference between total throu hfall and stemflow flux (23.1 g Cam' ) and rainfall flux (2.4 g Gem- 9 ), or 20.7 g corn-2. Algal productivity: Filamentous algae were noticeable in flooded areas of Creeping Swamp from January until about June, or until the swamp dried for an extended period. Generally algae did not reappear during summer or autumn spates, due probably to light limitation under the tree canopy and to the short duration of flooding. Algal growth became much less noticeable after mid- April and algae were commonly observed floating downstream at this time. Filamentous species observed in Creeping Swamp include Batrachiospermum sp., Bul bochaeta sp., Draparnaldia glornerata, Eunotia pztinal is, Hyalotheca dissil iens, Oedogonium sp., Spirogyra sp'. , Tri bonema sp., and Zygnema sp. General ly the fi1 imentous diatom, E. pectinal is appeared first in January with a more diverse flora developing in February, March, and early April. ass 0 3 s In om0 v- aJ C, +'In U U5 s rnh 3 rm By May only the red a1 ga Batrachiospermum sp. was observed In 1979 chlorophyll-a on the swamp floor totalled 9.1 (tl.O mg.mm2 on 15 March, 9.0 (t2.0) mgmm-2 on 16 April, and 4.6 (20.9) mg=m-1 on 22 May. In a 1974-75 study Kuenzl er, et a1 . (1 977) reported chlorophyll -a 1 eve1 s of ~0.4mg*m-3 in the water column in Creeping Swamp, much lower than 1979. Oxygen diffusion coefficients of 0.015 and 0.02 g 0~.m-2.hr-l (at 0% saturation and 16 C) were computed from measurements taken in April 1976 using the nitrogen-f i 1 1ed dome method of Cope? and and Duffer (1 964). Coefficients of from 0.03 to 0.08 g ~~*m-~*hr-l(at 0% saturation) were computed for normal streamflow conditions using stream physical and hydraul ic properties (Ts ivoglou and Wal lace 1972). For computations of gross primary productivity (9PP) in Creeping Swamp I used a diffusion coefficient of 0.035 g 02.m- ohr-1 (at 0% saturation). Due to the small size of the coefficient, GPP would be in error by less than 20% even if the coefficient was in error by a factor of 2. GPP rates of aquatic plants, primarily filamentous algae, but probably including benthic diatoms, mosses, liverworts, and submerged vascular plants, varied widely, ranging from 0.1 to 3.9 g ~~ern-~*da~-l(Fig. 12). Large winter and spring spates may have dislodged algal filaments and reduced productivity temporarily. In addition algal productivity in newly flooded areas was probably low until biomass could build up. There is likely some seasonal effect since the highest GPP values were recorded in March, just prior to canopy closure (early April ), and no values >0.9 g ~~em-z*da~-lwere measured during Apri l -June. Mean monthly algal GPP for January through June ranged from 0.25 to 2.00 g 02=m-2*day-l and mean monthly algal net primary productivity (NPP) ranged from 0.15 to 1.20 g ~~=m-~eda~-l(Mu1 holland 1979). Input of a1 gal NPP was 12.6 g C-m-2 in 1976 and 14.1 g cam-2 in 1977 when normal i zed for the en tire area of swamp. Biologic outputs Respiration rates: Sites of respiration measurements in Creeping Swamp were the swamp floor, in both the flooded and unflsoded conditions, and the water column. Organic carbon loss from Creeping Swamp by way of respiration activity on, or in, the swamp floor has three components: (1) aerobic terrestrial respiration and release of carbon dioxide (C02) when the swamp is unflooded, (2) aerobic aquatic respiration and release of C02, and (3) anaerobic respiration with the release of C02 and CHq, either in dissolved form or in bubbles. Tdgether these processes represent the principal non-hydrologic losses of organic carbon from Creeping Swamp. Terrestrial Respiration. Aerobic terrestrial respiration was measured for the 1i tter layer oi~ly(terrestrial 1 itter respiration, Rtl ), and for the total soil profile (total soil respiration, RtS). Rates of terrestrial litter respiration ranged from 3.6 to 180 mg ~-m-~.hr-', with temperature explaining about 66% of the variation (Fig. 13). For the subset of data gathered on 4 November 1977 and thereafter, inclusion of the amount of 1itter incubated (in AFDW) as an independent variable significantly improved (P<.05) the regression, explaining an additional 10.2% of the variation in respiration rate (Mulholland 1979). Litter moisture content was poorly correlated with respiration rate (r = .O4), and its inclusion did not significantly improve the strength of a multiple regression. Although some (Witkamp 1966, 1968) have reported a significant moisture effect, it may be less important in unflooded areas of Creeping Swamp due to the high water table and humid conditions. Total soil respiration rates ranged from 27 to 441 mg ~.rn-~.hr-1 , with temperature alone explaining 69% of the variation (Fig. 14). Roots are very abundant in the top soil layers of Creeping Swamp and much of the total soil respiration is likely to be root derived. Roots have been reported to contribute from 6 to > 50% of total soil respiration in forests (Col eman 1973; Edwards and Sol lins 1973; Edwards and Harris 1977; Wiant 1967; Witkamp and Frank, 1969).

Aerobic Aquatic Respiration. Respiration rates within the water column itself varied seasonally, with very low winter rates, slightly higher spring rates, and the highest rates measured in summer under low streamflow or stagnant conditions (Fig. 15). An early November peak in water column respiration rate corresponding to the release of readily leachable organic compounds from newly fallen leaves was not evident in 1977. The low rates of respiration in the water column were probably due to two factors: (1) the lack of particulate surfaces for bacterial colonization within the water column; turbidity and POC were very low most of the year, and (2) while concentrations of DOC in the water were large, most of it was 1 ikely refractory humic and fulvic compounds.

Aerobic aquatic respiration rates of the swamp floor in laboratory incubated co es (aquatic benthic respiration) ranged from near 0 to 194 rng 02.m-'.hr-l , with temperature explaining from 84% to 96% of the variation (Mu1 holland 1979). The data were divided into seasonal groupings with significantly different (P < 0.05) temperature-rate relationships hetween groups (Fig. 16). Temperature alone explained 93% of the cold season and 88% of the warm season variation in respiration rate, with QlO1s of 2.5 and 3.2, respectively. The aquatic benthic res iration rates reported here are somewhat higher than the 21 to 83 mg Op.m-5.hr-1 measured in a Louisiana cypress swamp bayou, but less than the 125 to 417 measured in a lake within the swamp (Day, et al. 1977). They fell within the lower range of the 29 to 896 mg 0~.m-2.hrq reported over an annual period for various North Carolina Piedmont and Coastal Plain streams (Hoskin 1959).

Anaerobic Respiration. Evolution rates of dissolved methane from Creeping Swamp ranged from near 0 to 9.9 mg ~~~.m-~-hr-l,and exhibited 088T Rate = 8.0 e*

TEMPERATURE (OC)

Figure 13. Effect of temperature on rates of terrestrial litter respiration in Creeping Swamp. Data collected in 1977-1978. w L =I +'ru L 6, Q Zod t'hm %-I- 0 l h +'om h wr- rc '4- 6 W 'r

Rate = 17.0 e' 093T

0 5 I0 15 20 25 TEMPERATURE ("C)

Figure 16. Effect of temperature on rates of aquatic benthic respiration in Creeping Swamp for the (a) cold season (November to February), and (b) warm season (March to October). no simple relationship to water temperature (Fig. l7b). Rates were generally higher in the spring months near Transect 1 (Fig. 17a), was inundated for longer periods and had slightly lower water velocities due to the diking effects of the state highway nearby. Funnel collectors were set out from February to June 1979 but no methane ebbulition was observed. Evidently if ebbul ition occurred, the water column was sufficiently undersaturated with methane that dissolution occurred rapidly. Rudd and Hamil ton (1978) measured methane evolution rates of 0.7 to 6.7 mg C~~-m-2*hr-lin a eutrophic Canadian shield lake and found they varied little throughout the year. Baker-Blocher, et al. (1977) measured methane evolution rates of from 4.2 to 50 CH4.m-2-hr-l in some Michigan wetlands and found them to increase with temperature.

Monthly and annual respiration:

Monthly and annual respiration rates were computed for water column respiration, aerobic terrestrial 1i tter and total soil respiration, and aerobic aquatic benthic respiration on a whole-swamp basis. Monthly water colu n respiration was low (<1 g C*rn-2*mo-l) and annual totals (3.2 g in 1976 and 4.5 g in 1977) were small compared to those for the swamp floor. Monthly rates of terrestrial litter and aquatic benthic respiration varied seasonally (Fig. terrestrial litter respiration rates (3 to 51 g C-m- exhibited a pattern of sharp summer peaks and low winter values, This reflected seasonal differences in both temperature and inundation. Monthly aquatic benthic respiration rates (0 to 7.4 g ~*rn-~*rno-l)were lower and more uniform throughout the year, reflecting the counteracting effects of low water temperatures and large inundation fractions in winter and early spring, and the reverse situation durin summer and fall. Terrestrial litter respiration totaled 275 g C*m-g in 1976 and 246 g ~*m-2in 1977, and represented 24% of total soil respiration in both years. This is in reasonable agreement with the 30% contribution by litter to total soil respiration in an English forest (Phillipson, --et a1 . 1975). Most of the remaining total soil respiration in Creeping Swamp is probably root-derived. Total soil respiration was 1145 g Cam-z in 1976 and 1026 g in 1967. Aquatic benthic respiration was 91 g ~~*rn-~in 1976 and 122 g 02-m- in 1977. Assuming an RQ of 1 0, organic carbon loss via aquatic benthic respiration totaled 34 g Corn-* in 1976 and 46 g c*mTin 1977.

Particulate organic carbon formation

Conversion of dissolved organic carbon to the particulate form -via DOC precipitation and possibly bacterial growth was measured from Novem- ber 1977 to M y 1978. Rates of POC formation ranged from 0.01 to 0.15 mg AFDW*l-l*day-P, with a mean of 0.10 mg AFDW*I-~*~~~-~; and were generally higher during the spring months (Mu1 hol and 1979). Using the mean POC formation rate of 0.10 mg AFDW*l-l*day-l (0.05 mg ~*l-l*da~-l)and mean water retention time of 30 h, computed from estimated swamp volumes (using topographic data from the five transects paced for inundation patterns) 5 10 15 20 25 TEMPERATURE (OC)

Figure 17. Methane evolution in Creeping Swamp from March 1978 to February 1979: (a) seasonal patterns, and (b) as a function of water temperature. Closed circles were near CP-14 and open circles near CP-10.

55 and streamflow rates at correspondin volumes, I computed annual swamp POC formation for 1977 of 0.6 g Cam- 9 . Leaching of DOC from ground litter Leaching of ground litter represents a transfer from POC to DOC and consists of three processes: (1 ) rapid, short-term loss of readily leachable material from submerged, newly fa1 1 en 1 i tter, mostly leaves, (2) removal of readily leachable material from newly fallen 1i tter by throughfall when the swamp is dry, and (3) slow, long-term leaching of submerged 1 i tter. The amount of readily leachable material in newly fallen leaves in Creeping Swamp, as measured by 7-day leaching experiments, ranged from 5.5% for swamp chestnut oak (Q. michauxii) leaves to 32.6% for tulip poplar (I.tul ipifera) 1eaves (T- For four of the seven 1eaf species studied, most of the initial leachinq was complete after 48 h of submersion but for tupelo gum (N. watica), black' gum (N. sylvatica var. biflora), and swamp chestnut oak significant leaching Took place between 48 and 168 h (Fiq. 19). Jn order to compute the readily leachable organic matter availabl e; total annual 1 eaf 1 itter input was apportioned among the seven most abundant tree species according to their basal areas. In 1976 41.4 g corn-2 and in 1977 32.3 g cam-2 of annual leaf 1 i tterfall was readily leachable. Much of the readily l eachable material removed from newly fa1 len 1 eaves has been shown to be rapidly utilized by bacteria in streams (Curnmins, --et a1 . 1972; Lock and Hynes 1976; McDowell and Fisher 1976; Wetzel and Manny 1972). In experiments with swamp floor cores, about half the DOC initially leached from red maple leaves was utilized (disappeared) within 12 h, from 55 to 70% within 24 h, and from 79 to 103%within 95 h (Table 8). The readily leachable material of newly fallen leaves which fall in unflooded areas can be leached and transported into the soil by throughfall. In concurrent studies of throughfall leaching of newly fallen red maple (A. rubrum) leaves from Creeping Swamp, incubated in the swamp and in a wooded area near Chapel Hi1 1 , North Carolina, approximately 10 to 12 cm of throughfall was required to extract the readily leachable fraction (Mu1 hol land 1979). Rates of slow leaching of ground 1 itter3 as measured from DOC increases in laboratory incubated swamp floor cores, ranged from 0 to 48 mg ~*m-2*hr-l(Fig. 20). For the November and January experiments exponential relationships (Y = A~BT)between leaching rate (Y) and temperature (T) described the data sl ightly better than did 1 inear relationships. Exponential equations were similarly used to fit the May and August data (Mulholland 1979). Temperature alone explained 74% of the variation in leaching rate in November, 81% in January, 86% in May, and 89% in August (cf. Fig. 20). If AFDW of core 1itter is included, the regression is significantly improved (P< .O5) and an additional 15% of the Table 7. Dry weight loss as a percentage of original dry weight during a 7-day leaching study of newly fallen leaves collected in autumn from several species of trees in Creeping Swamp. Leaves were leached in distilled water at 23 C.

48-hour Dry Weight Loss 7-day Dry Weight Loss Species (% of original weight) (% of original weight)

Quercus michauxii 2.5 Quercus nigra - Fraxinus spp. Liquidambar styraciflua 17.9 Nyssa sylvatica var. biflora 14.1 Nyssa aquatica 16.9 Acer rubrum 17.1 Liriodendron- tul ipifera 31.3

Table 8. Utilization of readily leachable material from newly fallen red map1 e 1eaves in swamp floor cores taken in November 1978.

Percent of Leachate Leaching Leachate Utilized Time* DOC

*Length of time leaves were submerged in distilled water to leach readily 1eachabl e fraction.

(b) Jon. 23,1979 t Y=l.8e.106T

(c) May 3,1978 C Y =0.4ee53T 4

TEMPERAT URE ('C)

Figure 20. Effect of water temperature on rates of slow leaching of ground litter during four experiments with Creeping Swamp f 1oor cores. variation is explained in November, and 4% in August. In May, the amount of litter in the cores was not measured and in January it did not significantly improve the regression. Leaching rates generally decreased from November to May, probably reflecting decreases in ground 1i tter. The relatively high leaching rates in August are contrary to the trend of decreasing rates with time after autumn litterfall. They may be explained by the relatively large amounts of 1i tter in the August cores (1140 g AFDW*~-~).Mean core litter in the November experiments was 793 g AFDW*~-2and in January 607 g AFDW.~-2. From inspection of ground litter levels during 1977 (Fig. 5), it appears that the amount of litter in the August cores was higher than mean swamp levels during August 1977, probably due to sampling bias-

Monthly and annual slow leaching rates were estimated using the leaching rate-temperature re1ationships (Fig. 20), mean monthly water temperatures, and mean monthly inundation fractions. The November relationship was used for November and December, the January relationship for January through March, and the May relationship for April through October. The August relationship was not used due to the large amounts of core litter. Slow leaching in December 1976 and November 1977 was assumed to occur over 24 and 23 days, respectively, because initial leaching during the first 7 days of inundation was already accounted for. Most of the slow leaching occurred during December through March and May through July in 1976, and in March, November, and December in 1977. Total annual slow leaching of ground litter was 16.6 g ~*m-2in 1976 and 24.3 g corn-2 in 1977.

Natural and channel ized streams in eastern North Carol ina

In addition to Creeping Swamp, organic carbon concentration was monitored in six other streams in eastern North Carolina, four of which have been channel ized (Fig. 1 ) . Weighted mean annual DOC and FPOC concentrations are presented for these seven streams (Table 9). Weighted mean annual DOC concentrations ranged from 8.5 mg*l-l in 1975 at Grindle Creek (GR- 10) to 19.7 mgel-1 in 1977 at Creeping Swamp (CP-70). Weighted mean annual FPOC concentrations ranged from 0.4 mg-1-1 in 1976 at Creeping Swamp (CP-10) to 3.2 mg*l-l in 1975 at Tracey Swamp (TR-10). Weighted mean annual FTOC concentration in Tracey Swamp was about 1.4 times higher in 1975 than in 1976. This difference may be due to the lack of winter high- streamflow samples in 1976. Streamflow was greater than 1 m3*sec-1 on four of seven sampl ing dates in 1975, but only two of ten sampl ing dates in 1976. In Tracey Swamp, a channelized stream, the lack of high-streamflow data may have biased 1976 weighted mean annual FTOC concentration toward lower values. In Creeping Swamp weighted mean annual DOC concentration increased by 40% from 1976 to 1977. This may be explained by large DOC concentrations during a very large spate just after peak leaf- fa1 1 in early November 1977.

Seasonal patterns of FTOC concentration are presented for four streams (Fig. 21 ). FTOC concentrations during 1976 in two natural swamp-streams, Chicod Creek (CH-20) and Palmetto Swamp (PM-10) , were somewhat higher, especially in the summer, than in two channelized streams, Clayroot (CY-10) and Tracey Swamps (TR-10). In Chicod Creek and Palme to Swamp FTOC concentrations during 1976 were generally 10 to 15 mgel-t whereas in Clayroot and Tracey Swamps they were normally 5 to 10 mg.1- 1.

Table 9. Weighted mean annual DOC and FPOC concentrations in natural and channelized streams in eastern North Carolina.

Number Weighted Mean Annual of Concentrations (mg.1-- ' ) Stream Year Sampl es DOC FPOC

Natural : Creeping Swamp (CP-10) 1975 16 12.6 1 .O 1976 29 14.8 0.4 1977 26 19.7 1 .O Pal me tto Swamp (PM-10) 1975 13 -- (10.1)* - -

Chicod Creek (CH-20) 1975 8 10.4 1.6 1976 9 -- (15.2)* - - Channel i zed: Cl ayroot Swamp (CY-10) 1976 26 12.0 2.5 Tracey Swamp (TR-10) 1975 7 14.0 3.2

Conetoe Creek (CO-10) 1975 7 11.1 1.7 Grindle Creek (GR-1 0) 1975 7 8.5 2.3

*Weighted mean annual FTOC concentration. Missing DOC samples from one or more high-streamflow sampling dates prevented computation of weighted mean DOC or FPOC concentrations. Figure 21. Seasonal patterns of fine total organic carbon (FTOC) concentration ( ) and streamflow (0) at (a) Chicod Creek (CH-201, (b) Palmetto Swamp (PM-I 0), (c) Clayroot Swamp (CY-lo), and (d) Tracey Swamp (TR-10) in 1976. 6 3 FTOC concentration was significantly correlated (P< .05) with log1 0 (streamflow) in all but Chicod and Grindle Creeks. For Grind1 e Creek the correlation was significant at P<.06. The correlations for the four channelized streams were all positive, whereas those for Palmetto and Creeping Swamps, both natural streams, were negative. Correlations between DOC concentration and log1 0 (streamflow) were significant and positive for Tracey and Clayroot Swamps, as was a correlation between DOC concentration and s treamf 1 ow for Conetoe Creek. FPOC concentration was not significantly correlated with streamflow or log1 0 (streamflow) in any of the streams. Effect of spates on organic carbon concentration in Creeping and Clayroot Swamps Concentrations of FTOC and DOC (not shown) increased slightly as streamflow increased, but were diluted slightly thereafter during spates in Creeping Swamp (CP-10 and CP-20) in January, June, and December 1976 (Fig. 22). Peaks in FTOC concentration at CP-10 during the June spate (3 June and 6 June) were due to large FPOC concentrations an these dates. FPOC concentrations were less than 3.0 mg*l-l on all other dates during the June spate. In Clayroot Swamp, a channelized stream within 5 km of Creeping Swamp (Fig. 1 ) , FTOC concentration during spates exhibited patterns simi 1ar to, but not as dramatic, as streamflow (Fig. 23). DOC concentration patterns (not shown) were not as sharply peaked as those for FTOC. FTOC concentration patterns were mostly a result of FPOC concentrations, which were generally high as streamflow increased, and returned to low levels as streamflow decreased.

DISCUSSION Organic carbon concentrations DOC concentrations reported here for three natural swamp-streams in eastern North Carolina are high compared to upland rivers and streams, with the exception of some Canadian streams reported by Lock, --et al. (1977), but are similar to those reported for other swamps (Table 10). High DOC concentrations found in swamps as compared to waters of upland forested watersheds are 1 i kely due to two factors. First, leaching processes are enhanced in swamps due to long contact times between water and ground 1 itter. In upland forests leaching takes place primarily by rainfall as it passes through the tree canopy and percolates into the soil. Contact time between water and ground 1itter is relatively short as rai nfall quickly penetrates vertically into the lower soil 1 ayers. There it is isolated from further contact with detritus, moves laterally, and eventually emerges in streams. In swamps, leaching takes place primarily by surface water. Swamp surface waters are shallow and velocities are low, resulting in relatively long contact time between water and ground lad

Figure 22. Patterns of fine total organic carbon (FTOC) concentration and streamflow at two stations in Creeping Swamp (CP-10 and CP-20) during spates in (a) January, (b) June, and (4December 1976. Sol id l ine depicts FTOC concentration and broken line shows streamflow. Dec 8 16 17 18 19 20 1976

Figure 23. Patterns of fine total organic carbon (FTOC) concentration and streamflow at two stations in Clayroot Swamp (CY-10 and CY-20) during spates in (a) January, (b) June, and (c) December 1976. Sol id 1 ine depicts FTOC concentration and broken line shows streamflow. I Tab1 e 10. Concentration of organic carbon in natural waters. Organic matter concentrations were converted to organic carbon by mu1 tiplying by 0.5.

DOC (mg*1-') FPOC (mg*1 -' ) CPOC [mg- l -') Ecosys tem range mean range mean range mean Reference

Rivers and streams : Fort R., Mass. 0.9-7.4 3.1 0.4-3.1 1.0 0.004-0.09 0.025 Fisher (1977) Bear Br., N.H. 0.4-3.8 1.3 0.0-1.5 0.13 0.0 -0.51 0.011 Fisher and Likens (1 973) Watersheds 2 and Hobbie and Likens 6, Hubbard Brook (1973) Expt. Forest, N.H. 0.3-4.8 <2.0 0.1-0.4 Augusta Cr. Wetzel and Manny Mi ch. 2.0-11.7 <8.0 0.2-7.0* (1977) Catahoula Cr., De la Cruz and Post Miss. (1 977) Nanaimo R. Nairnan and Sibert B.C. 6.0-14.0 6.4 (1978) 4 Guatamalean R. 1 .O-36 <10.0 0.0-10 4.0 Brinson (1976) 12 New Zealand Lock, --et a1 . (1977) Str. 0.7-4.6 5 Canadian Str. 4.9-24.6 Lock, --et al. (1977)

Swamps : Eastern N.C. 10.4-19.7 0.4-1.6 this study Louisiana bayous 10.2-12.2 1.4-1.9 Day, et a1 . (1977) Big Cypress Sw. , Fl a. 15.0-26.0 0 Carter, --et a1 . (1973) "estimate of POC which most closely approximates FPOC. litter. In Creeping Swamp, with a mean depth of about 0.3 rn, mean water velocity of about 0.1 mosec-1, and mean ground litter of about 450 g C-m-2, approximately 60 g ~-m-2-yr-l is leached from the soil surface. High swamp DOC concentrations are also due to the concentration effects of high rates of evapotranspiration in swamps as compared to upland watersheds. Annual evapotranspiration was computed to be 61% of annual precipitation in the Creeping Swamp watershed (Winner and Simmons 1977), but only 38% in the Hubbard Brook watersheds (Likens, --et al. 1977) and in the Fort River watershed (Fisher 1977). Evapotranspiration ranged from about 34 to 66% of precipitation in three Guatemalan watersheds (Brinson 1976). Burns (1 978) reported evapotranspiration of 71 % of precipitation in a Florida cypress swamp.

In a study from October 1978 to April 1979, L. P. Smock (personal communication) found that on average DOC was 86% dialyzable (<10,000 MW or <3.2 nm) at CP-14 and 75% dialyzable at CP-10 when flowing. However, in a stagnant pool in the floodplain during March, DOC was 32% dialyzable, and at CP-14 during October when the stream was intermittant, DOC was 56% dialyzable. Koenings (1976) reported that only about 15% of the DOC of a northern acid bog lake was dialyzable. It appears the size distribution of DOC molecules may be related to hydraulic properties of the aquatic environment as well as the source of the dissolved material.

Average concentrations of FPOC in the eastern North Carolina natural swamp-streams are low but within the range of those for most other reported rivers, streams, and swamps (Table 10). Low swamp FPOC concentrations are reasonable since water velocities are low and watersheds are relatively flat.

Productivity and biomass

Total annual aboveg ound net primary productivity in Creeping Swamp for 1977 (1450 g AFDW-m-' or 707 g was computed as the sum of that for trees and saplings (676 g shrub and herbaceous plants (16.7 g C-m-2) , and a1 gae (14.1 g Tree and sap1 ing abovegroun net productivity was comprised of net annual wood increment (324 g C*m-q), 1i tterfall (272 g ~*m-2), macr-0-1 itterfall (59 g em-2), and net throughfall and stemflow leaching (20.9 g corn-2) (Table 6; see also Mulholland 1979). Macro-litterfall may be a slight overestimate of net annual productivity since it may have included material from dead trees; however, this is probably minor.

Creeping Swamp was about as productive as some larger riverine swamps in Louisipna (Lac des A1 lemands Swamp) and North Carol ina (Tar River Swamp), and more productive than an Illinois riverine swamp (Cach River Swamp), an ombrotrophic Georgia swamp (Okefenokee Swamp), and most temperate-zone upland forests (Table 11). In general the flowing-water swamps (Creeping and Tar River Swamps, North Carolina, and Lac des Alle- mands, Louisiana) had higher total and litter productivity than did the bog swamps (in Minnesota and Georgia). Riverine and small er floodplain Table 11. Annual productivity and standing crop of biomass and detritus in swamps and upland forests. Values given in terms of organic matter (organic carbon taken as 50% of organic matter).

Annual Productivity Standing Crop Total Aboveground Shrub and Total Soil Detritus NPP Litterfall Herb Layer Biomass Litter Total* -2 -1 -2 -1 Location Type (g-m-2-yr-1) (gam .yr ) (g-rn -yr ) (kg-rn-') (g-m-2) (kg-m-*) - Swamps : Minnesota cedar bog 1 1030 488 11 16 -- 100 I11 inois cypress- tupel o, 2 riverine (Cach R.) 682 348 4 4 5 - - 15 North tupelo, riverine Carol i na (Tar R. )3 mi xed hardwood swamp-stream 4 (Creeping Sw. ) 1450 572 3 3 28 1000 16 Georgia cypress bog (Okefeno kee) Florida cypress 6 (Big Cypress) Louisiana cypress-tupel o, riverine (Lac des A1 1emands)7 13281 620 20 - - 422 - - mixed hardwood, riverine (Lac des A1 1ernands)7 1654~ 574 200

Bottom1and Forest : I11 inois map1 e-ash 8 (Sanqarnon R. ) 1250 - - 100 29 -- - - Table 11 (continued) Aboveground Shrub and Total Soil Detritus Location Type NPP Litterfall Herb Layer Biomass Litter Total *

Upland Forests: 9 Sweden beech 1540 9 spruce 1370 1 Minnesota oak 890 New Hamp- map1 e-beech-bir h shire (Hubbard Brook) h,ll 900 12 New York oak-pi ne (Brookhaven) 858 Tennessee tulip poplar (Oak ~id~e)'~1268 North lobloliy pine Carol ina plantation14 41 12 15 Georgia oak-hickory 600

World Averages : Temperate Rodin and Bazilevich (1967) 350-2000 Bray and Gorharn (1964) - - Whittaker (1965) 600- 3000 Schlesinger (1977) - - Tropical Bray and Gorham (1964) 2740 1990 ------Whi ttaker (1975) 1000-5000 (i = 2000) - - 6-80 (x = 45) - - * reported to a depth of 25 cm for Illinois and North Carolina swamps, 91.5 crn for the Okefenokee Swamp, 45 cm for the Hubbard Brook forest, 65 crn for the Swedish forests, 75 cm for the Tennessee forest, and 50 crn in Schlesinger's estimate. t middle of the reported range References : 1. Rei ners (1972) 6. Burns (1978) 11. Gosz, et a1 . (1976) 2. Mitsch, et a.(1977) 7. Conner and Day (1976) 12. hi ttakerand Woodwell (1969) 3. Brinson n977) 8. Johnson and Bell (1976) 13. Reichle, et al. (1973) 4. this study 9. Nihlgard (1972) 14. ~inerson,eta1 . (1 977) 5. Schlesinger (1978) 10. Whittaker, --et al. (1974) 15. Monk, --et ar n970) swamps may support some of the highest natural forest productivities in the temperate zone. However, swamps in cooler climates (Cache River Swamp, I11 inois) or those without abundant nutrient suppl ies (Okefenokee Swamp and the Minnesota cedar bog) had similar or somewhat lower productivity than upland forests.

Biomass in Creeping Swamp was intermediate compared to other swamps, but surpassed those for upland forests reported here (Tab1 e 11 ) . It was in the middle of the range given by Rodin and Bazilevich (1967) for temperate forests. A report on Southeastern forest biomass in good stands by Smith, et al. (1975) indicated that swamps similar to Creeping Swamp (termed branch bottom sites) possessed some of the highest biomass of any of the forests in this region, ranging from 24 kgem-2 in 20-year stands to 51 kg*rn-2 in 60-year stands.

Ground litter in Creeping Swamp was as great as that in Okefenokee Swamp and greater than that reported for other swamps, but within the range of upland forests (Table 11). This appears to be the result of rather large annual total litterfall in Creeping Swamp rather than slow decomposition as in Okefenokee Swamp. Total soil organic matter in Creeping Swamp was less than the Minnesota and'Georgia bog swamps and within the range of upland forests (Table 11). The fermentation and humus layer was thin and clay and sand layers were reached at shallow depths. Decomposition of ground litter appears to keep pace with total 1 itterfall over an annual period. Decsmposi tion is probably enhanced by alternating flooded and unflooded conditions, allowing periodic penetration of gaseous oxygen to ground litter and upper soil layers. Respiration

Aquatic benthic respiration:

Aquatic benthic respiration was measured in laboratory experiments with swamp floor cores (Fig. 16) and extrapolated to field conditions using --in situ mean monthly temperatures. Since this is an important organic carbon loss from Creeping Swamp, it is desirable to compare values measured in the laboratory with those measured in the field. Field measurements of aquatic respiration can be computed from die1 oxygen concentration data gathered to compute a1 gal productivity. Comparison between those field measurements and rates computed from the 1aboratory studies for similar temperature and season indicates there is reasonable agreement for much of the data (Table 12). The field measurements certainly have much greater error associated with them due to difficulties in estimating average water depth. The large discrepancy between laboratory and field-measured rates for March 1978 may be the result of increased algal respiration due to a large algal bloom observed at that time.

Aquatic benthic respiration was measured in terms of oxygen units. A mean RQ of 1.71 (40.10) was measured in 1aboratory experiments with swamp floor cores; however, they were probably overestimations of field values due to difficulties in measuring dissolved C02 accurately and possible artificial creation of increased anaerobic respiration within the lower parts of the core during incubation. An RQ of 1.0 was used to convert aquatic benthic respiration data from oxygen to carbon units because (1) most of the organic substrate was carbohydrate, and (2) anaerobic release of C02 was accounted for separately.

Table 12. Comparison of field and laboratory measurements of aquatic benthic respiration rates. Field measurements consisted of night respiration rates computed from die1 studies and laboratory measurements were computed from rates measured in swamp floor cores (Fig. 16).

Mean Fie1 d Aquatic Benthic Respiration Rates Temperature Field Laboratory

16-17 Feb. 1976 14.5 87 65 02-03 March 1976 17 46 54 03-04 March 1976 18 62 61 12-13 April 1976 11 25 25 24-25 May 1976 15 33 42 06-07 June 1976 16.5 33 50 07-08 June 1976 17 55 54 16-17 Feb. 1977 2 9 22 30-31 March 1977 17.5 53 57 12-13 April 1977 18.5 32 64 28-29 Apri 1 1977 17.5 29 57 14-15 Feb. 1978 3.5 34 24 14-15 March 1978 13 110 32 28-29 March 1978 14 78 37

Comparison of terrestrial and aquatic rates: The respiration rate-temperature relationships for terrestrial litter (Rtl) (Fig. 13) and aquatic benthic respiration (Ra) in November and January (Fig. 16a), converted to carbon units, were not statistically different from one another (PC .O5). The aquatic benthic respiration

72 rate-temperature relationship for the March, May, August, and September experiments (Fig. 16b), converted to carbon units, was statistically different from the terrestrial litter relationship for both coefficients. This implies that respiration was similar at all temperatures in both aquatic benthic and terrestrial 1 itter habitats in the cold season (Nov- ember-February) , but that respiration was greater in terrestrial li tter than in the aquatic benthic habitat at temperatures <35 C during the warm season (March-October). This behavior may reflect the dominance of terrestrial systems during the warmer periods of the year. Normally when water temperatures are relatively high, the aquatic system is much reduced in area. Aquatic microbes may be more cold-adapted, responding less strongly with increasing temperature than those in the terrestrial system.

Respiration of readily leachable DOC:

The microbial utilization sf DOC leached from newly fallen leaves was not accounted for in any of my respiration studies. This rapid utilization of leached DOC represented a Plush of intense metabolic activity immediately following submersion of newly fallen leaves, or in unflooded areas, during and immediately following rainfall events as throughfall leached newly fallen leaves and percolated through the litter and soil layers. Using approximate swamp floor topographic profiles from the five transects paced, surface water residence time within the study segment was estimated to be about 25-40 h when the swamp is >70% inundated. If all of the readily leachable material was leached by surface waters, about 60-70% of it would have been retained and presumably respired within the swamp (Table 8) and would not have appeared in my estimates for annual respiration. If throughfall leaching was important, the percentage that was not measured would have been still higher.

A large spate occurred on 6 November 1977,just after peak litterfall, flooding the entire swamp (Fig. 6). In most years much of the readily leachable fraction of swamp 1eaf 1i tter would have been leached by throughfall, percolated through the unflooded swamp litter and soil, and been retained and utilized there. However, in 1977 the size and timing of the November storm resulted in quick flooding and most of the readily leachable material was probably leached into surface waters. DOC concentrations measured at CP-10 during this spate were very large (weighted mean concentration of 21 mgel-I), suggesting that this was the case. I have estimated that 32.3 g cam-2 was 1 eached from soil li tter during this period, but due to its rapid utilization, only about 35% (11.3 g was accounted for as hydrologic DOC output at CP-10, the remainder (21.0 g cam-2) was retained and presumably respired within the swamp segment. Due to the lack of a significant increase in autumn water column respiration rates (Fig. 15) and the paucity of particles for bacterial colonization within the water column, it is probable that the utilization of the readily leachable material took place on the swamp floor by bacteria colonizing leaf and other detrital surfaces. Nevertheless, since the DOC was in the water column for some short period, the 21 g em-2 additional respiration in 1977 was assigned to the water column system component. In 1976 the situation was not quite as clear since the first large spate did not occur until the middle of December. Prior to this time throughfall undoubted1y 1eached some of the readily 1eachabl e material , which was then retained and eventually respired within the 1i tter and soil. Assuming equal portions of total readily leachabl e material (computed to be 41.6 g C-m-2 in 1976) were leached by throughfall and surface waters each, 20.8 g C-m-2 of unaccounted-for respiration took place in the terrestrial litter-soil system and about 65% of 20.8, or 13.5 g cam-2, was retained and presumably respired within the aquatic portion of the swamp segment. Total annual water column and terrestrial 1i tter respiration must be increased by 13.5 and 20.8 g corn-2, respectively, to account for the respiration of readily leachable matter lost from newly fallen leaves in 1976.

Anaerobic respiration:

Annual anaerobic respiratory conversion of organic to inorganic carbon involves both C02 and CHq production. Martens and Berner (1977) have stated that two major pathways 1ead to methane production: (1 ) C02 reduction (C02 + 4H2 -I. CHq + 2H20), and (2) fermentation of acetic acid (CH3COOH -+ CHq + 2H20). In fermentation C02 and CHq are produced in equal amounts. Claypool and Kaplan (1974) have reported that 30 to 50% of CHq production in anerobic environments is by way of C02 reduction.

While evolution rates of dissolved CHq in Creeping Swamp showed no striking seasonal or temperature effects (Fig. 171, season, temperature, and inundation together may explain some of the variation. The low rates measured in December, January, and February probably reflected the low soil temperatures during these months. The August values were low because the swamp had been flooded only a few days before the measurement and interstitial dissolved CHq concentrations had yet to build up. When unflooded for extended periods, gaseous oxygen can penetrate the litter and upper soil layers. Evolution rates were lower near Transect 4 than near Transect 1 probably due to more frequent drying in upstream areas away from the highway. It is plausible that methane evolution in Creeping Swamp is only important during the spring months when soil temperatures increase and flooded or water-logged conditions exist. It may also be important at times during summer and autumn when these conditions occur. However, the swamp is usually dry most of the summer and autumn allowing gaseous oxygen penetration and probably little anaerobic respiration.

Annual methane evolution in Creeping Swamp is dffficult to estimate from my data. I have only measured methane evolution under flooded conditions, and then primarily the dissolved form. Strayer and Tiedje (1978) reported ebullition transported from about one-half to as much methane out of the sediments of a small hypereutrophic lake as did diffusion. How- ever, in Creeping Swamp if ebullition from the soil-water interface occurs, rapid dissolution appears to preclude substantial loss of methane as bubbles and my technique for measurfng dissolved methane evolution would include ebull ition. Assuming that fermentation accounted for 60% of methane release, that methane evolution occurred only under flooded conditions from March through November, and that the mean annual methane diffusion rate was 3.0 mg ~~4.m-2.hr-1, annual release of inorganic carbon via anaerobic respiration in Creeping Swamp was 2.8 g C-m-2 as CHq and 1.7 g C-m-2 as C02 in 1976 and 6.1 g c-mW2 as CH4 and 3.7 g corn-2 as C02 in 1977. Total annual swamp floor respiration: Total annual soil respiration in Creeping Swamp, the sun of aerobic (Ra) and anaerobic aquatic benthic respiration and terrestrial soil respiration (Rts), was 1184 g corn-2 in 1976 and 1082 g C-m-2 in 1977. These annual soil respiration values are similar to those reported for an upland Tennessee forest , and greater than those reported for other forests and swamps, with the exception of a Puerto Rican mangrove swamp (Table 13). They are also at the upper end of the range given by Schlesinger (1977) for world temperate forests. If the difference between total soil and total litter respiration (Rts - Rtl) is assumed to be entirely root- derived, then total annual respiration due to swamp floor decomposition of detri tus was 333 g ~-m~~in 1976 and 334 g corn-2 in 1977. . - Table 13. Annual total soil respiration in various swamps and upland forests .

Total Annual Soil Respiration Location Forest type (g Reference

Forested Swamps: Minnesota cedar bog 739 Reiners (1968) North mixed hardwood Carol i na swamp-s tream 1184, 1082" this study Puerto Rico mangrove swamp 1350 Lugo and Sneda ker (1974)

Up1 and Forests : Minnesota oak 794 Reiners (1968) England beech 171 Phil ipson, --et a1 . (1975) Tennessee mixed deciduous 1036 Edwards and Sollins (1973) North loblolly pine Carol ina plantation 757 Kinerson, --et a1 . (1977) World Forests (ranges) : Temperate Zone 171-1414 Schl esinger (1 977) Tropical Zone 405-61 00 Schl esinger (1 977) * total terrestrial soil respiration, anaerobic and aerobf c aquatic respiration for 1976 and 1977. Production/respiration ratios for the aquatic ecosystem P/R ratios can be computed for the aquatic swamp-stream ecosystem (excluding higher plant production) on an annual basis and for the months of January through June. On an annual basis, Creeping Swamp is quite heterotrophic with P/R of 0.2 in 1977. However, inspection of monthly P/R indicates an increase in P/R from January to March and a decrease thereafter, with February and March near unity (Tab1 e 14). March offers the most favorable combination of sunlight and temperature for algal growth in Creeping Swamp since at this time leaf-out has yet to develop and water temperatures are increasing. While algal productivity was not measured from July to December, P/R ratios are thought to be extremely small and the ecosystem highly heterotrophic since 1ight intensity is low beneath the canopy and-flooding is usually of relatively short duration.

Table 14. Monthly P/R ratios for the aquatic ecosystem in 1 77 All production and respiration figures are g Cam-? of f ooded area only.

Month NPP

January February March Apri 1 May June

* sum of water column respiration ()aquatic benthic respiration (Ra), and anaerobic respiration.

Particulate organic carbon formation The formation of POC as a result of DOC precipitation is an insignificant process compared to other organic carbon flows in Creeping Swamp. Lush and Hynes (1973) reported that DOC precipitation in streams was positively correlated with turbulence, ionic strength, especially dissolved calcium, and pH, all of which are relatively low in Creeping Swamp. At CP-10 in 1976 mean soluble calcium concentration was 3.8 mg.l-l conductivity was usually 40 to 70 ilmho*cm-1, and pH usually les; than 5.8 (Kuenzler, et al. 1977). The low DOC precipitation rates measured in Creeping Swamp were not surp~ising. Annual organic carbon budget for Creeping Swamp Elemental cycling in ecosystems can be examined by an annual budget. The Creeping Swamp ecosystem was first treated as a "black box" with inputs and outputs quantified (Table 15). In a more detailed approach it was partitioned into its principal components with internal -system flows between components, and flows between components and the external environment (Fig. 24). I have partitioned Creeping Swamp into five components: three 1iving components (trees and sap1 ings, shrub and herbaceous plants, and algae) and two detrital components (water column and swamp floor). Ecosystem inputs and outputs were quantified (Table 15), and a1 1 components and flows of organic carbon presented in schematic form (Fig. 24). Through- fa1 1 and stemflow inputs were partitioned between the aquatic and terrestrial habitats by examining daily rainfall and inundation fraction data. Terrestrial 1 i tter respiration was assumed to include a1 l respiration of the swamp floor under terrestrial conditions , except that derived from roots . Complete budget data were available for 1977 (Table 15). Inputs exceeded outputs by 361 g corn-2 with a?1 but 38 g C-m-2 of this in annual aboveground net wood increment. With the exception of tree growth, Creeping Swamp is nearly in balance with respect to inputs and outputs of organic carbon. Further strengthening this contention is the absence of soil humus or peat layers of any size.

Water col umn : The components of the model can also be examined separately. Analysis of the water column indicates that it functions primarily as a flow- through or transport entity. Allochthonous hydrologic inputs of organic carbon totaled 196 g cam-2.yr-1, 76% of the 258 g C-m-2-yr-1 of total 1977 water column inputs (Fig. 24). Hydrologic outputs totaled 214 g ~-m-2*~r'f, 89% of the 240 g ~-m-Z*yr-lof total water column outputs,and respiration accounted for the remainder (Fig. 24). The swamp segment ecosystem studied here was about 7% efficient in retaining and presumably respiring total organic inputs (>95% DOC) to the surface water compartment, and even less efficient in i-etaining a1 lochthonous inputs (94% DOC). Swamp floor: The swamp f 1 oor detri tal component had inputs of 378 g ~em-~~~r-', of which 92% were inputs of terrestrial plant 1 itter (Fig. 24). Output from the swamp floor totaled 358 g ~-m-2-~r-l,of which 84% was as respiration and 16% as leaching (Fig. 24). While the leaching loss percentage was not large, it did represent a significant amount of organic carbon and is 1ikely the primary difference in organic carbon cycling between wooded swamps and up1 and forests , up1 and forests having 1 ess 1eachi ng exports. The swamp floor net input of 20 g ~-m-~*yr-l(about 5% of total input) is certainly within the limits of error for the estimation of plant litter inputs and respiration outputs, and is not significant. Inputs and outputs for the swamp floor were in approximate balance with no apparent storage or net consumption of detrital organic carbon in 1977. Table 15. Inputs and outputs of organic carbon in Creeping Swamp for 1976 and 1977. A1 1 values in g C-m-2.yr-1.

IN PUTS Hydrol ogic: Upstream (TOC) Tributary (TOC) Groundwater (DOC) Rai nfall Biologic: Tree and sapling aboveground net productivity - wood increment 1i tterfall macro1 i tter net throughfall and stemflow Shrub and herbaceous plant aboveground net productivity A1 gal net productivity Total Inputs

OUTPUTS Hydrol ogic: Stream (TOC) Biologic: Respiration - water col umn aquatic benthic 34.1 45.9 terrestrial 1i tter 296 246 anaerobic 4.5 9.8 Total Outputs 438 541

Whole ecosystem:

Total hydrologic output of organic carbon accounted for 37% of the 578 g cam-2 of total detrital inputs to the 8-km swamp-stream segment in 1977. The ecosystem was thus 63% efficient at retaining detrital carbon input, converting most, if not all of it, to C02 via metabol ic processes.

Organic matter retention efficiencies have been computed from budget data for 1.7- km segments of Bear Brook (Fisher and Likens 1973) and Fort River (Fisher 1977), two upland streams with much higher gradients than Creeping Swamp. Comparison of retention efficiencies between the upland streams and Creeping Swamp is not useful, however, since segment lengths were very different. Organic carbon retention is a function of stream processing and storage features which, in turn, are related to segment length. In addition retention efficiencies cannot be normalized to a unit stream length since much of the input is lateral and fluvial inputs discontinuous and non-uniform along stream segments.

In order to avoid the problems associated with comparing organic matter retention efficiencies of stream segments of different lengths or orders, Fisher (1977) has proposed basing indices on organic matter loading. Fisher's loading coefficient (k) expresses the rate of organic matter concentration change based upon an exponential model:

where Co and Ci are output and input concentrations for a particular stream segment. The loading coefficient can be normalized to a unit segment length (z) yielding: k'z Co=Cioe . If organic matter concentrations in lateral fluvial inputs are similar to the upstream concentration then non-hydrologic inputs, primarily 1i tter inputs, throughfall and algal production are responsible for stream organic matter loading. These inputs, normalized to a unit stream length, can be considered as a loading potential (LP). Fisher's stream metabolism index (SMI), computed as total ecosystem respiration divided by the amount of respiration needed to prevent stream organic matter concentration increase, reflects the ability of the ecosystem to prevent loading, given its organic inputs. Assuming a steady state with no annual organic matter storage, the metabolism index is a1 so a retention index.

Creeping Swamp, while loading at a rate higher than both Bear Brook and Fort River, has a metabol ism index (retention index) intermediate between these two, but closer to that of Bear Brook, a stream which ex- hibits no organic matter loading (Table 16). This ref1ects relatively efficient organic matter retention in Creeping Swamp despite its high loading potential (Table 16). Creeping Swamp was better able to retain its organic matter inputs than Fort River, and only slightly less able than Bear Brook. The relatively high metabolism index of Creeping Swamp may be attributed to long organic matter retention times due to physical retention features and the dominance of CPOM input, a more readily retained organic matter fraction. Although Creeping Swamp was closer in size to Fort River, it has retention features more similar to those of smaller, lower order streams such as Bear Brook (Bilby and Likens 1979). The presence of trees within the swamp-stream ecosystem increases the tortuosi ty of flow pathways and the potential for debris dams. In addition, unlike upland streams, the low gradient of Creeping Swamp produces low water velocities with 1 ittle erosive force.

Table 16. Characteristics of three stream segment ecosystems and indices describing organic matter dynamics in each. Data for Bear Brook are from Fisher and Likens (1973) and Fort River from Fisher (1977).

Parameter Bear Brook Fort River Creeping Swamp

Stream order 2 4 3 Study segment length (km) 1.7 1.7 8 Drainage at downstream station (km2) 1.3 107 80 2 Organic matter input (kg/m ) 1.36 30.8 1.16 Loading potential 1 (kglm) 2.3 14.3 154 Retention efficiencyZ 0.34 0.04 0.63 Loading factor3 1 .O 1.025 1.175 Loading coefficient (k1)4 0 0.0143/km 0.02021 km Steam metabol ism i ndex5 1.O 0.638 0.82 - ~#eteorologicand biological organic carbon inputs per unit stream length n L(~inputs - c outputs)/c inputs e actor by which organic carbon concentration increases in the study segment (after Fisher 1977) 4~omputedas Co = Ci*ek'z where Co = output concentration Ci = input concentration k' = coefficient for stream segment - length z '~cosystem retentionlretention necessary to prevent loading (after Fisher 1977).

Organic carbon export from watersheds Creeping Swamp watershed: Exports of organic carbon were computed for the entire watershed drained by Creeping Swamp. Monthly export of FTOC at CP-10 was computed for 1976 and 1977 by mu1 tiplying weighted mean monthly concentration by total monthly streamflaw. FTOC export exhibited dramatic seasonal variation (Fig. 25); it was highest in winter and early spring and lowest in summer and early autumn, a pattern similar to streamflow (Fig. 6). The large November export in 1977 was absent in 1976. November export of organic carbon was related to streamflow even more dramatically than in other months due to the large source of readily leachable organic carbon avai 1able just after peak 1eaf-fa1 1 when a spate occurred. Total organic carbon (TOC) flux at the upstream station, CP-20, was 3.10 g corn-2 of watershed in 1975, 1.89 g ern-2 in 1976, and 6.81 g in 1977. TOC export at the downs ream station, CP-10, was 4.15 g ern-2 of watershed in 1975, 3.37 g in 1976, and 8.55 g in 1977. TOC flux at both stations was in all years >93% DOC. Precipitation recorded in Greenville, North Carolina, about 25 km from Creeping Swamp, totaled 110 cm in 1975, 102 cm in 1976, and 126 cm in 1977 (National Oceanographic and Atmospheric Administration 1975, 1976, 1977). Precipitation in 1977 was equal to the 25-year average, while that in 1975 was 13% less, and that in 1976 24% less than the 25-year average (National Oceanographic and Atmospheric Administration 1952-1 978). Comparison with other watersheds: Exports of TOC from Creeping Swamp were high compared to other watersheds, mostly upland (Table 17). Brinson (1976) postulated a 1inear relationship (Y = 0.023X + 0.1 58, r = 0.91 ) between organic carbon export, Y (in g ~*m-2),and annual runoff, X (in cm), from upland watersheds. I have added two additional upland watersheds to those compiled by Brinson and found little change in the relationship (Y = 0.023X + 0.355, r = 0.91). However, an entirely different linear relationship best described organic carbon export from Creeping Swamp, other eastern North Carolina waters sheds with extensive swamp drainage, and a Louisiana swamp (Y = 0.123X + 0.490, r = 0.89) (Fig. 26). The slope of the l ine describing swamp- draining watersheds is about 5.3 times greater than that for upland watersheds, indicating much larger organic carbon export per unit runoff. There were also higher organic carbon exports for some swamp-draining watersheds than for upland ones, and many of the low exports from swamp- draining watersheds were 1i kely due to the be1 ow-average preci pi tation in 1975 and 1976. In the segment of Creeping Swamp under study here, the net export of organic carbon totaled 22 g corn-2 of swamp in 1976 and 18 g corn-2 of swamp in 1977, about 2.6% of ecosys tem NPP, (Tab1 e 18). These figures must be Table 17. Annual organic carbon export* and runoff from various upland and swamp-draining watersheds.

Annual Annual Organic Carbon Area Runoff Export Watershed h2) (cm) (g ~-rn-~-yr-') Reference

Up1 and Arctic: Char Lake, N.W.T., Can. 43.50 15.8 0.30 de March (1975) Cool Temperate: Hubbard Brook, N.H. watershed #2 (defor. )t 0.16 122.1 2.71 Hobbie and Likens (1973) Borman, et a1 . (1974) watershed #6 (forest. )t 0.13 96.2 1.51 Borman, et al. (1974) Bear Brook 1.30 72.0 1.95 Fisher and Likens (1973) Mirror Lake 0.85 64.7 1.89 Jordan and Likens (1975) Fort River, Mass. 107.30 79.8 3.29 Fisher (1977) Marion Lake, B.C., Can. 13 204.8 5.17 Efford (1972) Nanaimo River, B.C., Can. 894 167.8 14.60' Naiman and Sibert (1978) Humid Tropics: Rio San Marcos, Guat. 170 85 2.20 Brinson (5 976) Rio Sauce, Guat. 300 86 3.20 Brinson (1976) Rio Polochic, Guat. 5247 194 4.80 Brinson (1976)

Warm Temperate: Louisiana swamp Tab1 e 17 (continued) Annual Organic Carbon Watershed Area Runoff Export Reference

Swamp-draining Warm Temperate: Creeping Swamp, N .C. CP-10 (1975) 80 This Study (1 976) 80 This Study (1977) 80 This Study CP-20 (1975) 32 This Study (1976) 32 This Study (1977) 32 This Study Palmetto Swamp. N. C . (1975) 54 This Study (1976) 54 This Study

7 I racey Swamp, N.C. (1975) 141 This Study (1976) 141 This Study Chicod Creek, N.C. (1975) 132 This Study (1976) 132 This Study Clayroot Swamp, N.C. (1976) 110 This Study

*Export is as TOC for all upland watersheds except for watersheds 2 and 6 at Hubbard Brook and the Marion Lake watershed, in which export was as FTOC. Export was as FTOC for all swamp-draining watersheds except Creeping Swamp in which export was as TOC. +Export was computed by adding the average of 1965-1970 bedload measurements to the 1967-1969 average fine particu- +late and dissolved fractions. Runoff is the 1967-1 969 average. Watershed is mountainous but contains several bog lakes and lakes used for log storage. ++Annual runoff was computed by dividing total organic carbon export by the product of mean annual organic carbon ++concentration and watershed area. Runoff values are assumed to be equal to those at Creeping Swamp for that year.

treated with caution, however, because they were computed as the difference between two large values (total input and output), each of which has at least 5 to 10%of error associated with it. Nonetheless, the net export from Creeping Swamp seems to indicate that the swamp portion of the watershed exported more organic carbon per unit area than did the watershed as a whole (with exports of 3.37 and 8.55 g Corn-2 of watershed in 1976 and 1977).

Table 18. Net hydrologic flux of TOC (g of swamp-yr-') from Creeping Swamp for 1976 and 1977. Net flux computed as the difference between total hydro1 ogic inputs and outputs.

Stream Tributary Preci~i-Ground- Total ta tim water

Inputs 18.9 39.0 1.8 3.0 62.7 Outputs Net Output

Inputs Outputs Net Output 18.1

Exports of organic carbon are 1i kely influenced by low frequency runoff events as well as annual runoff. The timing and size of the 7-year spate that occurred in November 1977 was shown to influence considerably the 1977 dissolved organic export. Very low frequency spate events (>SO-100 yr) have substantial ly influenced both short and long-term export of particul ate material from mountainous watersheds in Oregon (Swanson, --et a1 . , in press) ; however, their effect is probably much less in lowland Coastal Plain swamp-streams. Nonetheless the effect of very low frequency spates on the export of particulate detritus in swamp-streams is uncertain and not accounted for in annual budgets such as the one presented here. Naiman and Sibert (1978) reported annual runoff of 168 cm and annual organic carbon export of 14.6 g Corn-2 by the Nanaimo River from an up1 and British Columbian watershed which includes several bog lakes and lakes used for log storage. When plotted on Fig. 26, this watershed falls in the intermediate region between upland and swamp-draining watersheds. This may reflect a combination of upland and swamp characteristics not present in the other watersheds plotted. While no data are available on the percentage of watershed in swamp habitat for the swamp-draining watersheds reported here, it is possible that they may be roughly similar and that watersheds draining less swampland or other types of wetlands would fall into the intermediate regions of Fig. 26, as does the Nanaimo River. The runoff-organic carbon export relationship may reflect different percentages of wetland drainage, or different types. Fate of exported organic carbon A1 1 watersheds export organic carbon regard1 ess of vegetation type or topography. This export represents an unavoidable leakage of the products of photosynthesis and decomposition from ecosystems due to hydrologic processes. While ecosystem productivity depends upon water for growth and nutrient transport, water is also responsible for the removal of some of that production to downstream aquatic ecosystems. The primary ecosystem carbon cycle is from the inorganic, gaseous state to the organic state via plant productivity, and back to the inorganic state via respiration processes within the ecosystem. A secondary cycle involves the same initial transfer of carbon from the inorganic to the organic state, but is followed by an export of organic carbon by way of hydrologic processes and subsequent breakdown of organic carbon to the inorganic state in downstream aquatic ecosys tems. Whi 1e the primary cycle dominates in most ecosystems, incl uding swamps, the re1 ati ve importance of the secondary cycle depends upon ecosystem hydrologic properties. Swamps and other wetlands have abundant surface water as well as relatively high rates of primary productivity. This combination of properties results in high organic carbon leakage compared to other ecosystems, especially terrestrial ones. Most of the organic carbon exported from watersheds is in the dissolved form and not readily utilized by consumer organisms (Wetzel and Manny 1972; McDowell and Fisher 1976). In order to complete the carbon cycle, transformation to a more readily utilizable form must occur prim to respiratory breakdown and release of inorganic carbon.

The fate of terrestrially produced and exported DOC has been recently investigated. While Lush and Hynes (1 973) reported that DOC precipitation occurred in freshwater, Lock and Hynes (1976) showed that it was of minor importance in natural hardwater streams. Shol kovi tz (1 976) reported that significant DOC precipitation occurred when filtered river water was mixed with filtered seawater. Gardner and Menzel (l974), in a study of river, estuarine, and offshore sediments, found that the greatest deposition of lignin-derived organic materials occurred at the interface of fresh and seawater. While most vascular plant-derived and exported DOC has a short retention time in flowina water ecosystems and is poorly utilized by heterotrophs, precipitated DOC may be much better utilized due to increased retention time, especially in vertically stratified estuaries, increased surface area avai 1able for bacterial colonization, and availabi 1i ty to a greater variety of organisms, particularly those which can only utilize particulate matter, such as f il ter feeders Estuaries have been generally viewed as among the most productive ecosystems and as serving important nursery functions (Lauff 1967; Odum 1971). Inputs of vascular plant-derived organic carbon from upstream may be extremely important to estuarine secondary productivity. Sibert, et a1 . (1977) and Naiman and Sibert (1978) have recently demonstrated a direct 1 ink between estuarine fish production and detrital carbon inputs from upstream terrestrial sources. In our study watersheds draining swamps have been shown to be greater exporters of organic carbon than upland watersheds. Swamp ecosystems, especially swamp-streams and larger floodplain swamps in coastal areas, may be very important in maintaining high secondary productivity in downstream estuarine ecosystems.

Comparison between natural and channel ized streams

There is no clear difference in weighted mean annual DOC and FPOC concentrations between natural and channelized streams in eastern North Carolina (Table 9). The natural streams may have slightly higher weighted mean annual DOC and lower FPOC concentrations than the channelized streams, however there is considerable overlap in the data. More intensive sampling, especially during spates, is needed to discern any significant differences in weighted mean annual concentrations.

Seasonal patterns of organic carbon concentration, however, differ between the natural and channel ized streams (Fig. 21 ) . Clayroot Swamp (CY-10) and Tracey Swamp (TR-10) , both channel ized streams, have low FTOC concentrations in summer and early fall, whereas Chicod Creek (CH-20) and Palmetto Swamp (PM-10) , both natural streams, have somewhat higher summer and fall concentrations, as did Creeping Swamp (Fig. 7). The channelized streams have relatively high FTOC concentration at high streamflow and low FTOC concentration at low streamflow, whereas the op- posite pattern appears to be true for natural streams. FTOC concentration was positively correlated with loglo (streamflow) for the four channelized streams and negatively correlated for two of the three natural streams studied (P<.05). This pattern was generally true for DOC concentration as well. Fisher and Likens (1973) reported a significant positive correlation between 1ogle (dissol ved organic matter concentration) and loglo (streamflow) for Bear Brook, New Hampshire, and attributed it to leaching of the canopy and soil litter and flushing of adsorbed organic matter from clay surfaces during rainfall events. No significant relationship was found for two other nearby New Hampshire streams over an entire year (Hobbie and Likens 1973).

The differences in organic carbon concentration-streamflow relationships between natural and channelized streams in eastern North Carolina may reflect differences in the source of runoff. At low streamflow the channelized streams, as well as upland streams, drain primarily groundwater with relatively low organic carbon concentration, and the natural streams drain swamp surface waters in which DOC has been concentrated due to high evapotranspiration and increased leaching of plant detritus. During periods of high streamflow the channelized streams have higher DOC concentrations, simil ar to those of natural streams, since they receive inputs f 1ushed from natural swamp-stream tributaries. The channel ized streams may also have higher FPOC concentrations at high streamflows, particularly as streamflows rise and peak during spates. The negative correlations between FTOC concentration and 1og10 (streamfl ow) in Creeping and Palmetto Swamps indicate a dilution effect in natural swamp-streams which may be tempered by increasing amounts of swamp inundated as streamflow increases, subjecting greater amounts of detritus to leaching. 3. PHOSPHORUS CYCLING IN THE FLOODPLAIN ECOSYSTEM AND EXPORTS FROM THE WATERSHED by Laura Anne Yarbro INTRODUCTION Genera l The semi-aquatic character of wetlands and their interactions with both terrestrial and aquatic ecosystems has 1 ed to many investigations of the capabil i ty of wet1 ands to process introduced materials so that these materials are not ultimately released to ecosystems downstream. However, the processes affecting elemental biogeochemistry in wetlands are poorly understood. For elements without significant gaseous forms, the hydrology of the wetland ecosystem is the driving force behind material transport into and from the system (Bormann and Likens 1979; Gossel ink and Turner 1978). The biota of wetlands are adapted to the hydrologic regime and its related material fluxes and, furthermore, can over time markedly influence the ecosystem's hydrology and export of materials (Richardson, --et al. 1978). Knowledge of the functioning of a wetland ecosystem in response to its characteristic hydrology and material fluxes can be gained by the study of the cycling or biogeochemistry of one or more elements in the ecosystem. I have chosen to examine the dynamics of the essential element, phospharus, in a small headwater stream floodplain swamp, Creeping Swamp, located in the Coastal Plain of North Carolina. The element phosphorus was chosen because (1) it is essential for the growth and maintenance of 1 ife; (2) it is generally described as being 1 imi ting in freshwater ecosystems (Voll enweider 1968; Schindl er 1977); and (3) its cycling is relatively easy to quantify: it has no gaseous phase significant to ecosystem biogeochemistry. Phosphorus cycling in wetlands Wetlands, by their ~uxtapositionto both terrestrial and aquatic ecosystems, may act as filters or traps for materials received from both types of ecosystems (Grant and Patrick 1970; Wharton 1970; Bender and Correll 1974; Steward and Ornes 1973, 1975; Kitchens, et a1 . 1975; Lee, --et al. 1975; Boyt 1976; Odum and Ewe1 1976, 1978; ~ichcdson,et al. 1976; Mitsch, --et al. 1977; Sloey, --et a1 . 1978). However, this-?je=ral- ization does not hold true for all wetlands; some wetlands export materials, particularly nitrogen and phosphorus, rather than conserve them (Crisp 1966; Gardner 1975; Heinle and Flemer 1976; Kemp 1978). Reten- tion is sometimes seasonally temporary (Bender and Correll 1974; Burke 1975; Axel rad, et a1 . 1976; Woodwell and Whi tney 1977; Richardson, et a1 . 1978; spang1 er ,etTl.-- 1976, 1977; Simpson and Whigham 1978; ~i 1 tonand Kadlec 1979), or occurs for a short period of time after which the ecosys tem becomes "saturated" (Steward and Ornes 1973, 1975; Val iel a, et a1 . 1973; McPherson, --et a1 . 1976; Valiela and Vince 1976), or significantly modifies ecosystem characteristics over successional or perhaps evolutionary time (Steward and Ornes 1973, 1975; Hartland-Rowe and Wright 1975; Richardson, --et a1 . 1978). Some ecosystems appear to retain a significant portion of their inputs of phosphorus (Burke 1975; Butler 1975; Viner 1975a; Hartland-Rowe and Wright 1975; Gaudet 1976; Odum and Ewe1 1976, 1978; Day, et al. 1976; Boyt, --et al. 1977; Kuenzler, --et al. 1977; Nessel 1978; ~onikaandLowry 1978). Thus, it is imperative to understand the factors controlling phosphorus cycling in wetlands in order to generalize about wet1 and biogeochemis try.

Factors controlling phosphorus cycling in wetlands:

The character common to all wetlands is that they are wet; they have water-saturated soils and are inundated periodically, seasonally, or continually. The nature and timing of water entering, remaining and leaving the wetland ecosystem is the hydrology of the system and consti- tutes the most important factor influencing wetland biogeochemistry (Gosselink and Turner 1978). It is also the 1 ink between wetlands and adjacent aquatic and terrestrial ecosys tems (Likens and Bormann 1974; Loucks 1975). Coupled with the amount and source of water entering the wetland ecosystem is the burden of materials carried in by the waters. The magnitude of the inputs and the amount and availability of phosphorus in the sediments are the second most important set of factors influencing wetland elemental cycling. These factors are most frequently and drastically modified by the activities of man. Thirdly, climate, as mani- fested by the timing of hydrologic inputs and seasonal responses of the biota, affects the cycling of materials in wetlands. Finally, the characteristic biota of a wetland and their evolutionary and successional response to elemental, hydrologic and energy inputs also exert control on elemental fl~xesand cycling in the ecosystem. Note that no one set of factors acts alone to control wetland biogeochemistry; rather all are affected by the others. It is the sum of these interactions which ultimately results in the characteristic functioning of a particular wetland. Yarbro (1979) gave a detailed review of the literature on phosphorus biogeochemistry in freshwater marshes and swamps and in salt marshes, particularly stressing the re1ationships to hydro1ogy, 1eve1 s and sources of inputs, storage of phosphorus within the ecosystem, climate, and influence of mankind.

Phosphorus cycling in Creeping Swamp:

A preliminary model of phosphorus inputs, cycling, and exports for the floodplain and surrounding watershed of Creeping Swamp has been developed (Fig. 27). Phosphorus cycl ing in the Creeping Swamp watershed is dependent primarily on precipitation inputs because of the highly weathered status of the surface soils (Soil Survey, Pitt County 1974). Additionally, a portion of the upper elevations of the watershed is covered with pocosins; the build-up of organic materials in the sediments of these systems has probably isolated them from mineral soils. The floodplain of Creeping Swamp receives phosphorus from incoming precipitation (Fig. 27) and streamwaters draining surrounding uplands. The fluxes of phosphorus in streamwaters draining both uplands and floodplain are dependent on stream volume. Anthropogenic sources of phosphorus may also contribute to streamwater fluxes.

Stream water phosphorus in the floodplain may pass through unchangedor be removed by seasonally abundant filamentous algae or by the forest floor. The forest floor includes leaf litter, the humus layer of the soil and associated microorganisms (Gosz, et a1 . 1976). Phosphorus in the algae may be exported downstream or it Eybe taken up by the forest floor follmring algal death. Phosphorus may be lost from the forest floor by leaching or loss from decomposing leaf litter to stream- or floodwaters, by uptake by roots of vegetation, or by percolation into the mineral subsoil. Phosphorus in vascular vegetation may be returned to the forest floor or floodwaters by leaching or litterfall. Subsoil phosphorus may be lost from the system by deep groundwater loss, or it may be taken up by roots. Phosphorus in precipitation falling on the floodplain may be intercepted by vegetation or fa1 1 directly on flood- or stream waters or on the dry forest floor (Fig. 27).

Losses of phosphorus from the floodplain and the watershed apparently occur predominantly in stream flow. A1 though rainfall is re1atively constant throughout the year, seasonal evapotranspiration rates are not and they exert a controlling force on the aquatic fluxes of phosphorus. During the cool, flooded season, fluxes of phosphorus are chiefly between floodwaters and forest floor. During the growing season when the floodplain is dry, phosphorus fluxes may be biological ly mediated between forest floor and soil and canopy vegetation with little or no export from the ecosystem (Fig. 27).

METHODS

Phosphorus budget for Creeping Swamp watershed

An initial objective of this study was to quantify a phosphorus budget for the entire watershed of Creeping Swamp (Fig. 1). This involved measurement of hydrologic inputs and outputs and associated phosrhorus fluxes with special attention to storm runoff periods when phosphorus fluxes in the watershed were greatest (Kuenzler, --et al. 1977). Phosphorus fluxes for two Water Years, 1977 (1 October 1976 to 30 Septem- ber 1977) and 1g7C (1 October 1977 to 30 September 1978), were measured.

Hydrologic measurements:

The quantity of precipitation received in Creeping Swamp watershed was estimated using daily measurements at the fire tower in Wilmar, N.C., about 5 km from the watershed (Fig. 1). In addition, some bi-weekly measurements were taken in an open field near CP-14 using a cylindrical rain gauge (Forestry Suppl iers) , particularly during the growing season when local thundershowers dominated rainfall. Corrections for evapora- tive losses were made. In most cases, monthly sums for the gauge at CP-14 agreed quite well with the fire tower data; therefore, Wilmar rain data was used in the calculation of precipitation fluxes of phosphorus. Methods for obtaining stream flow for Creeping Swamp and its tributaries were given by Mu1 holland (Chapter 2, above). For those tributaries whose individual watershed areas were uncertain, annual runoff values were obtained using slopes obtained from regressions of CP-10 or CP-20 daily discharges with measured discharges of a given tributary for all dates when hydrologic flux measurements were made. These slopes, mul- tiplied by CP-10 annual runoff, estimated annual runoff for each of the tributaries assuming that annual runoff was relatively constant per unit watershed for a given year. This was borne out by the comparison of annual runoff values for gaged watersheds of widely differing sizes (U.S. Geological Survey 1978, 1979), Water chemistry: Stream waters at CP-10, CP-14, CP-20, TB-01, TB-03, TB-07, TB-09, TB-10, TB-04 and TB-02 were sampled approximately biweekly in Water Year 1977 and monthly in Water Year 1978 when significant flow occurred (>0.05 m3.sec-1). In addition, water qua1 i ty measurements in Water Years 1975 and 1976 (Kuenzler, et a1 . 1977) for the main stream stations, CP-10, CP-14 and CP-20, provided further data to corroborate seasonal patterns. Finally, several series of daily measurements were made in winter (27 January - 4 February 1976 and 20-23 January 1978), early summer (3-8 June 1976), autumn (7-11 November 1977) and early winter (16-21 December 1976) to monitor changes in water quality during storm runoff (spate) periods. Sampling usually began on the first day of a large fr~ntalrainstorm and continued until 3-4 days after peak flow occurred. Field procedures were the same as those of Kuenzler, --et a1 . (1977). Field measurements of specific conductance (conductivity) and temperature were made with a YSI Model 33 S-C-T meter; conductivity was corrected for temperature, pH was measured using a portable Beckman Model N glass electrode pH meter. Grab samples for pbs phorus fractions, color, and turbidity measurements were taken from the center of the stream channels. Unfi 1 tered samples were stored at ambient temperature in polyethylene bottles for laboratory determinations of turbidity and color. Turbidity was measured on a Hach Model 2100 Turbidimeter calibrated against standard formazin suspension. Color was determined on filtered (0.45 um membrane) 50 ml samples using a Klett colorirneter, a #42 filter and a 4 cm cell . Measurements were cal i brated wi th standard chloropl atinate solutions (American Pub1 ic Heal th Association 1975). These 1 aboratory measurements were made wi thin 12-36 hours of sampl e coll ection. Water samples for phosphorus analysis were separated into filtered and unf il tered fractions in the field. Unf i1 tered samples were poured directly into clean, acjd-washed polyethylene bottles and stored on ice. Acid-washed membrane f i 1 ters (Gelman Metricel GA-6, 0.45 pm pore size) with glass fiber prefil ters (Gelman type A-E) in Mill ipore Swinnex filter holders were used with 60 ml syringes to fil ter water samples. Filtered samples were placed in polyethylene bottles and put on ice. Upon return from the field, all samples were frozen until analysis. On the average, about two weeks intervened between sample collection and chemical analysis. The following fractions of phosphorus were differentiated by filtration and chemical analysis : (1 ) filterable reactive phosphorus (FRP) ; (2) total filterable phosphorus (TFP); (3) total phosphorus (TP); (4) dialysable reactive phosphorus (DiRP); and (5) total dialysable phosphorus (TD~P) . Reactive fractions were measured either by the automated stannous chloride procedure on a Technicon Autoanalyzer (EPA 1971 ) or manually by the method of Strickland and Parsons (1972). Reactive samples run on the autoanalyzer using the 660 nm wavelength were corrected for absorption by the natural color of swamp water by the use of samples, to which the phosphorus mixed reagent minus ammonium molybdate had been added, as blanks. Total fractions were measured by persulfate digestion and the automated stannous chloride method (EPA 1971 ) or by the persulfate digestion method of Menzel and Corwin (1965) followed by the colorimetric determination of Strickl and and Parsons (1972). Reproduci bi 1i ty between the two methods of analysis was tested using standards and swamp water samples; no detectable differences were found. Standards and blanks were included with each analysis. Recoveries were checked by using internal standards intermittently. Dialyzable, or truly soluble, and colloidal fractions of phosphorus (Koenings 1977) were measured on main stream waters in Water Year 1977 and intermittently in Water Year 1978. Fractions were separated using Mill ipore Pel1 icon membranes (Type R), with a molecular weight cutoff of approximately 10,000, using methods of Kuenzler, et al. (5979). The following fractions were obtained by difference between total and reactive fractions : (1 ) particulate phosphorus (PP) ; (2) f i 1 terabl e unreactive phosphorus (FUP) ; and (3) dialysable unreactive phosphorus (DiUP) . Col- loidal reactive (CoRP) and colloidal unreactive phosphorus (COUP) concentrations were determined by differences between respective filterable and dialyzable fractions. Ground water quality in the Creeping Swamp study area was measured by sampling wells installed in 1975 and 1976 by Robert Sniffen and Patrick Mu1 holland. We1 1s consisted of 10 cm PVC pipe placed 15-740 cm into the ground with screened holes along the entire length of the pipe or with openings only at the bottom; see Yarbro (1979) for locations and depths of the wells. The tops of the wells were covered to prevent entry of precipitation and debris. We1 1s not inundated by stream or flood waters were sampled monthly from August 1976 through August 1977. Prior to sampl ing, we1 1s were pumped out and a1 1owed to refi 11. Only conductivity, pH and filterable phosphorus fractions were measured on well water samples. Because of suspended sediments in the water, only filtered water was analyzed. Data from Winner and Simmons (1977) was used to estimate phosphorus concentrations in deep ground water. Bulk precipitation collection was described above (Chapter 2). Samples for volume measurements were coll ected every two weeks from August 1976 through August 1978. Samples for phosphorus determinations were collected every four weeks over the same time period and consisted of intercepted rainfall of the previous two week period. Bottles used for collection were acid-washed and samples were preserved in the field using HgC12 with an average final concentration of 40 mg.1-1 (American Pub1 ic Health Association 1975). Phosphorus analyses on unfiltered samples were done by au toanalyzer as described above; some low-vol ume samples were diluted with deionized water because of the interference at higher concentrations of HgC1 2. Hydrologic fluxes of phosphorus: Annual stream water fluxes of phosphorus for each of the tributaries, CP-20 and CP-10 were obtained by multiplying annual runoff values (see above) by annual weighted mean concentrations of phosphorus. Concentra- tions were weighted by the flow occurring on the day of sampling so that concentrations measured on high flow dates were given more weight than concentrations on low flow dates. For the watershed budget, exports from CP-10 and CP-20 were normal ized for upstream watershed area resulting in units 0fm~-m-2.~r-l Fluxes into and from the 3.2 km2 floodplain study area were normalizei for the floodplain area inside the roads (Fig. 2) for the swamp phosphorus budget. Deep ground water loss of phosphorus from the watershed was estimated by mu1 tiplying average phosphorus concentrations in the wells by the volume of water loss (Winner and Simmons 1977). Phosphorus data from the three rain collectors were averaged and then weighted by the average volume of water collected during each sampling period so that higher volume samples were given greater weight. Annual weighted mean concentrations of precipitation phosphorus were mu1 tip1 ied by annual rainfall data to calculate annual fluxes of phosphorus in precipitation. Phosphorus cycling in Creep:nq Swamp floodplain The study of phosphorus cycling in Creeping Swamp floodplain included quantification of inputs and outputs, internal flows, and standing stocks. Measurements of stream water and precipitation fluxes have a1 ready been described. Internal flows examined in this study included 1 i tterfall , throughfall , s temf low, sedimentation of particulate water-borne phosphorus, and forest floor-water exchanges of phosphorus, The methods of the latter will be described in a separate section. Standing stocks of phosphorus in soil, ground 1itler, herbs, shrubs, vines, and bryophytes were measured. Standing stocks of phosphorus in canopy vegetation were estimated using 1i terature val ues and biomass data of Mu1 hol 1and (1979). Study area: In order to quantify the amount of phosphorus entering the floodplain 2 portion of the watershed , the 15 km area circumscribed by county and state roads,(fig. 2) was delineated. Approximately 3.2 km2 or 20% of this area was floodplain; the rest was crop lands and forests. Tributaries entering the study area were identified. Tributaries TB-01, TB-07, TB-09, TB-10, TB-02, and TB-04 (Fig. 2) were channel ized streams or drainage ditches; TB-03 was the only natural, unchannel ized tributary. The main stream of Creeping Swamp entered the study area at CP-20 and exited at CP-10. The portion of the floodplain selected for intensive study, CP-14 (Fig. 2), was about 350 meters wide, bordered on both sides by young pine plantations , and re1atively isolated from roads and ditches.

Floodplain hydrology:

The duration, extent, and frequency of floodplain inundation within the study area, particularly at the intensive study site, were estimated by recording depths of water at 5 m intervals along five transects of the floodplain (see Mu1 hol land, Chapter 2 above). Using a regression model which predicted the fraction of floodplain inundated as a function of water level at CP-10 (Mu1 hol land 1979) and daily Cp-10 water level data, I generated daily, monthly, and annual inundation fractions and areas for the whole floodplain of the Creeping Swamp study area for Water Years 1974-1 978.

Phosphorus in herbs, shrubs, vines and bryophytes:

The standing crop of phosphorus in herbs, shrubs, and vines was estimated from harvests made by P. Mulholland (see Chapter 2, above). The standing crops of bryophytes on the floodplain was estimated in August 1976 by removing all bryophytes in 1 m2 plots at 15 stations across the floodplain transects at CP-14. Plant samples were oven-dried at 80 C for at least 48 hours and then weighed to obtain dry weights. Rep1 icate subsamples were ashed at 500-550 C for 4 hours to obtain ash weights for calculation of ash free dry weights (AFDW). AFDW was assumed to be equivalent to organic matter. Phosphorus determinations on harvested plant materials, litter, and other detrital materials (see below) followed the methods suggested by Likens and Bormann (1970) and Lee et a1 . (1965). To 1-2 g rep1icate subsamples of dried material, 2 ml of 10% Mg(N0312 were added and the samples were ashed as described above. One ml of 11 fi H2SO4 was added to dissolve the ash. Residual ash (usually inorganic sediment associated with floodplain floor materials, see below) was removed by filtration. The filtrate was diluted and analyzed for phosphorus (Strick- land and Parsons 1972). Recovery of phosphorus following ashing and also fol 1owing addition of Mg (NO3) 2 determined by standard addi tions was quite good. The effects of the acidity of 11 fi H2SO4 on color development of the molybdenum blue complex was negligible due to the dilution necessary to read the samples (1 ml 11 N H2SO4 diluted 2100 to 4200-fold).

Phosphorus in ground litter:

Standing stocks of phosphorus in litter materials on the swamp forest floor were estimated bimonthly in Water Year 1978 and on occasional samples in W.Y. 1976 and 1977,including some collected by P. Mulholland. Ground I i tter was coll ected at random distances and directions from 15 stations across the same transect at CP-14 used to determine inundation patterns. All recognizable litter inside a 0.25 m2 circular metal frame placed on the forest floor at each station was collected and put into plastic bags. Smaller frames were used when the transect was flooded. Branches greater than 2 cm in diameter were omitted. In the laboratory, twigs and branches were separated from leaf materials for each station sample; then both fractions were dried at 80 C for about one week. Dry weights were determined; the samples were ground and then subsampled in replicate for the sshing and phosphorus procedures described above. Soil phosphorus: Organic and available phosphorus present in the top 25 cm of soil was estimated on 28-29 April 1977 and 30 May 1978, when the floodplain was dry. At each of 15 sites along a transect at CP-14, five cores (2.5 cm dia. ) of soil were taken. Each core was separated into a surface sample (0-5 cm soil depth) and a subsurface sample (6-25 cm). For each site, the respective depth subsamples were pooled and placed in plastic bags. Each site subsample was weighed, then dried at 80 C for 48-72 hours and reweighed to determine moisture content and dry weight. Each subsample was ground and homogenized and replicates were removed for ashing and phosphorus analysis. The possible loss of waters of hydration of the clays during ashing was investigated. Rep1 icate ashed samples of subsoil and litter material were weighed after cooling in a desiccator, then wetted, dried at 80 C and reweighed. Differences in weights were slight but significant at the d=0.10 level but not at the 0.05 level. No corrections were made. The sum of "available" phosphorus (phosphorus sorbed or complexed to soil particles but not chemically bound in inorganic crystal l ine 1attices) and organic phosphorus was determined using the methods of Legg and Black (l955), Saunders and Mil 1 iams (1955) and Lee, et al. (1965) on ashed materials.

Throughfall : Throughfall (canopy drip) was sampled using coll ectors described above (Chapter 2). Field sampling, duration and frequency of coll ection and laboratory protocols were the same as for bulk precipitation. For each col 1 ection date, average phosphorus concentrations of the 15 sampl es were weighted by the average volume of throughfall relative to the total average annual volume. These weighted values were summed to obtain annual weighted mean concentrations of phosphorus in throughfall. Based on the average difference between volumes of rainfall and throughfall, the average annual throughfall volume was estimated and mu1 tip1 ied by the annual weighted mean concentrations of phosphorus to obtain annual fluxes.

Quantitative measurements of the flux of phosphorus in stemflow were difficult to make because of problems with catching all the water that ran down tree trunks during rain storms. Qualitative estimates of phosphorus concentrations in stemflow were nade, however, on samples collected as described in Chapter 2 above. Observations during rain storms confirmed that the collars did not trap all the water; in addition, the garbage cans frequently overflowed. Stemflow collected over two-week periods was sampled from March 1977 through June 1978. Phosphorus analyses were done on the autoanalyzer. Quantitative fluxes of phosphorus in stemflow were estimated using concentrations measured in Creeping Swamp and literature estimates of the quantity of stemflow relative to precipitation (Helvey and Patric 1965).

Li tterfall : Litter less than 0.3 m in length (leaves, small twigs, fruits and flowers) of vegetation greater than 1.5 m in height was sampled as described in Chapter 2. The determination of dry weights, AFDW and phosphorus content fol l owed the methods for vegetation harvests. Phosphorus analyses were made on pooled samples for each collection date. Patrick Mu1 hol land did the field col lections and determinations of dry weight and AFDW during the first two years. I measured all phosphorus concentrations and also did the field sampling and dry weight and AFDW determinations during the third year. Additional ly, subsamples of P. Mu1 hol land's macro1 i tter (branch-fall ) collections (Chapter 2, above) were analyzed for phosphorus concentrations. Sedimentation: The amount of particulate phosphorus in floodwaters that sedimented out during floodplain inundation was assessed by measuring the increment of dry weight and phosphorus on tared squares of constant area (4 cm X 4 cm) of Quercus michauxii leaves placed in hardware cloth holders and located in triplicate on the CP-14 floodplain at four different elevations. These leaves decomposed slowly and provided a natural substrate for sedimenting materials. Leaves were collected immediately after falling in October 1977. Prior to being placed in the swamp, the leaf squares were soaked in deionized water for 48 hours to remove easily leachable materials and then were air-dried and weighed. Leaf squares were placed in the swamp for two week periods from January-May 1978. After two weeks, the leaf squares and holders were carefully removed from the floodplain and placed in plastic bags. In the laboratory, squares were air-dried for a week, weighed, oven-dried at 80 C for 48 hours and reweighed to determine air and dry weights, respectively. Rep1 icates for each elevation were then pooled and ashed to determine ash weights; phosphorus analysis of the ash followed the same protocol as in litter analyses except that Mg(N03)~was omitted. Earlier studies with standards demonstrated about a 5% reduction in phosphorus recovery in samples not treated with Mg (NO3)2. Rep1 icate control leaf squares were leached, air-dried, weighed, oven-dried, weighed, ashed, and analyzed for phosphorus content to account for air- to dry- weight changes, leaf ash, and phosphorus content. Only on rare occasions did some leaf squares lose weight while on the swamp floor; in those few cases, weight losses were a very small percentage of the total weight of the squares. Forest floor-water exchanges of phosphorus A third objective of this study was to measure the fluxes of phosphorus between the swamp forest floor and overlying floodwaters and to elucidate the factors controlling these fluxes. Experimental measurements were made in situ at CP-14 using additions of either inorganic phosphate or radioactive phosphorus-32 (32~)to chambers which isolated portions of the forest floor and associated floodwaters. Disappearance of the added phosphorus from the floodwaters was monitored for short periods of time (1-5 hours). Measurements were made throughout the winter and early spring months of 9977, 1978 and 1979 over a wide variety of temperatures, phosphorus concentrations, and floodplain el evations, and sometimes with the addition of biological inhibitors. Flowing-water chambers -- 1977: In the winter of 1977, phosphorus exchange experiments using flowing water chambers (Fig. 28A) were done simul taneously with aquatic respiration experiments conducted by P. Mulholland. The chambers were rectangular, made of clear plexiglass, isolated a 0.03 m2 area of floodplain, and could contain a maximum of 4.6 9 of water. Water was pumped in a closed loop through the chambers by Teal Marine submersible pumps driven by a 12 volt automobile battery. A reservoir was placed between the pump and chamber to replenish water lost due to sampling. Water flowed through the small inlet and outlet subchambers connected to the main chamber by a series of holes in order to maintain uniform circulation throughout the chamber . Chambers (2-3 per experiment) were pressed into the forest floor to a depth of 5-8 cm with the aid of a knife to cut through litter. Prior to the experiment, the injection ports were sealed and outside water was pumped through the chamber to clear out turbidity. The circulation system was then closed and allowed to flow for 5-10 minutes before initial sampling. The pump was turned off and initial (experimental time = 0) samples were taken from one of the water-circulating ports using a suction device. Pumping was resumed and known aliquots of phosphate standards (KH2P04 solution) and NaCl standards were injected into the chambers -via the injection ports. Enough phosphate was added to result in concentrations 15-20 pg-1-1 higher than ambient. Chloride was added as a conservative tracer to estimate the water volume in the chamber apparatus. In addition, some chambers were pretreated with 37% formaldehyde solution (60-180 ml) for about 30 minutes to inhibit biological activity. It was not evident, however, from measured dissolved 02 and C02 changes in the chambers that formaldehyde treatments effectively inhibited biological BATTERY

Water Surface G - Water

SUPPORTIVE RING

-PLASTIC TUBING INJECTION I 1

Figure 28. Design of chambers for in situ measurements of forest floor-water exchanges of phosphorus. A. Flowing-water chamber used in W.Y. 1977. B. Still-water chamber used in W.Y. 1978 and 1979. activity.

After a predetermined time, depending on ambient temperature, final samples were removed in the same fashion as initial samples. Initial and final sampl es were f il tered through 0.45 pm Gelman membrane f il ters and analyzed for FRPy TFP and C1 (Automated thiocyanate, EPA, 1974) on the autoanalyzer. For the purposes of a1 1 these experiments, standard phosphate and 32p04 were assumed equivalent to the FRP fraction measured by chemical analysis. In addition to chamber water samples, samples of nearby floodplain waters were also analyzed for the above chemical constituents. Experiments of this nature, using 2-3 chambers at a time, were conducted from February through June 1977; incubation times ranged from 1-4 hours depending on water temperatures. Experiments were conducted in several areas of the floodplain depending on water level. Fluxes of phosphorus between forest floor and floodwaters were expressed on a m2 basis.

Several serious problems were associated with this technique. Roots in the surface soil under the forest floor had to be cut to insert the chamber and often a tight seal between chamber and substrate was diffieul t to maintain. This was sometimes reflected in measured dilutions of the added C1 which were much higher than theoretically possible water volumes in the chambers. Despite the care taken to ensure uniform water circulation in the chambers, it was not clearly evident that this occurred. Finally, the weight of the 12 volt battery made field logistics extremely difficult. Because of these problems, a different experimental technique was used in 1978 and 1979.

Still-water chambers -- 1978 and 1979:

A. Field procedures: 1978

To obtain better rep1ication, facil 1tate experimental treatments, and alleviate problems of ineffective seals between exchange chambers and swamp floor, a different type of chamber without a pump was used in 1978 and 1979 (Fig. 28B). A1 though natural flowing water conditions were not mimicked, these chambers should provide a reasonable approximation of natural processes if incubation times were short enough to prevent the development of anoxic conditions. This chamber consisted of a base of 15 cm diameter polyvinylchloride (PVC) pipe which was pressed into the soil of the forest floor with the aid of a knife or keyhole saw. The pipe sections were about 15 cm long; following insertion, about 5-8 cm of pipe projected above the forest floor. To this base, a cylinder of clear polyethylene tubing (8.10 mm thickness), also 15 cm in diameter, was attached tightly by wrapping elastic around the base and the tube (Fig. 28B). The water within the chambers was between 10 and 15 cm deep, resulting in 2-3 1 of water being isolated. The top of the plastic tube projecting above the water surface was held open by a floating ring of styrofoam or rubber tubing permitting continual exposure of water in the chambers to the atmosphere. The simp1 icity of this design allowed as many as 6-10 chambers to be set up at one time. In 1978, all incubations used radioactive phosphate as a tracer of FRP fluxes. 20 pCi of sterile, carrier-free 32~0~(New England Nuclear) in 'I ml 0.01 HC1 were added to the water of each chamber at the beginning of the experiment. The water was stirred and an initial total activity sample of 10 ml was removed and placed in a scintillation vial. Additional samples were taken after careful stirring at 10-30 minute intervals over 1.5-4 hours, resulting in a total of 8-10 samples. Early in 1978, 32P samples were unfiltered. Later, because of the presence of algae or a floc associated with the forest floor, samples were filtered through 0.45 um membrane filters and the filtrate was saved for counting. The duration of the experiment and the frequency of sampling depended on ambient water temperatures (shorter duration and more frequent sampl ing at higher temperatures) and previous know1 edge of kinetics.

Several experimental treatments were tried in 1978 to assess the abiotic component of phosphorus fluxes. Formaldehyde was added to some chambers either in the evening or several hours prior to the beginning of the experiment in an amount sufficient to result in a 10% formalin solution. Sampling was as outlined above. From several chambers, all leaf 1itter was removed before the addition of 32~to examine the effects of the presence of litter on phosphorus fluxes. Finally, inorganic phosphate (as standard KH2P04 solution) was added to some chambers simul taneously with 32~to test the effects of enhanced phosphorus concentrations on fluxes. In all cases, replicate control (untreated) chambers were also monitored or had been sampled in different experimental series on the same day or within 24 hours.

During the period of each experiment, ambient water temperatures, weather conditions and floodplain elevations were recorded. Floodwaters were also sampled for phosphorus analysis. The depth of water in each chamber was measured so that volumes could be calculated. At the end of some experiments, ground litter in the chambers was harvested to determine whether fluxes were related to the amount of 1i tter present. Experimental measurements were made on 9 dates from March through June, 1978.

B. Field procedures: 1979

1. 32~kinetics

Further experiments on forest floorwater fluxes of phosphorus were carried out during the winter and spring of 1979. The chamber set-up remained the same (Fig. 288) but several different experimental manipul a- tions were tried. Data from 1978 demonstrated 1ittl e or no effects of floodplain elevation on phosphorus fluxes; therefore, careful attention was paid to the effects of temperature and phosphorus concentrations on phosphorus f 1uxes.

In 1979, phosphorus flux experiments with 32~followed the 1978 protocol with these modifications. A1 1 32~samples, except initial total activity samples, were filtered in the field through 25 mm Gelman glass fiber filters using Millipore Swinnex filter holders. Samples for determination of ambient FRP and FUP concentrations were taken from each chamber at the beginning and the end of each experiment. At the end of the 32~ experiments, 10 ml aliquots of water were removed from each chamber and submitted to isobutanol extraction (Kuenzler, et a1 . 1979) to remove FR~~Pto determine the amount of added 32~transformed to combined form ( FU~~P). In addition, at the completion of each experiment, chambers treated with 32P were sub 'ected to one of two treatments. To determine the distribution of 3d P on the forest floor in some chambers, a 3.8 cm diameter core tube was pushed into the sediments, isolating a portion of the floodwater and forest floor. Using a hand vacuum pump, all the water was pumped from the tube into a bottle to pick up the suspended and flocculant material associated with the forest floor. The volume was measured, the water was mixed and an a1 iquot was filtered through a 25 mm glass fiber filter. The filter was saved and placed in a scintillation vial to be counted. The core of ground litter was also saved for laboratory analysis of radioactivity. In chambers not subjected to this coring procedure, short-term regeneration of 32~from the forest floor was measured. The polyethylene tubes were removed from the PVC bases of these chambers, allowing free circulation of swamp floodwaters. After 5-10 minutes, the polyethylene tubes were replaced and sequential water samples were taken every half hour to measure changes in 32~in the water column. Because of time limitations, regeneration experiments lasted no longer than 2 hours. 2. FRP and FUP kinetics The effect of increasing FRP concentrations on fluxes of phosphorus between forest floor and floodwaters was studied on several dates in 1979 by adding increments of standard phosphate to a series of chambers. On 6 February 1979, 31POq (as KH2P04) was added in increments to 5 chambers which were separate from chambers in which 32~kinetics were being measured. After the addition of 31 PO^, 10-ml water samples were taken from each chamber every half hour for up to 3 hours for FRP analysis. Samples were filtered in the field through glass fiber fi 1 ters, and phosphorus reagent (Strickl and and Parsons, 1972) was added immediately. Following addition of reagent, samples were kept cool. On three other sampling dates, 20 PC$ of 32~04as well as increments of 31~04were added to a series of chambers. FRP and 32~samples were taken simultaneously and in the manner described previously. In addition, FRP samples were taken every half hour from a control chamber to which only S2p had been added. Initial and final samples for FUP analysis were also taken from each chamber to estimate FUP fluxes. Regeneration experiments were run on selected chambers with sampling every half hour for 32p04, FRP and FUP. Transformation of 32~into FU~~Pin the presence of enhanced FKP concentrations was measured using the isobutanol procedure out1 ined above.

3. The effects of biological inhibitors

A1 though formal in failed to show conclusive inhibition of biological influences on phosphorus fluxes in 1978, further experiments in 1979 attempted to differentiate biological uptake and release from chemical sorption and desorption processes. On 27 March, approximately 1 liter of 37% formaldehyde was added to a chamber, resulting in a 10% formalin solution. The bacterial inhibitors, penicillin G and streptomycin sulfate, were thoroughly stirred into the water of another chamber giving concentrations of 200 mg*l-l and 400 mg-1-1, respectively. To 4 other chambers, sodium arsenate (NaHAsOq) was added to result in 0.5, 2.0, 5.0 and 10.0 vM (74.9 vg ~s.1-1 = 1 vM) initial concentrations. These six chambers were preincubated for one hour, after which 20 fii of 32~and standard phosphate, sufficient to result in 10 initial concentration, were added to each chamber. In addition, arsenate and phosphate were added simul taneously to a separate chamber resulting in initial concentrations of 5 pM and 10 vM, respectively. Phosphate concentrations were increased over aibient so that changes in FRP concentrations over the period of the experiment could be easily measured. Samples for 32~and FRP analyses were removed at 0, 0.5, 1.0, 2.0, 3.0, and 4.0 hours. Control chambers were run simultaneously.

The Strickland and Parsons (1972) mixed phosphate reagent appeared to complex arsenate as readily as phosphate (Murphy and Riley 1962) ; no differentiation between these species was made on 27 March. On 11 May 1979, the effect of arsenate on phosphate fluxes was examined in a second set of experiments in which total initial As and P concentrations were held constant but in which the ratios of the concentrations of As and P were varied. The total initial As and P concentrations were 15 pH except for one chamber which had a total concentration of 30 pH. As:P ratios were as follows: !%:I,l3:l, 4:l, 1:1, 0.25:1, 0.075:1, and 0.01 6:l. A1 iquots of standard arsenate and phosphate solutions and 20 vCi of 32P04 were added simultaneously at the beginnin of the experiment. A control chamber received only 20 pCi of 32~0~.j2P was sampled at 0, 0.33, 0.67, 1.0, 1.5, and 2 hours. Samples for FRP and arsenate analysis (measured separately) were taken at 0, 1 .O, 1.5, and 2 hour intervals.

4. Other measurements

In addition to measurements of water temperature, site elevation, and ambient FRP and FUP concentrations during each experiment in 1979, ambient chl orophyl 1 concentrations were measured in floodplain waters in order to estimate a1 gal abundance. Along a transect immediately upstream of the experimental area, 10 samples 5 m apart were coll ected by pressing a 3.8 cm diameter core tube into the forest floor, and sucking the water and associated algae out of the tube. The volume was measured and a portion was placed in a clean polyethylene bottle for filtration and chlorophyll analysis. Samples for arsenate analysis were also taken from floodwaters. All samples for chemical analysis were placed on ice. C. Laboratory procedures 1. Work-up of field samples All filtered 32~samples were counted by measuring Cerenkov radiation (Haberer 1965) on a Packard Tri-Carb Model 3320 liquid scintillation counter. Filters (32~in floc) were counted as above after the addition of 10 ml of deionized water to each vial. The use of initial total activity samples made correction for color quenching (about 25% of Cerenkov) by the swamp water unnecessary. Ini tia1 total activi ty sampl es taken on 11 May 1979 contained large amounts of flocculant materials. Correction for quenching by these materials was made by counting known amounts of 32~added to filtered swamp water and the initial samples. In general, samples were counted twice and background and 1% standards were included to monitor the efficiency of the scintillation counter. In 1978, ambient water samples were analyzed for FRP and TFP on the autoanalyzer as described earlier. In 1979, all FRP and TFP analyses followed the methods of Strickland and Parsons (1972) and Menzel and Corwi n (1 965). Arsenate and phosphate concentrations were determined simul taneously fol lowing the protocol of Johnson (1971 ) . Chl orophyll s were analyzed according to Lorenzen (1967). Ground litter taken from chambers was dried at 80 C for 2-4 days, weighed, and subsampled for phosphorus analysis using the methods for 1i tterfa11 samples. Ground l itter cores (3.8 cm dia. ) taken from chambers in 1979 were dried, weighed and then digested with H2SO4 and Hz02 following the method of Lindmer and Harley (1942). The digested mixture was filtered to remove inorganic sediment and then both filter and filtrate (diluted to 10 ml) were counted on the 1 iquid scintillation counter. These samples were corrected for 32~decay. Adsorption of 32~by the polyethylene tubing used for the field chambers was measured in the laboratory by incubating 3 small pieces of tubing in 32~-spikedfiltered swamp water for 4 hours. After the incubation, the pieces of plastic were removed, shaken gently to remove water and placed in a vial with 10 in1 of deionized water for counting; initial aliquots of the spiked swamp water were also counted. After counting, each piece of plastic was air-dried and weighed, along with pieces having known areas, to estimate 32~adsorption per unit area of plastic. 2. Data analysis The flux of phosphorus estimated from sequential 32~Cerenkov counts was calculated using the model out1 ined by Koenings and Kuenzler (1979). 32~was removed from the floodwaters in an exponential fashion. The rate constant of removal , k, was calculated by:

where A, is the initial total activity of 32~,A is the 32~activity in the filtrate at time t, and t is the elapsed time. k32 was calculated by simple 1inear regression of in A with t. When remobal rates were high, only initial samples were used to estimate k32 . Three assumptions accom- pany this model : (1 ) the tracer is uniformly gixed with the stable phosphorus present and the tracer moves at the same rate as the stable species; (2) a steady state with respect to phosphorus kinetics exists between soluble and particulate-bound phosphorus; and (3) the rate of removal is proportional to the phosphorus concentration in the water.

In general, k (with units of t-') is defined as the fraction of FRP moving from the floodwaters to the forest floor per unit time. Biological and chemical sorption processes are included in the estimation of k. l/k is the turnover time of FRP in the water column. To obtain the gross removal rate of FRP from the water column (v3zp), k was mul tip1ied by the amount of FRP (in mg) present in the chamber at time 0 and this value was corrected to a unit area to obtain a rate having units of mg FRP*~-~*hr-1 . Phosphorus fluxes were calculated from time sequences of FRP measurements using the above model and this equation:

where Po is the FRP concentration at the start of the experiment and Pt is the concentration at time t. Turnover times and velocity were calculated as above. V31p is an estimation of the net removal of FRP from the swamp floodwaters to the forest floor. 32 The transformation of 32~into FU P was expressed on a mg*m'2=hr-1 basis using the following equation: ., m Fu3'p Flux (mg*m-2-hr-1 ) = cpm FIJJLp (f) Total FRP (i) cpm 32~(i) tx~ where i = initial, f = final, Total FRP = the total amount of FRP in the chamber in mg, t = duration of the experiment in hours, and A = the area of the forest floor (m2) isolated by the chamber.

A1 gal uptake experiments :

The contribution of the winter bloom of filamentous algae to phosphorus kinetics at the swamp floor-water interface was estimated by radiotracer experiments conducted in the field and the laboratory in March and April 1978. Algal samples for incubation were removed from the flooded swamp floor using a procedure similar to that for chlorophyll measurements except that a 5 cm diameter core tube was used. Samples were removed at 5 m intervals along a transect of a flooded portion of the floodplain; volumes were recorded and 50 m1 a1 iquots were placed into 125 ml Erlenmeyer flasks for incubation. The samples were diluted to 100 ml and 2-3 pCi 32~were added to each flask. Incubations were done either in the field with the flasks standing in swamp water on the floodplain or in the laboratory under constant temperature and light conditions. Sequential 10 ml 32~samples were removed from the flasks and fi1 tered through 0.45 pm membrane f il ters. The fi1 trate was saved for counting. Sampl ing intervals were very short initially (e.g., 0, 2, 5, 7, 10 minutes), but then lengthened over a period of 2-3 hours. Calculation of uptake rates followed the model for 32~kinetics given above.

Statistical analyses:

Statistical procedures were those of Stat stical Analysis Systems (Barr, et a1 . 1979). Correlations resulted in Pearson correlation co- efficieZs(r) with a level of significance of a = 0.05. Comparison of means used the Student's t-test with levels of significance of a = 0.10 or a = 0.05. Variability of data was expresse conservatively, as plus or minus one standard deviation from the mean x _+ s).

RESULTS

Wate'r chemistry and hydro1ogic fluxes

Hydrology and inundation patterns of Creeping Swamp:

Hydrologic fluxes in the Creeping Swamp watershed varied widely between Water Years (W.Y.) 1977 and 1978. The monthly distribution of rainfall showed no consistent patterns (Fig. 29A, 30A) but rather varied widely within and between years. During W.Y. 1977, 112 cm of precipitation fell, 10 cm below the long-term average of 122 cm (Sumsion 1970), while in W.Y. 1978, 135 cm of precipitation was measured, 13 cm greater than the annual average.

Total annual runoff from the Creeping Swamp watershed measured at the downstream station, CP-10, was nearly three times greater in W.Y. 1978 than in W.Y. 1977, a1 though annual precipitation was only 1.2 times greater (F?ig. ~9~30).This difference was a1 so ref1ected in the hydrologic budget where runoff was 21% of incoming precipitation in W.Y. 1977 and nearly 46% in W.Y. 1978. These values are quite different from, but bracket, the 37% value estimated by Winner and Simmons (7977). The relatively greater runoff in W.Y. 1978 may have been a result of a few intense rainstorms (Fig. 29C and 30C) or differences in the seasonal distribution of storms be tween years.

REEPING SWAMP, N.C. WATER YEAR 1978 I I I I I I I I I 1 I

CP- 10

ANNUAL TOTAL RUNOFF = 61.8cm

DATE

Figure- 30. Preci~itation,inundation and surface water discharge for creep\ng swamp during Water Year 1978. A. ~onthly-precipi tation measured at the Wilmar Fire Tower. B. Average monthly inundation pattern of Creeping Swamp floodplain estimated by paces of transects during high water levels. C. Daily discharge at CP-10 (U.S. Geological Survey, 1979). Open circles denote water quality sampling dates. In contrast to the monthly distribution of precipitation, daily discharge from CP-10 and floodplain .inundation showed consistent seasonal patterns (Fig. 295 and C, 30B and C). In both years, discharge was greatest during the cooler seasons and was characterized by sharp peaks during storm flow periods. There were four peaks or spate periods in W.Y. 1977 when discharge exceeded 4 m3osec-1 . In W .Y. 1978, there were eleven such periods, three of which exhibited extremely high discharge (219 m3~sec-1) . The greatest percentage of the floodplain (total area = 3.2 km2) was inundated during winter and early spring in both years (Fig. 295 and 3OB). The high per cent inundation in November 1977 was due primarily to a single heavy storm. When average inundation patterns for the five year period, W.Y. 1974-1978, were examined (not shown), November was usually a dry month. The entire floodplain was inundated >80% of the time during January 1978 due to frequent and heavy rain storms.

Comparison of monthly precipitation with monthly inundation (Fig. 29A and B, 30A and 0) shows the effects of watershed evapotranspiration during the warm season on the hydrologic budget. Precipitation remained variable but inundation decreased during the growing season to a minimum in July and August in both years. In general, inundation appeared to lag behind monthly precipitation patterns, resulting in a less variable pattern of inundation than would be predicted simply from precipitation alone.

For this study, the swamp floodplain was differentiated into four el evational regions relative to the water level stage measured at CP-14: (1 ) stream and pools (18-45 cm); (2) low floodplain (45-65 cm); (3) intermediate floodplain (65-75 cm); and (4) high floodplain (75-100 cm) (Table 19). The stream and pools category included the bed of the main channel, the area near CP-10 affected by the weir, and some very low floodplain downstream of CP-14. At CP-14, the stream became intermittent below a stage of about 20 cm and began to inundate the floodplain at 45 cm. Thus, the floodplain categories are representative of flowing water conditions. Using the relationship between water level at CP-10 and inundation of the floodplain derived by pacing transects of the floodplain and the daily water levels at CP-10, the volume of water in the swamp, the areas inundated and the overall average water depth for successive inundation of each floodplain region were estimated (Table 19). For each floodplain region, the percent of time that some portion was inundated during W.Y. 1977 and 1978 was also estimated (Table 19). The greatest error in these calculations probably exists in the estimations for the stream and pools region because of the damming effect of the weir at CP-10 when water levels were low. Relative to the low and intermediate floodplain, the high region was large in area and its inundation resulted in a large increase in total water volume (Table 19), even though it was inundated only a small percentage of the time. The stream and pools region was a1 so relatively 1arge in area, but when it alone was inundated, only a small volume of water was held in the swamp. Even during the relatively wet W.Y. 1978, the entire swamp was dry 24% of the time, primarily during the summer and fall. Table 19. Inundation patterns as different portions of the Creeping Swamp floodplain were sequentially flooded. Annual inundation periods are for Water Years 1977 and 1978.

- - - - Overall Stage Total Floodplain Total Average % heighth Volume of Average Inundated Time at CP-14 Water in Water Area Inundated Portion of Swamp (cm > (l06m3) Depth (cm) (106m2) 1977 1978 1 ) Stream and Po01 s 18-45 0-0.14 0-13 0-1.05 68 76 2) Low Floodplain 45-65 0.14-0.43 13-23 1.05-1.86 44 64 3) Intermediate Floodplain 65-75 0.43-0.64 23-28 1.86-2.26 25 41 4) High Floodplain 75-100 0.64-1.32 28-41 2.26-3.20 13 30

Phosphorus in bulk precipitation: The concentrations of phosphorus in bulk precipitation in the Creeping Swamp watershed were measured monthly in W.Y. 1977 and 1978. Phosphorus was expressed as reactive and total P because analyses were done on unfiltered samples. Annual weighted mean concentrations of reactive P were 0.042 mg*l-l for both water years (Tabl e 20 ). Concentrations of total P were slightly higher. Of the inputs of total phosphorus in bulk precipitation (60.5-79.7 mg*m-2*yr-l), 71 -78% were in the reactive phosphorus form (Tabl e 20 ). Table 20. Annual weighted mean concentrations and inputs of reactive phosphorus and total phosphorus in bulk precipi taticn during Water Years 1977 and 1978. Concentrations were weighted by volume of rainfall .

Reactive P Total P A. Annual weighted mean concentrations (mg*1-l ) Water Year 1977 .042 .054 Water Year 1978 .042 .059 B. Annual Fluxes (mgem'2* yr'l ) Water Year 1977 Water Year 1978 Surface water chemistry:

A. Annual average concentrations

Surface waters of the Creeping Swamp study area were typically low in concentrations of all forms of phosphorus, low in turbidity, conductivity, and pH, and were darkly colored (Tables 21 -24). The number of water samples for each year varied because stations were sampled only when the streams were flowing; in addition, storm event studies were conducted in W.Y. 1976-1978 but not in W.Y. 1975. In particular, in W.Y. 1975 most samples were taken during periods of declining discharge. This might partially account for the relatively low average phosphorus concentrations and low turbidities at the main stream stations during this period. Tributary sampling began in W.Y. 1976 with the exception of TB-09 and TB-10 which were sampled only in W.Y. 1978. Because of the intermittent nature of these small streams, most samples were taken during storm flow periods. Lack of agreement between total phosphorus concentrations and sums of the other forms is due to rounding errors.

During the study period, tributary TB-02 received waste from a hog farm until December 1976, These wastes markedly affected most of the parameters measured in the tributary waters and at the downstream station, CP-10, as we1 l (Tables 21-22). In particular, concentrations of filterable reactive phosphorus (FRP) and particulate phosphorus (PP) at TB-02 were elevated over unpolluted stream values (c.f., TB-04), and pH was increased, as were conductivity and turbidity. Water Year 1977, during which the hog farm ceased operating, was the period most strongly affected by the waste input. In W.Y. 1978, both CP-10 and TB-02 showed marked reductions in phosphorus concentrations and conductivity. Note that this year was also extremely wet.

In general, the sole tributary having a natural channel and floodplain, TB-03, had chemical characteristics more similar to the upstream main channel station, CP-20, than to any of the other streams (Table 23). In contrast to the other tributaries which were channelized or were field drainage ditches, turbidity and concentrations of phosphorus were consistently lower at TB-03.

Annual averages of conductivity remained relatively constant during the four year period at all stations except CP-10 and TB-02. The high variabil ity in conductivity, turbidity and color, as demonstrated by the standard deviation, was due to wide differences in these parameters seasonally and during storm flow periods. The high variability in turbidity was probably due to sharp increases in rising waters observed during storm flow periods. Average turbidities were lowest (5-9 JTU) in the unpolluted natural streams, CP-20 and TB-03 (Tables 21,23 ). The channelized tributaries exhibited the hi hest turbidities with annual averages ranging from 10-218 JTU (Tables 22,2 1). Color was also marked by high variability which could be accounted for primarily by seasonal variations (Kuenzl er, et a1 . 1977). The downstream station, CP-10, and the tributaries, 'ah hcc cn II "4 c 7J 7 n(D -'. 3w-' (PO* * dm

3 c.r II U3 r u ulcn

4 3ww II LD 1 +'UO NUN

3 w II U3 b-'w rum n -no M L. a-ha- %I- 5 aJ a-- 0 L UWV cn aJ 5 V) aJ L S'r Cd V) OL c ',--5x0 C, C, O-.r 53 C, LllW 5 s L s L 5-P aJC, s U La SaJOU 0-c- c UC,OO U U s s 50 "In aJ >5= E cn*L s-r-0 7 OUT 5 -r .r Q awn V) s 5 L 0 S+-'3T Q V)*a

TB-02 and TB-07, had the highest annual color values (105-177 Pt units) (Tables 20, 21, 23). Color generally increased from upstream to downstream. The high color at TB-02 may be partially due to filterable inorganic col loids which interfered with color determinations; this was visually observed on a sample with extremely high turbidity (2600 JTU, 18 April 1978). The high average color at TB-57 was probably truly indicative of natural conditions; the stream often looked orange. pH was moderately low at most stations and was relatively constant at a given station. The channelized tributaries, TB-01 and TB-07, had quite low pH's (3.7 - 4.6) (Tables 23, 24). Some of the drainage of these streams included pine plantations which may have contributed acidity.

Annual average concentrations of phosphorus fractions, estimated by weighting individual values by discharge, gave results skewed toward values measured during periods of high discharge and thus greater hydrologic fluxes. Annual weighted mean concentrations of all forms of phosphorus at the downstream station, CP-10, increased from W.Y. 1975-1977 and then decreased in W .Y. 1978 (Table 21). Of a1 1 forms, f i1 terabl e unreactive phosphorus (FUP) was lowest in concentration and remained relatively constant at CP-10 during the study period. Phosphorus concentrations at CP-20 (Table 21) were extremely low (0.002 - 0.025 mg I-?), about an order of magnitude lower than rainwater (Table 20), and showed little variation over the four year period. Concentrations of FUP and PP tended to be higher than FRP at CP-25.

FRP concentrations at TB-02 (Table 22) were three orders of magnitude higher than concentrations at CP-20 or TB-03 (Tables 21, 23) and between 2-3 orders of mangi tude greater than values measured at TB-04 (Table 22). TB-04 was similar to TB-02 hydrologically and also with respect to watershed characteristics. FRP was the predominant form of phosphorus at TB-02 during the period of hog farm discharge. Concentrations of FUP at TB-02 were only slightly elevated above those at other channelized tributaries. PP concentrations at TB-02 were much higher than those of the other streams for all years. The persistence of high PP concentrations at TB-02 in W.Y. 1978 when FRP showed a dramatic decrease suggests that these two phosphorus fractions were not similarly related to the hog farm operation. It is possible that, while there were no new sources of FRP to TB-02 in W.Y. 1978, the sediments carried by the stream waters were relatively enriched in phosphorus by earlier inputs when the farm was operating.

In the remaining channelized tributaries, TB-04, TB-01, TB-07, TB-09 and TB-10 (Tables 22-24), concentrations of all forms of phosphorus were similar to or slightly higher than concentrations in the unpolluted natural streams, CP-20 and TB-03. In particular, the concentrations of PP in unpol 7 uted channel ized tri bu taries were elevated over concentrations in the natural streams. B. Colloidal and dialyzable fractions of phosphorus The distribution of fil terable phosphorus (FRP, FUP) into do1 loidal (Co) and dialyzable (Di) forms was measured in stream waters at CP-10, CP-14 and CP-20 over a 14 month period from January 1977 through March 1978 (Table 25). On the average, FRP was 43%, 52% and 27% colloidal at CP-10, CP-14 and CP-20, respectively. Large standard deviations were associated with these average values; therefore, differences between stations were not significant. Overall, dialyzable fractions were larger than colloidal fractions. However, high variability was associated with these concentrations which were near the limit of detection for the analytical method used. More accurate analyses are needed for a precise determination of the distribution of filterable phosphorus into colloidal and dialyzable forms. No clear seasonal patterns of filterable, colloidal , and dialyzable forms of phosphorus were evident, a1 though total filterable forms tended to be higher in the early summer and late autumn. The lack of samples between June and November is due to the lack of flow during that period.

C. Seasonal patterns of surface water phosphorus concentrations The seasonal variations in concentrations of phosphorus fractions were examined at the mid-swamp station, CP-14 because it more closely approx- imated natural conditions than did CP-10 or CP-20 where weirs often caused pooling of the water; in addition, CP-14 was the site of intensive floodplain studies. Seasonally, FRP showed no consistent variations throughout the three year period and was the constituent lowest in concentration (Fig. 31). In contrast, FUP tended to be higher in summer and autumn than at any other time and was usually higher than FRP. PP concentrations were highly variable, but had a minimum in January or February of each year. Although runoff was highest in W.Y. 1978, PP concentrations were less variable and lower than in the other two years. In general, PP concentrations were similar to or greater than FUP concentrations (Fig. 31). Short-terfvariations in phosphorus concentrations and fluxes associated with storm events reported by Kuenzler, et al. (1977) and Yarbro (1979) are not shown, but are included in later cal- cul ations. - Surface water exports of phosphorus from the Creeping Swamp watershed: Annual exports of phosphorus from areas upstream of CP-20 and CP-10 (32 and 80 km2, respectively) were estimated by multiplying annual runoff values ( Fig. 29,30; Table 28) by annual weighted mean concentrations of phosphorus (Table 21) for Water Years 1975-1978. The fluxes calculated for W.Y. 1975 are probably low because relatively few storm dates were sampled in that year. Exports from the watershed upstream of CP-20 ranged from 2-1 5 mg ~~.m'~*~r-land were, with one exception (FRP, W.Y. 1978) lower than corresponding exports at CP-10 (Table 26). FRP was the smallest fraction

Table 25. Fractionation of filterable phosphorus into dialyzable and colloidal forms. Units are mg.1-1.

DATE FILTERABLE REACTIVE P FILTERABLE UNREACTIVE P FRP CoRP Di RP FUP COUP DiUP -- - Table 25 continued:

DATE FILTERABLE- REACTIVE P FILTERABLE UNREACTIVE P FRP CoRP Di RP FUP COUP QiUP

Table 26. Annual exports of phosphorus from the upstream watershed at Creeping Swamp stations CP-10 and CP-20 during Water Years 1975-1978. ------WATER ANNUAL 1 YEAR FRP FUP PP TP -2 -1 of total phosphorus export at CP-20 and ranged from 0.49-3.3 mg-m *yr . FUP exports were similar to or greater than FRP outputs and PP exports cons ti tuted the greatest fraction of total phosphorus export at CP-20. Exports in W.Y. 1978 were elevated over values in W.Y. 1975-1977; this is at least partially attributable to the larger volume of runoff in W.Y. 1978 (Fig. 29, 30).

Total phosphorus exports at CP-10 ranged from 9-29 mgem'2*yr-1 (Table 26) and were greatest in W.Y. 1977 when the hog farm had its greatest influence. With the exception of W .Y. 1978, FUP exports were the smallest fraction of outputs at CP-10 (Table 26). FUP exports increased by a factor of 2.8 from W.Y. 1977 to W.Y. 1978 while FRP exports dropped by nearly a factor of 5 and were smaller than FRP exports at CP-20. PP exports remained similar between these two water years. Thus, during the period of influence of the hog farm, FRP exports were the 1arges t component of phosphorus f 1uxes at CP-10. In W.Y. 1978, when the hog farm was no longer in operation and when runoff was high, FRP exports were small compared to the similar FUP and PP exports.

Ground water phosphorus concentrations and losses from the watershed:

Concentrations of FRP and FUP in water taken rom wells in the CP-14 study area were quite low (0.002-0.009 rng*l-') (Table 27A) and in the range of stream water concentrations (Tab1 es 21 -24). Conductivities, on the other hand, were higher than average conductivities at either of the stream stations, CP-10 or CP-20. The deeper wells located in or near the stream channel (Wells 4 and 5) had higher conductivities than did the shallow wells located in the floodplain (Wells 6-9).

Winner and Simmons (1977) estimated that 2% of the water entering Creeping Swamp watershed was lost from the watershed as deep ground water outflow. Using this value, precipitation data for W.Y. 1977 and 1978, and average concentrations in the deeper wells near CP-14 (We1 1s 4 and 5), the flux of phosphorus in deep ground water outflow was estimated (Table 27B). These fluxes were about 1% of the streamwater exports at CP-10 (Table 26) and thus represented a negligible loss of phosphorus from the watershed.

Phosphorus cycling in the swamp floodplain ecosystem

The cycl ing of phosphorus in the swamp floodplain ecosystem was studied by quantifying the inputs and outputs of the ecosystem, the standing stocks of phosphorus in both the living and non-living portions of the ecosystem and the fluxes of phosghorus within the ecosystem. The ecosystem was defined as the 3.2 kmL floodplain swamp located in the Creeping Swarip study area (Fig. 2). The vertical limits of the ecosystem were the tops of the forest trees and a depth of 25 cm in the soi 1. The inputs of phos~horusconsisted of bulk precipitation (Ta.::;ie 20) and CP-23 and tributary sgrface water inputs. Outputs occurred in surface water exports at CP-10 and in deep ground water losses. Table 27. A. Average conductivities and concentrations of filterable reactive phosphorus (FRP) and fi1 terable unreactive phosphorus (FUP) in water taken from wells in the Creeping Swamp study area. Standard deviation6 are in parentheses. B. Estimated deep ground water losses of phosphorus from the Creeping Swamp watershed.

A. Conductivity and Phosphorus Concentrations

Fl" 1 terabl e Fi1 terabl e Conducti vi ty Reactive Phosphorus Unreactive Phosphorus --Well n (umho *cm' ' ) (rng.1-') (ingel-')

B. Deep Ground Water Losses (mgem'2*yr-1) -1977 1978 Volume (cm)* 2.2 2.3 . Fi1 terable Reactive Phosphorus

Fi1 terabl e Unreacti ve Phosphorus

Total Soluble Phosphorus 0.27 0.32

"Estimated from Winner and Simmons (1977). Annual surface water imports and exports of phosphorus:

Annual fluxes of phosphorus in streamwaters flowing into and from the 3.2 km2 floodplain study area of Creeping Swamp (Fig. 2) were estimated by mu1 tiplying annual runoff values (Tab1 e 28) by annual weighted mean concentrations of phosphorus (Tables 21-24) for those water years where data were available. All fluxes were expressed on a mg*m-2 basis, which implies uniformity of processes within the ecosystem. This, however, was certainly not the case in view of the spatial heterogeneity of tributary inputs (Fig. 2), the elevation gradient within the floodplain and the accompanying species gradients, and other variability within the swamp.

A. A water budget for the swamp floodplain

An estimate of the mass balance of water entering and leaving the swamp ecosystem showed that during W.Y. 1977 and 1978 more water left as surface runoff at CP-10 and as evapotranspiration than entered as tributary runoff and precipitation (Table 28). Unmeasured runoff from up1 ands (1 1.8 km2) surrounding the floodpl ain within the Creeping Swamp study area (Fig. 2) was estimated by mu1 tiplying the annual runoff (cm) measured at CP-10 during W.Y. 1977 and 1978 (Fig.29,30) by the area of these up1 ands. Evapotranspiration in the swamp was estimated using the fraction of total annual precipitation for the whole watershed for each water year which did not leave CP-10 as runoff (Fig.29,30) and multiplying it times the volume of precipitation entering the floodplain. Annual deep ground water losses, which Winner and Simmons (1977) estimated to be quite small (2% of annual precipitation) were not included in the budget.

Two sources of error may account for the discrepancy between inputs and outputs. One is the uncertainty of the tributary runoff estimations as described in Methods. The area of the watershed drained by the tributaries was 33 km2. Using the annual runoff measured at CP-10 for each water year (Fig. 29,3O) and mu1 tiplying it by the tributary-drained area results in a separate tributary runoff estimate of 7.9 and 20.4 x 106 m3eyr-1 for W.Y. 1977 and 1978, respectively. These estimates are greater than the f low-regression estimates by factors of 1.04-1.5 and 1.2-1.4 for W.Y. 1977 and 1978, respectively. Use of the tributary-drained area estimates of runoff in the water budget results in total inputs (Corrected Total, Table 28) which were 96% of total outputs for both water years.

A second source of error in the swamp water budget may be from unmeasured ground water inputs to the floodplain. Winner and Simmons (1977) concluded that the portion of the swamp floodplain between CP-20 and CP-10 was an area where water from the Castle Hayne aquifer was discharged into the swamp. This discharge may account for the difference in inputs and outputs after all runoff corrections have been made. Table 28. A water budget for the Creeping Swamp floodplain. Runoff estimates for CP-10 and CP-20 were obtained from North Carolina Water Resources Data (U.S. Geological Survey, 1976-1 979) and were corrected for our estimate of upstream watershed area. 9- --- - 1 Water Flux (lo6 m3*yr ) Station or Area 1976 1977 1978

A. Inputs CP- 20 6.61 6.80 17.6

Tributaries: TB-07 0.38-1 .32 0.39-1.33 1.01-1.95 TB-02 1.87 1 .91 3.97 TB-03 1.07 1.10 3.23 TB-04 1.66 1.69 3.84 TB-07 0.13-1.52 0.13-1.53 0.35-1.74 TB-09 1.06

TOTAL Tributaries: Precipitation

Unmeasured Runoff

Total InpG Corrected Total (see text)

B. Outputs CP-10 20.7 Evapotranspiration

Total Outputs B. Phosphorus imports and exports The primary surface water sources of phosphorus to the floodplain were those imported at CP-20, TB-02, TB-04 and TB-09 (Table 29). A1 though runoff at CP-20 was the largest, the channelized tributaries TB-02, TB-04 and TB-09 were important because of their higher phosphorus concentrations. At CP-20, PP was the largest component and FRP the smallest component of total phosphorus imports during all four water years. Imports of a1 1 fractions of phosphorus at CP-20 increased sharply in W.Y. 1978. Phosphorus transported by TB-02 was the most significant surface water source to the floodplain (Table 29C). In W.Y. 1976 and 1977, imports of FRP represented the largest component of phosphorus transport by TB-02. In W.Y. 1978, FRP imports were greatly reduced over previous years but PP imports increased. During all three years, FUP imports were a small portion of the total at TB-02. A gross estimation of the potential import at TB-02 if it had not been polluted was made using the runoff at TB-02 (Table 28) mu1 tip1 ied by the annual weighted mean concentrations of phosphorus at TB-04 (Table 22). These two streams were similar hydrological ly and with respect to watershed characteristics. The "clean water" contributions were often an order of magnitude smaller than actual imports. Imports of phosphorus by TB-04 were similar to those of CP-20 (Table 29, E and A). PP was the most important component. Imports of phosphorus at TB-03 were smaller than those at CP-20 and PP imports were slightly larger than FRP or FUP imports. Phosphorus imports from the other tributaries were similar to or smaller than fluxes from TB-03 (Table 29). In all cases, PP was the most abundant form of phosphorus being transported Imports at TB-09 and TB-10, which were ditches draining crop lands, were large re1 ative to the other unpol luted tributaries , particularly wi th respect to PP. The discrepancies observed in the water budget (Table 28) result in an underestimation of phosphorus imports to the floodplain unless some correction is made. Phosphorus in runoff from uplands adjacent to the swamp ecosystem and in the difference between tributary-drained watershed estimates and flow-regression estimates of annual tributary runoff was calculated by summing the two runoff values (3.1 and 10.9 X lo6 m3-yr-1, respectively for W.Y. 1977 and 1978) and then mu1 tiplying these sums by the average of the annual weighted mean phosphorus concentrations in all tributaries except TB-02. The resulting annual imports of the four fractions of phosphorus are large compared to all other imports but those at TB-02 (Table 291). Floodplain-based exports of total phosphorus at CP-10 were large (216-734 mg*m-2.yr-l) (Table 29) but were always smaller than TP imports at TB-02 alone. With the exception of W.Y. 1978, FUP was the smallest component of surface water exports at CP-10. Total phosphorus exports 2 -1 Table 29. Annual surface water imports and exports (mg-a -yr ) from the Creeping Swamp floodplain. See text for explanation of "clean water" TB-02 imports. - WATER ANNUAL IMPORTS YEAR --FRP FU P PP TP CP-20 4.90 4.13 6.37 32.9 TB-01 0.59-2.1 2.2 -9.2 2.5 -4.9 TB-02 41 7 573 21 5 TB-02: "clean water" 6.43 11.9 9.92 TB-03 0.67 1.38 3.02 TB-04 5.70 10.6 9.61 TB-07 0.25-2.8 0.25-2.9 1.6 -8.1 TB-09 . 6.28 TB-10 1.57 Other Imports (see text for calculations) 12.8 6.89 34.0 30.6 CP-10 90.4 62.6 239 58.2 382 71.7 77.2 201 were greatest in W.Y. 1977 when FRP exports were greatest. In W.Y. 1978, FRP surface water exports at CP-10 dropped dramatically, FUP increased by nearly 3 times and PP exports dropped sl ightly. Overall exports at CP-10 were 75% and 46% of the total surface water imports, including "other imports", in !d.Y. 1977 and 1978, respectively. The contribution in bul k precipitation (60-80 mgd m-2.yr-l ) (Tabl e 20) and ground water inflow represents an additional source of phosphorus to the floodplain which was not included in the percentages given above.

Standing stocks of phosphorus in the floodplain ecosystem: A. Vegetation Phosphorus in above-ground biomass of trees, saplings, herbs, shrubs and vines (Table 30) was estimated from biomass data from Mulholland (1979). Assuming that the phosphorus content of wood was 0.01-0.02% (Duvigneaud and Denaeyer-de-Smet 1970; Cromack and Monk 1975; Likens and Bormann 1970), an estimated 3000-5500 rng ~~m-~was present in the branch s and stems of the Creeping Swamp canopy. Approximately 1100-1 300 mg P*m- h was present in actively growing leaves, assuming 0.1-0.2% phosphorus in 1eaves (Cromack and Monk 1975; Likens and Bormann 1970). The phosphorus standing crop estimate in leaves is nearly three times the amount of phosphorus in 1itterfall and may be considerably biased even when resorp- tion of phosphorus by trees prior to abscission is considered. The annual above-ground wood increment of phosphorus, 59-1 17 mg P- m-2, is based on measurements of Mu1 hol land (1 979) and 1 i terature estimates of wood phosphorus content (see above). The annual wood increment of phosphorus was about 2% of the standing crop in trees and branches. Approximately 34 rng am-' was present in herbs, shrubs and vines (Table 30). The amount of phosphorus in bryophytes, 8.4 mg P mm2 was less than 0.2% of the standing crop of phosphorus in the tree canopy, based on phosphorus analyses of harvested materials. B. Ground 1i tter The distribution of ground litter on the forest floor of the swamp floodplain was highly variabl e, particularly fol 1 owing s torm-i nduced inundation. Seasonal trends in the standing crop of ground litter were not clearly evident because of the large variability in distribution across the floodplain (large error bars in Fig. 32B). The consistent differences in standing crop data between Mu1 holland (1979) and this study may be due to different sampling procedures. Mulholland used a small quadrat (0.02 m2) with 30 sampling locations. I used a larger quadrat (0.25 m2) and 15 sampling locations. Phosphorus present in ground 1i tter appeared to peak in 1ate au tumn-early winter and decl i ne thereafter (Fig. 32A). Some of the variabil i ty associated with phosphorus in ground litter may result from phosphorus in materials sedimented onto the 1 eaves. On the avera e, 536 mg P. m-2 were present in ground 1 i tter in W.Y. 1977 and 368 mg P-m-g were present in W.Y. 1978 (Tabl e 31A). The difference between years is not statistically significant. Table 30. Biomass and phosphorus content of above-ground vegetation in Creeping Swamp. - BIOMASS* PHOSPHORUS (g dry wt~m-~) (mg*m-2)

A. Trees and Saplings Leaves 750 1100-1 300 Branches and Stems Total 27600 41 00-6800 Annual above-ground wood increment

B. Herbs, Shrubs, Vines 3 9 34 C. Bryophy tes 6.7 8.4

- ,, * from Mu1 hol land (1979) with the exception of bryophyte data.

C. Soil Phosphorus in organic and avail able forms (soil phosphorus) in the upper 25 cm of soil was estimated twice: on 28 Apri 1 1977 and 30 May 1978. Samples were split into surface (0-5 cm) and subsurface (6-25 cm) components. Phosphorus and organic matter varied widely across the transect of the floodplain and were highest at the lower elevations (Stations 6-9, 12) in both surface and subsurface layers (Fig. 33). Location of the sample plots on the transect across the floodplain was not identical for both years and determination of floodplain elevation in the sample plots was impossible because the floodplain was dry. Thus, there is some variability between years in the correlation of soil phosphorus with elevation. Soil phosphorus and organic matter were more concentrated in the surface layers than in the lower layers. However, statistical correlations between organic matter and phosphorus were weak. In the top 5 cm, soil phosphorus averaged 8700-8900 mg while in the subsurface layer, soi 1 P varied from 23,000-26,000 mg P-m-2 (Table SIB). The total amount of phosphorus in the soil column was 32,000- 35,000 mg pornm2, nearly 6 times the total amount of phosphorus in the living vegetation of the swamp floodplain and 75 times the amount present in ground litter.

Fluxes of phosphorus within the swamp floodplain ecosystem: Transfers of phosphorus within the floodplain ecosystem were measured in throughfall , stemflow, 1i tter- and branch-fa1 1 , sedimentation of water- borne particulates and in exchanges between the swamp forest floor and CREEPING SWAMe N.C. CP-14

W.Y. 1977 W.Y. 1978 DATE

Figure 32. Seasonal patterns of phosphorus and biomass in ground litter in Water Years 1977 and 1978, A. Phosphorus standing crop. Closed squares are estimates by the author. Open squares are estimates based on measurements by P. Mulholland. Error bars are one standard deviation. No standard deviations were available for Mulholland's data. 8. Dry weight of litter represented by solid symbols. Open symbols are ash-free dry weights. Solid lines are estimates by the author. Dashed 1 ines are estimates by P. Mu1 holland. CREEPING SWAMP N.C. CP- Id I I I I I I I PHOSPHORUS f 0-5cm 6-25cr-r

ORGANIC MATTER 0-5cm 6-25cm 1977 1p 0 1978 A A

2 4 6 8 10 12 I4 STATIONS ACROSS TRANSECT

Figure 33. Soil organic and available phosphorus and organic matter along a transect of the swamp floodplain on 28 April 1977 (dashed lines) and on 30 May 1978 (solid lines). floodwaters. Forest floor-water exchanges were the subject of intensive study and will be examined in a separate section.

Table 31. Phosphorus in ground litter and soil in the floodplain of Creeping Swamp.

BIOMASS PHOSPHORUS

A. Ground Litter (n = 7):

ORGANIC MATTER PHOSPHORUS (g AFDW~-~)(ilg P/g dry wt. soil) (mg 1977 1978 1977 1978 1977 1978- B. Soil (n=15) 0-5 cm: x 4980 4410 298 329 8660 8850

6-25 cm: x 12800 10800 152 123 26400 23300

0-25 cm: j;, 17800 15200 177 151 35100 32200 s 4430 3770 112 101 12200 13100

. . A. Throughfall and stemflow

Concentrations of reactive phosphorus (RP) and total phosphorus (TP) in throughfall varied widely over the two-year period (Fig. 34). Phos- phorus concentrations were very high in April of both years, coinciding with the period of forest pollen production. With the exception of these two April dates, phosphorus concentrations varied inversely with volume of throughfall. Concentrations of RP and TP in net throughfall (Through- fa1 1 concentration - precipitation concentration) (not shown) in the growing season (1 April - 31 October) were significantly higher (w = 0.05) than concentrations in the dormant season (1 November - 31 March). During the dormant season, average net throughfall concentrations were 0.012 mg ~~01-1(s.d. = 0.018) and 0.018 mg TP-1-1 (s.d. = 0.030). Average growing season concentrations were 0.091 mg ~~01'~(s.d. = 0.079) and 0.120 mg TP-1-1 (s.d. = 0.114). Figure 34. Seasonal patterns of throughfall volumes and phosphorus concentrations in Creeping Swamp during Water Years 1977 and 197%. RP: Reactive Phosphorus; TP: Total Phosphorus. Annual net fluxes of phosphorus from the forest canopy to the forest floor in throughfall (Table 32A) were roughly similar to phosphorus fluxes in bulk precipitation (Table 20) and were much higher in W.Y. 1978 than in W.Y. 1977. The increase in throughfall flux between years resulted from a greater volume and higher average weighted mean concentrations of phosphorus in W.Y. 1978.

Phosphorus concentrations in stemflow were measured 4 times in W.Y. 1977 and 5 times in W.Y. 1978. Concentrations in stemflow from different tree species varied wid ly; stemflow from -Acer rubrum was usually high in phosphorus (>0.10 mg-1-$). Accurate measurements of stemflow volumes were difficult; therefore annual average stemflow concentrations were arithmetic rather than weighted means (Table 32B). Concentrations of both forms of phosphorus in net stemflow were low compared to net throughfall concentrations (Tab1 e 32A). Fluxes of phosphorus in stemflow were calculated using net annual average concentrations measured in Creeping Swamp and the annual volume estimates of Helvey and Patric (1965) for hardwood forests of the eastern United States. Annu 1 net fluxes of phosphorus in stemflow varied between 0.68 - 1.12 mg RP*m-3*yr-1 and 2.18 - 2.63 mg ~~*m-~*~r-l for W.Y. 1977 and 1978 (Table 32B) and were 30 - 100 times smaller than net throughfall fluxes.

5. Litter and branch fall

Phosphorus in daily litterfall peaked sharply in October and November during W.Y, 1976 and 1977 and showed small increases in the spring, presumably because of flowering of the trees (Fig. 35A). Litterfall biomass had simi lar autumn peaks but no spring increases (Fig. 35B). Overall , l itterfall amounted to 613 g dry ~ei~htam-~in W.Y. 1976 and 589 g dry weightem-2 in C.J.Y. 1977. Total fluxes of phosphorus for these two years were 345 and 313 mg ~em-2*~r-l, respectively (Fig. 35). These fluxes are approximately 3-5 times greater than annual fluxes in throughfall and stemflow.

The fall of branches greater than 0.3 m in 1 ngth amounted to 121 g dry weightem-2 (Mulholland 1979) and 22.2 mg P-m-' in W.Y. 1977. This estimate did not include the fall of tree stems. Branch fall was 20% of litterfall on a dry weight basis but was only 14% of the phosphorus flux in 1i tterfall .

C. Sedimentation of particulate materials

Sedimentation of water-borne particulate materials was measured during the flooded season in W.Y. 1978. Phosphorus content of sedimented materials averaged about 0.013%; ash content was usually greater than 70% of dry weight. Sedimentation of phosphorus was relatively low during December- February and then increased slightly in March and April (Fig. 36). Tillage of agricultural fields began in late February and early March and greater sediment mobility as a result of this process may have influenced the amount of sedimentation observed. On the floodplain , sedimentation in terms of dry weight and phosphorus was greatest in the low areas and Table 32. Volumes, annual mean concentrations and fluxes of phosphorus in throughfall and stem flow in the Creeping Swamp floodplain

during Water Years 1977 and 1978. 3

1977 1978

A. Throughfall % of precipitation 90 % 86 % Vol ume 101.2 cm 116.4 cm

Total Throughfall wt'd mean (mg*l-l) 0.074 0.110 0.114 0.169 Net Throughfall wt'd mean (mg.1-l) 0.032 0.056 0.072 0.110 Net Flux (mg.m'2 oyr-l) 32.4 56.7 83.8 128

B. Stemflow 19-77 1978 % of precipitation* 5 % 5 % Vol ume 5.60 cm 6.75 cm

Total stemflow concentration - (mg*1") x 0.062 0.093 0.052 0.098 s 0.032 0.041 0.041 0.087 n 4 4 5 5 Net Stemfl ow concentration (mg. 1-l) 0.020 0,039 0.010 0 .039 Net Flux (mg.m-2 .yr'l) 1.12 2.18 0.68 2.63 *Helvey and Patric (1965)

decreased with increasing elevation (Table 33A). Overall, an average of 135 g.m-2 of sediment was de osited on the floodplain; this amounted to a phosphorus flux of 172 mgom' 9 .

Table 33. Sedimentation of particulate phosphorus onto the floodplain from floodwaters for the period 6 Dec 1977 to 30 May 1978.

A. Sedimentation at Different Floodplain Elevations

ELEVATION DRY NEIGHT PHOSPHORUS g ernm2 mg em-'

Low 305

Intermediate

High

Wt'd Average

B. Comparison of Sedimentation with Annual Particulate Phosphorus Fluxes

FLUXES PHOSPHORUS (kg)

Inputs 1840

Outputs

Sedimentation

Sedimentation represented a significant sink for particulate phosphorus entering the swamp (30% of inputs), but it was approximately half of the difference between inputs and outputs (Table 338). This disparity might be partially explained by the evidence cited earlier which suggested that particulate inputs from TB-02 may have been enriched. TB-02 was downstream of the site of sedimentation measurements and thus inputs from this polluted stream were not included in the estimate of sedimentation. Forest floor - floodwater exchanges of phosphorus The rates and mechanisms control ling the sorption and release of filterable forms of phosphorus at the interface between forest floor and floodwaters were subjected to intensive study. Over a three year period, several methods for estimating FRP and FUP fluxes were tried. All experiments were conducted at inundated floodplain sites at CP-14, mostly during the winter and early spring of W.Y. 1977-1979. Fluxes of phosphorus from the water column to the forest floor were expressed as negative rates; releases from the forest floor to the overlying floodwaters were expressed as positive rates. FRP exchanges:

In W.Y. 1977, estimations of fluxes were made by measuring changes in concentrations of FRP and FUP in the water fo lowing the addition of small amounts of FRP (to reach 0.015-0.020 mg.1- 1 ). In W.Y. 1978 and 1979 radiophosphorus-32 (32~)was added to chambers and the disappearance of 32~from the water column was monitored. In addition, in W.Y. 1979, I measured changes in concentrations of FRP and FUP in control chambers and in chambers to which incremental amounts of FRP had been added.

Average rates of the movement of FRP between the water column and the swamp forest floor ranged from 0.035 to -0.450 mgem-2* hr-1 over the three year period (Table 34). The average rate in W.Y. 1977 is higher than rates for the other years (but not significantly); this may represent the effect of small FRP additions. Annual average FRP fluxes calculated from 32~estimations in W.Y. 1978 and 1979 (Table 34) were significantly different (0.01 < P < 0.02) from each other but there are no apparent reasons for this difference. In W.Y. 1979, rates estimated by the two methods were significantly different as well (0.001 < P < 0.01). In fact, the average flux of FRP based on 31P measurements, a1 though small, moved from the forest floor to the floodwaters. At the low ambient concentrations of FRP in the swamp waters (<0.005 mg.1-I), estimations of changes in concentrations were subject to bigh variabilit . However, on all dates, comparison of 32~-estimatedrates with FR 3Y P-estimated rates in control chambers showed consistently that FR~~Prates were a small fraction of the 32~rates and that they were always close to zero. Turnover times of FRP in the water column, based on 32~estimations, were on t e average 3-4 hours in W.Y. 1978 and 1979 (Table 34). For those FR 3? P estimations where the flux was positive (i .e., from the forest floor), turnover times were about 450 hours. Such large differences between the two methods indicate that estimations using 3% included more than net fluxes of FRP, just as has been shown for algal cultures (Nalewajko and Lean 1978) and for phytoplankton populations (Kuenzler, --et al. 1979). Rates of FRP flux were highly variable between sampling days and from year to year. Therefore, the effect of several measured independent # n AN CON uu 0 e- '9 "9 00 I I

3: 3: -h WCO .--.I d ww cn Ln CO h . L 0 0 I I variables was investigated. Ambient (FRP < 0.015 mg*l-') fluxes of FRP, based on 32~and FR~~Pestimates, were correlated with ambient water temperature, elevation of the study site on the floodplain, the sequential date (1 January = date 1) of the experiment, ambient chlorophyll -a concentrations in the water (W.Y. 1979 only) and initial FRP concentrations in the chamber (Table 34). Because uptake by the forest floor was expressed as a negative rate, negative correlations of FRP fluxes would be expected with independent variables which had a stimulatory effect on the rates. Sequential date, which represents the time from early winter through late spring, incorporates the factors of increasing temperatures, longer daylight and probably greater chlorophyll-a concentrations on the floodplain.

In W.Y. 1977, ambient FRP fluxes were significantly and negatively correlated with temperature, sequential date and FRP concentration. Even under so-called "ambient" levels of FRP, small increments of phosphorus had been added to these chambers in order to estimate uptake rates. Thus, phosphorus added at any level had a stimulatory effect on the uptake rate by the forest floor. DespSte a wide range of floodplain elevations, little relationship was found between elevation and ambient FRP fluxes in W.Y. 1977.

In W.Y. 1978, ambient FRP fluxes, estimated using 32~changes, were significantly related to a1 1 the independent variables except temperature, but correlated most strongly with initial FRP concentrations (Table 34). Inclusion of data from chambers to which FRP had been added reduced the correlation coefficient between uptake and FRP concentration, but it remained significant. A1 though chl orophyl l-a measurements were not made in W.Y. 1978, a filamentous algae bloom developed during the flooded season and visually appeared to peak in biomass in mid-to-late March. If the algae associated with the forest floor were strongly influencing FRP fluxes, then the significant correl ation of FRP uptake with sequential date may have been due to the development of the algae bloom. The lack of a significant correlation with temperature is not readily explained. However, the excel lent correlation with elevation may have been due to higher algal standing crops in the lower, more frequently inundated portions of the floodplain.

Correlations of ambient FRP fluxes (32~estimates) with independent variables were much weaker in W.Y. 1979 and were insignificant with the exception of the correlation with chlorophyll-a (Table 34). This may have partially been due to the smaller number of experiments. FRP fluxes (32~estimates) including data from chambers which received FRP additions correlated significantly with initial FRP concentrations. FR~~P-based ambient FRP fluxes in W.Y. 1979 did not correlate significantly with any independent variable. Inclusion of data from chambers having FRP additions resulted in a significant correlation of FRP fluxes with initial FRP concentrations. Fluxes of FRP from the floodwaters to the swamp forest floor generally responded most strongly to changes in initial FRP concentrations. Although the data were insufficient, it appeared that the magnitude of FRP fluxes were also related to chlorophyll-a concentrations and thus probably to the algal populations on the flooded forest floor. In W.Y. 1979, careful investigation was made of the relationship of FRP fluxes to FRP concentrations by measuring changes in both 32~and FR~~Pfollowing the addition of different amounts of FRP to the chambers (Fig. 37). Regardless of the means of estimation, phosphorus fluxes inxeased with increasing initial FRP concentrations. Estimations based on FR~~Pchanges were usually slightly lower than rated based on 32~kinetics. Fluxes at ambient FRP concentrations (<0.015 mggl-1) were very small compared to fluxes at higher concentrations. Only on 27 March did fluxes appear to level off at the highest initial concentration (Fig. 37). Fluxes at a given FRP concentration appeared to increase through the flooded season. Over the period of measurement, water temperatures increased, day length increased and chlorophyll-a concentrations increased. In particular, from 28 February to 12 March, chlorophyll-a conc ntrations on the floodplain nearly t~ipled, going from 0.058 to 0.153 mgsrn-'; they then decreased to 0.129 mg-m- on 27 March.

The possible interaction of temperature and FRP concentrations on forest floor FRP uptake was investigated by plotting uptake rates measured during all three years as a function of both temperature and FRP concentration (Fig. 38). It is apparent that there was an interactive effect at relatively high concentrations and temperatures. At low (ambient) FRP concentrations, FRP fluxes were very small regardless of temperature.

FUP exchanges:

Fluxes of f i1 terable unreactive phosphorus (FUP) at the forest floor-water interface were estimated by measuring changes in FUP concentrations in chambers in W.Y. 1977 and 1979. On the average, FUP fluxes were positive (Table 34), moving from the forest floor to the water, and were, on an absolute scale, smaller than FRP fluxes. These fluxes did not correlate well with any independent variable. In W.Y. 1979, the rates of transformation of FR~~Pto FU~~P during the period of the experiment ranged from 0.013 to 0.050 mg=m'2*hr-l (Table 35) and were from 2.6 to 7.6% of the FR~~Pflux. The presence of FU~~Pin the chambers introduced this amount of error in the FR~~Pflux estimations because no differentiation of the forms of filterable 32~was made during scintillation counting.

Fate of 32~removed from the water:

The distribution of the fraction of 32~which disappeared from the water column of the chambers was examined by measuring the activity of forest floor litter and associated flocculant materials and algae. In addition, the possible sorption of 32~by the plastic walls of the chambers was investigated and found to constitute about 8% of the counts that QUc, S? >)F- aJ c, a u-r MS> -s 0.r b aJ U-v WE mU - L- aJM a v, IIWW S X LM L aJ 0 ss T aJ *r.r IIS m-v v, In 0 aJW II s DO, n.t- r s rg 5 aJ CnTT >sou .r 0 Q'r m13, U-v s s la 5 *I=- aJ L m v, LC, 3 3 Figure 38. The interactive effects of initial FRP concentrations and ambient water temperatures on the uptake of FRP by the swamp forest floor. r --W.Y. 1977 estimates; A--W.Y. 1978 estimates; m--W. Y. 1979 32~estimates; --W.Y. 1979 FR~~Pestimates. disappeared from the water column. This error was not compensated for in the estimations of FRP fluxes that used chan es in 32~activity. On five sampl ing dates in W.Y. 1979, 22-60% of the 3yP lost from the water during the period of the experiment was in the litter and 3.1-11% was in the flocculant material and a1 gae (Table 35). High variabi 1 ity was associated with these averag values. The amount of 32~in the litter did not correlate we1 1 (r5 < 0.2) with the amount of 1itter present or with the phosphorus content of the litter on any date. The fate of the 32~not accounted for (40-70%) in the FUP, litter, and floc, less 8%, is unknown. Some, however, may have moved into the soil.

Table 35. Transformation of FR 32P into FU "P and distribution of 32P removed from the water column during floor-\..rater ex- cnange experiments at ambient FRP concentrations in 1979.

Sampl ing % 32P in Date 1 itter

2-28-79 +0.014 (0.007) 60 (17) 5.2 (1.7)

3-12-79 t0.012 (0.002) 43 (21) 3.1 (2.6)

3-27-79 t0.014 (0.003) 36 * 8.5 * * 2 samples

Biotic and abiotic components of phosphorus fluxes:

Experiments in W.Y. 1979 were designed to investigate the factors control 1 ing fluxes at the forest floor-water interface, especial ly with respect to biotic and abiotic components of fluxes. Models of abiotic and biotic controls of phosphorus fluxes were developed and the assumptions and predicted outcomes were tested by examining prel iminary observations and in a series of experiments.

A. Prediction of the outcome of biotic versus abiotic control of f 1 uxes

If exchanges of FRP at the forest floor-water interface were abio- tical ly controlled, then the response of uptake as a function of FRP concentration could result in a curve in the shape of rectangular hyperbola as predicted by the Langmuir adsorption isotherm (Edzwald, --et al. 1976; Parfi tt 1978). Uptake rates would respond to increasing temperatures

147 as well. If, following incubation with high concentrations of FRP, the forest floor were then exposed to water having very low concentrations of FRP, desorption of bound FRP would occur to some extent depending on the affinity of the forest floor for phosphorus and the relative saturation of sorption sites (Carritt and Goodgal 1954; Parfitt 1978; Hingston, --et al. 1967, 1972). Under abiotic control of fluxes, the release of FUP from the forest floor would probably be the result of leaching processes that would remain constant at a given temperature.

Under biotic control of fluxes, the response of the FRP uptake rate to increased concentrations of FRP would be similar to that for abiotic sorption, i.e., a rectangular hyperbola, usually described in biological systems by the Michaelis-Menten equation. However, following sorption in the presence of high FRP concentrations, rapid release of FRP to water low in FRP would not be expected. Increases in temperature or biomass, indicated by chl orophyl 1-a measurements , for example, woul d resul t in increased fluxes. Finally inhibitors such as bacterial antibiotics or general poisons, would markedly reduce fluxes of FRP. Under biotic control, the flux of FUP would not be easily predicted.

B. Preliminary data

Partitioning of the fluxes of FRP at the forest floor-water interface into biotic and abiotic components was attempted by "killing" some chambers with formalin. In W.Y. 1977, formalin treatments resulted in a change in the direction of fluxes of FRP and FUP (Table 36). The differences in fluxes between control and formalin-treated chambers was statistically significant (0.05 < P < 0.10) for FRP but not for FUP. In W.Y. 1978, differences were also significant (0.05 < P < 0.10) between control and formal in-treated chambers on 3 May (Table 36). However, in this case, uptake of FRP by the forest floor in the treated chambers increased over uptake in control chambers. Two other sets of experimentim6) indicated a decline in uptake rate with formal in treatment. Insufficient rep1ication in these 1 atter experiments did not permit statistical testing. Overall, the response to formal in treatment was highly variable. This may have been due to different levels of biological activity during each experiment, or, more probably, to stressing rather than outright killing by the formalin treatment. Measurements of C02 production by the forest floor in treated chambers also gave inconsistent results (Mu1 ho1 land, personal communication).

Lack of confidence in the use of formalin or any other killing agent precluded a clear determination of biotic and abiotic components of the exchange process. Evidence from studies in W.Y. 1977 and 1978 suggested that fluxes may have the result of both abiotic and biotic components. Formalin treatments, except in one case, reduced or eliminated forest floor uptake of FRP, suggesting some biological activity. FRP fluxes correlated we1 1 with FRP concentrations (Table 34), indicating either biotic or abiotic control. Table 36. Effects of formalin solution on FRP and FUP floor-water exchanges. Rates are in mg*m'2@hr" .

CONTROL FORMALIN-TREATED FRP FUP FRP FUP

1978 a) ji -0.223* ------0.338* ----- s (0.012) (0.086) n 4 6

b -3.44 ------0.119 ----- 1979 -19.2 ------12.7 ------"significantly different at a = 0.10

Regeneration experiments in W. Y. 1979 indicated that sorbed FRP was released in small amounts upon the introduction of water low in FRP (Table 37); but, with one exception (28 February 1979) release was not proportional to the initial FRP concentrations. Either the affinity for FRP by abiotic sorption sites was very high, with very low equilibrium concentrations in the water, or biological activity predominated. The experiments on 28 February 1979 occurred after a high storm-flow period when visible amounts of sediment had been deposited on the floodplain. These sediments probably originated from the relatively phosphorus-rich surface soils of exposed fields and thus may have had a low phosphorus affinity at low concentrations. It is possible that FRP kinetics on this date were strongly influenced by the sediment on the leaves; chlorophyll-a at this time was relatively low. On those occasions when measurements were made, no regeneration of 32~from the forest floor was observed (Table 37). 3 2 31 With a few exceptions, regeneration of both FU P and FU P increased as initial FRP concentrations were increased in the experimental chambers (Table 37). These increased rates were observed in chambers which had been flushed and then filled with ambient low FRP water following the initial experiment. Whether this phenomenon was the result of enhanced biological activity due to the stimulatory effect of increased FRP concentrations or whether it was due to an irreversible sorption of FRP to fi1 terable materials is unknown. Table 37. Regeneration of FRP and FUP in chambers following experi - ments where FRP concentrations were increased over ambient. Rates are in mg m-2-hr'1. Where available, standard deviations are in parentheses.

Date Initial FRP FRP Flux FUP Flux (mg.1-') 31p -2P -- lP -- 32~

1.274 ---- 0.035 none

0.264 none 0.106 none 0.052 none 0.065 none C. Experimental testing

On 27 March 1979, experiments were conducted to examine the effects of the bacterial antibiotics, streptomycin and penicillin, formalin, and three levels of NaMAsOq (0.5, 5.0 and 10.0 NU) on FRP uptake at elevated FRP concentrations (0.3 mg*l-l or 10 NM). The addition of antibiotics appeared to have no effect on FRP uptake (Fig. 39). Formalin depressed uptake relative to untreated chambers, as did all concentrations of arsenate (As). Chemically, arsenate is very similar to phosphate and thus should respond similarly to sorption and compete with phosphate for sorption sites (Parfitt 1978; Hingston, et al. 1967, 1972). If chemical sorption controlled kinetics at the forest floor-water interface, then fluxes of As and FRP combined should be similar to fluxes of FRP alone at the same relative concentrations. Based on measurements of combined As and FRP changes, this was not the case, indicating a biological component which appeared uniformly inhibited by all three levels-of As (Fig. 39).

The relative sorption affinities of the forest floor for As and FRP were investigated by comparing the ratios of As to FRP initially present in the water of the chambers with the ratio of uptake of As and FRP (Table 38). Total concentrations of As and FRP were 15 yM in all chambers initially, except for chamber #1 kept at ambient concentr~tionsand #9 in which total As and FRP were elevated to about 30 pM. In the ambient chamber, As was re1eased and FRP was taken up during The experiment. In chambers where one species was kept at ambient levels and the other increased to 15 VM (chambers 2 and 3), the species at ambient level was released or showe?? no net movement. At an approximately 14:1 As:FRP ratio (chamber 4), FRP was released; at the inverse ratio (0.075, chamber 5), AS, the species present at lower concentration, was sorbed but to a smaller degree relative to FRP. In general terms, the latter result was found at ratios of 4:1 and 7:4 As:FRP. At l:1 concentration ratios, uptake ratios were similar to initial concentration ratios. These data suggest that at widely different ratios of concentrations, uptake of the more concentrated species was enhanced over the species at lower concentration. High As concentrations released FRP from the litter. At nearly equivalent ratios, uptake was in close proportion to concentration. Overall, this appeared to be a chemical sorption phenomenon rather than a biological process.

Annual estimates of floor-water exchanges of filterable phosphorus:

A good estimate of an-annual flux of filterable phosphorus at the forest floor-water interface is difficult to make because of the problems associated with the measurement techniques and because of the poor understanding of the apparent interactive effect af temperature and FRP concentrations on fluxes of FRP. Some crude estfmates of FRP and FUP fluxes for W.Y. 1977 and 1978 (Table 39) were obtained by multiplying average fluxes (Table 34) by a number called km2-days (Table 39) which represents the cumulative area and length of time of inundation of the floodplain for '""ICREEPING SWAMP, N.C. CP-I4

As and FRP Concentration (VM)

Figure 39. The effect of bacterial antibiotics, formalin and sodium arsenate on FRP uptake by the swamp forest floor at enhanced FRP concentrations. Experiments were done on 27 March 1979. Curves illustrate uptake as a function of initial FRP concentrations with no added arsenate. Bars indicate uptake in the presence of poisons as well as added FRP. Table 38. Comparison of the ratios of inltial As:FRP concentrations with the ratios of the uptake of As and FRP. Ratios are on a mole-to-mole basis and were estimated during experiments on 11 Nay 1979.

Chamber Initial FRP Initial As:FRP As :FRP Uptake (mg.1-' )

-11.7 As released

-1850 FRP released

-66.8 FRP released

Table 39. Estimates of the annual fluxes of FRP and FUP at the forest floor-water interface of the swamp floodplain.

Annual Annual Ma ter FRP Flux FUP Flux Algal Uptake Year krn2 -days (mg*m'2a yr-1) (mgem'2*yr'1) (mg*m'2eyr'l) each water year. Based on this calculation, the annual flufes of FRP from the floodwaters to the forest floor were 1300 and 710 mgem- -yr-l (Table 39) for W.Y. 1977 and 1978, respectively. FRP estimates in W.Y. 1977 were calculated based on rates measured in chambers to which small amounts of FRP had been added. The we1 1 -documented stimul atory effect (Tab1e 34) of increasing FRP concentrations on uptake rates may have contributed to the larger value in W.Y. 1977. W.Y. 1978 estimates were based on measurements using 32~which, under ambient conditions, appeared to result in a measurement of gross rather than net uptake rates; thus this technique probably overestimat FRP fluxes as well. Measurements in W.Y. 1979 demonstrated that FR 5? P fluxes under ambient conditions were small in absolute value but quite varying. Due to the low FRP levels present, the accuracy of these estimates is uncertain.

Algal uptake of FRP:

The uptake of FRP from floodwaters by a seasonal component of the forest floor, the filamentous algae, was estimated twice in W.Y. 1978. Similar estimates were planned for W.Y. 1979, but a clearly defined algae bloom did not develop on the floodplain at CP-14. All the above forest floor-water flux estimates include the effects of the algae present so that there is no missing component. Based on a 40-day bloom period in which the algae took up FRP at a rate of -1 .I8 mg.m-2*hr-l (based on two experiments with both 1ight and dark estimates), approximately 1100 mg ~~~*m-2-~r-l were removed from the floodwaters by the algae (Table 39). This is about 1.5 times the annual estimated flux of FRP at the forest floor-water interface during W.Y. 1978.

DISCUSSION

Phosphorus budget for the Creeping Swamp watershed

Annual inputs and outputs:

Elemental cycling and exports from watersheds can be examined on a fundamental 1eve1 by treating watersheds as "black boxes" and measuring only inputs and outputs. Using this approach, processes occurring within the watershed are not measured. If the Creeping Swamp watershed, including the swamp floodplain, is viewed in this perspective, inputs of total phosphorus, including natural and anthropogenic sources, greatly exceeded outputs during W .Y. 1977 and 1978 (Table 40). Wood1 ands constituted about 65% and crop lands about 25% of the Creeping Swamp watershed (Kuenzler, et a1 . 1977; Winner and Simmons 1977). The remaining area (10%) was newly clearcut or in young pine plantations. No urban centers were located in the watershed. Thus, the primary sources of phosphorus inputs to the watershed were precipitation and agricultural activities.

Phosphorus in bulk precipitation constituted the predominant natural source of phosphorus to the watershed. The contribution of phosphorus Table 40. Phosphorus budget for the Creeping Swamp watershed upstream of CP-10. Inputs and outputs were based on the entire watershed area of 80 km2 and are in units of mg P=m-28 yr-l.

FRP FUP PP TP 1977 1978 -1977 - 1978 1977 1978 1977 1978 A. Inputs Bulk Precipitation (I) 47.1 56.7

Ferti1 izers 250 250 250 250

Net TB-02 Excess

TOTAL (TI)

B. Outputs -I ~n Surface Water at CP-10 (0) 15.3 3.09 2.87 8.03 11.2 10.5 29.5 22.2 cn Deep Ground Water 0.07 0.08 0.20 0.24 0.27 0.32

Crop Harvests - - TOTAL (TO) 15.4 3.17 3.07 8.27 111 110 130122

C. Retention -- Natural Fluxes Bulk Precipitation - Streamflow (1-0) 32 54 -3.07 -8.27 -11.2 -10.5 31.1 57.5

Retention % (I-O)/I 68% 9 5% 51% 72% D. Retention -- All fluxes Total Inputs - Total outputs (TI-TO) 52 62 -2.12 -6.2 144 156 209 236

Retention % (TI-TO)/TI 78% 95% 56 % 59% 62% 66% from soil weathsring processes was probably quite small due to the highly leached nature of watershed soils (Soil Survey, Pitt County 1974). The annual weighted mean concentrations of total phosphorus in bulk precipitation in the Creeping Swamp watershed were at the high end of the range of concentrations measured at several locations in the eastern United States (Table 41 ) . Concentrations in mountainous areas (Hubbard Brook, Coweeta) or in uninhabited areas (Okefenokee Swamp) were quite low, whereas concentrations measured in watersheds with agricultural or urban activity (Tar River, Creeping Swamp, Rhode River, Walker Branch) were considerably higher. Phosphorus in bulk precipitation originates from air-borne particulates, such as smoke, field dust, etc. (Tamm and Troedsson 1955; Johnson, --et al. 1966; Winkler 1976; Miklas, --et a1. 1977). In regions having agricultural activity, significant amounts of particulate matter may enter the atmosphere. Rain wash-out may return these particulates and the nutrients borne on them to nearby ecosystems(Johnson, et a1 . 1966; Brezoni k 1972). Thus, the total amount of phosphorus broughtin5- the Creeping Swamp watershed in bulk precipitation may not represent a new source of phosphorus but rather a significant portion of this "input" may be derived from within the watershed itself (c.f., Kirshner 1975). The actual net input of phosphorus to the watershed is unknown but was probably small and similar to amounts (4-20 mg-m-2.yr-1) measured at Coweeta, Hubbard Brook and Okefenokee Swamp (Tab1 e 41 ) .

Anthropogeni c inputs of phosphorus to the watershed consi sted of loadings due to agricultural fertilization, animal feedlots and septic systems. Approximately 20,000 kg phosphorus or 250 mg-m-2.yr-l were applied to the whole watershed each year as fertilizer (Pi tt County Extension Service) (Table 40). This estimate allows for differences in fertilization practices among the four predominant crops grown in the watershed, corn, tobacco, soybeans and peanuts, and is presented as an input to the whole watershed, not just to the crop lands. Of all the estimated inputs to the watershed, fertil izers were the most important source of phosphorus. Animal-raising operations introduced phosphorus to the watershed as feed and removed it in the sale of hogs, chickens, etc, The net input of phosphorus from these operations is the difference between feed and sale of animals. The portion of this difference which reached the stream channels in Creeping Swamp watershed can be estimated for a poorly managed hog farm by the excess phosphorus transported in the pol luted tributary, TB-02 (Table 40). The TB-02 excess was calculated as the difference between the actual phosphorus export at TB-02 and the "cl ean-water" exports estimated from TB-04 phosphorus concentrations (Table 29). An estimated excess of 2200-2300 kg of total hosphorus was carried in TB-02 streamwaters which amounted to 28-29 mg*m-8*yr-l in W.Y. 1977 and W.Y. 1978 when averaged over the entire watershed. These inputs were 30-40% of precipitation inputs and were similar to CP-10 surface water outputs(Tab1e 40). At least one other hog farm was in operation in the watershed during the study period; therefore hog farm inputs were probably about twice the above estimate or about 60 mg*m-2.yr-1. Phosphorus inputs via septic systems were not estimated but represented another source of phosphorus to the watershed. Table 41. Comparison of annual mean concentrations of total phosphorus in bulk precipitation and annual inputs of phosphorus in bulk precipi tati on at different locations in the eastern Uni ted States.

AVERAGE ANNUAL CONCENTRATION INPUT LOCATION (mg* I-') (mg* m'2*y~'1)

Creeping Swamp, N. C. (1) 0.054-0.053 60-79 Piedmont, N. C. (2) 0.021 28

Tar River Swamp, N. C. (3) 0.053 49

Rhode River, Va. (4) 0.060-0.083 58-113

Coweeta, N. C. (5) 0.004 9-13

Walker Branch, Tenn. (6)

Okefenokee Swamp, Ga. (7)

Savannah River, S. C. (8) 0.033 30

Hubbard Brook, N. H. (3) 0.008 3.6

(1) this study; (2) We1 1s et a1 . , 1972; (3) Holrnes, 1377; (4) Mi klas et fl., 1977; (5) Swank and Douqaz, 1977; (6) Swank and Henderson, 1976; v)Schlesinger, 1978; (8) Polisini --et a1 . , 1970; (9) Likens et a1 . , 1977.

Losses of phosphorus from the watershed occurred primarily in crop harvests and surface water discharge at CP-10 (Table 40) In crops alone, the watershed yielded at least 8000 kg or 100 mg.m-2-yr-i of phosphorus (Pi tt County Extension Service) during W.Y. 1977 and 1978 (Table 40). In sharp contrast, phosphorus losses in stream discharge were 22-29 mg*m-2* yr-l or 20-30% of crop harvests (Table 40). A1 though approximately 25% of the watershed was in agricultural use, imports and exports due to this activity greatly exceeded natural fluxes. The low level of phosphorus exports in stream waters suggests that very little fertilizer phosphorus reached the stream channels, an observation made by other researchers (Taylor, --et a1. 1971 ; den; -et a1. 1973; Gburek and Heal d 1974; Gambrel 1 , --et a1 . 1974; Dunigan, --et a1 . 197g. A comparison of "natural " inputs (bul k precipitation) and outputs (CP-10 surface water discharge) (Table 40C) showed that net amounts of FRP and TP were retained within the watershed but that small amounts of FUP and PP were exported in stream waters. The influence of the TB-02 excess cannot be removed from the estimate of surface water discharge at CP-10. Therefore, exports at CP-10 may have been greater than would have occurred had no pollution source been present in the watershed (c.f. CP-20 watershed exports, Table 26). Overall, the percent of total phosphorus retained in the watershed increased from W.Y. 1977 to W.Y. 1978; greatly increased retention of FRP probably accounted for most of this difference. Comparison of natural and anthropogenic inputs and outputs from the watershed (Table 40D) shows that even with increased phosphorus loading the watershed released less phosphorus than it gained. Retention of total phosphorus (62-66%), which included inputs and outputs due to agricultural activities, was less efficient than FRP retention (78-95%), although much larger amounts of TP, 209-236 mg.m-2.yr-l, primarily in particulate form, remained within the watershed. Retention efficiencies of TP did not appear to differ between the two water years. Although the watershed budget was expressed on an areal basis, there was no assumption of uniformity of processes within the watershed. Crop lands and hog farms covered 25-30% of the watershed, yet phosphorus fluxes due to these activities predominated in the overall phosphorus budget. Agricultural areas tend to export more phosphorus than forested regions (Dillon and Kirshner 1975; Correll , --et al. 1977; and Browne and Grizzard 1979). Watershed studies in forested regions of eastern North America have shown that phosphorus exports in surface waters range from 1-67 mg*m-2e yr-l (Dillon and Kirshner 1975; Henderson, --et al. 1977; Likens, et al. 1977; Swank and Douglas~1977). Water-borne phosphorus expo ts from wzezheds primarily in agricultural use ranged from 10-460 mg*m-'*yr-l with a median range of 50-1 00 mg=m-2.yr-l (Castri 11 i and Dines 1978; Waniel ista 1978; Alberts, et al. 1978; Duffy, et al. 1978). Surface water exports from primarily wooded areas upstream of CP-20 were quite low (3-15 mg*m-2.yr-1 ; Table 26). With respect to surface water exports at CP-10, the Creeping Swamp watershed is more similar to forested than to agricultural watersheds. The Creeping Swamp floodplain is located at the lower end of the watershed and thus may act as a "receiving basin" for natural and anthropogenic sources of phosphorus in surface waters. The swamp may modify these inputs and contribute phosphorus to surface waters via processes peculiar only to the swamp floodplain. The overall result is a system which appears highly effective in minimizing phosphorus losses even under the pressure of agriculture and animal husbandry. Phosphorus cycling in the floodplain swamp ecosystem

The swamp floodplain ecosystem received and processed phosphorus derived from upland exports and direct precipitation. As a result of its location at the base of the Creeping Swamp watershed, phosphorus leaving the watershed in stream waters had entered, cycled and passed from the swamp floodplain ecosystem. Therefore, the floodplain probably played a central role in the amount, nature and timing of phosphorus exports from the watershed as a whole.

Phosphorus concentrations in swamp waters:

Stream water phosphorus concentrations were predominanatly in filterable form except during rising discharge (Tables 21-24). Upstream of the influence of he hog farm, phosphorus concentrations were quite low (r0.010 mg 1- f ) throughout the year with a slight seasonal maximum during the summer months (Fig. 31). Similar trends were observed by Kuenzler, -et -a1 . (1977) for another unpolluted stream swamp, Palmetto Swamp, located southeast of Creeping Swamp. As in Creeping Swamp, FUP concentrations were usually higher than FRP concentrations and had the most consistent seasonal maxima in August and September of 1975 and 1976. Phosphorus concentrations in Chicod Creek, a natural swamp stream which received inputs from a hog farm, were much higher and more erratic than those observed in Creeping or Palmetto Swamps (Kuenzler, --et a1 . 1977). In two unpolluted channelized swamp streams in the same region, Tracey Swamp and Conetoe Creek, FUP and FRP concentrations were usually very low, except for scattered maxima in summer. PP concentrations, on the other hand, were much higher and more erratic than in the natural swamp streams.

Phosphorus concentrations in waters draining swamps or lowlands in other parts of the eastern United States were quite variable. Holmes (1 977 found total phosphorus concentrations to vary from 0.045 to 0.410 mg 1- 1 in the Tar River, N.C., swamp surface waters. Similar variations were found in riverine swamps in Illinois (Mitsch, --et al. 1977) and in Louisiana (Day, et a1. 1976; Seaton 1979). Total phosphorus concentrations averaged 0.21 mg7-in streams in poorly-drained areas of the Coastal Plain of northeastern North Carolina (Overcash, --et al. 1977). Bloxham (1976) measured, for the most part, very low concentrations of phosphate in streams draining the upper Coastal Plain of South Carolina as did Kitchens, --et al. (1975) for the swamp-draining Santee River. It appears, then, that streamwater phosphorus concentrations in Creeping Swamp are similar or lower than concentrations measured in other Coastal Plain streams.

The chemical and size speciation of filterable forms of phosphorus in Creeping Swamp stream waters indicated that significant amounts of FRP and FUP were actually colloidal rather than dissolved (Table 25). It has become increasingly evident that very little phosphorus exists in orthophosphate form in natural waters (Kuenzler and Ketchum 1962; Rigler 1968; Sinha 1971 ; Lehmusloto and Ryhanen 1972; Minear 1972; Chamberlain and Shapiro 1973; Lean 1973; Jackson and Schindler 1975; Lean and Nal ewaj ko 1976; Brown, --et a1 . 1978; Downes and Paerl 1978). In addition, often large percentages of FRP or "DIP" chemically determined by the molybdenum-bl ue method (Strickl and and Parsons 1972) have been shown to be colloidal in size (This study; Lean and Nalewajko 1976; Downes and Paerl 1978; Peters 1978). The biological availabil ity of colloidal reactive phosphorus (CoRP) is uncertain but it has been shown that colloidal phosphorus is produced by algae and zooplankton (Lean 1973; Peters and Lean 1973; Ferrante 1976; Lean and Nalewajko 1976; Peters 1978) and that colloidal phosphorus can be used by algal populations -via phosphatases (Paerl and Downes 1978). The biological availability of CoRP in Creeping Swamp waters is unknown; therefore the portion of FRP actually directly involved in inorganic phosphate fluxes within the ecosystem is also unknown but is probably smaller than the amount estimated by FRP measurements. The fractions of large molecular weight FRP and FUP that can be broken down by extra-cellular phosphatases is also unknown. Francko and Heath (1979) found that large molecular weight FUP from darkly stained bog waters was not phosphatase sensitive, but that it did release phosphorus following low level ultraviolet radiation. The reactivity of colloidal forms of filterable phosphorus deserves study if the true nature of stream water phosphorus cycling is to be elucidated.

The nature of large molecular weight phosphorus-bearing compounds is not completely understood. Evidence suggests that phosphorus may be chemically bound or sorbed to the humic and fulvic acid fractions of refractory organic matter (Levesque and Schni tzer 1967; Schni tzer 1969; Sinha 1971; Lehmusloto and Ryhanen 1972; Lean 1973; Jackson and Schindler 1975). With respect to sorbed or compl exed forms of col 1oidal phosphorus, iron (~e3+)or aluminum molecules bound to organic matter are thought to complex phosphate molecules (~evesqueand Schni tzer 1967; Jackson and Schindler 1975; Koenings and Hooper 1976). Due to its sensitivity to the acid-molybdate procedure of Strickland and Parsons (1972), the phosphate in CoRP measured in Creeping Swamp waters may be bonded by easily hydro- lyzable bonds to organic matter, or it may be sorbed to the organic material. Overall, it appears that very little inorganic phosphate exists in stream waters of Creeping Swamp. Rather, of the small amount of phosphorus which is present in stream and flood waters, most appears to be bound to organic matter or inorganic particles. Finally, ambient arsenate concentrations were similar to or slightly less than measured FRP concentrations in Creeping Swamp waters (see also Winner and Simmons 1977) and also were similar to concentrations measured in South Carolina and Georgia Coastal Plain rivers (Waslenchuk 1978; Sand1 er and Nelson 1979). The acid-molybdate technique for FRP measurement does not distinguish between arsenate and phosphate (Murphy and Riley 1962; Johnson 1971 ). This adds a further complication to the assessment of the true abundance of phosphorus in Creeping Swamp waters. Standing stocks of phosphorus in the Creeping Swamp ecosystem:

In Creeping Swamp, a total of about 5600 mg porn-' was present in above-ground 1iving biomass, 97% of which was in trees and sap1 ings (Table 30). About 450 mg P-m-2 was in the ground litter and nearly 34,000 mg ~.m-2was present in the first 25 cm of soil as available and organic phosphorus (Table 31). Most of the phosphorus in the ecosystem was in the soil, an observation in other forest ecosystems (Duvigneaud and Denaeyer-de-Smet 1970; Likens, et-- a1 . 1977; Lang and Forman 1978). The amount of phosphorus present in the below-ground biomass of roots was not measured but it was probably a significant portion of the total standing stock. Lugo, et al. (1978) estimated that 3500 mg ~.m-2were located in the roots of acypess-strand ecosystem in Florida. The a1 ternate wetting and drying of the ecosystem forest floor was postulated to account for the high biomass and turnover rate of roots in this and other seasonally inundated forests (Lugo, et al. 1978; Hook and Brown 1973). Whittaker, et -al. (1979) found that 60%?fthe phosphorus in trees o the Hubbard Brook forest was below-ground; this amounted to 5300 mg Porn" 5 . Rolfe, et al. (1978a) found phosphorus in root biomass (1200 mg porn-2) in an IlFnois oak-hickory forest to be less than that in above-ground biomass but it was still a significant component of the total living standing stock of phosphorus (4100 mg porn-2).

In Creeping Swamp, the amount of phosphorus in above-ground living biomass was large compared to values found for other forested ecosystems (Duvigneaud and Denaeyer-de-Smet 1970; Roc how 1976; Schl esi nger 1976; Rolfe, --et al. 1978a; Whittaker, --et al. 1979). Most of the phosphorus in above-ground living biomass was in tree stems and branches. Leaves were relatively rich in phosphorus and were the second largest category. Phos- phorus in herbaceous vegetation, shrubs, vines and bryophytes was a very small component of the total but was similar to values found in other forested ecosys tems (Duvigneaud and Denaeyer-de-Smet 1970; Rochow 1976; Whi ttaker, --et al. 1979). The net above-ground increment of 59-1 17 mg ~*m-z*yr-lin the Creeping Swamp floodplain is similar to values (100-200 mg ~*m-Z*~r-l)measured in other forests less than 100 years old (Duvigneaud and Denaeyer-de-Smet 1970; Schlesinger 1976; Whi ttaker, --et a1 . 1979). On the forest floor of the swamp floodplain, the fermentation and humus layers traditional ly associated wi th forest floors (Gosz, --et a1 . 1976) were very thin or absent entirely. Therefore, ground 1i tter predominated in the forest floor and it will be used as a basis for comparison with other studies. The average amount of phosphorus in ground litter in Creeping Swamp (Table 31 ) was smal 1 compared to other ecosystems (Duvig- neaud and Denaeyer-de-Smet 1970; Reiners and Reiners 1970; Yount 1975; GOSZ, --et al. 1976; Rochow 1976; Lang and Forman 1978; Rolfe, et al. 1978a). There was a slight seasonal maximum in ground litter phCpGrus (Fig. 32) which was associated with the autumn maximum in litterfall.

Day (1978) found that ground litter was scarcest in the more frequently inundated portions of the Dismal Swamp in Virginia. The smaller accumulations in seasonally flooded ecosystems may be due to enhanced decomposition rates and increased nutrient release occurring during alter- nate wet and dry periods (Day 1978; Schlesinger 1977). In addition, Sniffen (in prep. ) observed relatively high populations of macroinverte- brates in the more frequently inundated portions of the Creeping Swamp floodplain. Detrital processing by these animals in frequently inundated areas may be the cause of lower average standing crops of ground litter. During the growing season, roots grew into the ground litter on the low floodplain and often by late July or early August, ground 1i tter had completely disappeared from the lower elevations of the floodplain.

The relatively low standing stock of phosphorus in ground litter, the lack of significant fermentation or humus layers and the observed colonization of litter by roots are characteristics seen in tropical forested ecosystems (Luse 1970; Stark and Jordan 1978). Experiments using 32~have demonstrated that phosphorus in the ground litter of tropical forests is rapidly transferred to living biomass upon mineraliza- tion (Luse 1970; Stark and Jordan 1978). As a result, phosphorus cycling in these ecosystems is often described as being tightly coupled. This study of Creeping Swamp, including measurement of very low concentrations of phosphorus in flood and stream waters, suggests that the Creeping Swamp ecosystem is also very effective in retaining and recycling phosphorus. Similar observations were made in a temperate beech-oak forest in Maryland (Correll and Miklas 1975).

Phosphorus measured in the mineral soil of Creeping Swamp averaged 32,000 - 35,000 mg porn-2. Concentrations were higher in the top 5 cm of soil; organic matter content was higher as well. Across the floodplain, soil phosphorus and soil organic matter were higher at lower elevations and near stream channels (Fig. 33). Ground 1i tter phosphorus was typically low at low elevations. While litter breakdown may have been faster at lower floodplain elevations, accumulation of phosphorus and organic matter in the soil was more pronounced in the lower, more frequently inundated areas of the floodplain.

Intrasystem transfers of phosphorus:

The major transfers of phosphorus within the Creeping Swamp floodplain ecosystem occurred in 1 itterfall , throughfall, stemflow, sedimentation, floor-water exchanges, and undoubtedly root uptake. A bloom of filamentous a1 gae on the floodplain contributed another set of seasonal ly important transfers of phosphorus.

Canopy losses of phosphorus throughfall + stemflow + 1i tterfall ) totaled 372 and about 460 mg Pmm-heyr-1 for W.Y. 1977 and 1978 (average litterfall value used for W.Y. 1978) (Table 32, Eig. 35). Litterfall dominated canopy losses and averaged 330 mg P-m' *yr-1 for W. Y. 1976 and 1977. Litterfall was about 30% of the estimated standing stock of phosphorus in leaves (Table 30) suggesting that this leaf compartment was either over-estimated or that much of the leaf phosphorus was removed to tree branches an3 boles before abscission in the fall Phosphorus in large branch fall was estimated to be 22 mg ~*m-2-yr-i and was less than 10% of phosphorus in litterfall. Inclusion of this measurement with litterfall results in 367 and 335 mg ~em-~*~r-lin total litterfall during W.Y. 1976 and 1977, respectively. The annual flux of phosphorus in litterfall in Creeping Swamp was similar to values measured in Okefenokee Swamp (Schlesinger 1976) and in temperate deciduous forests (Duvi gneaud and Denaeyer-de-Smet 1970; We1 1s , --et a1. 1972; Rochow 1976; Rolfe, et al. 1978b; Whittaker, et al. l979), but was less than fluxes in a northern swamp (620 mg ~*m-2*~r--h-Einers and Reiners 1970), in the Tar River, N. C., swamp (540 mg Pgm-2*yr-1; Holmes 1977) and in seasonally inundated tropical forests (Golley, et al. 1975; Ewel 1976). The amount of phosphorus in throughfall differed by more than a factor of two between the two years when measurements were made (Table 32). Greater rainfall occurred in W.Y. 1978, when throughfall fluxes were greater, but the difference is better explained by the doubling of the net weighted mean concentrations of phosphorus in throughfall between W.Y. 1977 and 1978. The rainfall volumes measured in W.Y. 1978 were much more uniform than those in W.Y. 1977 (Fig. 34). Two storms in W.Y. 1977 comprised more than 25% of the total throughfall volume measured that year and phosphorus concentrations in these samples were quite low. Data from these dates lowered the W.Y. 1977 annual weighted mean concentrations of phosphorus. Phosphorus fluxes in throughfall were similar to or less than estimates made for other upland and swamp forests (Duvigneaud and Denaeyer-de-Smet 1970; Wells, --et al. 1972; Ewel, et al. 1975; Golley, --et -al. 1975; Schlesinger 1976; Henderson, et al. 1977a;Holmes 1977; Rolfe, --et al. 1978b; Whittaker, et a1 . 1979). - Stemflow (2.2 - 2.6 mg ~.rn-~*~r-')accounted for less than 1% of the phosphorus transport from the tree canop to the forest floor. In other forests, val ues ranged from 1-30 mg Pom-s*yr-l , wi th most measurements greater than 5 mg P .mm2 eyr-l (We1 1s , --et a1 . 1972; Ewe1 , et a1 . 1975; Holmes 1977; Jordan 7 977; Rol fey et a1 . 1978b; Whi ttakerTea1.-- 1979). Particulate phosphorus sedimentation was a significant component (27%) of phosphorus inputs to the forest floor (canopy losses + sedimentation). Greater sedimentation occurred at lower floodplain elevations; this was reflected in greater ash weights of ground litter at lower elevations (data not shown). Sedimentation probably included primarily the settling out of phosphorus bound to silt and clay particles. Forest floor-water exchanges of phosphorus:

The floodplain of Creeping Swamp provided many potential sites for exchange of water-borne phosphorus with the flooded forest floor during the cool season when inundation was greatest. Exchanges at this interface could modify phosphorus concentrations and speciation in floodwaters passing through the swamp. A large portion of the loss of particulate phosphorus from surface water could be attributed to physical sedimentation (Table 33). Examination of the rates of fi1 terable phosphorus exchange at the forest floor-water interface and the factors controlling these rates constituted an important part of the overall study of phosphorus cycling in Creeping Swamp.

A. Interpretation

This study was hampered by the small amounts of phosphorus in the water, making measurements of fluxes at ambient fi1 terable phosphorus concentrations difficult. The highest average "ambient" fluxes of phosphorus were measured in W.Y. 1977 when small increments of KH2P04 were added to flowing-water chambers so that changes in FRP concentrations duri ng the experiment could be measured (Tabl e 34). The excel 1ent correlation of the uptake rates with initial FRP concentrations suggested a stimulatory effect of added FRP on the rate of FRP uptake by the swamp forest floor in W.Y. 1977 as well as in the other two water years (Table 34). Efforts to use 32~04as a tracer of FRP kinetics led to the estimation of gross FRP uptake rates, which, at ambient concentrations, were demonstrated in W.Y. 1979 to be much greater than net uptake rates as measured by changes in FRP concentrations. Simi 1ar resul ts have been found by other researchers studying a1 gal phosphorus kinetics (Nalewaj ko and Lean 1978; Brown, --et al. 1978; Lean and Nalewajko 1976; Kuenzler, --et al. 1979). If FRP concentrations change during the course of the experiment, then the assumption of steady state in the model of phosphorus kinetics (see METHODS) is violated. Therefore, the 32~-estiaateduptake rates were not necessarily equal to the unmeasured "efflux" rates, but were gross uptake rates of which some portions were equal to the rates of FRP efflux and the remainder were net removal by the forest floor. It is apparent, then, that when 32~04is being used as a tracer for FRP kinetics, one must know whether the experimental system is at steady state. This is not easily determined a priori in natural systems. In particular, in Creeping Swamp, a flux of FRP was always measured under ambient conditions, but it varied from uptake to release between chambers and on different occasions, even though the measured ambient FRP concentrations always remained quite low. The swamp forest floor undoubtedly had a tremendous micro-spatial and temporal heterogeneity that may have contributed to the variability in rates measured at ambient FRP concentrations. It remains uncertain whether, over periods of time longer than the duration of the floor-water exchange experiments, FRP kinetics were in steady state or whether there was a net uptake of FRP by the forest floor. The paralleled release of FUP during these experiments (Tabl e 34), the re1ative high concentrations of FRP in precipitation (Table 20) and the overall retention of surface water FRP by the swamp ecosystem (Table 29) strongly indicate that there was, over the period of seasonal floodplain inundation, uptake of FRP by the swamp floodplain. When phosphorus concentrations were elevated experimentally in the chambers, then rates based on 32~0~kinetics were similar but usually slightly lower than rates based on changes in ~~FRP.This has also been observed for batch cultures of algae (Nalewajko and Lean 1978; Perry 1976) and sug ests that, under non-steady state conditions, flux rates estimated by g2P04 kinetics may be a reasonable indicator of the net movement of FRP from the water. 3 1 Lack of aqreement between 32~and FR P means of estimating FRP fluxes and the difficil ty in ascertaining the factors control ling FRP fluxes at the forest floor-water interface call attention to technical inadequacies, poor or violated assumptions in the mathematical model used in calculating fluxes, and imperfect understanding of the molecular kinetics of phosphate in natural systems. Measurement of rates --in situ, while resulting in rates occurring under natural conditions, introduced such uncontrol led factors into the experiments as spati a1 and biologi ca? heterogeneity of the leaf matrix sf the forest floor and lack of knowledge of the phosphorus "history" of the floor-water interface (recent sedimentation during a storm event, status of the algal community, inundation duration, and so on). The heterogeneity of the forest floor probably prevented complete mixing or diffusion of added phosphate, As, antibiotics, or formalin to all the sites in the litter layer that were active in phosphorus kinetics. The duration of the experiments may have resulted in the measurement of diffusion or sorption phenomena rather than biological activity due to the apparent under-saturation of the forest floor for phosphorus. The kinetics of diffusion or sorption to empty sites under these conditions may have overwhelmed the slower uptake rates of the biological community. This problem may have been particularly acute when the assessment of seasonal trends was attempted. Finally, lack of knowledge of the portion of FRP which was phosphate rather than arsenate and which was actively involved in exchange reactions precludes precise definition of the amount and speciation of FRP involved in floor-water exchanges.

Determination of bi01 ogical versus- chemical control of phosphorus kinetics at the forest floor-water interface would necessitate experimen- tation in the laboratory, where mixing, temperature, light and forest floor character (tree leaf species, presence or absence of algal or soi 1) could be careful ly control led. In addition, arsenate and phosphate would need to be distinguished and the concentrations of both accurately determined. If such careful determinations could be relied upon, then the need for the 32~tracer could be eliminated and FRP kinetics could be measured directly. Disregarding the somewhat uncertain estimations of arsenate concentrations in swamp waters, it is possible to use the measurement of FRP fluxes as an estimation of the movement of a chemically defined species at the forest floor-water interface of the swamp. When compared with other measured FRP fluxes in the swamp ecosystem, the floor-water exchange measurements are useful in gaining an overall picture of phosphorus cycling.

The factors controlling FRP fluxes at the forest floor-water interface were not clearly biological or chemical and there were no distinct seasonal patterns ,of biological versus chemical control. At "ambient" FRP concentrations, FRP fluxes correlated significantly with water temperature only in W.Y. 1977 when small increments of FRP were added to the experimental chambers. When FRP fluxes measured at both ambient and incremental levels of FRP were graphed according to both temperature and initial FRP concentrations (Fig. 38), it was apparent that temperature did influence FRP uptake at elevated FRP concentrations. This was particularly evident in W.Y. 1979 (Fig. 37). In the swamp, increasing water temperatures occurred simultaneously with increased daylength and growth of filamentous algae on the flooded forest floor. FRP fluxes correlated more strongly with sequential date and chlorophyll-a (W.Y. 1979 only) than with water temperature (Tab1 e 34). However, the strongest and most consi sten t correlations were with FRP concentrations, either under ambient or elevated FRP conditions. Regardless of the controlling mechanisms of FRP fluxes, the swamp forest floor appears to be below saturation with, and to have a high affinity for, FRP. The lack of any FR~~Pregeneration and the low rates of FR~~Pefflux (with the exception of 12 March 1979) regardless of initial FRP concentrations further support this idea. Effluxes of FR~~P on 12 March 1979 were measured in chambers having visible amounts of freshly deposi ted sediments. If these sediments originated from the surface soil of agricultural fields, then they may have been more nearly saturated with FRP and therefore had a low affinity for the incremental FRP in the chambers. Similar observations have been made by researchers working with freshly deposited river sediments (Taylor and Kunishi 1971; McCal 1is ter and Logan 1978).

The addition of bacterial antibiotics or the general poisons, formalin and sodium arsenate, did not result in a clear reduction or inhibition of FRP uptake by the forest floor. On the average, formalin and sodium arsenate additions resulted in decreased FRP uptake at elevated FRP concentrations (Table 36, Fig. 39), suggesting that biological activity was only partially responsible for FRP fluxes. However, the spatial heterogeneity of the 1eaf-covered forest floor probably prevented ki11 i ng the entire biological community. Statistical correlations suggest that biological activi ties were probably responsible for the seasonal increase in FRP uptake rates seen in W.Y. 1979 (Fig. 39). However, the extremely low ambient FRP concentrations and the distinct and repeatable response of uptake rates to increased FRP concentrations indicated that the forest floor biological community or chemical sorption sites may have been severely under-saturated and thus potentially (biological ly) 1imi ted with respect to FRP.

In general, FUP moved fromthe swamp forest floor to overlying floodwaters (Table 34), and FUP release usually increased in chambers following the addition of FRP (Table 37). In addition, the transformation of 32~04 into FU~~Poften accounted for 10-25% of the total FUP efflux from the forest floor (Table 37). At ambient FRP concentrations, the measured efflux of FUP and the transfOrmation of 32~04into FU~~Pindicated that the swamp forest floor was a site for the change of reactive phosphorus into unreactive forms. FUP was a large component of swamp surface water concentrations and exports (Tables 21-24, 29; Fig. 31). Net inputs of phosphorus to the Creeping Swamp watershed in bul k precipitation were primarily in reactive form (Table 20). To account for the FUP concentrations and exports in swamp surface waters, transformation of FRP into FUP must have been an important process both in the entire watershed and in the stream channels and floodplain of the swamp ecosystem.

The factors controlling FUP fluxes at the forest floor-water interface were not easily defined. FUP fluxes did not correlate significantly with any measured inde endent variable in W.Y. 1977 or in W.Y. 1979. The transformation of 92PO4 into FU~~Pduring the period of the experiments and the increases in FU~~Pwith increasing initial FRP concentrations (Table 37) suggest that some portion of the process occurred over very short lengths of time and was dependent on FRP concentrations. However, when FRP uptake was compared with FUP efflux for each isotopic fraction of phosphorus, no clear relationship emerged between the rate of FRP uptake and the subsequent release of FUP at varying initial FRP concentrations . 32 FU P efflux rates were usually less than 25% of the total F P efflux rates (Table 37). It is possible that a large portion of the FU 3Y P release was attributable to displacement of FUP, bound through the phosphate molecule to sorption sites, by FRP (Parfi tt 1978; Meyer 1979); this would account for the increased FUP release with increased FRP concentrations. The importance of biological mediation of this process is not at all clear because no signi f icant correlations were found between FUP f 1uxes and temperature, chlorophyll-a, or sequential date, all of which might indicate biological influence. In conclusion, the efflux of FUP from the forest floor to overlying floodwaters was a significant component of the phosphorus cycle of the swamp ecosystem in that losses of FUP from the forest floor represented a potentially important pathway for the loss of phosphorus from the ecosystem as a whole.

5. Comparison wi th measurements in other ecosystems

Phosphorus fluxes at the sediment-water interface of aquatic and semi-aquatic ecosystems have received a great deal of attention since the classic work of Mortimer (1941, 1942) which demonstrated that in a eutrophic lake an oxidized "microzone" at the surface of the sediments prevented the release of phosphate to overlying waters under aerobic conditions. When waters and sediments became anaerobic, as commonly happens in eutrophic lakes during summer stratification, the oxidized zone disappeared and phosphorus along with several other elements were released to the water. This release was and continues to be considered as an important source of phosphorus to phytoplankton foll owing 1ake mixing. The study of sediment- water exchange of phosphorus has broadened into estuaries (Pomeroy, et al. 1965; Hale 1975; Fisher and Carlson, unpubl. data), oceans (Hartwig -7q Rowe, --et al. 1977), swamps and marshes (Banoub 1975a; Nixon, et al. 1976; Holmes 1977), and 01 igotrophic lakes (Glass and Poldoski 1974d~Gp- Nielsen 1974, 1975; Bannerman, --et al. 1975; Viner 1975b, c; Schindler, --et al. 1977) and has focused primarily on FRP exchanges. In aquatic ecosystems which undergo a1 ternate periods of aerobic and anaerobic conditions at the sediment-water interface, some researchers (Fi 110s and Molof 1972; Kamp-Niel sen 1974, 1975; Patrick and Khal id 1974) have found phosphorus fluxes to vary in the same manner as that reported by Mortimer (1941, 1942). However, many others have found that phosphorus kinetics appear to be more closely related to fluctuations in temperature (Banoub 1975a; Hale 1975; Kamp-Nielsen 1974; Gallepp 1979; Fisher and Carl son unpubl . data). Fluxes have been found to vary from net uptake to net release or to increasing release with increasing temperature. Increasing temperatures may stimulate decomposition and increased oxygen consumption in the bottom sediments, leading over sufficient time to the release of phosphorus from mineraliza- tion reactions and from ferric hydroxides following the reduction of iron to the ferrous state (Patrick and Mahapatra 1968).

In Creeping Swamp, where the floor-water interface remained aerobic under flowing water conditions, temperature may have stimulated FRP uptake at measureable FRP concentrations. In this and other aquatic ecosystems poor in phosphorus, movement of phosphorus at the sediment- or floor- water interface was typical ly from overlying waters to sediments (Hayes 1955; Viner 1975b, c; Schindler, --et al. 1977).

The average rates of FRP uptake measured in Creeping Swamp were in the same absolute range of rates measured in other ecosystems (Table 42). Holmes (1977) estimated a relatively high flux of phosphorus from the water to the forest floor of the Tar River swamp. These rates were based on 32~04measurements and average total phosphorus concentrations in swamp waters; as such they may have been an overestimation of what actually occurred. In other studies, uptake of FRP at the sediment-water interface was measured most often under low temperature, oxygenated conditions. These were the conditions that existed for the most part during the flooded period in Creeping Swamp. If the Creeping Swamp floodplain were subject to long-term inundation during the summer months, perhaps then FRP would be re1eased from the forest floor to floodwaters under low oxygen conditions. This is supported by the results of Kuenzl er, --et a1 . (1977) who showed that FRP concentrations in Creeping and Palmetto Swamps were higher during summer stagnation when dissolved oxygen was low. In effect, the dry conditions, which typically persist during the warm months, may be indirectly quite important in the overall retention of phosphorus by the ecosystem.

A budget of phosphorus cycling in Creeping Swamp:

A budget of phosphorus cycling in Creeping Swamp was developed by partitioning the ecosystem into five components and into pathways between components and pathways leading to and from the ecosystem. Components were 1) the swamp floodwaters, which were further subdivided into FRP, FUP, PP and LPP (particulate phosphorus > 0.25 mm) , 2) trees and sap1 ing, 3) herbaceous vegetation, shrubs, vines and bryophytes, 4) the swamp forest floor, including the seasonally abundant filamentous a1 gae, and 5) the Table 42. Rates of phosphorus exchange at the sediment-water interface of various aquati c and semi -aquatic ecosystems. (-1 indi - cates uptake by sediments.

INTERFACE FRP FLUX ECOSYSTEM CHARACTERISTICS -(mg P mF2 hr-') Creeping Swamp (1) aerobic, 4-25 C -0.450 to 0.035 Tar River, N.C. swamp (2) * P measurement -2.21 to -7.81 Bodensee marsh (3) aerobic, 4-25 C 0.05 to 0.39 Narragansett Bay embayment (4) aerobic, summer -0.23 to -1.16 South River, N.C. estuary (5) aerobic, 0-15 C -0.32 to 0.0

aerobic, 15 C 0.0 to 0.96

Narragansett Bay estuary (6) aerobic, 4-25 C -0.29 to 1.29 Doboy Sound estuary (7) + 0.001 Pacific Ocean floor (8) aerobic 0.11

Lake Superior (9) 1600 hr experiment 2.2 x

Lake Ontario (10) aerobi c 0.008

Swedish lake (11) eutraphic, 4-15 C -2.08 to 2.08

, Danish lakes (12) aerobi c, cool season -0.04 to -0.29

aerobic, warm season 0.13 to 0.62 eutrophic, aerobic -0.06 to -0.08 eutrophic, anaerobic 0.51 to 0.72 dystrophic, aerobic 0.01

oligotrophic, aerobic 0.03 Lake microcosms with chironomids ( 13) aerobi c 0.01 to 0.39

Lake Kinneret (14) anaerobic 0.03 1) this study; (2) Holmes, 1977; (3) Banoub, 1975a; (4) Nixon &. aJ., 1976;((5) Fisher and Carlson, unpublished data; (6) Hale, 1975; (7) Pomeroy et. a1 . , 1965; (8) Hartwig, 1974; (9) Glass and Poldoski, 1974; (10) Bannermi% et. a1 . , 1975; (11) Bengtsson, 1974; (12) Kamp-Nielsen, 1974, 1975; (13) Gallcp,3979; (14) Serruya -et. -al., 1974. mineral soil, subdivided into surface (0-5 cm) and subsurface layers (6-25 cm). Standing stocks of phosphorus and flows of phosphorus within the ecosystem and in ecosystem inputs and outputs were quantified and, where possible, were divided into the appropriate phosphorus fraction for W.Y. 1977 and 1978 (Table 43). The overall model is presented in Figure 40. A1 1 standing stocks and fluxes are based on the 3.2 km2 area of the swamp floodplain ecosystem (Fig. 2).

The source of phosphorus to the ecosystem was hydrologic. Surface water inputs (Table 29) of phosphorus ranged from 980 - 1200 mg ~=m-~*~r-l and were primarily in the form of FRP and PP (Table 43). Inputs of large particulate phosphorus (LPP), were small (< 1%) compared to total surface water inputs of phosphorus. Phosphorus inputs in bulk precipitation were 6-7% of surface water inputs, half of which entered floodwaters and half of which fell on the dry forest floor (Fig. 40). Total inputs to the swamp ecosystem ranged from 1050 - 1290 mg ~*m-2*~r-land were slightly greater in W.Y. 1978 than in W.Y. 1977.

Outputs from the ecosystem were also hydrologic, occurring almost entirely in surface water exiting at CP-10 (Table 43). The loss of phosphorus as LPP remained quite small compared to other surface water fractions of phosphorus but increased slightly over LPP imports. Overall, total exports ranged from 560 - 735 mg ~.m-z.yr-l and were smaller than total imports. Total exports also decreased from W.Y. 1977 to W.Y. 1978 in contrast with the increase in imports over the same period. This resulted in a two-fold increase in the amount of total phosphorus retained by the ecosystem from W.Y. 1977 to W.Y. 1978. In W.Y. 1977, FRP and PP comprised the largest fraction of total exports, whereas in W.Y. 1978, FUP and PP were the major components. This was due to a nearly five-fold decrease in FRP exports and nearly a three-fold increase in FUP exports from the swamp between the two years. Lessening of the influence of the hog farm upstream of TB-02 (imports of FRP in W. Y. 1978 were about ha1 f of imports in W.Y. 1977) probably lowered FRP exports. Runoff in W.Y. 1978(Fig. 29,30)was almost three times greater than in W.Y. 1977; if water volume-dependent leaching controlled FUP release from the forest floor, then the dramatic increase in FUP release in W.Y. 1978 could be explained by the increased volume of water leaving the swamp ecosystem in that year.

During both water years, there was a net export of FUP from the swamp ecosystem (Table 43). In W.Y. 1977, there was a small net export of PP from the swamp. In W.Y. 1978, the trend was reversed with a net retention of PP by the ecosystem. This appeared to be due to greatly increased imports of PP during this relatively wet year. Even with greatly increased imports and runoff, the swamp ecosystem maintained a relatively constant level of PP exports. The potential damage of increased erosion and increased suspended sediment concentrations in surface waters may have been damped out by the swamp ecosystem.

Phosphorus cycling in the Creeping Swamp floodplain ecosystem was

Table 43 continued 11. STANDING STOCKS IN THE ECOSYSTEM TOTAL PHOSPHORUS COMPONENT 1977 -- 1978 I. Above-ground living biomass: Trees and saplings: Leaves Wood TOTAL Herbaceous vegetation, Shrubs, Vines Bryophytes 8.4 A1 gae 0-400 AVERAGE TOTAL 5490 2. Forest Floor and Soil : Ground Litter 536 3 68 Surface Soil (0-5 cm) Subsurface Soil (6-25 cm) AVERAGE TOTAL

111. INTRASYSTEM CYCLING REACTIVE P TOTAL P FLUX 1977 1978 1977 1978 Canopy Return: Litterfall Branchfall TOTAL Litterfall Throughfall Stemflow TOTAL CANOPY RETURN Tree Wood Increment Estimated Root Uptake (Wood Increment + TOTAL Canopy Return) Floodwater Losses to the Forest Floor Sedimentation FRP Uptake by Algae FRP Uptake by Algae and Forest Floor FUP Release by Forest Floor WATERS (0-9.5) I SURFACE WATERS

FUP 61.5- 185 FUP 0-3.8 4 FUP 71.7-201

FRP 0-1.9 n FRP 315 -616 FRP 77.2-382

TOTAL 1 TOTAL INPUTS = 1 OUTPUTS = 1 ' 1050- 1290 I 559- 736

SUBSURFACE SOIL (6-25cm) C

GROUND WATER 1 I ,DEEP LOSS INFLOW 0.32

Figure 40. A phosphorus budget for the Creeping Swamp floodplain ecosystem. Standing stocks are in units of mg porn-2 and fluxes are in mg ~*rn-Z-~r-l.Any expression in parentheses was estimated by difference. Sizes of components do not represent an accurate relationship according to standing stocks because of the magni tude of fluxes between compartments. dominated by fluxes occurring through the floodwaters (Fig. 40). Standing stocks of phosphorus in floodwaters, calculated from average concentrations at C 14 Fig. 31) and volumes presented in Table 19, varied from 0 mgm2 P-m"-wheh there was no water in the swamp to an average of 9.5 mg P-m when the entire floodplain was inundated. The small size of the floodwater phosphorus compartment indicated that the residence times of surface water inputs in the floodwaters must have been extremely short. Major pathways leading to and from the floodwaters were those of imports and exports and transfers between the floodwaters and the forest floor. In W.Y. 1978, a measured net transfer of 172 mg ~~*m-~-~r-l occurred from floodwaters to the forest floor; this was slightly greater than half the estimated amount of surface water PP retained by the swamp ecosystem (Table 43). As mentioned earlier, the greatest contributor of PP, TB-02, was located downstream of the sedimentation measurement site. This may account for the underestimate. The single greatest transfer of phosphorus from the floodwaters to the forest floor was in the form of FRP and most of the transfer could be attributed to algal uptake. Measurements of the uptake' of the total forest floor were subject to error and were probably greater than what actually occurred due to methodological problems. Algal uptake estimates were conservative, and suggest that algal dynamics were very important in phosphorus fluxes between the forest floor and floodwaters. However, algae were present only during the winter and early spring when the floodplain was inundated. As a result, the phosphorus accumulated by the algae must have been released back to the floodwaters and forest floor or lost downstream as LPP. The LPP algal loss was incorporated into the LPP estimated exports and w s very small. Based on an annual algal productivity of 16.5 g C-m-' (Mu1 hol land 1979) and the2Redfield (1958) ratio of 106 C: 1 P in algal biomass, about 400 mg Porn- was incorporated into algal biomass during a bloom period. The difference (700 mg P-m-2) between uptake and incorporation was probably returne to the floodwaters either as FRP or FUP (Fig. 40). Because the use of 9P as a tracer of phosphorus uptake at low or limiting concentrations results in gross uptake rates, it is conceivable that a large amount of the phosphorus taken up by the algae was returned to the surface waters. The distribution of the returned phosphorus into FUP and FRP is not known; the distribution in Figure 40 is for the sake of illustration only. With reference to the algae and the total forest floor, about 3-5 times more FRP was removed from the floodwaters (Fig. 40) than was retained by the swamp as a whole over an annual period (Table 43). This suggests that considerable recycling occurred between the total forest floor and the floodwaters. In the spring, leafing out of the tree canopy and drawdown of floodwaters resulted in the senescence and death of the algal gloom. Upon decomposition of the a1 gae on the floodplain, mineral ized phosphorus was released either to floodwaters or was leached into the forest floor. Typically, the floodplain was drying during these periods so-that the latter more likely occurred. The net result of the whole process was that the filamentous algae acted as a trap and temporary storage of phosphorus brought in with winter floodwaters until the onset of the growing season. Algal biomass was probably richer in phosphorus and more easily broken down into its mineral elements than the leaf litter of the forest floor. Therefore, decomposition of a1 gal biomass and subsequent re1ease of mineral ized phosphorus wou1 d coordinate we1 1 with the increased mineral needs of the canopy during leaf-out. The spring wildflower, Erythronium americanum, has been observed to play a similar role in the Hubbard Brook -w'ller and Bormann 1976).

The presence of algae in stream waters has been shown to influence dissolved phosphorus dynamics in stream ecosystems. Studies of the trans- 1ocation of phosphorus in wood1 and stream ecosystems have demonstrated that phosphorus added to stream waters is quickly removed by the benthos and that periphyton associated with stream bottoms are usually responsible for a major portion of the removal (Davis and Foster 1958; Garder and Sku1 berg 1966; Nelson, --et a1 . 1969; Elwood and Nelson 1972). In contrast, however, Meyer (1979) found that the inorganic sediments of a mountain stream removed most of the phosphorus added to the water. Little algae was associated with this stream bottom, however. Periphytic or filamentous algae increase biomass and phosphorus uptake with increasing stream water velocity (Whitford 1960; Whitford and Schumacher 1964; Schumacher and Whi tford 1966; McIntire 1966), implying greater phosphorus removal under flowing water conditions. These studies concur with the estimated importance of the filamentous algae bloom in phosphorus removal in Creeping Swamp.

Estimated FUP releases from the forest floor were 10-20 times greater than the net release of FUP from the ecosystem in downstream surface waters. FUP release measurements were based on changes in FUP concentrations in water overlying the swamp floor over periods of time and are therefore, not subject to the same criticism as FRP estimates using 32~. This implies that the FUP fraction was also recycled between the forest floor and floodwaters. Keup, et al. (1970) found that leaching of phosphorus from sediments associatawith slow moving Coastal Plain rivers occurred primarily from the leaf litter layer of swamp floors. The soil layer contributed little phosphorus to overlying waters. In Creeping Swamp, the source of FUP is also likely to be the leaf litter layer because the soils were so low in organic matter. However, on the whole, net leaching was not large, implying recycling or precipitation as suggested by Meyer ( 1979).

Apart from floodwater-forest floor exchanges, phosphorus also cycled between forest floor and vegetation (Fig. 40). In relation to fluxes at the forest floor-water interface and between the forest floor and the vegetation, the standing stocks of phosphorus in the forest floor were quite small. Comparing simply inputs from litterfall and standing stocks of phosphcrus in the ground 1 itter suggests a maximum turnover time of a year or a little more for leaf phosphorus. Therefore, in a year's time about 300-400 mg P-m-2 should be mineralized and available for root uptake in the upper layers of the mineral soil. Such tight coupling between canopy return, decomposition and root uptake is further indicated by the abundant growth of surface roots observed during the growing season in Creeping Swamp. These roots invade the ground litter and thus may quickly absorb any phosphorus released through decomposition. Similar functioning has been observed in a tropical rain forest ecosystem (Stark and Jordan 1978). Root uptake (530 mg ~*m-~-~r-')by the canopy was estimated by summing the amount of phosphorus in annual above-ground wood increment and in canopy return. Root uptake was about 10% sf the standing stock of phosphorus in above-ground trees and sap1 ings but represented a loss of less than 10% of the phosphorus present in the top 5 cm of soil layer. Overall, the soil was the largest reservoir of phosphorus in the ecosystem. Phosphorus in above-ground biomass of trees and saplings was the second largest pool in the ecosystem. The contribution of the herbaceous vegetation, shrubs, vines and bryophytes to the phosphorus cycle of the ecosystem was very small due to their low combined productivity (Mu1 holland 1979) and small stocks of phosphorus. In conclusion, phosphorus was retained by the swamp ecosystem and cycling was dominated by transfers to and from the swamp floodwaters. Measured rates implied that recycling and rapid turnover of phosphorus in the floodwaters and forest Floor were predominant features. The importance of the floodwater-forest floor interaction is further accen- tuated by the seasonality of the interchange. Typically, the forest floor was inundated during the cool season when the vegetation was dormant and heterotrophic processes dominated. During this period, a filamentous algae bloom appeared to serve as a temporary storage for stream water phosphorus which may have been incorporated into vegetation at the onset of the growing season. In sharp contrast to the magnitude of fluxes occurring between them, the stocks of phosphorus in the floodwaters and forest floor were small. The major reservoir of phosphorus in the ecosystem was in the mineral soil. Net inputs to the ecosystem ranged from 0.8-1.6% of the amount of phosphorus in the soil. However, because these inputs were carried in by surface waters, they were probably quickly transferred from one compartment to another over the period of a year. The large stock of soil phosphorus in relation to root uptake indicated a long turnover time of that compartment. The net increment of phosphorus in above-ground wood was less than the amount of phosphorus retained by the ecosystem. The remainder may have been in new root biomass or ultimately stored in the soil. The forest floor with its relatively low stocks of phosphorus and high rate of turnover certainly does not represent a reservoir of phosphorus as it does in some other wetlands (Crisp 1966; Novitski 1978; Richardson, --et al. 1978). Comparison to other wetland and forested ecosystems:

A. Wetlands comparison

Phosphorus cycling in different ecosystems can be compared most simply by examining ecosystem inputs and outputs to determine whether net import or export of phosphorus occurs over an annual period. Knowledge of specific structural and functional characteristics of the ecosystems may faci 1i- tate explanation of differences between ecosystems. Such comparisons can be made at two levels: the level of the ecosystem and the level of the watershed, with careful consideration being given to areal differences among ecosystems and among watersheds.

Although most wetland ecosystems appear to retain a large portion of their inputs of phosphorus over an annual period (Table 44), some tidally influenced brackish and salt marshes export net amounts of phosphorus in the range of 190-340 mg porn-2*yr-1. Annual inputs and exports from floodplain and riverine swamps and freshwater, brackish, and saline marshes are large compared to fluxes in bogs. Bogs are restricted with respect to both water flow and phosphorus fluxes, whereas swamps and marshes are more open hydrologically and thus potentially chemically as well. Imports and exports measured in the Creeping Swamp ecosystem are in the same general range of those of other swamps and freshwater and brackish marshes with the exception of the Cache River swamp (Mitsch, et al. 1977). The large fluxes in this swamp resulted from a single large flood- of the Cache River during which very large amounts of phosphorus flowed through the ecosystem. This similarity among ecosystems is remarkable in view of the differing hydrologic regimes, vegetation, and sediment characteristics of the ecosystems studied. Those wet1 and ecosystems where the source of phosphorus is restricted to precipitation (bogs) have 1ow imports and exports of phosphorus (Tab1 e 44) and accumulate significant portions of the inputs within the ecosystem. Swamps and marshes which receive phosphorus primarily in surface water inputs, are characterized by much larger inputs and outputs. Losses also occur in surface waters. These ecosystems, for the most part, appear capable of retaining much greater amounts of phosphorus than do bogs; this may simply be the result of greater supply because exports are correspondingly larger as well. Exceptions are the salt marshes which appear to export phosphorus over an annual cycle. There are large stores of phosphorus in the sediments of salt marshes (Pomeroy, --et al. 1969) which may be more susceptible to loss because of the seasonal increases of sulfide concentrations in the ecosystem. Sulfide precipitates out ferrous iron, thus removing the iron which may be active in binding phosphorus in the sediments. With rising sea level, it is likely that previously fresh and brackish marshes have become more saline. The increased salinity, and therefore sulfate concentrations, may make the phosphorus in the sediments more vulnerable to loss. Table 44. Comparison of phosphorus inputs and outputs in wetlands, upland forests and agricultural watersheds.

DRAINAGE INPUT EXPORT.. SYSTEM AREA (km2) (mg*m-2*yr" (tng*m-2ayt--1 - ) ) I. Wetland ecosystems Floodplain Swamps (1) 3.2 984-1 1 30 560-735 Riverine Swamps: Louisiana (2) Illinois (3) -Bos: Minnesota (4) Britain (51

Freshwater Marshes Wisconsin (7) pol 1 uted Brackish Marshes: Maryland (8) 0.06-0.13 Maryland (9) 0.13 Virqinia (10) 0.14

Saline Marshes Flax Pond, N.Y. (11) 0.54 South Carolina (12) I I. Up1 and Watersheds Watersheds containing swamps Creeping Swamp, CP-10 (1) 80 89-1 08 22-29 CP-20 (1) 32 60-79 5-1 5 Palmetto Swamp (13) 54 55-66 5-1 5 Channelized swamps (13) 56-1 76 55-66 13-31 Temperate forests, hardwoods Hubbard Brook (14) 0.1-0.4 Walker Branch, TN (15) 1 .O Rhode River, MD (16) 0.3-2.5 Coweeta, NC (17) 0.1-0.6 Ontario (18) 3-73 Temperate--Agricultural Rhode River. MD (15). , Chowan, NC (19) 9-98 Florida (20) 6.1 Tropical Forests Rio Negro, Brazil (21) Panama (22) 259 *net fluxes for ecosystem. (1) this study; (2) Day, et al. 1976; (3) Mitsch, et al. 1977; (4) Verry 1975; (5) Crisp 1966; (6) Burke 1975;77T~rentki,et a1 . 1978r - (8) Stevenson, et al. 1976; (9) Heinle and Flemer 1976; (10) Axelrad, et a. 1976; (11) Woodwell and Whitney 1977; (12) Settlemeyre and Gardner 1975; (13F~uenzler,e_t c. 1977; (14) Likens, et a1. 1977; (15) Henderson, et al. 1977b; (16) Correll, et al. 1977; (17) Swank and Douglass 1977; (18) Dillon and Kirshner 1975; (19) Overcash, et a1. '1977; (20) Burton, et a?. 1975; (21) Ungemach 1967,1970; (22) Golley, et al. 1975. There is not a wide enough latitudinal spread in the wetlands that have been studied to allow a quantitative assessment of the effects of climate on phosphorus cycling. Within the scope of the studies presented in Table 44 outputs and retention capabi 1 i ties appear closely re1 ated to the hydrologic character of the ecosystems. Net retention of phosphorus by wetland ecosystems suggests that these ecosystems are not at steady-state but rather that they are accumulating biomass (Vitousek 1977). Such retention and the age of the Creeping Swamp forest (40-50 years) indicated that the ecosystem had not yet reached maturity. However, retention (or export) of the different phosphorus forms varied widely from year to year (Table 43). These wide variations and the error associated with the measurements suggest that an accurate assessment of both phosphorus retention and ecosystem maturity cannot be made unti 1 after 10-15 years of measurements (cf. Bormann, --et a1 . 1979). B. Watersheds comparison Watershed-based inputs and outputs of the Creeping Swamp watershed were ,on the average, slightly higher than values measured in other forested, primarily upland, watersheds. For the sake of this comparison, fertilizer inputs and crop harvest outputs (Table 40) were omitted because in most studies precipitation and surface water fluxes were the measured variables. The watershed-corrected TB-02 excess (Table 40) was included in inputs because it represented a net release to surface waters in the watershed. Completely forested and undisturbed watersheds (Hubbard Brook, Wal ker Branch, Coweeta , Ontario, Rio Negro, Panama) a7 1 exhibited very 1ow 1eve1 s of phosphorus exports. Mi th one exception (Wal ker Branch), inputs in bulk precipitation were quite small as well. Exports from the relatively undisturbed watershed upstream of CP-20 and the Palmetto Swamp watershed were similar to the low levels measured leaving upland forest watersheds. In watersheds drained by channelized swamp streams and other watersheds having some measure of human disturbance (Creeping Swamp at CP-10, Rhode River, Chowan and Florida), imports and exports were elevated. A1 1 watersheds, however, exhibited net retention of phosphorus. From this comparison, the effect of land-use changes in the Creeping Swamp watershed (at CP-10) become evident. It is also very clear that the Creeping Swamp watershed, as well as the others, are capable of lessening the influence of agriculture and ditching on phosphorus concentrations in downstream waters. In view of the typical phosphorus-1 imited nature of freshwaters, this functioning is extremely important in the maintenance of the quality of streams and rivers. The effect of land-use changes in watersheds does not appear to be completely ameliorated; therefore, monitoring of the effects of these changes in wooded watersheds should continue.

4. SEASONAL CHANGES IN FORMS AND SPECIES OF IRON AND MANGANESE IN SWAMP WATER AND SOILS bY Leonard A. Smock

INTRODUCTION

The physical and chemical characteristics of freshwater swamps are such that the forms of Fe and Mn may differ from the forms found in more typical freshwater systems. In particular, the high acidity, high concentrations of organic matter, and intermittent flooding with periods of low dissolved oxygen especially within waterlogged soils can have a wide range of effects on the concentrations, species, forms and export of Fe and Mn from a swamp. This in turn can affect the biogeochemical cycling of other metals and nutrients in the swamp, including in particular phosphorus. The species, or oxidation state, in which Fe exists is important in determining the reactions into which it will enter. The ferric species predominates in oxidizing water and unflooded swamp soils, occurring mainly as either ferric oxide or oxyhydroxide coatings on particulate matter or as a discrete, colloidal, oxide particle such as Fe(OH)3 (Bear 1964; Jenne 1968). Through adsorption and coprecipi tation reactions , these oxides (and also those of Mn) are a major fxtor controlling the concentrations and movement of trace metal s in soi l s and water (Jenne 1968). The compl exation or precipitation of ferric Fe with phosphorus can also be a major factor regulating the availability of the phosphate ion (Mortimer 1941; Hutchinson 1957; Stumm and Morgan 1970). Concentrations of uncomplexed, ionic Fe are normally quite low in natural waters. However, soluble ferric concentrations considerably higher than those predicted from solubility equilibria are often found due to the formation of soluble ferric - organic matter complexes. Organic acids have functional carboxyl, phenolic and hydroxyl groups which complex with the Fe ion. Besides increasing soluble Fe concentrations, these complexes inhibit both the formation of ferric oxides and the reactions of Fe with phosphorus, thereby affecting the concentrations, reactivity and transport of Fe and Fe-associated trace metals and phosphorus. Under reducing or low pH conditions the ferric ion is converted to the more soluble ferrous species. This causes the dissolution of inorganic ferric oxides and precipitates, thereby re1 easing associated trace metals or phosphorus. This increases the residence time of the Fe, trace metals, and phosphorus in the water column, increasing their bio-availabil ity and il ity of their transport from the swamp (Verry 1975). Besides occurring in the ionic state, ferrous Fe also occurs adsorbed to particulate matter, including expecially colloids (Stumm and Morgan 1970), or complexed with colloidal organic matter. Complexation of colloidal organic matter with the ferrous ion occurs either through a direct complexation reaction or through reduction of ferric Fe previously complexed with organic matter (Bloomfield 1952; Hem 1960; Shapiro 1966).

Changes in the forms or species of Fe can be important in terms of phosphorus cycling and availability. Precipitation or complexation of phosphorus with ferric Fe reduces the concentration of reactive phosphorus in the water column through either sedimentation to the swamp floor or through its association with suspended particulate matter. Reduction of ferric Fe, particularly in water-logged soils, re1eases phosphorus, increasing its concentration in the water column. Complexation of Fe with colloidal organic matter retards formation of Fe - phosphorus associations, thereby a1 lowing more PO4-P to remain biological ly available (Koenings 1976).

Mn can also affect the forms and concentrations of phosphorus, particularly through the formation of insoluble precipitates or complexes with the manganous ion (Stumm and Morgan 1970). Mn is somewhat more soluble than Fe, being solubilized sooner than Fe as the water becomes increasingly more reducing. In low pH waters, it exists primarily as the free, divalent manganous ion (Gotoh and Patrick 1972). It also commonly occurs associated with particulate matter either as a Mn oxide or copre- cipitated with Fe oxides. Colloidal organic matter - Mn complexes also occur (Hem 1965; Zaj eci ko and Pojasek 1976). A previous study showed that the amounts of total Fe and Mn in the water of Creeping Swamp varied seasonally, being generally more concentrated in summer when the stream was flowing very slowly or was stagnant (Kuenzler, et a1 . 1977). Soluble Fe was significantly correlated (Pi0.002) with the soluble species of Ca, MCJ, Na, K, and Si at the upstream station (CP-20), and with total organic N, with filterable unreactive P, and with soluble Mn at the downstream station (CP-10). Soluble Mn was correlated with the soluble species of Ca, Mg, Na, K, and Si at both the upstream and downstream stations. Soluble Fe also increased with color and decreased with dissolved oxygen at both stations (Kuenzler, --et a1 . 1977), as was soluble Mn at the downstream station. These correlations suggest important biological and chemical interactions in swamp waters and soils.

In order to better understand the potential effects af Fe and Mn on phosphorus cycling, a study of the species and forms of Fe and Mn in the swamp was undertaken. Concentrations of both ferrous and ferric Fe and of Mn in swamp water were measured and their concentrations in particulate, colloidal and ionic forms were determined. Of particular interest was the differentiation of organically and inorganically bound forms of colloidal Fe and Mn. Given the potential significance of soils and suspended particulate matter on the cycling of Fe and Mn, the concentrations of Fe and Mn in the exchangeable, crystalline mineral, oxide and organically-bound forms associated with particulate matter were measured. Data on the forms of phosphorus and organic carbon were collected on the same water samples from which Fe and Mn data were obtained. These data were used in an attempt to examine interrelationships and controls on the cycling of these elements.

METHODS

Sampling sites and schedule

Samples were collected from three sites within Creeping Swamp over a six-month period. These included samples from the stream channel at both CP-14 and CP-10 and samples from the floodplain at CP-14. Seven sets of samples were collected, beginning in October 1978 and continuing to May 1979. These samples encompassed several distinct hydro1ogic perf ods. During the October sampling period, the only water in the swamp was in shallow stream pools. Three sets of samples were collected during periods of flooding: late November, February and April. Low to moderate water levels occurred in the floodplain during sampling in January and in mid and late March.

Water sampl es

Collectiorl and particle size separation:

Any sampling program which attempts to examine the s ecies of Fe in water must take into a count the unstable nature of Fe5+ in the presence of oxygen. Contact of Fe5+ with oxygen during either sample collection or processing (e. g. fi1 tration) can result in considerable underestimation of ~e~+(Troup, et a1 . 1974). Aerobic filtration of even oxygenated water can also rzunin shifts of the chemical form of Fe, causing, for example, a decrease of ionic Fe and an increase in inorganic colloidal Fe (Koenings 1976). In order to circumvent these potential changes, all sample collections and filtrations were performed with equipment purged of oxygen with nitrogen gas,

Water samples were collected with a mouth-operated suction device into one-1 iter polyethylene bottles fi1 led with nitrogen gas. Samples were collected from the mid-depth of the water column at each station and from within the leaf litter at the CP-14 floodplain. Most samples were collected in duplicate. All sample bottles had been cleaned with detergent, 10% HC1 and deionized water and were wrapped with tinfoil to reduce potential photoreduction of ~e3+to Fezt. Samples were transported to the laboratory on ice. Further processing was begun within 8 hours of sample collection. Subsamples were removed and placed into oxygen-free 100 ml polyethylene bottles for future determination of total Fe, Mn and P concentrations, Subsamples for total organic carbon analysis were placed into glass bottles. In the laboratory, further subsamples from the one-liter bottles were membrane filtered (<0.45 pm) in a nitrogen atmosphere to obtain the particulate fractions. Millipore filters were used through the sampling of February 28, 1979, after which Nuclepore filters were used due to their better particle size selectivity urnis is on 1975). A1 1 filters had previously been washed with 50 m1 of 10% HC1 and 250 ml of deionized water, dried at 60 C for eight hours, dessicated, and then weighed. The filters were saved for later determination of suspended solids and associated Fe and Mn concentrations. A portion of the filtrate was placed into either oxygen-free polyethylene or glass bottles for future determination of filterable Fe, Mn, P and C concentrations. The remaining filtrate was then filtered through Millipore Immersible Molecular Separators (Pel 1 icon Type PTGC) which retain molecules with a nominal molecular weight of 10,000. All sample containers and tubing used for this procedure were filled with nitrogen gas prior to use. The resulting f i 1 trate was placed into oxygen-free polyethylene or glass bottles for determination of dialyzable Fe, Mn, P and organic C concentrations. This series of size separations a1 lowed analysis of total, filterable (<0.45 pm) and dialyzable (<10,000molecular weight or <3.2 nm) Fe, Mn, P and organic C concentrations. By subtraction, the particulate (total minus filterable) and co1 loidal (filterable minus dialyzable) fractions of these elements were also determined. Iron analysis: The scheme for the analysis of Fe was designed to determine the chemical forms of Fe in the particulate, colloidal and dialyzable states. The scheme was similar to that of Koenings (1976). It takes advantage of the specificity of bathophenanthroline (BPN) in complexing+only with the ferrous ion, and thus a1 lowing separation of Fez+ from Fe . Further, the procedure differentiates between reactive (to BPN) and unreactive forms of Fez+ and Fe3+. BPN-reactive Fe exists in a free dissociated (hydrated) form, as an Fe-organic acid complex, or in a weakly adsorbed form which develops a bond with BPN that is stronger than the electrostatic bond between the Fe and the sorbing particulate matter. As operationally defined by the particle size separation performed on the water samples, the BPN-reactive dialyzable Fe exists in the first form, BPN-reactive colloidal Fe exists in the second form and BPN-reactive particulate Fe exists in the last form. BPN-unreactive Fe is mainly inorganically complexed, including in particular Fe oxyhydroxides. It may also include any Fe-organic matter complexes which are stronger than the Fe-BPN complex. BPN-reactive ~e'+ was measured by adding 4 ml of BPN to a 10 ml water sample, shaking the sample for 30 sec, and then adding 4 ml of an Fe-free, 10% sodium acetate (N~cOOCH~)$01 ution. Go1 terman and Clymo (1969) gave procedures for preparation of reagents necessary for this analytical scheme. Following addition of the acetate, the pH of the sample was approximately 4.0 and color development of the BPN-Fe complex was thus at a maximum. The sample was then placed into a 125 ml separatory funnel and 10 m1 of hexanol (reagent grade) was added. The sample was shaken for 30 sec and then allowed to separate for 5 min. The bottom, water layer was run out of the separatory funnel and 10 ml of the hexanol layer was pipetted into a 50 ml flask. The solution was diluted to 25 ml with 95% ethyl alcohol (reagent grade). Its absorbance was read at 540 nm on a Klett-Sumerson colorimeter with a 4 cm light path and compared to a standard curve to give the concentration of BPN-reactive Fez+ in the sample.

The second procedure measured the total Fe2+ in the sample, including both the previously measured reactive ~e2+and a1 so the ~e2+in unreactive forms. A sufficient volume of 1 N HC1 was added to a 10 ml water sample to lower the pH to 1.5. ~xtractionof unreactive Fez+ was allowed to proceed for 10 min after which BPN and acetate were added. The pH of the sample was then raised to 4.0 with 4 N ammonium hydroxid (NH40H). Hexanol extraction and measurement of absorbance due to total Fe5+ were performed as above. This procedure was performed under very low light conditions from just prior to the addition of the BPN to just after the pH had been raised with the NH40H. McMa hon (1 967) and experiments performed during this study showed an interaction between acidity and 1 ight which inhibited color development. Different buffering capacities of the sampl e water and the water used for the standard solutions caused differences in the acidity of the solutions which were of a sufficient magnitude to differen- tially affect the color development in the samples and standards. Working in low light eliminated this problem.

The third procedure measured total reactive Fe (Fe2+ plus Fe3+). One ml of an acid-free, 10% hydroxyl ami ne hydrochloride (NH20H. HC1) solution was added to a 10 ml water sample. This reduced the reactive Fe3+ to Fez+, which was then able to complex with BPN. After 5 minutes, BPN and acetate were added and extraction and measurement of the absorbance due to the total reactive Fe was performed.

Finally, total Fe (reactive and unreactive Fez+ and Fe3+) was measured. Sufficient 1 N HC] was added to lower the pH of a 10 ml water sample to 1.5. One ml zf NH20HuHC1 was then added and the sample was heated for 15 minutes at 100 C. The sample was then cooled and its volume made up to 10 ml with deionized water. Again, in dim light, the procedures for total Fez+ color development, extraction and measurement of absorbances were followed.

Two separate blanks were necessary with each of these procedures. The colorimeter was initially zeroed with a reagent blank consisting of 10 ml of deionized water which was treated identidally to the water samples as described above. A color blank consisting of 10 ml of the water sample and all pH changing compounds (but not BPN) used in the given procedure was also used to correct for absorbance due to the color of the water (Golterman and Clymo 1969). This absorbance was subtracted from the absorbance of the actual sample measured with BPN. Standard curves were developed for each set of samples using ferrous ammonium sulfate hexahydrate ( (NH4)2S04= FeS04*6H20) according t Go1 terman and Clymo (1969). Sensitivity of the procedure was 10 pg Feel-?, similar to that reported by other investigators using this procedure. Spikes of (NH4)2SO4e FeS04-6H20 and ferric perch1 orate (Fe (C1 04)3)general ly gave recoveries of from 96 to 101%. The (NH4)2S04.FeSO4-6H20 reacted as BPN-reactive while the ~e(~10~)~fol lowing reduction, reacted as BPN-reactive ~e3". The above steps provided measurements of BPN-reactive and total Fe2+ and BPN-reactive and total Fe. BPN-unreactive ~e~+and BPN-unreacti ve total Fe concentrations were calculated by subtracting reactive concentrations from total concentrations. Ferric concentrations were then calculated by subtracting ~e~+concentrations from total Fe concentrations. By using the above outlined procedure on raw, 0.45 pm filtered water, and molecular filtered water, the BPN-reactive and unreactive forms of both ~e~+and ~e3+ in the particulate, colloidal and dialyzable forms were determined. Duplicate water samples were usually run from each station on each sampling date. The values reported are means of those duplicates.

Manganese analysis : A portion of the water samples was acidified with Ultrex HNO3 immediately following 0.45 pm filtration and molecular filtration. Mn concentrations of these samples were measured using a Perkin-Elmer 303 Atomic Absorption Spectrophotometer. Particulate (total minus f i 1 terabl e) , colloidal (filterable minus dialyzable), and dialyzable concentrations are reported. Forms of iron and manganese associated with suspended particulate matter: Analysis of the forms in which Fe and Mn were associated with suspended particulate matter was also performed. This provided information on the chemical forms and the mechanisms of transport of these two metals both within and from the swamp. The analytical scheme, adapted from Gibbs (1973), utilized a series of chemical extractions to determine the concentrations of Fe and Mn which existed in an ion-exchangeable (=adsorbed) form, as oxides and carbonates (=reducible), as sulphides and complexed with organic matter (=oxidizable), and as inert crystal 1 ine mineral (=crystal1 ine) . Caution must be exercised in interpreting the data, since separation into these chemical forms is not necessarily absolute, and these forms are defined operational ly by the separation techniques employed rather than by their actual chemical form. Furthermore, the total Fe and Mn concentrations measured by the following extraction procedures do not necessarily equal the "particulate" Fe and Mn concentrations calculated in the above section. Both correspond to Fe and Mn associated with suspended particles >0.45 urn in size. However, the procedures-outlined above used 1 N HC1 to ultimately determine "particulateu concentrations. In the procedures-described below, Fe and Mn associated with "suspended particulate matter" are determined using stronger and more special ized extractants. These procedures removed more Fe and Mn from the particulates, resul ting in overall higher concentrations.

The amount of solids on the filters and the concentrations of suspended sol ids in the water column were first determined. The 0.45 pm membrane filters used to remove suspended particulates from the water samples were again dried at 60 C for 8 hours and weighed. The concentrations of suspended solids was determined as the difference between the pre- and post- filtration weights of the filters divided by the volume of water filtered.

Each filter was then placed for two hours into a 50 ml flask with 15 ml of 1 M MgC12 which had been adjusted to pH 5 with NaHC03 to maintain the approxiKate pH of the water samples, Samples were hand shaken every 15 minutes during the 2-hour extraction period. After 2 hours each sample was filtered through a 0.45 pm filter. The flask was rinsed with 10 ml of MgC12 and this was also filtered. The filtrate, containing exchangeable (adsorbed) Fe and Mn, was acidified with Ultrex HN03 and transferred to a polyethylene bottle for later analysis.

The filter and retained solids were then transferred back to the flask (containing the original filter) with 10 ml of 0.3 M Na-citrate. One ml of NaHC03 was added to the sample, which was then heated to 80 C. One- tenth gram of sodium dithionite (NazS204) was stirred in and the solution was kept at 80 C for 15 rnin. At this point the Fe and Mn oxides were dissolved and complexed with the citrate. The solution was filtered through a 0.45 um filter. The flask was rinsed with 15 ml of deionized water which was also passed through the filter. The filtrate, containing the Fe and Mn which had been in the form of oxides and carbonates, was acidified with Ultrex HNO3 and transferred to a polyethylene bottle.

The third separation involved the oxidation of organic matter. The filter and solids were transferred back to the original flask with 10 rnl of water and the sample was heated to 95 C. Three drops of bromcresol green indicator solution were added to the flask, and then 0.1 N HC1 was added dropwise until the solution turned yellow, indicating pH of 5.4 or less. Ten ml of 30% hydrogen peroxide (H202) was added and the oxidation allowed to proceed at 95 C until only 5-10 ml of sample remained. Then 5 ml of Hz02 was used to rinse down the sides of the flask and the oxidation continued for 15 minutes longer. After cooling, the solution was filtered through a 0.45 urn filter. The filtrate, containing the Fe and Mn which had been complexed with organic matter or any sulphides present, was acidified with Ultrex HNO3 and transferred to a polyethylene bottle. The last step in the separation scheme involved determination of the Fe and Mn found as crystalline mineral. The four filters and any solids in the flask were washed into 100 ml Teflon beakers with deionized water. The samples were heated to dryness, cooled, and then 10 ml of 9.6 N HF and 2.5 ml of 41 N Ul trex HNO3 were added. The samples were heated to-1 l5 C until dry, afTer which the above acid additions and heating to dryness were repeated. Finally, 3 ml of HC1 and 15 ml of HNO3 were added and the sample boiled for 15 minutes to dissolve any remaining solids. After cooling, the solution was brought up to 25 ml with deionized water and transferred to polyethylene bottles. Triplicate filter blanks, processed in a manner identical to the samples, were used to correct for the Fe and Mn added to the samples from the reagents or the filters. Fe concentrations were determined using a Perkin-Elmer Model 305 Atomic Absorption Spectrophotometer equipped with an HGA-2200 Graphite Furnace. Mn concentrations were determined on a Perkin-Elmer Model 303 Atomic Absorption Spectrophotometer. Phosphorus and organic carbon analysis: Portions of the same water samples which were analyzed for Fe and Mn content were also analyzed for various fractions of P and organic C. Thus the concentrations of Fe, Mn, P and organic C are directly comparable with each other. Both total- and reactive-P concentrations in the raw, 0.45 Urn filtered, and molecular filtered water were determined by the automated stannous chloride procedure described above (Yarbro, Chapter 3 above). This provided measurements of particulate-P, col l oidal reactive- and unreactive-P, and dialyzable reactive- and unreactive-P. Organic C concentrations in unfiltered, 0.45 vm filtered, and molecular filtered water were determined using the procedure outlined in Chapter 2 above.. From this, concentrations of particulate, colloidal and dialyzable organic csrbon were determined. Fie1 d measurements Several water qua1 ity characteristics were measured in the field during each sampling period. Water temperature was measured using a field Tel-Tru thermometer. The pH of the water was determined with an Orion Research Ionalyzer, Model 407A, standardized with pH 4.01 and 7.00 buffer solutions. Duplicate water samples for measurement of dissolved oxygen (DO) were collected directly into BOD bottles with a mouth-operated suction device. The DO was measured by the Winkler method' (American Pub1 ic Health Association 1975) and reported as the mean of the two measurements. Dupl i cate water samples were a1 so col 1 ected to determine water turbidity, which was measured in the laboratory on a Hach Model 2100 Laboratory Turbidimeter calibrated with a standard formazin reference suspension. The concentrations of suspended solids in the water column were determined as part of the procedure for metal analyses, as discussed above. Soil samples and pore water

Soi 1 sampl es:

Soil cores were col lected during several sampl ing periods at different locations at CP-14 in order to determine the forms of Fe and Mn which existed in the soil and to relate the concentrations of these two metals with P and organic C concentrations. Cores were collected in October 1978 from a stream pool (CP-14) and the middle floodplain (MF). Cores collected in November 1978 and February 1979 came from the stream and the low (LF), middle and high (HF) floodplains. The distinction between the three floodplain areas was on the basis of elevation ,the low floodplain I / area being about 19 cm below the middle floodplain area and about 46 crn below the high floodplain area.

A 34 mm (I.D.) plastic corer was pushed or pounded at least 15 cm into the soil, retrieved, and the core extruded. Cores were frozen until processed further. In the laboratory, the cores were partially thawed and then sectioned according to depth. Generally, the top leaf litter, a section of amorphous organic matter from the top 1-2 cm depth, and a section of clay from the 8-9 cm depth were removed. These were dried at 60 C for 24 hours. A 0.1 to 0.2 g subsample was removed from each sample and weighed. Each subsample was then analyzed for exchangeable, oxide, organically bound, and crystalline Fe and Mn according to the procedures outlined above for suspended solids. The organic C content sf separate subsamples were determined by loss upon ignition (see page ) The total-P content of still other subsamples were determined according to the procedures described for soil P on page . Pore water:

Soil pore water was sampled using 20 mm lengths of Spectrapor dialysis tubing with a 14.6 mm diameter and a 6,000 to 8,000 molecular weight cutoff. The bags were acid washed (10% HC1) and rinsed with deionized water before being filled with distilled water and tied at each end. Triplicate sets of bags were placed within both the leaf litter and the soil at the location where the low floodplain soil cores were collected. The bags in the leaf litter were held in place by being tied to dowel rods driven into the ground. To place the bags into the soil, a trench slightly larger than the bags was dug by hand and then the bags were covered with 1-2 cm of soil, followed by leaf 1itter. This method undoubtedly disturbed the normal physical , chemical and biological characteristics of the soil. However, the bags were left in place from 28 to 43 days, a period probably sufficiently long to enable conditions to return to normal and certainly long enough to allow an equilibrium to be established between the chemical characteristics of the pore water and the water within the dialysis bags.

Upon removal from the soil or leaf litter, a syringe, with its needle evacuated of oxygen with nitrogen gas, was immediately inserted into the bag and the water drawn into the syringe. The water sample was then squeezed into a nitrogen-filled test tube with a rubber stopper for storage until processed further. Contact with oxygen was thus minimal, decreasing possible oxidation of ~e2+to ~e3+. The concentrations of BPN-reactive and unreactive Fe2+ and ~e3+and reactive P were measured in the laboratory using the procedures described above. These data provided information on the forms, concentrations and interactions of Fe and P in the swamp soil.

RESULTS

Water quality characteristics

Water temperature, pH and turbidity, and concentrations of DO and suspended solids showed temporal changes (Table 45). The lowest observed DO concentration was 3.8 mg-1-1, occurring in a small pool at CP-10 in October 1978. The water column was thus never anoxic during any sampl ing period. The concentrations of suspended sol ids (>0.45 pm dia. ) reflected changes in flow patterns. Highest concentrations occurred in November, due to the initial flood of the season suspending the particulate matter which had accumulated on the swamp floor over the previous dry season. Temperature followed the normal seasonal pattern and pH was typically between 4.6 and 5.5.

Iron

The concentrations of particulate, colloidal and dislyzable BPN- reactive and unreactive ~e2+and Fe3+ varied considerably at CP-14. Concentrations and trends in the iron data at the other sampling stations were similar to those found at CP-14 except where noted in the text. In general , total Fe3' concentrations were higher than total ~e2' concentrations and total unreactive concentrations were higher than total reactive concentrations (Fig. 41, 42), Particulate Fe concentrations were usually higher in unreactive forms (Fig.41B, 42 ), but colloidal and dialyzable Fe differed ccording to the oxidation state of the Fe. There was more colloidal Fe4' in the reactive than the unreactive form (Fig. 41), while colloidal and dialyzable Fezt occurred exclusively in the unreactive state (Fig. 42). Concentrations of reactive ~e3"varied considerably less over the year than concentrations of unreactive Fe3+ (Fig. 41).

Nearly all of the reactive Fe present over the sampling period was in the ~e3+form. The only reactive ~e~~ occurred in the particulate form in the stream pools at CP-14 and CP-10 in October (Fig 42) and in the floodplain leaf litter at CP-10 in February (0.06 mg*l-1, not shown). Highest concentrations of reactive Fe3+ occurred in the October stream pools, totaling 0.32 mg.l-l at CP-14; the predominant reactive form at this time was colloidal, although particulate and dialyzable concentrations were also high (Fig. 4lA). The reactive Fe concentrations decreased with the first flooding of the swamp in November. They remained low and Table 45. Physical and chemical characteristics of the swamp water during each sampling period. Station designations CP-14 and CP-10 are for measurements made in the stream channel; LF = floodplain, LL = leaf litter, and PL = isolated pool near CP-14.

Dissolved Suspended Date Station Temperature Oxygen pH Turbidity Solids (C> (mg*l -7) (JTU) (mg 1-1 ) BPN-REACTIVE FERRIC IRON 9 e PARTICULATE o COLLOIDAL s DIALYZABLE

BPN-UNREACTIVE FERRIC IRON cb PARTICULATE o COLLOIDAL DIALYZABLE

Figure 41. Concentrations of BPN-reactive and unreactive ferric Fe in particulate, colloidal and dialyzable forms at CP-14.

3+ fairly stable until the heavy April spate, when particulate reactive Fe increased at CP-14 (Fig. 41A) and all three forms of reactive ~e3+increased at CP-10 to 10-12 mg-1-1 each (not shown).

Changes in total filterable reactive concentrations (colloidal and dialyzable) were due mainly to fluctuations in colloidal reactive Fe, the correlation coefficient between these two forms being very high (r = .96, P<.0001). Dialyzable reactive Fe exhibited a steady decline from November until mid-March, then an increase in April (Fig. 4lA). Except for the floodplain pool (PL) sampled in March, differences in reactive Fe concentrations between stations were sl ight and inconsistent. The floodplain pool had the highest observed reactive Fe concentrations, with colloidal reactive ~e3+being the predominant form (0.30 mg-1-11, and particulate (0.15 mg.1-1) and dialyzable (0.11 mgml-1) concentrations a1 so being very high (not shown).

Total unreactive Fe concentrations were general ly higher than total reactive Fe concentrations, especial ly during periods of high flow. Much of this unreactive Fe was ferric until the February (CP-10, LF and LL) or early March (CP-14) sampl ing periods (Fig. 41) when unreactive ferrous Fe predominated (Fig. 42). Particulate and dialyzable unreactive ~e3+ concentrations (Fig.41B) were usually higher than respective ~e2' concentrations (Fig. 42). Unreactive dialyzable ~e~'occurred only in the October stream pools and during the early March sanipl ing (Fig. 42), but some unreactive dialyzable ~e3+was present during most sampl ing. periods in at least low concentrations (Fig.41B). Conc ntrations of Fe in this latter form were quite high (0.19 and 0.30mg.1' 7 )in 0 tober at CP-14 and CP-10. Unreactive colloidal Fe occurred mainly as Fe5+; ~e3+concentrations were very low except in October (Fig. 418). Unreactive colloidal ~e2+concentrations general ly increased from November on, with concentrations reaching 0.18 mg*l-l at CP-10 in April (Fig. 42). this was the major reason for the predominance of unreactive ~e2+over ~e3+during the later months of the sampling program.

Manganese

The particulate, colloidal and dialyzable Mn concentrations in the water also varied markedly (Fig. 43). As with Fe, concentrations and trends were similar between all sampling stations; therefore, only the data from CP-14 are presented. Particulate Mn concentrations were usually at or below 0.01 mg*l-l except in the stream pools in October and during the heavy November spate when concentrations reached a maximum of 0.08 mg*l-l (Fig. 43). Concentrations of colloidal Mn were extremely high at both CP-14 (0.58 mg=l-1) and CP-10 (0.51 mg-l-l) in October. Colloidal Mn decreased to 0.16 mgel-1 during the November spate and thereafter was present only in very low concentrations. Dialyzable Mn was the dominant form during all sampling periods except October. Concentrations were highest in November and decreased to stable, low levels after January (Fig. 43). NGANESE @ PARTICULATE 0 COLLOlDAL 0 DIALYZABLE

Figure 43. Concentrations of Mn in the particulate, colloidal and dialyzable forms at CP-14. Forms of Fe and Mn associated with suspended particulate matt^ The concentrations of Fe associated with suspended particulate matter in the exchangeable, reducible, oxidizable and crystalline mineral forms showed large changes. Trends and concentrations were usual ly simi 1ar at all sampling stations but data are only shown for CP-14. Total suspended Fe concentrations were highes in October and November. Concentr tions at CP-14 ranged from 92 pgb1-? in January to a high of 3260 pg.lq7 in November (Fig. 44). A1 though the range in concentrations was somewhat smaller at CP-10 (not shown), it always had higher total Fe concentrations than CP-14 beginfiing with the January sampling. There was little difference in Fe concentrations between CP-14 and LF except during the November spate when Fe concentrations reached 5390 vg-l-l at the latter station. Total Fe was positively correlated with the concentration of suspended solids in the water column (r = .80, P<.0001). Suspended Fe existed mainly in the crystal 1 ine mineral form during most sampling periods Concentrations of Fe in this form at CP-14 ranged from 17 to 2860 eg-l-i in January and November, respectively. As with total Fe, the range in suspended crystalline Fe at CP-10 was somewhat less than at CP-14. Crystalline Fe usually comprised 60% or more of the total Fe in the water column at all sampling stations. Only during October, when reducible Fe concentrati~nswere extremely high, and during January, when the water was very clear, was some other form of Fe dominant in the water column. The ran e in concentrations at LF was considerably greater (25 - 5090 pg.1' I) but again the trends were similar to those at CP-14. Crystalline Fe concentrations were positively correlated with suspended solids (r = .80, P<.0001), suggesting the importance of spates on this form of Fe. This was the only form of suspended Fe (other than total Fe) which was significantly correlated (k0.05) with suspended sol ids. Most sf the suspended Fe not present as crystalline mineral was in the reducible form. Concentrations of reducible Fe were very high in the stream pools in October, reaching 1450 pg*l'l at CP-10. They were also high during the November and April spates. When the water in the swamp contained 1i ttle suspended sol ids (e.g . January), reducible Fe was the dominant form, with concentrations of approximately 70 1-1. Reduci bl e Fe concentrations were 1.3 to 2.6 times higher in the water in the leaf litter (LL) than in the overlying floodplain water (LF). This was the only consistent difference between Fe partitioning in the water at these two sampling sites. Reducible Fe constituted 28% of the suspended Fe in the highly stained floodplain pool (PL) sampled during March (not shown). Unlike crystal 1ine and reducible Fe, exchangeable and oxidizable Fe exhibited little variation In concentration over the sampling period (Fig. 44). These forms of Fe usually each comprised less than 4% of the total suspended Fe, with concentrations generally being less than 10 eg-1-1. Oxidizable Fe was usually higher in the floodplain water than at CP-14 and existed in fairly high concentrations at CP-10 (36 and 60 pg.1) in February and April. LE OXIDIZABLE CRYSTALLIN

Figure 44. Concentrations of the forms of Fe associated with suspended particulate matter >0.45 pm . at CP-14.

. 197 The amounts of Mn associated with suspended particulate matter at CP-14 a1 so varied seasonally. There were two distinct periods in total suspended particulate Mn (Fig. 45). Concentrations were very high in fa1 1, with concentrations during November ranging up to 264 ug-l-l at CP-14 and 471 ug.l-l at LF. A1 though somewhat lower in the stream pools in October, total Mn concentrations were still considerably higher than during later sampling periods when they ranged only from 1.3 to 11.5 ug.l-l at CP-14.

Nearly a1 1 of the total Mn was in the crystal1ine form during the first two sampling periods, especially in November when it comprised no less than 85% of the total Mn at any sampling station. Crystalline Mn concentrations were positively correlated with suspended sol ids (r = .66, P<.001), which were high in the turbid October and November water. Oxi- dizable Mn was almost always below the detection limit of 0.3 vg=l-l; when not detectable, one-half the detection 1imit was shown on Figure 45. Exchangeable Mn was found at low levels exc pt during October and November when concentrations ranged up to 12.2 pg=l-Pin the floodplain leaf litter during November. Also, most of the suspended Mn in the floodplain pool during March was in the exchangeable form, with a mean concentration of 36 vg-1-1. Reducible Mn was found in generally low nd uniform concentrations, with a maximum concentration of 10.8 pgol-Tbeing measured at CP-14 in October.

Phosphorus and organic carbon concentrations

Concentrations of PP, CoRP, COUP, DiRP and DiUP at CP-14 also tended to be low in winter (Fig. 46). These values are similar to those found throughout the in-depth study on P described earlier (Yarbro, this volume). Station-to-station differences over the sampling period were minimal.

Concentrations of particulate, colloidal, and dialyzable organic carbon at CP-14 are shown in Figure 47. Of the filterable (~0.45urn) organic matter, most was dialyzable rather than colloidal. In the highly stained floodplain pool in Oct ber, however, 68% of the filterable organic matter was colloidal (26 mg-1' P ). Partitioning of organic carbon between the three size fractions at the other sampling stations was similar to that found at CP-14. Pore water

Iron and phosphorus levels in soil pore water were determined from dialysi s bags. Both total BPN-reactive Fe and BPN-reactive ~e2tconcentra- tions increased from January through April (Table 46). Total BPN-reactive

Fe concentrations reached 2.9 ug*l-l in April, with 2.0 f1~0l-l of this being BPN-reactive ~e2t. BPN-reactive ~e3' concentrations varied over the four sampling periods, but were highest during periods of flooding. Reactive phosphorus concentrations did not show a distinct seasonal trend, but may have varied with periods of flooding in the swamp. Reactive P concentrations were highest during January and March (low flow) and were lowest during February and April (high flow) (Table 46). CHANGEA DUClBLE A E 10 INE

Figure 45. Concentrations of the forms of Mn associated with suspended particulate matter >0.45 pm at CP-14. PHOSPHORUS PARTICULATE CQLLOlDAb REACTIVE COLLOIDAL UNREACTIVE DIALYZABLE REACTIVE DIALYZABLE UNREACTIVE t

Figure 46. Concentrations of the forms of phosphorus in the water at CP-14.

Table 46. Concentrations of dialyzable Fe (mg.l-') and P (~~01-')in soil pore water from the low floodplain. ------BPN-Reacti ve Fe Reactive P Date ~e3+ ~e2+

Forms of Fe and Mn associated with swamp soil<

The concentrations of exchangeable, reducible, oxidizable, and crystalline mineral Fe and Mn associated with swamp soils are tabulated according to the three sampling dates, location of sampling within the swamp (stream channel or low, medium, or high floodplain), and two or three depths from each soil core. The dominant form of Fe in a11 samples was crystalline mineral (Table 47). Much of the remaining Fe existed as Fe oxides (reducible form), while the smallest fraction of Fe usual ly occurred in the exchangeable form. The more reactive forms of Fe (exchangeable, reduci ble, oxidizable) had their highest concentrations in the top 1-2 cm of the soil. This was especially true for the Fe oxides, suggesting long-term leaching and upward movement of the Fe from the deeper soils under reducing conditions, with deposition of the oxides occurring under the prevailing oxidizing conditions at the soil-water interface. Reducing conditions in the top-most soil probably did not exist during any of the soil sampling periods due to the flow-through nature of the swamp. Periods of standing water resulting in waterlogged soils probably would have decreased the concentration of Fe oxides.

The dominant form of Mn depended on the depth in the soil (Table47 ). In the lower depths, or in the stream channel (CP-141, where the soil was mainly clay, Mn in the crystalline mineral form was predominant, Highest Mn concentrations occurred, however, in the leaf 1 itter on top of the soil, the Mn being present mainly in the reducible and exchangeable forms. As with Fe, there was a decrease in total Mn concentration with depth in the soil, suggesting a leaching and removal of the Mn from the lower soil particles. Table 47. Concentrations (,g Fe/g dry weight) of the forms of iron associated with swamp soils, concentrations (pg Mn/g dry weight) of the forms of manganese associated with swamp soils, and percent organic matter in swamp soils.

Forms of Iron* Forms of Manganese* %Organic Date Station Depth Type Exch Red Oxid Cryst Exch Red Oxid Cry= Katter (cm)

10/20/78 CP-14 1-2 C** 3 1 51 2 90 1195 106 33 < 3 70 29.3 4-5 C 9 332 32 6 90 48 41 < 1 46 23.2 MF 1-2 0** 149 31 1 21 772 4 7 94 8 76 31.3 8- 9 C 13 168 14 61 3 11 8 <1 56 11.2 11/30/78 CP-14 1-2 C 4 101 16 270 2 1 16 < 1 37 3.3 4-5 C 0 108 33 83 1 11 27 < 1 43 2.5 L F Surface L** 4 96 3 5 237 310 192 4 5 35 79.4 1-2 0 5 175 - - 822 40 25 < 1 60 26.3 8- 9 C 10 73 20 71 6 60 5

* Exch = Exchangeable; Red = Reducible; Oxid = Oxidizable; Cryst = Crystalline ** L = leaf litter; 0 = amorphous organic matter; C = clay DISCUSSION

Tot concentrations in the water column were dependent on flow, periods of flooding increasing the amount of Fe-bound suspended particulate matter. This relationship was also found by Kuenzler, --et al. (1977) from studies on changes in iron concentrations during spates. Most of this Fe was inert, crystal 1 ine mineral , but high concentrations of Fe oxides were also measured. High Fe oxide concentrations also occurred in the swamp soils. During periods of low suspended solids concentrations (e.g. January) most of the particulate-bound Fe in the water existed as Fe oxides, probably on clay particles. Very high Fe oxide concentrations were noted in the October stream-bed pools, probably as a result of a bioturbation effect. High densities of various species of fish, amphi- bians and aquatic beetles existed in these shallow stream-bed pools. Their activity at the sediment-water interface suspended oxide-rich particulates and probably also released pore water Fe into the overlying oxic waters. Under the oxidizing conditions, this Fe would have rapidly formed either as oxide coatings on particulate matter or as discrete col 1oi dal FeOOH particles (Gambrel 1 , --et a1 . 1976). Ferrous ion concentrations were higher in the floodplain pore water than were ferric ion concentrations, indicative of reducing conditions commonly found in waterlogged soils. Ferrous concentrations increased in the pore water through April as longer periods of standing water probably increased reducing conditions in the soil. Both ferrous and ferric Fe were present in the pore water, indicating that long-term, totally reducing conditions did not occur. Attempts to measure the redox potential of the pore water with a platinum electrode indicated very heterogeneous redox conditions, with oxidizing and reducing conditions occurring within centimeters of each other at the same depth in the soil. The complex nature of the physical characteristics of the swamp soi 1 and swamp hydrology apparently cause this varied redox condition, resulting in the simultaneous presence of both the ferrous and ferric ion. The only BPN-reactive forms of Fe in the swamp water throughout the duration of the study were ferric. Any reactive ferrous Fe transported from the pore water to the overlying water column must have been rapidly oxidized or transformed into an unreactive ferrous form. Reactive ferric concentrations, and especially ionic ferric Fe, were low and fairly con; stant during the sampling period, consistent with observations of other investigators showing that most variation in total stream metal concentrations is due to particulate rather than filterable concentrations. The somewhat higher concentrations of reactive ferric Fe in the October stream pools were likely due to bioturbation releasing ionic Fe from pore water . No one form of reactive ferric Fe was predominant during the sampling period. Low concentrations of particul ate reactive ferric Fe suggest that once ferric Fe becomes associated with particulate matter in oxic waters it is not readily exchangeable. The low concentrations of exchangeable Fe on the suspended particulate matter and in the soils further suggest strong binding (as oxides) of the Fe onto the particulates.

Concentrations of unreactive Fe varied more than did reactive concentrations, mainly as a result of Fe bound to particulate matter suspended during periods of high flow. Most of the unreactive particulate (>0.45 Fe occurred as ferric oxides, with little ferrous Fe being bound to particles in this size range. However, the unreactive Fe in the colloidal size-range occurred mainly as ferrous Fe, with very 1ittle colloidal- sized ferric oxide being present. This inorganic colloidal ferrous Fe was an important component of the total Fe in the water column throughout the entire sampl ing period, increasing in concentration during the 1ate spring. The exact form of this ferrous Fe is unknown, but it may represent ferrous Fe released from the soil undergoing a slow oxidation to the ferric species. The rate of oxidation to the ferric species can be very slow since in waters with a pH below 6, such as found in this swamp, oxidation rates are nearly independent of the dissolved oxygen concentration (Stumm and Lee 1961, Stumm and Morgan 1970).

Some unreactive Fe, mainly ferric, was detected in the dialyzable size fraction. Operationally, all dialyzable Fe should be ionic and thus reactive. This unreactive dialyzable Fe may have been "colloidal" Fe which was small enough to pass through the molecular filters. Another possibility is that the Fe existed originally as ionic Fe, passing through the filters, but then precipitated as unreactive Fe(OH)3. In this case, the concentrations of ionic Fe in the water column were actually somewhat higher than reported.

Most of the Mn in the water column was either dialyzable or colloidal. The sum of these two fractions, the filterable Mn concentration, ranged from 3.01 to 0.63 rncj.1-1 (Fig. 43) while suspended particulate Mn concentrations ranged from <0.01 to 0.08 mg.l-l. The ,suspended particulate Mn existed mainly as inert, crystalline Mn, a constituent of soil minerals. Concentrations of Mn present in this form were directly related to the amount of suspended solids in the water, and thus were affected by flow patterns. Increases in Mn due to increases in suspended sol ids following rains were also found in South Carolina coastal plain streams by Giesy and Briese (1978). The very high suspended particulate Mn in the October

I stream pools was due to biological activity resuspending sediments, as discussed for Fe.

I The predominant form of filterable Mn was dialyzable (except in October), suggesting that Mn was present as a free divalent, manganous ion. This is the expected form given the characteristics of the water and inorganic solubi 1i ty characteristics of Mn (Stumm and Morgan 1970; Nakhashima 1975). Concentrations of the manganous ion were low (0.01 to 0.10 mg el-]) and varied little after the November sampling. The high manganous concentrations in late November were probably due to the initial suspension and solubilization of Mn which had collected on the swamp floor over the summer and fall. Colloidal Mn was virtually undetectable during most sampling periods except in October and to a lesser extent in November. Mn in this size range could exist either complexed with organic matter or associated with colloidal, suspended particulate matter. Organo-manganese complexes have been found to be locally important in some soils (Patrick and Turner 1968). Mn is known to complex with tannic acids (Hem 1965) and fulvic acids (~ajacekand Pojasek 1976). The amount of Mn complexing with organics is inversely proportional to the molecular weight of the organic matter, with most organically complexed Mn in the South Carolina coastal plain streams examined by Giesy and Briese (1978 j being associated with organics with a molecular weight of approximately 10,000. Complexation of Mn with organic matter was not necessarily demonstrated here, however. There was no significant correlation (P>. 05) between concentrations of colloidal Mn and colloidal organic matter. Colloidal Mn was nearly undetectable (0.01 mg.l-7) in the floodplain pool (PL) in arch when the highest colloidal organic matter concentration (26 mg.1- !" ) was measured. This suggests that the colloidal Mn was actually associated with colloidal particulate matter, probably as a Mn oxide, rather than being bound in an organic complex. Supporting this contention is the observation that concentrations of suspended particulate (>. 45 vm) Mn oxides were high in the October stream pools. Furthermore, much of the Mn associated with the top-most soil particles in the swamp was present in an oxide form. It is likely that colloidal-sized particles with Mn oxides would have been suspended due to the observed bioturbation activities. The data on the concentrations of Fe, Mn, P and organic carbon during each sampling period were all determined from the same water samples and thus are directly comparable to each other. Therefore, an attempt was made to relate changes in the concentrations of the species of phosphorus to changes in Fe, Mn and organic carbon concentrations. Data from all sampling dates and stations were analyzed using a series of stepwise regressions and correlations. Each form of phosphorus (including parti - culate and both reactive and unreactive colloidal and dialyzable P) was separately included as the dependent variable in the regression analyses. All forms of ferric and ferrous Fe, Mn and organic carbon were used as independent variables in each analysis. Few meaningful regression equations or correlations resulted from the analyses. The regressions and correlations did indicate that particulate P concentrations were positively related to unreactive particulate ferrous and ferric concentrations. Kuenzler, --et al. (1977) also found a positive correlation between particulate P and particulate Fe concentrations. This is as expected, since disturbance of the swamp floor or turbid inputs from the watershed will increase the amount of suspended particulates, including particulate P and Fe. Some of this particulate P may be copre- cipitated or complexed with Fe oxides coating particulates. Furthermore, there were high correlations between crystal 1ine Fe and crystal 1 ine Mn (r = 0.91) and between total suspended forms of Fe and Mn (r = 0.89). Final ly, suspended Fe and suspended Mn were positively correlated (r = 0.80 and 0.68) with total suspended solids, mostly because of the crystalline forms of these two elements. No strong relationships between the reactive forms of P and Fe, Mn or organic carbon could be discerned from this statistical treatment of the data. It seems likely that the highly dynamic nature of the swamp, and especially the variability of its chemical characteristics caused by fluctuations in water flow, obscure interrelationships between the forms and species of these elements such as found by Koenings and Hooper (1976). Rather than using data from infre- quent grab samples, future studies should probably rely on in situ radio- isotope studies designed specifically to investigate possible =tionships between these el ements.

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