Scientific Investigations Report 2004–5171

U.S. Department of the Interior

U.S. Geological Survey

Hydrology and Cycling of Nitrogen and Phosphorus

in Little Bean Marsh: 

1996–97

Prepared in cooperation with the Missouri Department of Conservation

A Remnant Riparian Wetland

along the Missouri River in Platte County, Missouri,
Hydrology and Cycling of Nitrogen and 


Phosphorus in Little Bean Marsh: 


A Remnant Riparian Wetland along the 


Missouri River in Platte County, 


Missouri, 1996–97


By Dale W. Blevins 

Prepared in cooperation with the 

Missouri Department of Conservation 

Scientific Investigations Report 2004-5171 

U.S. Department of the Interior

U.S. Geological Survey
U.S. Department of the Interior 

Gale A. Norton, Secretary 

U.S. Geological Survey

Charles G. Groat, Director 

U.S. Geological Survey, Reston, Virginia: 2004 

For sale by U.S. Geological Survey, Information Services


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Suggested citation: 

Blevins, D.W., 2004, Hydrology and cycling of nitrogen and phosphorus in Little Bean Marsh: A remnant riparian wetland 

along the Missouri River in Platte County, Missouri, 1996–97: U.S. Geological Survey Scientific Investigations Report 

2004-5171, 80 p. 
 iii 

CONTENTS


Abstract. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1


Background and Problem . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2


Purpose and Scope . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4


Study Area Description. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5


Methods and Materials . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5


Hydrology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12


Hydrologic Inputs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12


Ground-Water Interactions. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13


Hydrologic Outputs. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15


Water Chemistry . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15


Specific Conductance. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15


Dissolved Oxygen . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21


Temperature . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21


pH . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 29


Nitrogen . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 30


Phosphorus. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 34


Turbidity and Total Suspended Solids. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 34


Dissolved Organic Carbon. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 36


Trophic Indicators. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 36


Nitrogen Cycling and Denitrification . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 38


Sinks and Sources of Nitrogen . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 39


Atmospheric Contributions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 39


Contributions from Storm Runoff. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 45


Ground-Water Contributions. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 45


Denitrification . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 47


Phosphorus Cycling . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 53


Correlations. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 53


Sinks and Sources of Phosphorus. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 55


Comparisons with the Missouri River, nearby Wetlands, and Missouri Reservoirs . . . . . . . . . . . . . . . . . . . . . . . . . . 56


Summary and Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 58


References. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 60

iv 

Figures 

1.	 Map showing location of Little Bean Marsh and monitoring sites . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .3


2.	 Graph showing mean monthly air temperatures in 1997 compared to 30-year means recorded


at the Kansas City International Airport . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .5


3.	 Graph showing monthly precipitation measured at Little Bean Marsh in 1997 compared to 


30-year means recorded at the Kansas City International Airport. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .6


4.	 Maps showing historical evolution of Little Bean Marsh . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .6


5.	 Map showing bathymetry of Little Bean Marsh . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .8


6.	 Photograph showing denitrification chambers in Little Bean Marsh. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .9


7.	 Charts showing inflow and outflow water budgets for Little Bean Marsh during 1997. . . . . . . . . . . . . . 13


8.	 Graph showing specific conductance, marsh water level, mean ground-water level, 


precipitation, and days with mean air temperatures below 0 degrees Celsius at 


Little Bean Marsh. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14


9.	 Maps showing altitude of potentiometric surface near Little Bean Marsh during 

(A) low water-table conditions on February 17, 1997, and (B) high water-table 


conditions on October 24, 1996. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16


10.	 Graph showing Missouri River stage, marsh water level, mean ground-water level, and 


precipitation at Little Bean Marsh. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18


11. Boxplots of constituent concentrations in samples collected from Little Bean Marsh in 1997 . . . . . . 22


12–22. Graphs showing:


12.	 Daily maximum and minimum dissolved-oxygen concentrations at site M3 in Little 


Bean Marsh. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 25


13.	 Dissolved-oxygen concentrations at site M3 in Little Bean Marsh during March 1997 . . . . . . . 25 

14.	 Water temperatures near the top and bottom of Little Bean Marsh at site M3 and


average daily wind speed at the Kansas City International Airport. . . . . . . . . . . . . . . . . . . . . . . . . . . 26


15.	 Daily mean air and water temperatures at site M3 in Little Bean Marsh . . . . . . . . . . . . . . . . . . . . . 27


16.	 Hourly water temperatures 0.5 meter below water surface at site M3 in Little Bean 


Marsh during 1996 and 1997 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 28


17.	 Median daily fluctuations of pH at site M3 in Little Bean Marsh, 1996–1997 . . . . . . . . . . . . . . . . . . 29


18.	 Hourly fluctuations of pH at site M3 in Little Bean Marsh during May 1997. . . . . . . . . . . . . . . . . . . 30 

19.	 Total nitrogen and dissolved ammonium nitrogen concentrations with time at Little


Bean Marsh. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 31


20.	 Chlorophyll a and nitrite plus nitrate concentrations with time in Little Bean


Marsh. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 32


21.	 Total phosphorus and soluble reactive phosphorus concentrations with time at 


Little Bean Marsh. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 35


22.	 Ratios of particulate nitrogen to particulate phosphorus with time at Little Bean


Marsh . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 37


23.	 Logarithmic plots of geometric means of chlorophyll a concentrations with (A) total


phosphorus concentrations and with (B) total nitrogen concentrations for scour 


and oxbow lakes along the Missouri River and Little Bean Marsh. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 39

24.	 Diagram showing conceptual compartments and dominant processes involving nitrogen 


in Little Bean Marsh . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 40

 v 

25–28. Graphs showing: 

25.	 Eh in bottom sediment in relation to depth at six sites in Little Bean Marsh, 1997 

and 1998. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 48 

26.	 Concentrations of nitrous oxide in denitrification chambers after 48 hours showing 

differences between chambers with ambient concentrations of nitrate and those 

spiked with nitrate. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 52 

27.	 Denitrification rates in relation to temperature in denitrification chambers spiked with 

nitrate in Little Bean Marsh . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 52 

28.	 Denitrification rate in relation to dissolved-oxygen concentration in denitrification 

chambers spiked with nitrate in Little Bean Marsh on July 28, 1997. . . . . . . . . . . . . . . . . . . . . . . . . . 52 

29.	 Diagram showing conceptual compartments and dominant processes involving phosphorus 

in Little Bean Marsh . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 54 

Tables 

1.	 Specific conductance and concentrations of selected nutrients in ground-water samples 

collected from monitoring wells near Little Bean Marsh. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11 

2.	 Sample analyses and storm loads of surface inflow to Little Bean Marsh . . . . . . . . . . . . . . . . . . . . . . . . . 19 

3.	 Analyses of surface-water samples collected from Little Bean Marsh . . . . . . . . . . . . . . . . . . . . . . . . . . . . 67 

4.	 Means and ranges of selected constituent concentrations and properties of water samples 

collected from the Missouri River, Little Bean Marsh, and other wetlands along large rivers 

of the Midwestern United States . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 33 

5.	 Concentrations of selected ions in water samples collected from Little Bean Marsh . . . . . . . . . . . . . . 36 

6.	 Correlation coefficients (r), coefficients of determination (r2), and levels of significance (p) for 

the relation between chlorophyll concentrations and concentrations of total nitrogen, total 

phosphorus, and total suspended solids in Little Bean Marsh . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 37 

7.	 Classification of the trophic state of Missouri reservoirs used by Jones and Knowlton (1993); 

May–September means of total phosphorus, total nitrogen, chlorophyll a, and total suspended 

solids for Missouri reservoirs by physiographic region (Jones and Knowlton, 1993); means 

of monthly samples of four oxbows along the Missouri River (Knowlton and Jones, 1997) 

from 1994–1996; and 1997 annual means for Little Bean Marsh. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 38 

8.	 Loads of selected constituents in surface outflows from Little Bean Marsh . . . . . . . . . . . . . . . . . . . . . . . 41 

9.	 Concentrations of selected nutrients with depth in bottom sediment cores collected from 

Little Bean Marsh in 1997. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 46 

10.	 Amounts of denitrified nitrogen generated in in-situ experiments. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 50 

11.	 Spearman’s rho values and levels of significance (p) for correlation of phosphorus with mean 

daily dissolved oxygen at five sites in Little Bean Marsh. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 55 

12.	 Spearman’s rho values and levels of significance (p) for correlation of phosphorus with 

chlorophyll concentrations at five sites in Little Bean Marsh . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 55 
vi 

Conversion Factors and Datum 


Multiply By To obtain 

Length 

centimeter (cm) 0.3937 inch (in.) 

meter (m) 3.281 foot (ft) 

kilometer (km) 0.6214 mile (mi) 

Area 

hectare (ha) 2.471 acre 

hectare (ha) 0.003861 square mile (mi2) 

Volume 

cubic meter (m3) 264.2 gallon (gal) 

Flow rate 

cubic meter per day (m3/d) 35.31 cubic foot per day (ft3/d) 

cubic meter per day (m3/d) 264.2 gallon per day (gal/d) 

kilometer per hour (km/h) 0.6214 mile per hour (mi/hr) 

Mass 

gram (g) 0.03527 ounce, avoirdupois (oz) 

kilogram (kg) 2.205 pound avoirdupois (lb) 

Density 

gram per cubic centimeter (g/cm3) 62.4220 pound per cubic foot (lb/ft3) 

Hydraulic conductivity 

meter per day (m/d) 3.281 foot per day (ft/d) 

Hydraulic gradient 

meter per kilometer (m/km) 5.27983 foot per mile (ft/mi) 

Application rate 

kilograms per hectare per year 

[(kg/ha)/yr] 0.8921 pounds per acre per year [(lb/acre)/yr] 
 vii 

Temperature in degrees Celsius (°C) may be converted to degrees Fahrenheit (°F) as follows: 

°F = (1.8 x °C) + 32 

Vertical coordinate information is referenced to the National Geodetic Vertical Datum of 1929 

(NGVD 29). 

Horizontal coordinate information is referenced to the North American Datum of 1927 (NAD 27). 

Altitude, as used in this report, refers to distance above the National Geodetic Vertical Datum of 

1929 (NGVD 29). 

Specific conductance is given in microsiemens per centimeter at 25 degrees Celsius (μS/cm at 

25°C). 

Concentrations of chemical constituents in water are given either in milligrams per liter (mg/L) or 

micrograms per liter (μg/L). 
Hydrology and Cycling of Nitrogen and Phosphorus in 


Little Bean Marsh: A Remnant Riparian Wetland along 


the Missouri River in Platte County, Missouri, 1996–97


By Dale W. Blevins 

Abstract 

The lack of concurrent water-quality and hydrologic data 

on riparian wetlands in the Midwestern United States has 

resulted in a lack of knowledge about the water-quality functions that these wetlands provide. Therefore, Little Bean Marsh, 

a remnant riparian wetland along the Missouri River, was investigated in 1996 and 1997 primarily to determine the magnitude 

and character of selected water-quality benefits that can be produced in such a wetland and to identify critical processes that 

can be managed in remnant or restored riparian wetlands for 

amelioration of water quality. 

Little Bean Marsh averages 69 hectares in size, has a maximum depth of about 1 meter, and the majority of the marsh is 

covered by macrophytes. In 1997, 41 percent of the water 

received by Little Bean Marsh was from direct precipitation, 14 

percent was from ground-water seepage, 30 percent from watershed runoff, and 15 percent was backflow from Bean Lake. 

Although, Little Bean Marsh was both a ground-water recharge 

and discharge area, discharge to the marsh was three times the 

recharge to ground water. Ground-water levels closely tracked 

marsh water levels indicating a strong hydraulic connection 

between ground water and the marsh. Reduced surface runoff 

and ground-water availability are stabilizing influences on 

marsh hydrology and probably contribute to the persistence of 

emergent vegetation. The rapid hydraulic connection between 

Little Bean Marsh and ground water indicates that the hydrologic regime of most wetlands along the lower Missouri River 

is largely a function of the altitude of the marsh bottom relative 

to the altitude of the water table. 

More water was lost from the marsh through evapotranspiration (59 percent) than all other pathways combined. This is 

partially because the transpiration process of abundant macrophytes can greatly contribute to the evapotranspiration above 

that lost from open water surfaces. Surface outflow accounted 

for 36 percent and ground-water seepage accounted for only 5 

percent of the losses. Large residence times allows the marsh to 

greatly affect water quality before water escapes as groundwater recharge or surface outflow. 

The shallowness of Little Bean Marsh and ion exclusion 

during ice formation caused the highest specific conductances 

of 1,100 to 1,300 microsiemens per centimeter at 25 degrees 

Celsius to occur during the winter. This concentration of dissolved solutes under ice can make wetlands more vulnerable to 

toxic contaminants than deeper surface-water bodies. 

Dissolved oxygen was less than 5 mg/L (milligrams per 

liter) for 3 to 4 months and near 0 mg/L for about 1 month in 

summer. Despite depths of less than 1 meter, temperature stratification persisted more than 3 months during the summers of 

1996 and 1997, preventing mixing and contributing to periods 

of anoxia. Shallow depths and extended periods of anoxia in the 

marsh limit the ability of some organisms to escape high-temperature stress. 

Turbidity in Little Bean Marsh usually was low for several 

reasons: sediment loadings from the largely flood-plain drainage were low, emergent vegetation shade out algae and shield 

the water from wind, and high concentrations of bivalent cations increase flocculation rates of inorganic suspended material. The high concentrations of bivalent cations was largely 

because of a substantial amount of ground-water seepage into 

the marsh. 

Dissolved organic nitrogen was the dominant nitrogen species in Little Bean Marsh. Denitrification and biotic uptake kept 

more than 62 percent of nitrate (NO3) and 43 percent of ammonium (NH4) concentrations in marsh samples less than a detection limit of 0.005 mg/L. This contrasts with the Missouri River 

where inorganic NO3 dominates. Consequently, artificial floodplain drainage that bypasses riparian wetlands likely deliver 

substantially more biotically available inorganic nitrogen to 

receiving waters than surface water that has been routed 

through a remnant wetland. Average total nitrogen concentrations in Little Bean Marsh were substantially less than those at 

other Missouri River wetlands, roughly one-half the mean concentrations in the Missouri River, but roughly twice the average 

nitrogen values in reservoirs of the glaciated plains of Missouri. 

The largest concentrations of nearly all species of nitrogen 

and phosphorus and the most intense period of hypereutrophy 

coincided with a phytoplankton bloom and senescence of River 

Bulrush (Scirpus fluviatilis) and common cattail (Typha latifolia) in September 1997. The rapid leaching of nitrogen that 
2 Hydrology and Cycling of Nitrogen and Phosphorus in Little Bean Marsh: A Remnant Missouri-River Wetland 

occurs soon after macrophyte senescence combined with a 

recent destratification of the marsh probably provided nitrogen 

to the nitrogen-limited open-water areas and triggered a phytoplankton bloom. Despite the rarity of runoff events, surface 

runoff from the watershed, combined with atmospheric deposition, contributed more than seven times the 530 kg (kilograms) 

of nitrogen that escaped Little Bean Marsh in surface outflow 

during 1997. Atmospheric deposition alone was more than 530 

kg. Seepage to ground water contained less than 1.5 percent of 

the nitrogen leaving the marsh in surface outflow. The slow 

decay rate of Scirpus fluviatilis and reducing conditions in bottom sediments make burial of organic nitrogen a substantial 

sink of nitrogen. 

Denitrification experiments indicate that denitrification 

rates were limited by NO3 in the water column. Consequently, 

decomposition and nitrification of NH4 and organic nitrogen 

are the rate limiting steps of nitrogen removal in Little Bean 

Marsh. The NO3-limited rates of denitrification also indicate 

that Little Bean Marsh has a large unused capacity for nitrogen 

removal. These data indicate that the vast extent of riparian 

marshes along the Missouri and Mississippi Rivers may have 

had a substantial role in limiting NO3 loads to the Gulf of Mexico before agricultural development of flood plains. Drainage 

and removal of riparian marshes may be a major cause of the 

increased NO3 loads to the Gulf of Mexico. 

Periods of anoxia had much larger effects on phosphorus 

release than the other variables. The largest concentrations of 

phosphorus occurred in late summer and corresponded with 

senescing macrophytes, periods of anoxia, and a large algal 

bloom in Little Bean Marsh. Low water levels prevented the 

escape of phosphorus in surface outflow during these periods of 

highest phosphorus concentrations. Dry weather in late summer 

is typical and probably makes the correspondence of low water 

levels, anoxia, and consequent low phosphorus release a common occurrence in marshes along the Missouri River. Little 

Bean Marsh retained more than 95 percent of the phosphorus it 

received. The amount of phosphorus in surface inflows to the 

marsh were more than one order of magnitude greater than that 

escaping in surface outflows. The long hydraulic residence time 

of the marsh and large contributions of iron from ground water 

(that provide many sorption sites for phosphorus) make the 

marsh an effective sediment and phosphorus trap. 

Background and Problem 

Little Bean Marsh is a riparian wetland located along the 

Missouri River in northwestern Missouri (fig. 1). Surface runoff 

flows through this remnant native wetland. More than one-half 

of Little Bean Marsh and some surrounding land [totaling 100 

ha (hectares)] was purchased by the Missouri Department of 

Conservation in 1981, and 61 ha of this tract was designated as 

the Little Bean Marsh Natural Area. The marsh is characterized 

by shallow and fluctuating water levels and an abundance of 

aquatic and marsh plants. 

Water-quality benefits commonly are ascribed to remnant 

native wetlands to justify their preservation or restoration. 

These benefits typically include removal or trapping of nitrogen 

(N), phosphorus (P), and sediment from surface waters; however, few limnological studies of wetlands in the Midwestern 

United States have been undertaken to document or quantify 

these benefits. Water-quality benefits, which have been demonstrated in a small number of coastal and northern wetlands 

(Mitsch and others, 1979; Kemp and Day, 1984), have been 

assumed to occur in Missouri and other parts of the Midwest. 

Also, wetland scientists and managers frequently have overstated the ability of wetlands to retain sediment and nutrients 

(Johnston, 1991). Wetland nutrient studies have a wide range of 

nutrient removal rates and processing characteristics depending 

on wetland characteristics—especially hydrology (Gosselink 

and Turner, 1978; Mitsch and Gosselink, 1993). Numerous 

studies have shown that some wetlands actually are net nutrient 

sources or nutrient transformers (Peverly, 1982; Elder, 1985; 

Baillie, 1995). Only three water-quality studies have been 

undertaken in wetlands along large rivers in Missouri, and none 

of these studies collected hydrologic data needed to explain 

many nutrient cycling phenomena. The first study was done in 

the Mississippi Embayment (Wylie and Jones, 1986) where the 

soils, vegetation, and hydrology are substantially different than 

those present elsewhere in Missouri. These factors (especially 

the altered hydrology) may substantially affect the water quality 

of wetlands. The second study (Knowlton and Jones, 1997) 

focused on nutrients in flood-plain water bodies created along 

the Missouri River by the 1993 flood. Although most of these 

water bodies were immature and too deep to be considered wetlands, four of these water bodies were remnant oxbows with 

monthly samples that are very comparable to Little Bean Marsh. 

However, few hydrologic data and no continuous data are available for these Missouri River oxbows. Hoyer and Reid (1982) 

analyzed samples from 10 seasonally flooded waterfowl refuges in December 1981, but the sites had neither natural hydrology nor vegetation, and the short sampling period limited comparisons with data collected in this study at Little Bean Marsh. 

High N concentrations in many agricultural streams of the 

Midwestern United States (Goolsby and others, 1993) have 

highlighted the need for opportunities to reduce N loads in surface runoff before they are received by larger rivers and 

streams. The primary mechanisms for removal of N by wetlands 

are plant uptake, denitrification, and volatilization of ammonia 

(NH3) from soil, plant, and water surfaces. Generally, physical 

and microbial processes are more important than vegetative 

uptake in controlling nutrient retention in wetlands (Johnston, 

1991). Because most N uptake by plants is recycled annually in 

a mature wetland and NH3 volatilization has not been demonstrated to be substantial, denitrification is considered the most 

important mechanism for N removal in remnant riparian wetlands (Brinson and others, 1984; Gale and others, 1993; 

Johnston, 1993; Hanson and others, 1994; Lowrance and others, 

1995). Denitrification has been shown to remove substantial 

nitrate (NO3) from overlying surface water (Lindau and others, 

1994) as well as from sediments. Consequently, denitrification 
Background and Problem 3 

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4 Hydrology and Cycling of Nitrogen and Phosphorus in Little Bean Marsh: A Remnant Missouri-River Wetland 

rates usually are the primary mechanism of removal of N from 

surface water by wetlands. Denitrification rates can be limited 

by temperature, pH, redox potential, NO3 availability, and 

carbon availability. Dissolved organic carbon (DOC) concentrations may be greatest soon after senescence of rushes and cattails (Klopatek, 1978) and winter temperatures may severely 

limit denitrification rates, causing seasonal patterns in concentrations of N species. Nitrate availability commonly is identified 

as limiting denitrification in native wetlands (Broderick and 

others, 1988; Cooper, 1990) because oxidizing conditions 

required for nitrification are sometimes limited (Zak and Grigal, 

1991). Therefore, alternations of oxidizing and reducing conditions or the presence of an aerobic surface zone overlying anaerobic sediments, which are common in wetlands, are conducive 

to denitrification (Brinson and others, 1984; Patrick, 1991). 

Some wetlands just transform inorganic N to organic N (Elder, 

1985) without substantially decreasing the total N export. This 

transformation also can be beneficial, however, because refractory organic N generally is not available for uptake by algae. 

Phosphorus availability often limits primary productivity 

and trophic state in surface-water bodies. Therefore, the ability 

for wetlands to sequester particulate P by sedimentation and 

burial (Mitsch, 1978; Patrick, 1991) can be important to receiving streams. Particulate P is both sorbed on particulates and precipitated as iron and aluminum compounds (Cooke, 1992). 

However, large amounts of particulate P commonly are released 

to overlying water under various biogeochemical conditions 

common to wetlands and shallow lakes. Although release of 

sorbed P is most commonly associated with anoxic conditions 

that reduce and solubilize ferric iron compounds (Mortimer, 

1941a, 1941b; Cooke and others, 1992), numerous biogeochemical conditions can cause a release of P from bottom 

sediments. In fact, aerobic release of P sometimes exceeds 

anaerobic release over an annual cycle (Bostrom and others, 

1982). High temperatures, high pH, bioturbation, macrophytes, 

wind-induced resuspension of sediment, “P saturation” of the 

sediments, and carbonaceous sediments are all factors in aerobic processes or characteristics that can release P to overlying 

aerobic water. Sondergaard and others (1992) determined that 

wind induced resuspension of sediment in a shallow eutrophic 

lake released 20 to 30 times more P than seeped from undisturbed sediments. It also is possible for denitrification to at least 

indirectly affect P in two ways: removal of NO3 allows ferric 

iron (Fe+3) to serve as an electron acceptor with subsequent P 

release, and high rates of denitrification can increase rates of 

mineralization of organic matter that can release P to the water 

column (Bostrom and others, 1988). P-mobilization processes 

in surface sediments commonly result only in a change in mineralogical species and not a release to the water column. For 

example, P mobilized by mineralization of sedimented organic 

matter may be rapidly adsorbed by iron oxyhydroxides under 

aerobic conditions (Bostrom and others, 1988). Nearly every 

process that mobilizes P in the sediment (both aerobic and 

anoxic) and every mechanism that transports P to overlying 

water is maximized in wetlands compared to deeper lakes (Bostrom and others, 1982; Baccini, 1985). Therefore, wetlands may 

not be efficient removers of P (as reported by Peverly, 1982) 

and relations between P concentrations and temperature, wind, 

invertebrate density, macrophyte populations, oxidation/reduction potential (Eh), pH, and other characteristics have implications for wetland restoration and management of water quality. 

To address the need for water-quality and hydrologic data 

to gain knowledge about nutrient cycling in remnant riparian 

wetlands, the U.S. Geological Survey (USGS), in cooperation 

with the Missouri Department of Conservation, conducted a 

study of Little Bean Marsh in 1996–97. 

Purpose and Scope 

The purpose of this report is to describe results of a study 

of hydrology and cycling of nitrogen and phosphorus in Little 

Bean Marsh. The general goals of this study were to determine 

the magnitude and character of selected water-quality benefits 

that may occur in a remnant riparian wetland along the Missouri 

River and to identify critical processes such as denitrification 

and P release that can be managed in remnant or restored riparian wetlands for water-quality improvement. Specific objectives to achieve these goals were: 

1.	 Quantify the sources and losses of water to a remnant 

riparian wetland and determine the affects of hydrologic 

variation on water quality. 

2.	 Characterize concentrations and speciation of N and P in 

a remnant riparian wetland to identify the role of the 

wetland as a sink, source, or transformer of N and P for 1 

year. 

3.	 Identify the factors limiting denitrification and determine 

whether or not denitrification rates are large enough to 

remove most of the N received by Little Bean Marsh. 

4.	 Determine whether sedimentation within the wetland 

results in a net removal of P from runoff and atmospheric 

deposition by the wetland. 

5.	 Relate P concentrations in water to various 

biogeochemical conditions such as temperature, pH, Eh, 

dissolved oxygen, macrophyte life cycles, wind, and 

external loadings to identify the primary mechanisms of 

P release and the consequent ability of the wetland to 

retain P. 

6.	 Compare N and P concentrations and speciation in a 


remnant riparian wetland with similar data from other


wetland studies.


Hydrologic and water-quality data used in this report were 

collected from November 1996 through December 1997. Sampling and measurement locations included four stream sites, six 

marsh sites, eight wells, and six denitrification sites. Continuous water-level, water-temperature, dissolved-oxygen, specific 

conductance, and pH data were collected from January 1996 

through December 1997 at one site. 
5 

Study Area Description 

The drainage area of Little Bean Marsh is 952 ha. About 

80 percent of this drainage area is bottomland in row-crop agriculture, 5 percent is the bordering loess bluffs, about 7 percent 

(69 ha) is the marsh, and the remaining 8 percent is uncultivated 

flood plain. The surrounding terrain is flat with low ridges and 

a large terrace forms the north and west boundary of the marsh. 

Discharge from Little Bean Marsh flows directly into Bean 

Lake to the south. The area is protected from small Missouri 

River floods by levees, but large floods in 1952, 1973, 1983, 

and 1993 breached levees so the marsh is not entirely isolated 

from flooding disturbances, which can prevent an ecosystem 

from forming a stable community. Bean Lake and Little Bean 

Marsh also are occasionally flooded by local runoff. The outflow from Bean Lake is controlled by a flood gate in the Missouri River levee. When the gate is closed during high river 

stage, water can back all the way into Little Bean Marsh if large 

amounts of local runoff enter Bean Lake. Stage and discharge 

from Little Bean Marsh usually are controlled by a mud riffle 

between Bean Lake and Little Bean Marsh. 

The average annual air temperature at Little Bean Marsh is 

12.0 °C (degrees Celsius) with average monthly temperatures 

ranging from -3.5 °C in January to 24.5 °C in July (National 

Oceanic and Atmospheric Administration, 1997) (fig. 2). The 

largest departure from the 30-year mean monthly temperature 

was -3.2 °C in April, and the mean annual departure was -0.2 

°C. Average annual rainfall at Kansas City International Airport 

(KCI) is 95.5 cm/yr (centimeters per year), with the greatest 

amount occurring in the summer [30.5 cm (centimeters) from 

June through August], and the least in the winter (10.1 cm from 

December through February) (Missouri Department of Natural 

Resources, 1986). Annual evaporation averages 152 cm/yr. 

During 1997, the total precipitation was 98.7 cm with 25.6 cm 

Background and Problem 

from June through August and 17.2 cm during January, February, and December. The 30-year mean monthly precipitation 

values at KCI and monthly precipitation at Little Bean Marsh 

are shown in figure 3. Although the summer was a little drier 

than average, the winter was wetter than average, and the spring 

was a little cooler than average, 1997 was not an unusually wet, 

dry, cool, or warm year. 

In 1997 about 42 ha (60 percent) of Little Bean Marsh was 

dominated by River Bulrush (Scirpus fluviatilis), common cattail (Typha latifolia), smartweed (Polygonum coccineum), and 

arrowhead (Sagittaria brevirostra). There also were about 28 ha 

of bottomland hardwood forest on areas of higher ground. 

About 40 percent of the marsh is open water with a maximum 

depth of about 1 m (meter) at typical water levels. A detailed 

description of the vegetation and natural history of Little Bean 

Marsh since 1850 is in Castaner and LaPlante (1992) (fig. 4). 

Methods and Materials 

Concurrent hydrologic and hydrochemical data were collected from November 1996 through December 1997 at six 

marsh sites (M1 to M6, fig. 1). Continuous water-level, watertemperature, dissolved-oxygen, specific-conductance, and pH 

data were collected for 2 years from January 1996 through 

December 1997 at site M3. Miscellaneous data were collected 

before and after these dates in support of several experiments. 

Sampling, measuring, and experiment sites were located along 

two flow paths to investigate changes in nutrient concentrations 

and speciation as water passed through the marsh. The first flow 

path begins with inflow sites I1 and I3 (fig. 1) at the northeast 

end of the marsh and continues through marsh sites M1, M2, 

and M5. The second flow path passes through marsh sites M3, 

M4, and M5. Sites M2, M3, and M5 were sampled 26 times 

from November 1996 through December 1997. Vegetated sites 

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6 Hydrology and Cycling of Nitrogen and Phosphorus in Little Bean Marsh: A Remnant Missouri-River Wetland 













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M1 and M4 were sampled at 2-week intervals from April 

through December 1997 when standing water was present. 

Open-water site M6 was sampled six times from September 

through December 1997 when site M1 was dry and was located 

along the northern flow path between sites M2 and M5. The 

southern end of Little Bean Marsh receives little surface runoff 

and mostly is an open-water area that includes outflow site M5. 

Marsh water samples were synoptically collected twice per 

month at three to six locations (sites M1 to M6 depending on the 

area inundated) in high-density polyethylene bottles. A few 

additional samples were collected within a few days after floods 

inundated the marsh. When water depths were sufficient, the 

bottles were raised and lowered through the water column at an 

equal transit rate during filling to make sure the sample was representative of the entire depth. Samples were refrigerated near 

4 °C immediately after sample collection. The day after sample 

collection, samples were poured into a churn splitter and subsamples of suspended solids, turbidity, and total N, P, and DOC 

were split into separate bottles for laboratory analysis. The 

remaining water was then filtered through 0.45-micron polycarbonate filters into separate bottles for laboratory analysis of dissolved N and P species. All laboratory analysis were done at the 

University of Missouri Limnology Laboratory in Columbia, 

Missouri. Mean concentrations of N and P in the marsh samples 

were multiplied times the volume of water in the marsh to compute the change in mass stored in the marsh between sample 

events. 

Laboratory analyses of water samples for total N were 

done using a persulfate digestion as described by Crumpton and 

others (1992), for dissolved orthophosphate (PO4) using a colorimetric method and NO3 and nitrite (NO2) using a cadmiumreduction method as described in U.S. Environmental Protection Agency (1979), for ammonia nitrogen using a phenate 

method as described in Fisheries and Environment Canada 

(1977), for total and dissolved P using a persulfate reduction as 

described in Prepas and Rigler (1982), for suspended chlorophyll using a flourometric determination described in Knowlton 

(1984) and Sartory and Grobbelaar (1984), for total and volatile 

suspended solids using methods described in American Public 

Health Association (1980), and for DOC using a wet persulfate 

oxidation described in Crowther and Evans (1978). Suspended 

ammonium N was calculated by subtracting dissolved from 

total ammonium N. A single analysis for calcium and magnesium was performed with atomic adsorption spectrometry, for 

sodium and potassium with flame emission spectroscopy, and 

for alkalinity by titration according to methods described in 

American Public Health Association (1980). Laboratory measurements of turbidity and color were made with a colorimeter 

using methods described by Hach Company (1987). 

Field measurements of water temperature, specific conductance, and dissolved oxygen were made following methods 

given in Wildke and others (1998). Measurements of water temperature and dissolved oxygen were made in situ and measurements of specific conductance and pH of water samples were 

made in the field at the time of sampling. Dissolved oxygen was 

measured using an electronic meter that had been calibrated in 

Background and Problem 

an air chamber saturated with water vapor near the ambient temperature of the marsh. Temperature was measured using a thermistor attached to the dissolved-oxygen meter. The thermistor 

occasionally was checked against a mercury thermometer. Specific conductance and pH were measured with electronic meters 

that were calibrated with standard solutions that bracketed measured values. 

The significance of correlations between site characteristics and various water samples was determined statistically. 

Tukey’s multiple comparison tests (Tukey, 1977), Wilcoxon 

rank-sum tests (Wilcoxon, 1945), and Kruskal-Wallis tests are 

nonparametric statistical tests that were used to analyze the spatial variability of water chemistry (Ott, 1993). Correlation and 

significance testing also was performed between various watersample constituents. In this report, “significant” denotes that the 

p-value of a comparative statistical test is less than 0.05 or that 

there is a 95 percent chance that statistical relations were not 

false. 

A bathymetric survey (fig. 5) was conducted (Heimann 

and Femmer, 1998) to calculate temporal changes in surface 

area and volume for hydrologic and nutrient budgets and to 

determine hydraulic residence time, which is defined as the 

average time that water remains in an impoundment, and is 

determined using the following equation (modified from Mitsch 

and Gosselink, 1993): 

Rt = S/(DaR) (1) 

where 

Rt = hydraulic residence time, in years; 

S = storage volume, in cubic meters; 

Da = the drainage area of the watershed, in square 

meters; and 

R = the runoff from the local watershed, in depth 

(meters) per year. 

Approximately 1,000 points were surveyed along 20 transects 

through the marsh. The bathymetric survey was tied to the 

water-level gage so that surface area and volume could be determined for any gage height. 

Marsh water levels were measured at site M3 with a pressure transducer and recorded to the nearest 0.3 cm on a data logger equipped with a real-time satellite radio transmitter. The 

pressure transducer was set and calibrated to an outside staff 

gage anchored into the marsh bottom with concrete. Altitudes 

were surveyed into the gage using standard surveying techniques from a U.S. Army Corps of Engineers’ reference mark 

located on the Missouri River levee. 

Inflow site I1 (fig. 1) was located on the upstream side of 

a concrete box culvert with a concrete bottom. Stage was measured with a pressure transducer to the nearest 0.6 cm and set to 

an outside staff gage mounted on the culvert wing wall. Discharge measurements were made over a range of gage heights 

to obtain a rating curve (Rantz and others, 1982). The rating 

curve was programmed into an automatic sampler, which was 

connected to a pressure transducer. The autosampler was pro­
8 Hydrology and Cycling of Nitrogen and Phosphorus in Little Bean Marsh: A Remnant Missouri-River Wetland 

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grammed in such a way as to sample after every 1,130 m3 (cubic 

meters) of runoff passed the gage. 

Discharge at the outflow site M5 (fig. 1) was computed 

using discharge measurements and marsh water levels to prepare a rating curve. However, the frequent presence of beaver 

dams on the outflow control point made rating curves unusable 

for some periods. During these periods, outflow was estimated 

with water-budget calculations made from marsh water-level 

data, inflow data, and estimates of evapotranspiration and 

ground-water seepage. These estimates were checked and constrained with frequent discharge measurements of the outflow. 

Fortunately, daily discharge fluctuations usually were a small 

percentage of average daily streamflow, which minimized the 

error induced by these estimates. Long periods of no outflow 

were documented. 

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masses of N or P removed, added, or transformed by the marsh 

and for comparison of volumetric and areal removal rates with 
------ -------

9 

those of other wetlands. Wet atmospheric deposition was estimated with on-site precipitation measurements, and concentrations of NO3 and NH4 were collected and published by the 

National Atmospheric Deposition Program (1997) at two sites 

approximately 120 km (kilometers) southwest and northeast of 

Little Bean Marsh. Dry atmospheric deposition was estimated 

from data collected as a part of the U.S. Environmental Protection Agency’s Castnet network at Alhambra, Illinois (http:// 

www.epa.gov/castnet/data.html, accessed 2002). Wet deposition of P was estimated from data collected at Centralia, Missouri, by the Agricultural Research Service (written commun., 

1997). 

Inflow from the easternmost tributary to Little Bean Marsh 

were measured and sampled at site I3 (fig. 1) when wetland 

water levels were low (usually base-flow conditions) and at site 

I1 when marsh water levels were high (usually during runoff 

events). About 70 percent of the drainage area of the tributary 

at site I3 is bottomland in row-crop agriculture and 30 percent 

loess uplands in mixed agriculture and forest. The land use of 

the other tributaries is almost 100 percent bottomland agriculture. Scant runoff was observed from these tributaries after rains 

as large as 3.8 and 7.6 cm. Substantial runoff was observed in 

these bottomland tributaries only after the 14.8-cm rain on April 

10 and 11, 1997. Therefore, automatic samples and continuous 

water-level and discharge data were collected only at inflow site 

I1. Data collected at I1, changes in marsh concentrations of N 

and P (sites M1 to M6), and marsh volume measured immediately before and after rainfall events were used to compute the 

mass of N and P received by the marsh in runoff. Inflow sites 

received some ground-water seepage (base flow) when groundwater levels were high. Therefore, all inflow sites were measured and sampled twice per month (when flow was present). 

Sites I3, I4, I5, and I6 also were sampled during the largest runoff events. Water-budget data also were computed using 

changes in marsh volume and specific conductance values, 

which were diluted by surface runoff [between 50 to 200 μS/cm 

(microsiemens per centimeter at 25 degrees Celsius)] and 

increased by ground-water seepage (800 to 1,000 μS/cm) (equation 2): 

Mc Gc Gv 

Mv - Sc Sv 

Mv = × + × (2) 

where 

Mc = specific conductance of marsh, in microsiemens 

per centimeter at 25 degrees Celsius; 

Gc = specific conductance of ground water, in microsiemens per centimeter at 25 degrees Celsius; 

Gv = ground water volume, in cubic meters; 

Mv = marsh volume, in cubic meters; 

Sc = specific conductance of surface water, in microsiemens per centimeter at 25 degrees Celsius; 

Sv = surface water volume, in cubic meters; and 

Gv + Sv = Mv. 

The primary process removing N from most wetlands is 

denitrification (Johnston, 1993), which seldom occurs below 8 

cm deep in wetland sediments (Broderick and others, 1988). 

Background and Problem 

Measurements of denitrification rates were made using in situ 

acetylene-block methods every other month from April through 

November when temperatures were high enough to support 

denitrification. In situ acetylene-block experiments are considered the most accurate way to measure field denitrification 

rates. Acetylene causes denitrified N to accumulate as nitrous 

oxide (N2O) [instead of nitrogen gas (N2)] that can be detected 

and quantified with a gas chromatograph (Yoshinari and others, 

1977; Ryden and others, 1979a, 1979b). This method consisted 

of setting a capped 15-cm diameter Polyvinylchloride (PVC) 

pipe 10 cm into the bottom sediments leaving 20 cm of water in 

the enclosure (fig. 6). The overlying water inside the enclosure 

was then saturated with acetylene (modified from Ryden and 

others, 1979b) by bubbling acetylene through the overlying 

water through needle tubes that extended down through the top 

of the enclosure. Water samples were collected in syringes 

through the needle tubes at 24-hour intervals for N2O analysis. 

These experiments lasted 48 to 72 hours. Measurements were 

made at 3 locations in the open water and three locations in the 

macrophytes at sites corresponding to sampling sites M1 to M6 

(fig. 1). 

&IGURE$ENITRIFICATIONCHAMBERSIN,ITTLE"EAN-ARSH

In situ acetylene-block experiments were used to determine whether NO3 concentrations in the water column limit 

denitrification rates by artificially adding NO3 to the overlying 

water of a second chamber. If more N2O were to be produced in 

the spiked chamber than the unspiked chamber, then denitrification was NO3-limited. Wetlands with NO3-limited denitrification rates have unused N-removal capacity. Because NO2 is 

typically very small compared to NO3 in surface waters, 

NO2+NO3-N will be referred to as NO3 in this report. 

Eight wells (W1 to W7 and W9) were hand augered to 3 m 

deep; screened with 1.6 m of slotted, 5-cm diameter, PVC pipe, 

and cased with solid PVC pipe for a separate study by Walter 

(1998). Altitudes of measuring points on wells were determined 

by running levels from the staff gage on Little Bean Marsh 

using a transit, rod, and conventional surveying methods. 

Ground-water levels were measured at 2 to 4 week intervals 

from September 1996 through December 1997 with an electric 

tape to the nearest 0.3 cm. Daily ground-water levels were estimated for this report by straight-line interpolation between 

measurements. Darcy’s Law for vertical flow (equation 3) was 
10 Hydrology and Cycling of Nitrogen and Phosphorus in Little Bean Marsh: A Remnant Missouri-River Wetland 

used to determine ground-water seepage into and out of the 

marsh (Darcy, 1856; Freeze and Cherry, 1979; Carter, 1986): 

Q = KA(dh/dl) (3) 

where 

Q = flow through the bottom sediments, in volume per 

unit time (can be positive or negative); 

K = hydraulic conductivity of the bottom sediments, in 

length per unit time; 

A = inundated area of bottom sediments, in length 

squared; 

dh = the head difference between the marsh and the 

underlying water table, in length (can be positive 

or negative); and 

dl = the distance between the marsh bottom and the 

underlying water table. 

Area of inundation was determined from the bathymetric 

survey for every daily marsh water level. The term dh was 

determined by subtracting the average ground-water altitude in 

wells W1, W4, W5, and W6 from the altitude of water in the 

marsh. The term dl was the distance between the average 

ground-water altitude and the average altitude of the marsh bottom determined for every stage. A value of K was selected that 

minimized error in the annual marsh water budget. 

In the Walter (1998) study, ground-water samples were 

collected from 8 monitoring wells (W1 to W7 and W9, fig. 1) 

every other month from September 1996 to April 1997, and 

analyzed at the University of Missouri at Kansas City for dissolved NO3+NO2-N, P, magnesium, calcium, sodium, potassium, sulfate, chloride, bicarbonate, aluminum, iron, manganese, nickel, lead, and zinc. For this report monthly samples 

collected by USGS personnel after April 1997 through December 1997 at four of the wells (W1, W4, W5, and W6) were analyzed for dissolved total N, dissolved ammonium (NH3+NH4-

N), NO2+NO3-N, PO4, and total P (table 1) at the University of 

Missouri Limnology Laboratory in Columbia, Missouri. The 

mass of N and P in ground-water seepage was then calculated 

by multiplying the seepage volume times the average N or P 

concentration of ground water if flow is into the marsh, or the 

negative seepage volume times N and P concentrations of the 

marsh if seepage is out of the marsh as shown in equation 4: 

Mnp = Cnp × S (4) 

where 

Mnp = mass of nitrogen or phosphorus entering or leaving 

the marsh through ground water, in milligrams; 

Cnp = concentration of nitrogen or phosphorus in the 

marsh, in milligrams per liter; and 

S = the volume of seepage entering or leaving the 

marsh into or from ground water, in liters. 

The Eh and pH of bottom sediments were measured at each 

of the marsh-sampling sites (sites M1 to M6) by collecting a 

core in a 2.54-cm diameter soil-sampling tube. Eh and pH 

probes were inserted into the bottom end of the core while it was 

still in the tube and the meter readings were allowed to reach 

temporal stability before being recorded. Sampling depth intervals were shorter near the sediment-water interface and increase 

away from the sediment-water interface, and ranged from 0.5 to 

15 cm. Visual observations of the color, wetness, and organic 

material in the core were recorded as each section of the core 

was extruded from the tube. The Eh meter was calibrated with 

standard Zobell solution before each series of measurements 

and checked against Zobell solution again at the end of the measurements to check for drift. The pH meter was calibrated with 

pH 4, 7, and 10 buffer solutions according to methods described 

in the USGS National Field Manual for the collection of waterquality data (Wildke and others, 1998). 

Hourly measurements of temperature, dissolved oxygen, 

specific conductance, and pH were made with a multiparameter 

water-quality monitor at a single point in the marsh. This point 

was at site M3 in open water in one of the deepest parts of the 

marsh so that measurements were possible at low water levels. 

The hourly data were stored in a data logger. A mechanical stirrer operated during a 3-minute warm-up period to ensure proper 

water flow across the dissolved-oxygen membrane. The hourly 

circulation also minimized biofouling of the sensors. The monitor was protected in a 15-cm diameter PVC pipe with numerous 

holes to provide for water exchange. Two horizontal floats were 

attached to the side of the vertical PVC pipe to allow the monitor sensors to float at a fixed depth of 0.5 m below the surface. 

The monitor was checked against field meters, cleaned, and calibrated with standard solutions at 2-week intervals to determine 

if corrections were necessary and to recalibrate the monitor. 

Precipitation was collected at site M3 with a tipping bucket 

rain gage at hourly intervals and recorded on the data logger. 

Precipitation data collected by the National Weather Service 

(NWS) at the KCI (National Atmospheric and Oceanic Administration, 1996 and 1997) were used to verify the data collected 

on site and to estimate short periods of missing daily rainfall. 

KCI is 34 km southeast of Little Bean Marsh. Daily average 

wind-speed data, also collected at KCI, were obtained from the 

NWS for correlation with P concentrations. 

Evapotranspiration from November 1996 through April 

1997 was estimated using climatological data collected at DeSoto, Kansas; St. Joseph, Missouri; and Powhatten, Kansas 

(National Oceanic and Atmospheric Administration, 1996 and 

1997) and the Penman-Montieth method (Penman, 1948; Montieth, 1965) of estimating potential evapotranspiration. The 

Penman-Montieth method has been successfully used in wetlands where water tables are high and abundant macrophytes 

keep evapotranspiration near potential values (Souch and others, 1998). Potential evaporation estimates from all three sites 

were averaged to produce estimates at Little Bean Marsh. Penman-Montieth evapotranspiration values are calculated from 

solar radiation, wind, humidity, and air temperature measurements using the Penman equation (equation 5). Pan evaporation 

data collected at Smithville Reservoir (39 km east of Little Bean 

Marsh) by the National Oceanic and Atmospheric Administration (1996 and 1997) were used from May through October 

along with a “pan coefficient” to estimate evapotranspiration 
-- --

-- -- --

-- -- -- -- --

-- -- -- -- --

-- -- -- -- --

--

--

-- -- --

-- -- --

-- -- -- -- --

-- -- -- -- --

-- -- -- -- --

-- --

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-- -- -- -- --

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--

--

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--

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-- -- -- -- --

-- -- -- -- --

--

--

--

--

Background and Problem 11 

Table 1. Specific conductance and concentrations of selected nutrients in ground-water samples collected from monitoring wells 

near Little Bean Marsh. 

[N, nitrogen; mg/L, milligrams per liter; P, phosphorus; μS/cm, microsiemens per centimeter at 25 degrees Celsius; --, no data; <, less than] 

Dissolved 

Well Total Nitrite plus ammonia N as Soluable Specific 

number dissolved N, nitrate as N, N, Dissolved P, reactive P, conductance, 

(fig. 1) Date in mg/L in mg/L in mg/L in mg/L in mg/L in μS/cm 

W1 01/20/1997 <0.02 <0.02 807 

W1 02/17/1997 <.005 .006 768 

W1 02/27/1997 584 

W1 03/09/1997 671 

W1 04/13/1997 482 

W1 04/25/1997 0.25 <.005 <0.005 .013 0.007 

W1 08/12/1997 .68 <.005 .065 .148 .140 

W1 12/02/1997 .33 <.005 .042 .019 .010 796 

W4 01/20/1997 <.02 <.02 645 

W4 02/17/1997 <.005 .002 779 

W4 02/27/1997 639 

W4 03/09/1997 625 

W4 04/13/1997 639 

W4 04/25/1997 <.005 .049 .285 .286 

W4 08/12/1997 <.005 .127 .760 .816 

W4 12/02/1997 .010 .142 .874 .866 

W5 01/20/1997 .04 .04 763 

W5 02/17/1997 .010 .012 728 

W5 02/17/1997 739 

W5 03/09/1997 709 

W5 04/13/1997 720 

W5 04/25/1997 .19 <.005 <.005 .024 .025 

W5 08/12/1997 .90 <.005 .098 .124 .125 

W5 12/02/1997 .28 .024 .052 .025 .024 

W6 01/20/1997 <.02 .03 775 

W6 02/17/1997 <.005 .012 731 

W6 02/27/1997 755 

W6 03/09/1997 749 

W6 04/13/1997 794 

W6 04/25/1997 .10 .019 <.005 .020 .019 

W6 08/12/1997 .07 <.005 .013 .014 .022 

W6 12/02/1997 .14 <.005 .025 .060 .055 761 

W9 01/20/1997 <.02 .17 1,112 

W9 02/17/1997 936 

W9 02/27/1997 1,068 

W9 03/09/1997 1,034 

W9 04/13/1997 1,017 
---------------------------------------------------------

12 Hydrology and Cycling of Nitrogen and Phosphorus in Little Bean Marsh: A Remnant Missouri-River Wetland 

from May through October. A coefficient of 1.0 was determined 

by selecting a value that gave the best match with monthly 

Penman-Montieth estimates: 

s Q*-^Q ( s) + @ CpD/Ra Qe = a

s+g 1 Rs Ra ( + ⁄ ) - - (5) 

where 

Qe = the latent heat flux of evaporation, in megajoules 

per square meter per day; 

s = the slope of the saturation vapor pressure and 

temperature curve, in millimeters of mercury per 

degree Celsius; 

Q* = the net all wave radiation, in megajoules per square 

meter per day; 

Qs = the net storage heat flux, in megajoules per square 

meter per day; 

@a = the density of air, in kilograms per cubic meter; 

Cp = the specific heat of moist air, in joules per kilogram; 

D = the vapor pressure deficit, in millimeters of 

mercury; 

Ra = aerodynamic resistance, in seconds per meter; 

Rs = the surface resistance, in seconds per meter; and 

g = the density of water, in kilograms per cubic meter. 

Stages on the Missouri River near Little Bean Marsh were 

estimated using continuous stage data collected by the USGS at 

St. Joseph, Missouri, and 59 daily river-stage readings at Atchison, Kansas, as measured by a U.S. Army Corps of Engineers 

observer and recorded by the National Weather Service River 

Forecast Center in Pleasant Hill, Missouri (written commun., 

2000). The Atchison data were collected between October 6, 

1996, and July 4, 1997. Concurrent stage readings at the two 

sites were used to develop average altitude drops and river 

slopes between St. Joseph and Atchison. The altitude drop averaged 7.11 m with a range of 6.77 to 7.43 and the slope averaged 

0.173 m/km (meter per kilometer) and ranged from 0.165 to 

0.180. The average altitude drop was subtracted from the continuous stage data at St. Joseph to estimate stage at Atchison for 

days where Atchison stage data were unavailable. The average 

river slope was then used to extrapolate stage data 9.7 km downstream from Atchison to Little Bean Marsh. The ranges of altitude drop and river slope indicate stage estimates at Little Bean 

Marsh usually were within 1 foot of actual stages. Given the 

large fluctuations in river stage, this error was considered 

acceptable for use in comparing stage with ground-water and 

marsh levels in the vicinity of the marsh. 

Hydrology 

Although hydrologic characteristics such as source of 

water, hydraulic residence time, and duration and timing of 

inundation are the primary determinants of many wetland functions (Moustafa, 1997) and are necessary for improved understanding of wetland limnology and ecology, hydrologic measurements are rarely made in wetland studies (Kibby, 1978; 

Carter, 1986; Mitsch and Gosselink, 1993). Therefore, a hydrologic analysis of Little Bean Marsh was made to better measure 

nutrient cycling processes and to quantify inputs and outputs of 

nutrients to the marsh. 

Hydrologic Inputs 

In the absence of active management at Little Bean Marsh, 

marsh vegetation has been fairly stable (Castaner and LaPlante, 

1992) since acquisition by the Missouri Department of Conservation in 1981. Because emergent vegetation responds fairly 

quickly to changes in inundation frequency, duration, and water 

level (Carter, 1986), hydrologic conditions probably have been 

fairly stable as well. In 1997, 41 percent of the water received 

by Little Bean Marsh was direct precipitation and 14 percent 

was ground-water seepage (fig. 7). Surface-water inflow constituted 45 percent of the marsh inputs, although about one-third 

of surface inflow (15 percent of marsh inputs) were backflows 

from Bean Lake. Nearly all the backflow occurred after two 

large rainstorms in April and May when the outflow from Bean 

Lake was blocked by closed flood gates in the Missouri River 

levee. The relatively high percentage of direct precipitation is 

because of the low ratio of drainage area to water-surface area 

(12:1). Like many flood-plain water bodies, more than 90 percent of the marsh’s drainage area is flood plain. The nearly flat 

slopes, sandy soils, and large shallow depressions in the flood 

plain keep runoff coefficients low. No runoff was observed 

from several storms that had more than 3.8 cm of rain; consequently, surface-water contributions from the watershed were 

small compared to upland watersheds. The tributary that most 

frequently produced runoff was one that contained a small percentage of upland soils (site I1, fig. 1). 

Minimal surface runoff and a substantial hydraulic connection with ground water tend to stabilize water levels in the 

marsh. Ditches created for agricultural drainage cause runoff 

from the loess hills to the east to bypass Little Bean Marsh, and 

thus bypass its potential for flood attenuation and nutrient 

removal. Flood-plain ditches and river levees have converted 

Little Bean Marsh from an open to a closed system. This conversion is not an unusual occurrence in the Missouri River flood 

plain as landowners have drained valuable agricultural bottomlands by routing runoff from the uplands straight across the 

flood plain into the river. Consequently, the hydrology, water 

quality, and ecological functions of remnant wetlands located 

along today’s Missouri River usually are vastly different from 

the predevelopment condition. Open wetlands often are large 

sinks for sediment and nutrients, whereas closed wetlands primarily rely on precipitation for nutrients (Craft and Casey, 

2000). 

The large percentage of precipitation as an inflow source 

has diluting effects on water quality, as indicated by low specific conductance values and concentrations of nutrients immediately after large rainstorms and during periods of low groundwater levels (fig. 8). If precipitation did not make up a large proportion of the inflow, concentrations of most constituents would 
Hydrology 13 



PERCENT



PERCENT



PERCENT

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have been greater, especially during periods of intense evapotranspiration and ice formation. 

Ground-Water Interactions 

Although the bottom of Little Bean Marsh is composed of 

fine organic material mixed with sand, underlying sands and 

gravels have large hydraulic conductivities that typically range 

from 126 to 326 m/d (meters per day) (Emmett and Jeffery, 

1969; Kelly and Blevins, 1995). Also, Little Bean Marsh is 

incised 3 to 4 m below the surrounding flood plain, which 

allows rapid hydraulic communication between the marsh and 

ground water. The bottom of Little Bean Marsh is only 1 to 2 m 

above the Missouri River stage at median discharge (Missouri 

River Bottoms Work Group, written commun., 1979). The 

marsh bottom was between 1.0 m below and 0.7 m above the 

median stage of the river in 1997 and the water surface in the 

marsh averaged only 0.5 m above the river stage. Higher-thannormal river stage in 1997 (Hauck and others, 1998) probably 

increased ground-water levels above normal, which may have 

caused greater-than-normal ground-water contributions to Little Bean Marsh. 



PERCENT



PERCENT



PERCENT

Attempts often have been made to justify wetland protection on the basis that wetlands are recharge areas. However, 

many wetlands have been shown to be ground-water discharge 

areas (Carter, 1986). Although Little Bean Marsh was both a 

recharge and discharge area, discharge to the marsh was three 

times greater than recharge to ground water during the 12 

months of monitoring. Ground-water levels were measured in 

two shallow (3 m) wells along the northern edge of Little Bean 

Marsh (W4 and W5, fig. 1), and two shallow wells along the 

southern edge (W1 and W6, fig. 1). The water-table gradient 

slopes to the southwest, downstream, and toward the river (fig. 

9) as typically occurs in the Missouri River alluvial aquifer 

(Emmett and Jeffery, 1969). Ground-water levels on the south 

side of the marsh typically were 6 to 18 cm below the marsh 

water level, whereas ground-water levels on the north side 

(wells 4 and 5) typically were 6 to 18 cm above the marsh water 

level. Accumulations of iron deposits on the north side of the 

marsh bottom were frequently observed, indicating groundwater seepage from the alluvial aquifer, which typically has 

large concentrations of iron (Ward, 1973; Ziegler and others, 

1993). Also, three marsh tributaries, incised into the terrace 

scarp that comprises the northern and western boundaries of the 

marsh, were observed to receive small amounts of ground-water 
14 Hydrology and Cycling of Nitrogen and Phosphorus in Little Bean Marsh: A Remnant Missouri-River Wetland

seepage during periods of high ground-water levels and wet 

weather. These factors indicate that ground water typically 

recharges the marsh on the north side and seeps out of the marsh 

bottom on the south side. 

Ground-water levels were more closely related to water 

levels in Little Bean Marsh than to stages in the Missouri River 

(fig. 10), indicating the marsh has good hydraulic connection 

with ground water. Although the interaction between the marsh 

and ground water is close, it also is complex. Even in the southern, downgradient wells, the lowest ground-water levels were 

always above the lowest parts of the marsh bottom. The groundwater connection of Little Bean Marsh indicates that the hydrologic regime of most wetlands along the lower Missouri River 

largely is a function of bottom altitude relative to water-table 

altitude. In the past, when Little Bean Marsh was much deeper, 

the connection between the marsh and ground water would have 

been even greater. Ground water would likely have caused most 

of the marsh to remain permanently inundated. In the future, as 

Little Bean Marsh continues to fill with sediment and detritus, 

the ground-water connection will decrease, slowly changing the 

vegetation to terrestrial plants. Although the marsh was inundated during this study, local farmers reported some years when 

the marsh has been nearly dry. After large runoff events, all 

ground-water levels are less than the marsh level, and the entire 

marsh recharges the aquifer. At these times, the marsh can 

remove selected nutrients and contaminants before they enter 

the aquifer where they are difficult to remove. Conversely, the 

marsh also can be an area of aquifer vulnerability where soluble 

surface-water contaminants, such as agricultural pesticides, can 

be directly recharged into the aquifer. 

During long periods of high evapotranspiration rates, 

marsh water levels often are sustained by ground water, providing conditions favorable for the cattails (Typha latifolia) and 

bulrushes (Scirpus fluviatilis) that are so abundant in the marsh. 

This interaction with shallow ground water indicates that lowering the water table by less than a meter by pumping would 

likely make Little Bean Marsh drier and could substantially 

affect the marsh ecosystem. Specific conductance values of 

ground water in the Missouri River alluvium typically are about 

1,180 μS/cm (Ziegler and others, 1993); somewhat greater than 

values in shallow ground water near Little Bean Marsh, which 

average about 750 μS/cm (table 1). Well W9 is located about 









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1.6 km upgradient from Little Bean Marsh near the bluff line, 

and has a more typical (Ziegler and others, 1993) ground-water 

conductance of about 1,000 μS/cm. The lesser conductance values of ground water near the marsh may reflect surface recharge 

from the marsh itself. Conversely, the alternating effects of precipitation/runoff and ground-water seepage on specific conductance in the marsh are evident (fig. 8). Except during extended 

ice periods, specific conductance is a good indicator of water 

source in the marsh. When mean ground-water levels are higher 

than marsh water levels, specific conductance increases in the 

marsh; the opposite is true when ground-water levels decline. 

Hydrologic Outputs 

More water was lost from the marsh through evapotranspiration than by all other pathways combined (59 percent, fig. 7). 

This partially is because of abundant macrophytes that transpire 

water, in addition to that lost from evaporation of open-water 

surfaces. Jensen and Rahman (1987) determined that evaporation from the free water surface of a Bangkok wetland rice 

paddy was only 13 percent of the total evaporative loss. In Little 

Bean Marsh surface outflow accounted for 36 percent and 

ground-water seepage accounted for only 5 percent of the water 

losses (fig. 7). Consequently, the marsh seems to serve the 

hydrologic function of surface-runoff reduction, and may even 

lower the water table when ground-water levels are high in the 

summer. The percentage of outputs from evapotranspiration 

would have been less, and surface outflow would have been 

greater, had Short Creek still run through Little Bean Marsh. 

This evaporative loss obviously is not present where marshes 

have been drained and surface water is routed directly to the 

river. Evaporation leaves behind dissolved constituents and 

causes higher specific conductances during the summer than 

would be expected from a precipitation-dominated system. The 

shallowness of the marsh, ground-water inflows, and large 

hydrologic-budget components of precipitation and evapotranspiration lead to large fluctuations in specific conductance and 

dissolved-constituent concentrations (fig. 8). The shallowness 

of wetlands makes open-water evaporation in wetlands 

extremely responsive to solar radiation (Parkhurst and others, 

1998; Souch and others, 1998). Consequently, the amount of 

surface outflow also is sensitive to solar radiation. Although 

macrophytes in Little Bean Marsh transpire much water, they 

also substantially decrease evaporation by shading. This factor 

is most important during the early spring and fall when transpiration from Scirpus fluviatilis is low, but senescent plants still 

shade most of the water surface. 

The importance of precipitation and evapotranspiration in 

a Midwestern riparian cypress swamp (Taxodium distichum) in 

southern Illinois also was noted in a study by Mitsch (1978) 

where precipitation accounted for 70 percent of inflow and 

evapotranspiration accounted for almost one-half of the outflow. The ratio of drainage area to surface area for this Illinois 

swamp was 3.5 to 1, and total annual precipitation was only 

74.3 cm. Low drainage area to surface area ratios, annual preWater Chemistry 15 

cipitation averaging 75 to 130 cm/yr, average potential evapotranspiration values of 100 to 175 cm, and high percentages of 

flood plain in the drainage area likely make precipitation and 

evapotranspiration the largest components of the hydrologic 

budgets of most riparian wetlands in the Midwest. Ground 

water interactions are more likely to have increased importance 

along larger rivers associated with highly productive alluvial 

aquifers, such as the Missouri River alluvium. 

Dividing the average storage of Little Bean Marsh in calendar year 1997 (200,000 m3

) by the average rate of inputs [480 

m3

/d (cubic meters per day)] gives a hydraulic residence time of 

42 years. Although ground-water inflows and evapotranspiration are not accounted for in the formula for hydraulic residence 

time and may decrease the real residence time somewhat, the 

average amount of time that water remains in the marsh is still 

very long (probably years or decades) and allows the marsh to 

greatly attenuate peak flows. In fact, the marsh has the capacity 

to substantially attenuate much larger peak flows, but drainage 

ditches, designed to remove water as quickly as possible, have 

routed flows around Little Bean Marsh and deprive it of this 

function. The large residence time also allows the marsh to 

greatly affect the water quality of outflow (Moustafa, 1997). 

Water Chemistry 

The lack of water-quality data on remnant native wetlands 

requires that certain parameters be characterized before specific 

processes can be investigated. For example, denitrification and 

P release occur during low redox conditions, but the temporal 

and spatial distribution of reducing conditions in remnant wetlands along the lower Missouri River are not known. Therefore, 

the central tendencies and seasonal variations of N, P, pH, Eh, 

specific conductance, temperature, and dissolved oxygen were 

computed on data measured in Little Bean Marsh for 12- or 24­

month periods. The intermixing of emergent macrophytes and 

open water is expected to create substantial spatial variation in 

these parameters (Rose and Crumpton, 1996). Therefore, samples collected at six sites (M1 to M6, fig. 1) were used to quantify the spatial variability and to investigate statistical differences in water quality between vegetated and open-water sites. 

Specific Conductance 

Daily fluctuations of specific conductance were less than 

50 μS/cm and seemingly nonsystematic except during large rain 

storms. However, longer-term variability in specific conductance ranged from 180 to 1,300 μS/cm and was related to the 

source of water in the marsh, the amount of evaporation, and the 

amount of ice (fig. 8). The primary sources of water to the 

marsh were precipitation (specific-conductance values usually 

less than 25 μS/cm) (National Atmospheric Deposition Program, 1997), runoff (50 to 240 μS/cm, table 2), and groundwater seepage (about 750 μS/cm, table 1). The large difference 

between specific conductance of ground water and precipita­
16 Hydrology and Cycling of Nitrogen and Phosphorus in Little Bean Marsh: A Remnant Missouri-River Wetland 

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18 Hydrology and Cycling of Nitrogen and Phosphorus in Little Bean Marsh: A Remnant Missouri-River Wetland 

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tion/runoff, and the short periods of input from precipitation and 

runoff, allow the use of specific conductance as an indicator of 

the primary source water in the marsh. For example, two storm 

events in late February and early April 1997 substantially 

decreased specific conductance to about 400 μS/cm. These 

events were followed by a rapid increase in specific conductance that were likely caused by ground-water seepage as these 

periods correspond with periods of low evapotranspiration and 

high ground-water levels. Conversely, a late May 1996 storm 

decreased specific conductance to 200 μS/cm with only a small 

rebound (Heimann and Femmer, 1998), when ground-water 

levels were too low to be a substantial source of water to the 

marsh. 

Another important factor affecting specific conductance 

was ice formation. Both the 1996 and 1997 annual peaks in specific conductance corresponded with ice formation (Heimann 

and Femmer, 1998; fig. 8). These peaks were 1,140 and 1,300 

μS/cm, which were substantially higher than those associated 

with ground water (about 800 μS/cm) and about 500 μS/cm 

larger than any peaks not associated with ice formation. These 

extremely large increases probably are a result of ion exclusion 

from the ice and the concentration of excluded solutes in shallow water under the ice. This is likely not an uncommon phenomenon in high-latitude wetlands, where ice can make up a 

large percentage of the water in shallow water bodies. As the 

volume of ice in the marsh increases, solute concentrations 

would be expected to increase geometrically. The concentration 

of solutes could have significant effects on biota that survive 

under the ice, especially for solutes with high toxicity. Therefore, concentration of solutes by ice formation could make wetlands more vulnerable to toxic contaminants than other water 

bodies. 

During 1997, the mean specific conductance at site M2 

was more than 50 percent larger than the mean conductance at 

site M5 near the outflow. Tukey’s multiple comparison test 

indicates specific conductance at sites M1 and M2 on the north 
-- --

-- -- 

-- --

-- --

Table 2. Sample analyses and storm loads of surface inflow to Little Bean Marsh. [mg/L, milligrams per liter; kg, kilograms; μS/cm, microsiemens per centimeter at 25 degrees Celsius; m3, cubic meters; --, no data available; <, less than] 

Total Total ammonium Total Total dissolved 

Site nitrogen nitrogen phosphorus nitrogen 

number Concentration, Load, Concentration, Load, Concentration, Load, Concentration, Load, 

(fig. 1) Date in mg/L in kg in mg/L in kg in mg/L in kg in mg/L in kg 

I3 April 5, 1997 1.14 1.90 0.26 0.43 1.14 1.90 I1 April 10–11, 1997 3.62 104 .82 24 3.05 8.73 I4 April 10–11, 1997 2.03 47.7 .71 17 1.97 46.2 I1 May 8, 1997 22.26 228 4.78 48.8 9.32 95.3 11.39 116 

Total 381 136 252 

Total suspended 

Site Dissolved ammonium nitrogen Total dissolved phosphorus Soluable reactive phosphorus solids 

number Concentration, Load, Concentration, Load, Concentration, Load, Concentration, Load, 

(fig. 1) Date in mg/L in kg in mg/L in kg in mg/L in kg in mg/L in kg 

I3 April 5, 1997 <0.005 <0.008 0.17 0.29 0.15 0.25 I1 April 10–11, 1997 <.005 <.14 .26 2.4 .26 7.3 429 12,300 I4 April 10–11, 1997 .18 4.3 .37 8.6 .35 8.3 120 2,810 I1 May 8, 1997 2.34 23.9 .53 5.4 .050 .51 9,640 98,500 

Total 28 22 16 113,600 
-- -- --

--

Table 2. Sample analyses and storm loads of surface inflow to Little Bean Marsh.—Continued 

[mg/L, milligrams per liter; kg, kilograms; μS/cm, microsiemens per centimeter at 25 degrees Celsius; m3, cubic meters; --, no data available; <, less than] 

Volatile suspended 

Site solids 

number Concentration, Load, Specific conductance range, Storm runoff volume, 

(fig. 1) Date in mg/L in kg in μS/cm in m3 

I3 April 5, 1997 1,670 I1 April 10–11, 1997 36 1,000 147–241 28,600 I4 April 10–11, 1997 17 400 52–151 23,500 I1 May 8, 1997 720 7,360 10,200 

Total 8,790 64,000 
side of the marsh to be significantly higher than at sites M3 and 

M5 on the south side of the marsh (p<0.05). Whereas M4, on 

the south side of the marsh, was not significantly different from 

the north or south sites (p>0.05), specific conductance was less 

at M4 than at either one of the northern sites (fig. 11). These statistics indicate that areal differences were both large and nonrandom. The higher conductance values on the north side of the 

marsh are likely related to the higher ground-water levels on the 

north side that provided a constant source of high-conductance 

ground-water seepage compared to the south side, where 

ground-water levels usually were below marsh water levels. No 

consistent statistical significance was observed between specific conductance in open-water and vegetated sites. The areal 

distribution of specific conductance seems to be a good locator 

of areas where ground-water seepage is prevalent. 

Dissolved Oxygen 

Dissolved-oxygen concentrations showed both seasonal 

and daily patterns. Low dissolved-oxygen concentrations are 

expected in the summer because of decreased water levels, 

lower solubility of oxygen, and increased rates of respiration 

and decomposition at higher temperatures. However, summertime dissolved-oxygen values in Little Bean Marsh were less 

than most streams and the epilimnions of most lakes in the 

region, which typically range from 7 to 10 mg/L (milligrams per 

liter) (Davis, 1991). The water-quality standard for dissolved 

oxygen for protection of aquatic life in Missouri streams and 

lakes is 5 mg/L (Missouri Department of Natural Resources, 

1999). Dissolved oxygen was near zero for about a month in the 

summer of 1996 and 1997 (fig. 12). Dissolved-oxygen concentrations in both years remained less than 5 mg/L for 3 to 4 

months each summer. Additionally, dissolved oxygen 

decreased to less than 5 mg/L under the ice in February 1996. 

These extended periods of anoxia indicate reducing conditions 

probably were present in the water column as well as in the bottom sediments. Reducing conditions in the water column would 

inhibit nitrification, enhance denitrification, and possibly cause 

the release of P from sediments through the reduction of iron. 

The periods of least dissolved oxygen corresponded with the 

period of highest water levels in the marsh in 1996 and with the 

lowest water levels in 1997. Apparently, anoxia is not limited to 

periods of low water. 

The decreased availability of oxygen in areas of emergent 

macrophytes likely has profound effects on biochemical processes in a marsh. For example, removal of oxygen from the 

water column decreases rates of nitrification, which is the rate 

limiting step for N removal in the marsh. Oxidizing energy flow 

paths are replaced by reducing flow paths such as denitrification, sulfate reduction, and methanogensis, and anoxic water 

favors invertebrates adapted for low oxygen conditions and, 

therefore, affects rates of secondary production and consumers 

higher up the food chain. 

Although a short period of anoxia occurred under ice, ice 

also contributed to extremely high concentrations of dissolved 

Water Chemistry 21 

oxygen in the underlying water. Concentrations as high as 26.5 

mg/L (nearly 200 percent of saturation) occurred when heavy 

coats of algae were both in and on the undersides of the ice. Evidently, while algae was producing oxygen, the ice was preventing escape to the atmosphere and low temperatures increased 

solubility and minimized respiration and decomposition. Also, 

ice formation may have concentrated some oxygen into the 

underlying water through ion exclusion as happened with other 

dissolved constituents. Once again, the shallowness of the wetland contributes to large fluctuations in water quality as all 

changes in constituent mass are confined to a shallow and, thus, 

small volume of water. 

Dissolved-oxygen concentrations followed typical diel 

patterns caused by algal photosynthesis (fig. 13). The largest 

diurnal changes were about 5 mg/L that typically occurred during periods of low dissolved-oxygen concentrations. 

Median dissolved-oxygen concentrations ranged from 

10.1 to 11.2 mg/L at open-water sites M5 and M3 (fig. 11) and 

were significantly higher than the three sites with emergent vegetation (M1, M2, and M4) (median concentrations ranged from 

4.1 to 7.5 mg/L) (p<0.05) in four of the six possible pairwise 

comparisons. Open-water sites (M5 and M3) also had high dissolved-oxygen concentrations in the other two pairwise comparisons as well, but those differences were not statistically significant (p>0.05). Although all measurements were made 

during the day and within 6 hours of each other, some of the 

scatter in these data might be from the lack of simultaneous 

measurement, which would decrease the sensitivity of statistical tests. Nevertheless, these data confirm what has been noted 

by others (Reddy, 1981), that emergent and floating macrophytes decrease dissolved-oxygen concentrations by shading 

out algal photosynthesizers and decreasing contact with the 

wind. Rose and Crumpton (1996) reported that in a prairie pothole wetland, emergent macrophytic vegetation (primarily 

Typha spp and Scirpus fluviatilis) causes much lower concentrations of dissolved oxygen than those in open water areas. 

Macrophytes also contribute to the oxygen demand through 

production of large amounts of detritus. Macrophyte limitation 

of dissolved oxygen can be an important wetland characteristic, 

as it can decrease the wetland’s ability to generate dissolved 

oxygen for fish and other aquatic fauna during summer periods 

of anoxia. 

Temperature 

Continuous temperature measurements were made at one 

of the deeper parts of the marsh in an area of open water (site 

M3). Water depth at the measuring site was about 1 m at a typical stage. Daily mean water temperatures were computed from 

continuous data measured at the marsh bottom and 0.5 m below 

the water surface. Despite shallow depths and the closeness of 

the two temperature sensors, vertical temperature differences of 

3 to 9 °C persisted during June, July, and August of 1996 and 

1997, indicating persistent stratification in summer (fig. 14). 

Persistent stratification also occurred in April and May of 1997. 
	

	 	

	

	

	

  

	

	 	

	

	

	



 



	

	 	



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Water Chemistry 25 















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26 Hydrology and Cycling of Nitrogen and Phosphorus in Little Bean Marsh: A Remnant Missouri-River Wetland 

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Temporary stratifications were common in the spring and fall Heimann and Femmer (1998) at two other wetlands in Missouri. 

months (fig. 14). Had the upper temperature measurement been The greatest differences occurred above 20 °C where the density 

made at the water surface, the extent of stratification would have of water changes more rapidly with temperature, thus creating 

been even more evident. Persistent temperature differences stable stratification able to withstand wind mixing. Seepage of 

were not expected in such shallow open water, where wind and 13 oC ground water into and through the bottom of the marsh 

night-time temperatures were expected to make stratification a may have contributed to stratification in some places. The pertemporary phenomenon. At low latitudes, shallow lakes rarely sistence of stratification also limits the movement of dissolved 

have persistent stratification (Dye and others, 1980). However, oxygen to the bottom of the marsh, making the bottom sedipersistent temperature differences also were observed by ments and deeper water susceptible to anoxia. The period of 
lowest dissolved-oxygen concentration corresponded with the 

period of maximum stratification (fig. 14, June and July 1996). 

In July 1997, a strong period of stratification also corresponded 

with a period of reduced dissolved-oxygen concentrations. Earlier and cooler periods of stratification in spring 1997, however, 

did not correspond with periods of low dissolved oxygen. The 

ability of open-water marshes to stratify may increase areas and 

periods of reducing conditions, which inhibit nitrification and 

potentially allow denitrification and P release from iron reduction. Chlorophyll data indicate that much of Little Bean Marsh 

underwent a major algal bloom in September 1997 (table 3, at 

the back of this report) shortly after summer stratification 

abruptly ended. The bloom may have been triggered by the 

mixing of the water column that made ammonia nitrogen from 

the hypolimnion available to a N-limited algae population. 

Daily mean water temperatures 0.5 m below the surface 

closely corresponded to the daily peaks of air temperatures in 

the summer (fig. 15). Rarely did the water temperature fall as 

far as air temperature on cool days. The higher mean water temperatures were likely caused by solar heating of the open water 

areas that add heat in addition to conductive heat from the air. 

Summer stratification held much of this heat in the upper 0.5 m. 

Water Chemistry 27 

The annual pattern of water temperature was similar for both 

years with temperatures peaking at 25 to 30 oC for almost 2 

months during the summer each year. The mean July temperatures at KCI (34 km southeast of Little Bean Marsh) were 24 °C 

in 1996 and 25.4 oC in 1997, which are 1.8 and 0.4 degrees 

below normal (National Oceanic and Atmospheric Administration, 1996 and 1997). However, the close relation between air 

temperature and water temperature indicates that abnormally 

warm summers would increase water temperatures (and thermal 

stresses on biota) in direct proportion to higher air temperatures. 

Shallow depths and extended periods of anoxia in the marsh 

would limit the ability of some organisms to escape high-temperature stress. 

Hourly near-surface water temperatures (0.5 m below the 

surface) at site M3 showed seasonal patterns of diel variability. 

The diel variability was substantially larger in the spring and 

fall than the summer and winter (fig. 16). The low variability of 

summer temperatures closely corresponded to periods of persistent stratification, whereas periods of maximum temperature 

variation occurred during mixed conditions or short periods of 

temporary stratification (figs. 14 and 16). The lower summer 

variations likely result from a lack of mixing. 

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28 Hydrology and Cycling of Nitrogen and Phosphorus in Little Bean Marsh: A Remnant Missouri-River Wetland

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Water Chemistry 29 

Surface temperatures near 30 °C were recorded at the 

open-water sites (table 3, fig. 14). Qualitative observation of 

other open water areas indicate even higher surface temperatures where dark algal masses, just below the water surface of 

Little Bean Marsh, added solar heat to the top 1 cm. High surface-water temperatures contribute to high evaporative losses 

and help account for the large part of the hydrologic outflows 

attributed to evapotranspiration in Little Bean Marsh. For all of 

1997, sample sites with dense emergent vegetation had cooler 

temperatures than open water sites in 7 out of 10 pairwise comparisons of 5 sites where at least 10 concurrent samples were 

available. However, a Tukey’s multiple comparison test did not 

show these differences to be statistically significant at the 95 

percent confidence level (p>0.05) because the maximum difference of mean temperatures between sites was only 2.3 oC. This 

difference is much less than seasonal variability, which greatly 

limits the ability of statistical tests to identify small daily differences. 

pH 

Daily median values of pH ranged from 6.8 to 10.2 during 

a 2-year period (fig. 17). The lowest values occurred in winter 



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during extreme ice and during large backflow from Bean Lake 

in June 1996. A rapid, one-unit rebound in February 1997 

occurred during and immediately after a large, ice-melting 

storm. Evidently, the dominant source of water to the marsh 

plays an important role in determining the pH. During periods 

when high-bicarbonate ground water constituted the majority of 

water in the marsh, pH fluctuations were buffered between 7.6 

and 8.4. Higher pH values with much larger fluctuations 

occurred during late summer and fall of 1997, when Little Bean 

Marsh was filled with low specific-conductance (less than 350 

μS/cm) surface water with little buffering capacity. Diurnal 

fluctuations of pH usually ranged from 0.2 to 0.5 units, with the 

minimum values typically occurring between 0800 and 1000 

hours, and the maximum values around noon (fig. 18), although 

this pattern occasionally was concealed by other factors such as 

storm events. When most of the water in the marsh was from 

lightly buffered rainfall and runoff, the larger fluctuations of pH 

caused by algal carbon dioxide were much larger. 

The median of synoptic pH measurements in Little Bean 

Marsh ranged from 7.4 at site M1 (vegetated) to 8.2 at site M3 

(open water). The median pH of vegetated sites was significantly less than those of open-water sites. All six pairwise comparisons of vegetated and open-water sites had significantly difAND ,OWBUFFERING

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ferent median pH values (p<0.05). This difference can be 

explained by daytime sampling, when algal photosynthesis 

increases the pH (Horne and Goldman, 1994) at the open-water 

sites more than vegetated sites. Only 1 of the 4 pairwise comparisons of same-cover sites had significant differences at the 

95 percent confidence level (p<0.05). 

Nitrogen 

Although most emergent macrophytes receive 95 to 99 

percent of their nutrients from the soil (Wetzel, 1983), enrichment of N and other nutrients in the water column can affect the 

plant composition of a marsh. For example, nutrient enrichment 

can cause algal blooms, decrease water transparency and 

impede growth or kill macrophytes (Stewart and Ornes, 1973; 

Carter and Haramis, 1980; Carter and others, 1985; Carter, 

1986). Average total N concentrations in Little Bean Marsh 

ranged from 0.6 mg/L at site M1 to 1.7 mg/L at site M6 (fig. 11). 

Concentrations of N would have been larger except for the near 

absence of inorganic N (NH4+NO3+NO2-N) in Little Bean 

Marsh (figs. 19 and 20). Apparently, denitrification and biotic 

uptake keep dissolved inorganic N concentrations low. The 

average percentage of total N in the dissolved organic form at 

the five sites ranged from 61 percent at site M5 to 78 percent at 

site M4. Most of the remaining N was particulate (attached to 

particles) organic N (table 3). Therefore, most N lost from the 

marsh was organic N, and most of that was dissolved. This situation contrasts with those of the Missouri River and other 

streams in the region where oxidizing conditions dominate and 

the dominant N species is NO3 (table 4). Nitrogen in the organic 

form generally is less usable by plants than inorganic N. Therefore, artificial flood-plain drainage that routes water directly to 

the river and bypasses wetlands are more likely to contribute to 

excess inorganic N in the river where deep water and high turbidity can limit uptake of the readily available forms of N. 

Generally, only the largest storms (greater than 5 cm) produced enough runoff to measurably affect the concentrations of 

nutrients in Little Bean Marsh. Concentrations of both total N 

and total dissolved N in the marsh generally were unchanged by 

storms (figs. 19 and 20, table 3) at outflow site M5. At site M2, 

concentrations of total N and NO3 also increased after the second largest storm (fig. 20), which occurred February 20 to 22, 

1997. This site was located near the largest tributary of Little 

Bean Marsh (fig. 1). N concentrations at site M3, which is 

somewhat isolated from tributaries, showed little response to 

any runoff events (figs. 19 and 20). The reduced quantities of 

surface runoff limits the Marsh’s allochthonous sources of 

nutrients, and concentrations of N and P in precipitation usually 

are much less than those present in surface runoff. Consequently, the response of N concentrations in the marsh to rainfall events is erratic and none systematic. 

No statistically significant differences were seen among 

any sites for dissolved NO3 (p = 0.15), dissolved NH4 (p = 

0.22), and total NH4 (p = 0.30) using the Kruskal-Wallis test on 

ranks of the data. This lack of relation primarily was because of 

the large number of NO3 concentrations that were less than the 

laboratory reporting limit, which were estimated to be one-half 

of the reporting limit. Consequently, the full variability of NO3 

concentrations between sites was not available for the Tukey 

test. Nevertheless, the large demand on inorganic N from nitrification, denitrification, and biotic uptake keeps concentrations 

and spatial variability of NO3 minimal within the marsh. 

No significant differences were detected among sites M2, 

M3, M4, and M5 for total N and total dissolved N (p>0.05) 

using Tukey’s multiple comparison test. However, concentrations of total N and total dissolved N were significantly less at 

site M1 (p<0.05) than all other sites despite being a heavily vegetated site. This difference may have been because site M1 is 

near the largest surface inflow and is in the area of greatest 

ground-water inflow. Consequently, the water at site M1 may 

not have had a chance to dissolve or suspend large quantities of 

organic N from the marsh vegetation. Also, any NO3 seeping 

into the marsh from ground water may be removed in the anoxic 

bottom sediments (Chen and others, 1972) before entry into the 

water column. 
Water Chemistry 31 

















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32 Hydrology and Cycling of Nitrogen and Phosphorus in Little Bean Marsh: A Remnant Missouri-River Wetland 





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Table 4. Means and ranges of selected constituent concentrations and properties of water samples collected from the Missouri River, Little Bean Marsh, and other wetlands 

Midwestern United States. 

[mg/L, milligrams per liter; μS/cm, microsiemens per centimeter at 25 degrees Celsius; --, no data; <, less than] 

Total organic Total 

Number of Period plus ammonia ammonium Total Ortho- Specific 

sample of data Total nitrogen, nitrogen, nitrogen, Nitrate nitrogen, phosphorus, phosphorus, conductance, 

Site Statistic dates collection in mg/L in mg/L in mg/L in mg/L in mg/L in mg/L in μS/cm 

Missouri River Mean 14 Dec. 1996– 1.1 0.1 1.1 0.23 0.06 718 

at St. Joseph, MOa Range Dec. 1997 0.36–2.50 0.03–0.21 0.32–1.7 0.10–0.57 0.030–0.11 650–830 

Heron Pond, Mean 6–9 Aug. 1976– 1.64 1.00 <0.01 0.47 0.16 

Johnson Co., ILb Range July 1977 

Eagle Bluffs wetland Mean 5 Dec. 1996– 2.6 c0.88 2.2 0.97 0.72 899 

Boone Co., MOd Range Dec. 1997 1–3.9 c0.19–1.43 0.41–5.1 0.8–1.21 0.53–0.91 774–1,070 

Eagle Bluffs scour Mean 3 Dec. 1996– 0.98 c0.54 0.05 0.082 <0.01 801 

Boone Co., MOd Range Dec. 1997 0.88–1.1 c<0.015–0.92 <0.05–0.068 0.05–0.13 <0.01–0.01 763–801 

Little Bean Marsh Mean 12 March 1996– 0.58 0.54 0.03 0.047 0.21 0.1 730 

Site M1 Range Dec. 1997 0.41–0.77 0.40–0.77 <0.05–0.10 <0.005–0.24 0.074–0.74 0.031–0.61 530–1,150 

Little Bean Marsh Mean 27 Dec. 1996– 1.12 1.09 0.086 0.029 0.16 0.033 694 

Site M2 Range Dec. 1997 0.35–2.77 0.35–2.76 0.003–0.32 <0.005–0.39 0.034–0.37 0.006–0.074 294–1,220 

Little Bean Marsh Mean 26 Dec. 1996– 0.97 0.96 0.077 0.013 0.12 0.025 560 

Site M3 Range Dec. 1997 0.44–1.90 0.43–1.84 0.009–0.41 <0.005–0.064 0.030–0.23 0.004–0.16 250–1,210 

Little Bean Marsh Mean 11 March 1996– 1.03 1.03 0.062 <0.005 0.14 0.052 510 

Site M4 Range Dec. 1997 0.67–1.29 0.67–1.29 0.018–0.13 <0.005–<0.005 0.076–0.30 0.007–0.23 320–700 

Little Bean Marsh Mean 30 Dec. 1996– 1.16 1.13 0.18 0.026 0.16 0.017 500 

Site M5 Range Dec. 1997 0.46–3.00 0.42–3.00 0.003–1.32 <0.005–0.20 0.037–0.60 0.002–0.072 260–910 

aHauck and others, 1999. bMitsch, 1978. cDissolved phase only. dRichards, 1999. eEleven to 13 samples collected between November 15, 1996, and June 25, 1997, only. 
34 Hydrology and Cycling of Nitrogen and Phosphorus in Little Bean Marsh: A Remnant Missouri-River Wetland 

Phosphorus 

The rerouting of flow from Short Creek around Little Bean 

Marsh in the 1960’s undoubtedly decreased the amount of P 

entering the marsh. The fine sediments present in a loessal 

stream such as Short Creek likely added substantial quantities 

of attached P (Craft and Casey, 2000) in the past; however, that 

stream now bypasses Little Bean Marsh. Concentrations of total 

P increased in the marsh in response to the large storm of April 

10 and 11 (fig. 21, table 3) at outflow site M5. At site M2, concentrations of total P also increased after the second largest 

storm, which occurred on February 20 to 22, 1997. This site was 

located near one of the larger tributaries. The minimal amounts 

of runoff from all other storms generally precluded increases in 

P concentrations from storm runoff. 

Dissolved to total P ratios for the five marsh sites ranged 

from 0.30 to 0.43, indicating that the majority of the P usually 

was in the particulate phase (table 3), although these ratios varied widely for individual samples (range = 0.04 to 1.00) with 

time. Average annual ratios of soluble reactive P (free PO4-P 

plus metal-bound PO4-P) to dissolved P for five sites (M1, M2, 

M3, M4, and M5) ranged from 0.50 to 0.85, indicating that most 

of the dissolved P usually was PO4. These ratios also ranged 

widely (from 0.12 to 1.0) for individual samples with time. 

These variabilities were likely because of varying amounts of 

algae, sediment, and senescing plants in wetland water in 

response to season, algal blooms, hydrologic events, and other 

sources of particulates. 

Although mean concentrations of total P and total dissolved P were both larger at vegetated sites (M1, M4, and M2) 

than at open-water sites (M3 and M5), no statistically significant difference was detected in either constituent between vegetated and open-water sites for 13 concurrent samples collected 

between April and December 1997 (p>0.05). Soluble reactive P 

was significantly greater at site M1 (p<0.05) than at all other 

sites, despite the fact that it was a heavily vegetated site. Those 

greater concentrations may have been because site M1 is 

located in the area of largest surface- and ground-water inflow 

to the marsh and the lack of time for emergent vegetation to 

remove substantial amounts of PO4 from the water. Emergent 

vegetation is slower to take up PO4 from the water column than 

algae and nonrooted vegetation that are shaded out by emergent 

plants at site M1. 

Turbidity and Total Suspended Solids 

The annual median turbidities for sites M2 and M3 were 10 

formazine turbidity units (FTU) and 17 FTU for site M5. The 

largest turbidities were associated with the two largest storm 

events and a September algal bloom, which increased turbidity 

to 115 FTU (table 3). Excluding two September 1997 samples 

collected during an algal bloom, the total range of turbidities 

from all three sites was only 5 to 35 FTU. Turbidities usually 

were low in Little Bean Marsh (table 3) for several reasons: sediment loadings from the largely flood-plain drainage area were 

minimal, emergent vegetation shade out algae and shield the 

water from wind, and high concentrations of bivalent cations 

increase flocculation rates of inorganic suspended material. The 

clarity of water in Little Bean Marsh contrasts with northern 

Missouri reservoirs where inorganic material often limits light 

penetration and subsequent algal growth (Jones and Knowlton, 

1993). Akhurst and Breen (1988) determined that floculation 

rates of inorganic material in shallow African lakes increased 

from 0.6 to almost 1.7 percent per hour as specific conductance 

increased from 370 to 622 μS/cm. The settling rate of clays at 

1,054 μS/cm was nearly 2 percent per hour. Settling rates of 

clays are preferentially increased by bivalent ions such as calcium and magnesium that dominated the ionic content of the 

Little Bean Marsh (table 5). The high divalent cation content is 

enhanced by ground water received by Little Bean Marsh, 

which indirectly increases the trapping efficiency of sedimentaffinitive constituents such as P. Also, ground water generally 

is much less turbid than surface water when they first enter the 

wetland. Although emergent vegetation increases water clarity 

in shallower areas, high ionic-content water can decrease the 

duration of high-turbidity episodes (caused by wind, runoff 

events, and backwater through the flood gates) in open-water 

areas by increasing the flocculation rate. For example, approximately one week after the 12.5 cm rain on April 10–11, 1997, 

turbidities had decreased to 5, 35, 22, and 30 FTU at sites M1, 

M2, M3, and M5 (table 3). Consequently, the amount of ground 

water received by other wetlands in the lower Missouri River 

flood plain may have a substantial effect on water clarity. A 

more incised, ground-water-fed wetland may be expected to 

have better water clarity than a shallow surface-water-fed wetland. In addition to enhancing algal growth, increased water 

clarity also can enhance the growth and establishment of submerged and emergent vegetation. 

The average concentration of volatile suspended solids 

expressed as a percentage of total suspended solids for each 

marsh site ranged from 28 to 56 percent with an average of 35 

percent. Assuming chlorophyll represented 2 percent of the living algal biomass, mean concentrations of algae at the marsh 

sites ranged from 10 to 22 percent of the volatile suspended solids, indicating most of the organic seston was detritus. 

The turbidity and suspended-solids differences between 

open-water and vegetated sites was likely caused by light limitation of algae in areas shaded by emergent macrophytes. The 

spatial variability of turbidity was similar to total suspended 

solids in that vegetated sites M1, M2, and M4 had lesser turbidities and suspended solids concentrations than open-water sites 

M3 and M5. Tukey’s multiple comparison test indicated differences in concentrations of total suspended solids at vegetated 

sites M1 and M4 and open-water sites M2, M3, and M5 were 

statistically significant (p<0.05). Turbidity at vegetated sites 

M1, M2, and M4 was significantly less than at the two openwater sites M3 and M5 (p<0.05). Site M2 was less densely vegetated than sites M1 and M4, and both turbidities and suspended 

solids concentrations at site M2 fell in between values for the 

open-water sites (M3 and M5) and the more densely-vegetated 

sites. 
Water Chemistry 35 

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36 Hydrology and Cycling of Nitrogen and Phosphorus in Little Bean Marsh: A Remnant Missouri-River Wetland 

Table 5. Concentrations of selected ions in water samples collected from Little Bean Marsh. 

[mg/L, milligrams per liter; --, no data] 

Site Dissolved Dissolved Dissolved Dissolved Dissolved Dissolved 

number Alkalinity, bicarbonate, chloride, calcium, magnesium, sodium, potassium, 

(fig. 1) Date in mg/L in mg/L in mg/L in mg/L in mg/L in mg/L in mg/L 

M2 02/05/1997 554 688 

M2 02/17/1997 688 840 

M2 11/19/1997 532 4.6 168.0 47.4 24.5 16.1 

M3 01/20/1997 681 849 

M3 02/05/1997 598 747 

M3 11/19/1997 280 5.3 80.5 30.9 16.5 13.1 

M4 11/19/1997 295 5.7 86.4 30.7 17.0 14.0 

M5 01/20/1997 456 556 

M5 02/05/1997 255 317 

M5 02/17/1997 480 586 

M5 10/28/1997 273 

M5 11/19/1997 3.8 74.7 29.0 16.8 11.7 

Total suspended solids were not significantly correlated 

with windspeed at open-water sites M2, M3, and M6 (Spearman’s rho ranged from 0.30 to 0.54 and p ranged from 0.10 to 

0.13) or at the vegetated sites (maximum Spearman’s rho = 

0.02, minimum p = 0.49) M1 and M4. This result contrasts with 

the open-water oxbows sampled by Knowlton and Jones (1997) 

where a good relation existed. The minimal effect of wind on 

total suspended solids was likely because of the wind-breaking 

affects of emergent macrophytes. Wind affects were limited 

even in open water areas by emergent macrophytes that break 

these areas up into small patches with short fetches. Relations 

between windspeed and total suspended solids also may have 

been masked, at least in September 1997, by algal blooms that 

were not related to windspeed. 

Dissolved Organic Carbon 

Dissolved organic carbon (DOC) samples were collected 

from Little Bean Marsh in the winter and spring of 1997 only. 

Concentrations averaged 6.2 mg/L, and ranged from 2.9 to 12.1 

mg/L (tables 3 and 4). These values are large enough that they 

probably do not limit denitrification rates in Little Bean Marsh. 

These values also are somewhat larger than the narrow range of 

DOC measured in samples collected from the Missouri River at 

Herman, Missouri, in 1997 (Hauck and others, 1998), and are 

evidence that wetland outflows from vegetated marshes, such as 

Little Bean, increase concentrations of DOC in the Missouri 

River. 

Trophic Indicators 

Particulate N:P ratios were nearly always less than 22 (fig. 

22), which is indicative of N deficiency in phytoplankton 

according to Hecky and Hendzel (1980). This relative deficit in 

N is likely because of the removal of N from the marsh through 

denitrification rather than an overabundance of P. N:P ratios in 

the inorganic fractions were less than 10, which also is indicative of N-limited growth of phytoplankton (Horne and Goldman, 1994). 

Chlorophyll concentrations were significantly correlated 

with total N at open-water sites (Spearman’s rho = 0.69 to 0.94, 

p = 0.003 to 0.03) in Little Bean Marsh (table 6). Correlations 

with total P usually were not as strong. These relations also indicate that N typically is more limiting than P. N limitation is not 

uncommon in other types of wetlands (Kibby, 1978). The lack 

of correlation between total N and chlorophyll at vegetated site 

M4 is evidence of light limitation. 

Jones and Knowlton (1993) classified the trophic state of 

Missouri reservoirs using concentrations of chlorophyll a, total 

N, or total P (table 7). Concentrations of these constituents were 

determined in samples from the six sampling sites in Little Bean 

Marsh collected in 1997 (M1 to M6). Comparison of these criteria with concentrations in Little Bean Marsh indicate the 

marsh typically is eutrophic or hypereutrophic, regardless 

which of the three criteria are used. If the high primary productivity of the water column indicated by the trophic state was 

added to that of the macrophytes, Little Bean Marsh is indeed a 

highly productive system. Four Missouri River oxbow lakes 

sampled by Knowlton and Jones (1997) were even more 

eutrophic than Little Bean Marsh (table 7). The mean concentrations of total N and total P were less than those of the four 
Water Chemistry 37 















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Table 6. Correlation coefficients (r), coefficients of determination (r2), and levels of significance (p) for the relation between chlorophyll 

concentrations and concentrations of total nitrogen, total phosphorus, and total suspended solids in Little Bean Marsh. 

[N, number] 

Observations Total nitrogen Total phosphorus Total suspended solids 

Spearman’s Spearman’s Spearman’s 

Site N rho p rho p rho p 

M2 (semi-open water) 10 0.94 0.003 0.71 0.018 0.26 0.16 

M3 (open water) 10 .69 .02 .44 .095 .59 .015 

M4 (vegetated) 7 -.32 .80 -.39 .842 .09 .52 

M5 (open water) 10 .76 .01 .76 .002 .64 .005 

M6 (open water) 6 .89 .03 .71 .063 .83 .011 
--

--

--

--

38 Hydrology and Cycling of Nitrogen and Phosphorus in Little Bean Marsh: A Remnant Missouri-River Wetland 

Table 7. Classification of the trophic state of Missouri reservoirs used by Jones and Knowlton (1993); May–September means of total 

phosphorus, total nitrogen, chlorophyll a, and total suspended solids for Missouri reservoirs by physiographic region (Jones and 

Knowlton, 1993); means of monthly samples of four oxbows along the Missouri River (Knowlton and Jones, 1997) from 1994–1996; and 

1997 annual means for Little Bean Marsh. 

[μg/L, micrograms per liter; mg/L, milligrams per liter; <, less than; --, no data; >, greater than] 

Trophic state or Chlorophyll a, Total nitrogen, Total phosphorus, Total suspended solids, 

water body in μg/L in mg/L in mg/L in mg/L 

Trophic State 

Oligotrophic <3 <0.3 <0.01 

Mesotrophic 3–7 0.3–0.5 0.01–0.025 

Eutrophic 7–40 0.5–1.2 0.025–0.100 

Hypereutrophic >40 >1.2 >0.10 

a

Missouri Reservoirs 15 0.6 0.045 11 

a

Glaciated Plains Region 15 0.7 0.047 10 

a

Osage Plains Region 23 0.8 0.067 20 

bFour oxbows along 113 2.3 0.371 103 

Missouri River 

Little Bean Marsh c56 d0.9 d0.114 e20 

a

Samples collected in the summers only from 1978 through 1984.


b

Samples collected at monthly intervals between June 13, 1994, and June 18, 1996.


c

Samples collected at biweekly intervals from July 29 through December 19, 1997.
 dSamples collected at biweekly intervals for all of 1997.
 e

Samples collected at biweekly intervals from April 17 through December 19, 1997.


oxbows (fig. 23), and the ratios of chlorophyll a to both N and 

P were greater than these ratios in the oxbows, perhaps indicating more detritus in the oxbows. The mean concentrations of all 

three constituents were larger in these open-water oxbows than 

in Little Bean Marsh. Periods of hypereutrophy in Little Bean 

Marsh, as indicated by concentrations of chlorophyll and total 

P, occurred for much of the summer at open-water sites M3, M5, 

and at partially open site M2 (figs. 20 and 21). Vegetated sites 

had less persistent periods of hypereutrophy, especially as indicated by chlorophyll a concentrations. The most intense periods 

of hypereutrophy occurred in September, coinciding with both 

an algal bloom and senescence of Scirpus fluviatilis and Typha 

latifolia. Even total N concentrations indicated hypereutrophic 

conditions during this period. These two events may not have 

been coincidental. The rapid leaching of N that occurs soon 

after macrophyte senescence may have provided a load of N to 

the N-limited open-water areas, which triggered a phytoplankton bloom and an intense period of hypereutrophy. However, the 

time interval between sample collection was too long to document this sequence of events. The triggering of fall periods of 

hypereutrophy by senescing macrophytes could be a common 

phenomenon in remnant riparian wetlands along the Missouri 

River where emergent vegetation is abundant. The period of 

intense hypereutrophy was ended abruptly in early October by 

a 2-inch rain and cold temperatures (figs. 10 and 15). 

Eutrophication and hypereutrophication occurred in Little 

Bean Marsh in spite of the fact that most of the marsh is shaded 

by macrophytes and is N limited. However, the somewhat lesser 

concentrations of trophic indicators in Little Bean Marsh, compared to the four open-water oxbows of Knowlton and Jones 

(1997) (table 7, fig. 23), are most pronounced in the large areas 

of emergent macrophytes and may be caused, in part, by shading by that vegetation. This interpretation indicates some light 

limitation in addition to the N limitation identified earlier. Chlorophyll a concentrations were significantly correlated with total 

suspended solids at open-water sites M3, M5, and M6 (table 6), 

but were not significantly correlated at vegetated sites M2 and 

M4. Therefore, at open-water sites, a substantial part of the particulate matter probably is phytoplankton, but at light-limited 

vegetated sites is either detritus or mineral sediment. 

Nitrogen Cycling and Denitrification 

One of the primary beneficial functions commonly 

ascribed to wetlands is the removal of N from surface waters. 
Nitrogen Cycling and Denitrification 39 

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Therefore, the following section addresses N cycling and denitrification in Little Bean Marsh. 

Sinks and Sources of Nitrogen 

Nitrogen limitation of algal growth and delivery of excess 

N into the Missouri River, and ultimately the Gulf of Mexico, 

make quantification of N cycling and denitrification in riparian 

wetlands an important element in the management of trophic 

problems such as Gulf Coast hypoxia (Goolsby and others, 

1999). However, the compartments and processes of N in wetlands are many and complex (fig. 24). Therefore, detailed analysis and special denitrification experiments were done on Little 

Bean Marsh to assess its effectiveness in removing N from surface water before it enters the Missouri River. 

The amount of N leaving the marsh in surface outflow is 

critical to evaluating the nutrient removal function of a marsh. 

Outflow loads can be used to evaluate the impact of the marsh 

on receiving water bodies, such as the Missouri River, as well 

as evaluating the relative impacts of various N sources to the 

marsh. Approximately 530 kg (kilograms) of total N left Little 

Bean Marsh in surface outflows in 1997 (table 8). Dividing this 

value by the 952 ha drainage area gives an areal yield of 0.56 

kg/ha (kilogram per hectare), which is less than one-half of the 

1.4 kg/ha areal yield of the Missouri River at St. Joseph, Missouri, in 1997 (Hauck and others, 1998) and slightly more than 

one-half of 0.9 kg/ha average total N yield of the entire Missouri 

River Basin in 1980–1996 (Goolsby and others, 1999). The 

majority of the N in Little Bean Marsh was dissolved organic N; 

very little was inorganic N (table 3). The largest N loads leaving 

the marsh in surface outflows occurred in response to the two 

largest storms on February 20–22 and April 10–11, 1997. 

Atmospheric Contributions 

Direct atmospheric inputs of N to wetlands generally are 

small (Peterjohn and Correl, 1984; Bowden, 1987; Morris, 

1991). Concentrations of N in rainfall generally are less than 

those present in the Missouri River (Hauck and others, 1998; 

http://nadp.sws.uiuc.edu/nadpdata), or in runoff from its tributaries in Missouri and Kansas. However, because precipitation 

constituted 41 percent of the inflow to Little Bean Marsh and 

other sources of N were relatively small, precipitation was the 

second largest source of N to Little Bean Marsh. Data from the 

National Atmospheric Data Program (http://nadp.sws.uiuc.edu/ 

nadpdata, accessed 2000) show deposition of NO3 and NH4-N 

at two sites within 322 km and on either side of Little Bean 

Marsh in 1997 were 4.7 and 6.0 kg/ha respectively. Using an 

annual precipitation loading of 5.4 kg/ha and an average surface 

area of 81 ha, the estimated precipitation load of NO3-N and 
$IFFUSION )NFILTRATION

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Table 8. Loads of selected constituents in surface outflows from Little Bean Marsh. [Bold figures are total ammonium (NH4) estimated from ratio of total to dissolved phase of NH4; m3/s, cubic meters per second; g/d, grams per day; g/period, grams per period; kg, kilograms] 

Total Total Dissolved Suspended 

Site Mean Total dissolved suspended ammonium ammonium ammonium Nitrite plus 

number 

(fig. 1) Date 

discharge, 

in m3/s 

Period 

(days) 

Total nitrogen, 

in g/d 

nitrogen, 

in g/d 

nitrogen, 

in g/d 

nitrogen, 

in g/d 

nitrogen, 

in g/d 

nitrogen, 

in g/d 

nitrate, 

in g/d 

M5 11/15/1996 0.0054 17 651 465 186 160 93 67 63 M5 12/02/1996 .0003 17 21 12 9 5 0 5 0 M5 12/19/1996 .074 17 5,853 4,708 1,145 732 16 719 16 M5 01/03/1997 .013 15 1,093 518 575 426 3 423 3 M5 01/20/1997 .0048 17 337 229 108 1 1 0 1 M5 02/05/1997 .002 14 79 55 24 7 0 7 8 M5 02/17/1997 .000 8 0 0 0 0 0 0 0 M5 02/21/1997 .0037 3 188 99 89 19 1 18 8 M5 02/24/1997 .146 12 12,097 4,536 7,561 1,096 32 1,071 769 M5 03/17/1997 .057 18 573 206 367 9 1 8 1 M5 03/31/1997 .002 10 257 115 142 11 0 11 0 M5 04/05/1997 .002 6 231 106 125 0 1 0 0 M5 04/10/1997 .042 1 4,478 2,973 1,505 0 440 0 73 M5 04/11/1997 .747 3 51,031 43,280 7,752 0 3,294 0 4,974 M5 04/17/1997 .000 13 0 0 0 0 0 0 0 M5 05/05/1997 .000 17 0 0 0 0 0 0 0 M5 05/21/1997 .000 18 0 0 0 0 0 0 0 M5 06/10/1997 .0085 17 646 499 147 12 6 7 2 M5 06/25/1997 .018 15 1,256 1,161 95 159 4 156 4 M5 07/08/1997 .0300 16 1,245 2,179 0 44 23 21 6 M5 07/29/1997 .013 18 1,208 943 265 46 18 28 3 M5 08/12/1997 .001 15 148 131 17 14 10 4 0 M5 08/27/1997 .000 14 0 0 0 0 0 0 0 M5 09/10/1997 .000 15 0 0 0 0 0 0 0 M5 09/24/1997 .000 16 0 0 0 0 0 0 0 M5 10/14/1997 .0051 17 595 599 0 258 238 20 21 M5 10/28/1997 .002 18 123 126 0 19 10 9 6 M5 11/19/1997 .000 18 0 0 0 0 0 0 0 M5 12/02/1997 .001 14 94 66 28 2 1 1 1 M5 12/19/1997 .015 21 1,216 872 344 37 3 34 3 
Table 8. Loads of selected constituents in surface outflows from Little Bean Marsh.—Continued 

[Bold figures are total ammonium (NH4) estimated from ratio of total to dissolved phase of NH4; m3/s, cubic meters per second; g/d, grams per day; g/period, grams per period; kg, kilograms] 

Dissolved Soluable Total Volatile 

Site Total organic organic Total Dissolved reactive Suspended Chlorophyll suspended suspended 

number nitrogen, nitrogen, phosphorus, phosphorus, phosphorus, phosphorus, mean, solids, solids, 

(fig. 1) Date in g/d in g/d in g/d in g/d in g/d in g/d in g/d in g/d in g/d 

M5 11/15/1996 428 309 53 8 8 45 0 0 0 M5 12/02/1996 15 12 2 1 0 1 0 0 0 M5 12/19/1996 5,109 4,676 439 108 13 331 0 0 0 M5 01/03/1997 665 512 43 17 9 25 0 0 0 M5 01/20/1997 335 227 29 8 3 21 0 0 0 M5 02/05/1997 64 47 9 3 2 6 0 0 0 M5 02/17/1997 0 0 0 0 0 0 0 0 0 M5 02/21/1997 161 90 41 5 4 36 0 0 0 M5 02/24/1997 10,232 3,743 1,096 202 139 895 0 0 0 M5 03/17/1997 562 203 69 83 2 0 0 0 0 M5 03/31/1997 245 114 70 4 1 67 0 0 0 M5 04/05/1997 231 104 40 5 2 35 0 0 0 M5 04/10/1997 4,404 2,459 551 139 73 411 0 141,672 20,921 M5 04/11/1997 46,058 35,011 4,457 2,196 1,227 2,261 0 258,387 148,573 M5 04/17/1997 0 0 0 0 0 0 0 0 0 M5 05/05/1997 0 0 0 0 0 0 0 0 0 M5 05/21/1997 0 0 0 0 0 0 0 0 0 M5 06/10/1997 632 492 43 38 21 5 0 18,645 3,964 M5 06/25/1997 1,094 1,153 159 70 40 89 0 35,467 7,952 M5 07/08/1997 1,196 2,150 399 80 49 319 0 141,354 25,937 M5 07/29/1997 1,159 922 166 41 25 124 38,871 33,351 7,475 M5 08/12/1997 133 120 20 4 3 16 4,906 3,426 648 M5 08/27/1997 0 0 0 0 0 0 0 0 0 M5 09/10/1997 0 0 0 0 0 0 0 0 0 M5 09/24/1997 0 0 0 0 0 0 0 0 0 M5 10/14/1997 315 340 70 14 7 57 6,034 15,503 3,171 M5 10/28/1997 98 110 16 3 1 13 1,864 2,349 587 M5 11/19/1997 0 0 0 0 0 0 0 0 0 M5 12/02/1997 91 64 17 8 7 8 1,390 1,370 587 M5 12/19/1997 1,176 865 145 66 26 79 19,423 14,402 4,228 

13-month total in kg 12-month total in kg 
Table 8. Loads of selected constituents in surface outflows from Little Bean Marsh.—Continued 

[Bold figures are total ammonium (NH4) estimated from ratio of total to dissolved phase of NH4; m3/s, cubic meters per second; g/d, grams per day; g/period, grams per period; kg, kilograms] 

Total Dissolved 

Total Total Dissolved Suspended ammonium ammonium 

Site Total dissolved suspended ammonium ammonium ammonium Nitrite plus plus organic plus organic Total organic 

number nitrogen, nitrogen, nitrogen, nitrogen, nitrogen, nitrate, nitrogen, nitrogen, nitrogen, 

(fig. 1) Date in g/period in g/period in g/period in g/period in g/period in g/period in g/period in g/period in g/period 

M5 11/15/1996 7,903 3,161 2,719 1,581 1,138 1,067 9,998 6,836 7,279 M5 12/02/1996 204 146 91 1 90 1 349 203 258 M5 12/19/1996 80,032 19,467 12,437 270 12,221 270 99,282 79,815 86,845 M5 01/03/1997 7,763 8,625 6,383 43 6,348 43 16,353 7,728 9,971 M5 01/20/1997 3,889 1,839 18 18 0 18 5,714 3,875 5,692 M5 02/05/1997 767 336 103 6 98 106 998 662 894 M5 02/17/1997 0 0 0 0 0 0 0 0 0 M5 02/21/1997 296 267 56 2 54 25 538 271 482 M5 02/24/1997 54,437 90,729 13,156 378 12,853 9,224 135,942 45,213 122,787 M5 03/17/1997 3,700 6,606 167 22 150 22 10,289 3,682 10,121 M5 03/31/1997 1,145 1,424 115 4 112 4 2,566 1,142 2,452 M5 04/05/1997 637 750 16 8 8 3 1,385 635 1,385 

M5 04/10/1997 2,973 1,505 880 440 440 73 4,404 2,900 4,404 

M5 04/11/1997 129,840 23,255 19,766 9,883 9,883 14,922 138,173 114,918 138,173 

M5 04/17/1997 0 0 0 0 0 0 0 0 0 M5 05/05/1997 0 0 0 0 0 0 0 0 0 M5 05/21/1997 0 0 0 0 0 0 0 0 0 M5 06/10/1997 8,486 2,496 212 100 112 31 10,956 8,461 10,744 M5 06/25/1997 17,415 1,431 2,386 60 2,338 60 18,799 17,368 16,413 M5 07/08/1997 34,859 0 705 373 332 104 19,836 34,776 19,131 M5 07/29/1997 16,974 4,761 828 331 497 52 21,694 16,933 20,866 M5 08/12/1997 1,964 257 217 154 62 5 2,217 1,960 2,000 M5 08/27/1997 0 0 0 0 0 0 0 0 0 M5 09/10/1997 0 0 0 0 0 0 0 0 0 M5 09/24/1997 0 0 0 0 0 0 0 0 0 M5 10/14/1997 10,183 0 4,388 4,043 344 359 9,749 9,823 5,361 M5 10/28/1997 2,273 0 341 188 153 108 2,111 2,164 1,771 M5 11/19/1997 0 0 0 0 0 0 0 0 0 M5 12/02/1997 918 397 27 15 12 14 1,302 904 1,274 M5 12/19/1997 18,313 7,214 777 69 721 69 25,472 18,258 24,695 

13-month total in kg 405 175 66 18 48 27 538 379 493 

12-month total in kg 379 164 62 16 46 25 502 353 461 
Table 8. Loads of selected constituents in surface outflows from Little Bean Marsh.—Continued 

[Bold figures are total ammonium (NH4) estimated from ratio of total to dissolved phase of NH4; m3/s, cubic meters per second; g/d, grams per day; g/period, grams per period; kg, kilograms] 

Soluable Total Volatile Dissolved 

Site Total Dissolved reactive Suspended Chlorophyll suspended suspended organic 

number phosphorus, phosphorus, phosphorus, phosphorus, mean, solids, solids, carbon, 

(fig. 1) Date in g/period in g/period in g/period in g/period in g/period in g/period in g/period in g/period 

M5 11/15/1996 901 142 142 759 0 0 0 56,114 M5 12/02/1996 30 9 0 21 0 0 0 2,829 M5 12/19/1996 7,462 1,839 216 5,624 0 0 0 735,425 M5 01/03/1997 638 259 138 380 0 0 0 115,577 M5 01/20/1997 495 141 57 354 0 0 0 57,985 M5 02/05/1997 127 43 24 84 0 0 0 9,592 M5 02/17/1997 0 0 0 0 0 0 0 0 M5 02/21/1997 124 16 13 108 0 0 0 0 M5 02/24/1997 13,156 2,419 1,663 10,736 0 0 0 0 M5 03/17/1997 1,242 1,497 35 0 0 0 0 0 M5 03/31/1997 705 44 15 667 0 0 0 0 M5 04/05/1997 242 32 13 210 0 0 0 0 M5 04/10/1997 551 139 73 411 0 141,672 20,921 0 M5 04/11/1997 13,372 6,589 3,682 6,783 0 775,161 445,718 0 M5 04/17/1997 0 0 0 0 0 0 0 0 M5 05/05/1997 0 0 0 0 0 0 0 0 M5 05/21/1997 0 0 0 0 0 0 0 0 M5 06/10/1997 736 649 349 87 0 316,965 67,386 0 M5 06/25/1997 2,386 1,050 596 1,336 0 532,006 119,284 207,554 M5 07/08/1997 6,391 1,286 788 5,104 0 2,261,671 414,985 0 M5 07/29/1997 2,981 745 455 2,236 699,671 600,310 134,552 0 M5 08/12/1997 294 61 40 233 73,589 51,384 9,726 0 M5 08/27/1997 0 0 0 0 0 0 0 0 M5 09/10/1997 0 0 0 0 0 0 0 0 M5 09/24/1997 0 0 0 0 0 0 0 0 M5 10/14/1997 1,198 232 120 966 102,577 263,555 53,909 0 M5 10/28/1997 293 61 24 233 33,561 42,282 10,570 0 M5 11/19/1997 0 0 0 0 0 0 0 0 M5 12/02/1997 233 118 99 115 19,457 19,183 8,221 0 M5 12/19/1997 3,052 1,387 555 1,665 407,885 302,445 88,791 0 

13-month total in kg 57 19 9 38 1,337 5,307 1,374 1,185 12-month total in kg 53 17 8 36 929 5,004 1,285 1,126 
NH4-N (inorganic N) to Little Bean Marsh in 1997 was approximately 437 kg; which was greater than 70 percent of the N 

leaving Little Bean Marsh in surface outflows. The 437 kg computation does not include organic N or N in dry atmospheric 

deposition. 

Lawrence and others (2000) reported that the dry deposition rate of NO3-N and NH4-N usually is about one-half of the 

wet deposition rate, and that the dominant form of dry N deposition in southern Minnesota, Iowa, and northern Missouri is 

NHx, which originates primarily from senescing crops within 

50 km of deposition sites. The closest dry deposition monitoring station to Little Bean Marsh was a U.S. Environmental Protection Agency Clean Air Status and Trends site at Alhambra in 

western Illinois (http://www.epa.gov/castnet, accessed 2000). 

The average annual NO3 and NH4 deposition rate at this site 

was 2.0 kg/ha during 1989 to 1999 and 1.7 kg/ha in 1997. When 

organic N and loadings from dry deposition are added to wet 

deposition of inorganic N, it is likely that atmospheric deposition alone could account for more than the entire 530 kg of N 

leaving the marsh in surface outflow. Consequently, it is likely 

that Little Bean Marsh removes much inorganic N, probably 

through denitrification. 

Contributions from Storm Runoff 

Storm loads from four storm events were measured at three 

sites in the spring of 1997. Loads were measured from two 

storms at site I1. These storms contributed 64,000 m3

 of surface 

runoff to Little Bean Marsh (table 2), which was approximately 

12 percent of the total surface runoff into the marsh in 1997. 

These four storm events contributed approximately 381 kg of N 

to Little Bean Marsh. If the loading of total N from these storms 

is prorated to the total annual amount of surface runoff (516,000 

m3), then surface runoff contributed several times the amount of 

N that was released in surface outflow from the marsh. When 

the storm runoff loads of N are added to the amount of N contributed by direct precipitation, the amount of N contributed to 

Little Bean Marsh from storms is almost seven times more than 

the amount of N leaving the marsh in surface outflow. Unlike N 

speciation in the marsh and in the surface outflow, a substantial 

amount, roughly one-third, of the N in storm loads was in the 

form of NO3. Another one-third was dissolved organic nitrogen, and the remaining one-third was suspended nitrogen that 

probably also is in the organic form. Little of the N in storm runoff entering the marsh was in the NH4 form. The small concentrations of NO3 in Little Bean Marsh and the small loads of NO3 

leaving the marsh are strong evidence of substantial N removal. 

Ground-Water Contributions 

Ground-water seepage usually is a small nutrient source to 

freshwater marshes (Mitsch and Gosselink, 1993). Even though 

the marsh receives substantial quantities of ground-water seepage, Little Bean Marsh receives little N from ground water 

because ground water has small concentrations of N (table 1). 

Nitrogen Cycling and Denitrification 45 

Nineteen of 21 ground-water samples collected from 5 monitoring wells around Little Bean Marsh (fig. 1) had concentrations 

of NO3 less than 0.02 mg/L in 1996–1997. Ziegler and others 

(1993) determined that more than one-half of the shallow wells 

in the Missouri River alluvium had concentrations of NO3 less 

than the detection limit of 0.05 mg/L. Ammonium (NH4-N) 

concentrations in Little Bean Marsh wells averaged only 0.05 

mg/L, but total dissolved N averaged 0.3 mg/L. Most of the dissolved N likely was organic N that entered ground water from 

the marsh during ground-water recharge periods. Ground water 

away from the marsh would likely have lesser concentrations of 

organic N. Using a net ground-water seepage of 157,400 m3 

(fig. 7) in 1997 and assuming total dissolved N concentrations 

in ground water averaged as high as 0.3 mg/L, ground-water 

seepage would have contributed less than 1 percent of the total 

N leaving Little Bean Marsh in surface outflow. However, NO3 

concentrations have been shown to decrease as ground water 

moves through anoxic lake-bottom sediments (Chen and others, 

1972). Consequently, the anoxic sediments of the marsh likely 

remove NO3 that may be present in ground water as it moves 

into the marsh, thus providing another pathway for N removal 

in the riparian corridor. Therefore, most wetlands along the 

Missouri River probably receive only small amounts of N from 

ground water. 

Reduction of NO3 to NH4 can occur under intensely reducing conditions, as bacteria can rapidly immobilize and then 

ammonify N (Chen and others, 1972), but this process generally 

is insignificant compared to denitrification (Reddy and others, 

1978) which actually removes N from the marsh. Reddy and 

others (1980) determined that 92 to 96 percent of isotopically 

labeled NO3-N escaped as gaseous N in laboratory incubations 

lasting 21 days, leaving 4 to 8 percent as organic N and NH4. 

Several authors have noted decreases in NO3-N in water passing through wetlands (Kibby, 1978); Little Bean Marsh shares 

this characteristic, as NO3 concentrations in the surface outflow 

commonly were less than 0.005 mg/L. 

Nitrogen fixation can supply about one-half of the total N 

input to a wetland wholly dependent on precipitation as a source 

of water. Asymbotic N fixation in a Scirpus atrovirens marsh in 

Massachusetts ranged from 0.4 to 2.0 g/m2/yr (grams per square 

meter per year) (Kana and Tjepkema, 1978), and a rate of 1.8 g/ 

m2

/yr was measured in situ in a Typha latifolia stand in Minnesota (Biesboer, 1984). However, N inputs from fixation generally are small compared to those from surface-water inflows, 

and are not a substantial source of N to downstream waters 

(Johnston, 1991). Also, the macrophytes that dominate Little 

Bean Marsh do not symbiotically fix N, but do shade out floating N fixers such as bluegreen algae. Even if the greater fixation 

rates detected in the Massachusetts and Minnesota marshes are 

assumed for Little Bean Marsh, the total amount of N supplied 

to the marsh by fixation would be less than 0.5 percent of the 

amount of total N escaping the marsh in surface outflows. 

Therefore, N fixation rates, which commonly vary widely over 

short distances, were not measured in this study. 

Whereas the input-output analysis in this report treats the 

whole marsh as a storage compartment, it is useful to review the 
46 Hydrology and Cycling of Nitrogen and Phosphorus in Little Bean Marsh: A Remnant Missouri-River Wetland 

wetland interdistributions and fluxes of N in other Scirpus/ 

Typha wetlands and compare their inputs and outputs with Little Bean Marsh. Standing stocks of Scirpus fluviatilis were 

0.066 kg/ha in an Iowa prairie pothole (Davis and Van der Valk, 

1978) and 0.154 kg/ha in a Wisconsin wetland (Klopatek, 

1978). Both of these densities are less than the mean (2.07 kg/ 

ha) of all nonwoody emergent vegetation in 30 wetlands around 

the United States (Johnston, 1991). The above-ground part of N 

in the Wisconsin wetland was about three times that present in 

the roots and rhizomes, and two times that present in the litter, 

for a total of 28 g/m2 (grams per square meter) (Klopatek, 

1978). At this rate, the amount of plant material at Little Bean 

Marsh was about 20,000 kg of N, which is much larger than the 

annual loss or gain of N to the marsh from inflows or outflows. 

Consequently, much more of the N in Little Bean Marsh is 

internally recycled every year than is lost to surface outflows. 

Perhaps Scirpus fluviatilis has adapted to the low inorganic N 

content of Little Bean Marsh by recycling N. Annual leaching 

fluxes of N in Scirpus fluviatilis were measured as 1.49 g/m2

/yr, 

and annual litterfall fluxes were 8.09 g/m2

/yr in a study by 

Davis and Van der Valk (1978). Klopatek (1978) reported an 

annual leaching rate of 7.34 g/m2 for Scirpus fluviatilis, which 

is 42 percent of the total uptake in the Wisconsin wetland. 

Applying these leaching rates to Little Bean Marsh would indicate that several times the amount of N were leached to the 

marsh that was lost in surface outflow in 1997. Because most 

leached inorganic N is rapidly oxidized to NO3 and subsequently lost to denitrification, macrophytes in Little Bean 

Marsh likely facilitate the removal of soil N from the marsh. 

The largest concentrations of total N occurred during the September to early October period of senescence (figs. 19 and 20). 

However, Scirpus fluviatilis is slow to decay (Whitley and others, 1990) and in the Iowa pothole, Scirpus fluviatilis retained 

more than 50 percent of its dry weight after 525 days of decomposition. Klopatek (1978) determined that 26 percent of the N 

uptake by Scirpus fluviatilis was retained during winter months, 

with the rest lost to leaching and litterfall. Consequently, much 

of the N bound in Scirpus fluviatilis is prone to burial as evidenced by the large quantity of N in bottom sediments of Little 

Bean Marsh (table 9). Even though most (95 to 99 percent) of 

the nutrients used by emergent macrophytes are from the soil 

rather than the overlying water (Wetzel, 1983), the amount of N 

brought back into the water column by emergent vegetation 

would likely be a tiny fraction of the amount in the soil. The 

ratio of NH4-N to total N in the bottom sediments would indicate that decomposition in bottom sediments is slow, and burial 

of N in bottom sediments also may be a substantial sink of N, 

despite the ability of Little Bean Marsh to remove inorganic N 

through denitrification. Once vegetation is buried in anoxic sediments below the rhizospheric root zone, decomposition would 

be slow and organic N would be more stable. 

Soils contain, by far, the largest stocks of nutrients of any 

wetland storage compartment (Johnston, 1991). In a Scirpus 

fluviatilis marsh in Wisconsin, soils between 3 and 15 cm deep 

contained two orders of magnitude more N than that present in 

the Scirpus fluviatilis alone (Klopatek, 1978). The mean total N 

concentration in the top 15 cm of bottom sediments collected 

from four locations in Little Bean Marsh was 1,230 parts per 

million (table 9), which is near the low end of the range of total 

N concentrations in the mineral soils of six wetlands in WisconTable 9. Concentrations of selected nutrients with depth in bottom sediment cores collected from Little Bean Marsh in 1997. 

[cm, centimeters; ppm, parts per million; >, greater than] 

Site Depth, Ammonium nitrogen, Nitrate nitrogen, Total nitrogen, Available phosphorus, 

(fig. 1) in cm in ppm in ppm in ppm in ppm 

M1 0–4 9.1 1.7 1,053 20 

M1 4–10 8.4 .5 644 24 

M1 10–21 26.3 3 2,084 24 

M1 21–35 12.9 2.4 1,718 27 

M2 0–8.5 16 .3 580 35 

M2 8.5–16 25 .3 1,257 29 

M2 16–18 36.9 1 1,775 26 

M3 0–3 59.4 4.1 3,900 22 

M3 3–8 37.4 1 1,517 27 

M3 8–16.5 51 .5 1,470 29 

M3 >16.5 20.6 1.7 1,223 15 

M5 0–6 40.6 .4 835 26 

M5 6–12.5 35.7 .4 494 31 

M5 12.5–19 67 .4 1,736 25 

M5 19–29.5 55.8 2.5 1,897 27 
-----------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------------

Nitrogen Cycling and Denitrification 47 

sin, North Dakota, and North Carolina [range = 900 to 11,000 

mg/kg (milligrams per kilogram); Johnston, 1991]. However, 

assuming a bottom-sediment density of 1.5 and a mean concentration of total N in bottom sediments of 1,230 parts per million, 

the top 15 cm of Little Bean Marsh contains 470 kg/ha of total 

N. This amount of N is still more than one order of magnitude 

greater than the amount of N entering the marsh in surface runoff, and more than two orders of magnitude greater than the 

amount of N in surface outflow during 1997. Consequently, 

burial of organic N is likely a substantial sink of N; however, 

most (95 to 99 percent) of the nutrients used by emergent macrophytes are from the soil, rather than the overlying water (Wetzel, 1983) so much of the N feeding emergent vegetation is 

recycled. Most of the bottom-sediment N exists in the organic 

form (table 9). NH4-N concentrations in bottom sediments were 

minimal probably because of plant uptake and nitrification in 

rhizospheres. However, NH4-N concentrations were significantly correlated with total N concentrations (Spearman’s rho = 

0.54, p = 0.022), indicating some new NH4 was being supplied 

through slow decomposition of organic N. The extremely low 

concentrations of NO3 could be expected, considering the 

reducing conditions present in all but the rhizospheric parts of 

the bottom sediments. No detectable trend in N with depth can 

be discerned from the data in table 9 and, surprisingly, no clear 

differentiation in total N concentration can be made between 

open-water sites (M3 and M5) and vegetated sites (M1 and 

M2). 

Because of the large percentage of N in emergent vegetation, macrophytes can increase nutrient concentrations at senescence when nutrients are leached into the water column. In fact, 

time trend plots of all nondry marsh sites (M5, M4, and M3) had 

peak concentrations of nearly every N species during the 

senescing period of September and early October (figs. 19 and 

20). Consequently, this is a period in which nutrients are vulnerable to escape. However, this period usually follows a period of 

low water levels, which also can prevent escape of nutrients in 

surface outflows, as happened at Little Bean Marsh in 1997. 

Denitrification 

In the absence of oxygen, many soil and water microbes 

can use NO3 as an electron acceptor for respiration. This process converts NO3-N to gaseous N2O and N2, and is thus 

described as denitrification (equation 6). Because anoxic conditions are prevalent in the inundated bottom sediments of Little 

Bean Marsh, widespread denitrification is possible. The chemical equations in the nitrification-denitrification system are: 

Measurements of Eh (oxidation/reduction potential) in 

bottom-sediment cores from Little Bean Marsh consistently 

were less than the 0.25-volt maximum Eh for denitrification 

(Mitsch and Gosselink, 1993) (fig. 25). Eh typically increased 

with depth, probably because of the increased organic matter 

and moisture content that was observed (not measured) near the 

top of the cores. Usually, the Eh ranged between 0.0 and 0.1 

volts in the top 10 or 15 cm, and increased more rapidly at 

depths below 15 cm. Because the water column usually had 

concentrations of dissolved oxygen greater than 0.3 mg/L, a 

thin oxidized layer likely exists on top of the sediment column 

(Patrick and Reddy, 1976; Mitsch and Gosselink, 1993). Eh 

measurements were not usually made at depths less than 0.5 cm 

because of the size of the Eh and temperature probes used. Evidently, the oxidized layer usually was less than 0.5 cm thick at 

Little Bean Marsh as oxidizing conditions were detected only 

occasionally near the top of the sediment column (fig. 25). 

Redox conditions in the open-water areas of Little Bean 

Marsh seem to follow a conceptual model of N cycling that has 

been observed in the laboratory (Patrick and Reddy, 1976) and 

seems to provide a large capacity for N removal. The thin oxidizing layer at the sediment/water surface and in rhizospheres 

oxidizes NH4 to NO3 and creates an NH4 gradient within the 

reduced sediments, which allows NH4 to diffuse upward toward 

the zone of nitrification. NO3 then diffuses back down into the 

reduced sediments from the oxidized sediments and water 

where NO3 is converted to N gases in the reduced sediments by 

denitrification, and is lost to the atmosphere. Diffusion of NO3 

into the sediments was shown to be the rate-limiting step in laboratory experiments and mathematical theory of denitrification 

in flooded soils by Reddy and others (1978) and Phillips and 

others (1978). Consequently, denitrification occurred at a first 

order rate (logarithmic with time) relative to soil-NO3 concentration, typical of diffusion processes. NO3 removal was faster 

in the absence of overlying water, and denitrification reverted to 

a zero-order rate (linear with time); however, increasing the 

water depth from 3 to 6 cm made little difference in NO3 

removal rates. In laboratory experiments conducted by Reddy 

and others (1980), NO3 diffused less than 3 cm into bottom sediments after 96 hours and diffusion rates were only minimally 

affected by temperature. Soil denitrification increased the diffusion rate compared to diffusion rates into sterilized soil by 

increasing the concentration gradient across the soil water interface. These experiments were conducted in anoxic conditions 

and did not examine the rate of nitrification that is likely the 

rate-limiting step at Little Bean Marsh as evidenced by the very 

low concentrations of NO3. 

- Nitrification 6NH4+ + 12O2 → 6NO3 + 6H2O + 12H+

Denitrification 6NO3

- + 1.25C6H12 O6 + 6H+ → 3N2 + 7.5CO2 + 10.5H2O (6)


Combined 6NH4

+ + 1.25C6H12O6 + 12O2 → 3N2 + 7.5CO2 + 16.5H2O + 6H+
48 Hydrology and Cycling of Nitrogen and Phosphorus in Little Bean Marsh: A Remnant Missouri-River Wetland $%04().#%.4)-%4%23



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Whereas the laboratory experiments described in the previous paragraph were designed to simulate open water systems, 

about 60 percent of Little Bean Marsh is covered with Scirpus 

fluviatalis and Typha latifolia. Bottom-sediment cores from 

macrophyte areas contained large quantities of rootlets that 

were easily identified by red ferric oxyhydroxide deposits that 

precipitated in the oxidizing rhizospheres around the rootlets. 

The oxidizing rhizospheres provide a massive intertwining of 

oxidizing and reducing zones, and many more chemical gradients over shorter distances. These rhizospheres are likely essential to the removal of soil N at depth. Phillips and others (1978) 

have shown that without rhizospheres, NO3 is rapidly [about 60 

μg/g (micrograms per gram) of soil per day where NO3 was not 

limiting] removed from saturated soil, and is present only in the 

top few centimeters after a few days. Without rhizospheres, the 

only source of NO3 is diffusion down from the water column. 

NH4 also has to diffuse up into the surface oxidizing zone. 

Whereas emergent macrophytes bury large amounts of organic 

N in the form of root stocks and detritus, rhizospheric plants 

substantially increase the rate of N removal from bottom sediments by increasing the rate of organic N decomposition, 

increasing the rate-limiting step of nitrification in rhizospheres, 

decreasing the length of chemical diffusion gradients, and substantially increasing the number of gradients. Thus, removal of 

soil N from vegetated areas is likely much more rapid and efficient than from open water areas, because the amount of soil N 

generally is much larger than in the overlying water. 

Denitrification is inhibited greatly or terminated when dissolved-oxygen concentrations exceed 0.3 mg/L, temperatures 

are below about 5 o

C, or pH is less than 5 (Bremmer and Shaw, 

1958; Stanford and others, 1975; Reddy and others, 1980; Broderick and others, 1988). Denitrification rates increase about 

10-fold between 5 and 10 °C, double every 10 °C between 11 

and 35 °C, and maximum denitrification rates occur at 60 to 65 

°C (Bremmer and Shaw, 1958) and, therefore, occur at the fastest rate during the highest temperatures of the summer. At pH 

values less than 6, reduction of N2O to N2 is strongly inhibited 

(Wijler and Delwiche, 1954). The entire range of pH values 

measured (concurrently with Eh) in the cores from Little Bean 

Marsh was 6.1 to 7.3, which is greater than these denitrificationlimiting values. Water pH was never less than 5 units (fig. 18), 

water temperatures were greater than 5 °C for about 9 months 

each year (fig. 14), and dissolved oxygen was less than 0.3 mg/ 

L only for a month or less each summer (fig. 12) in Little Bean 

Marsh. Therefore, denitrification is strongly inhibited by dissolved oxygen in the water column for about 11 months each 

year, and most denitrification must occur in the bottom sediments. 

The results of in situ denitrification experiments are presented in table 10 and summarized as means in figure 26. 

Apparently, the small ambient concentrations of NO3 were the 

primary factor limiting spring, summer, and fall denitrification 

rates as all ambient incubations yielded little or no detectable 

N2O. However, incubations with added NO3 had large accumulations of N2O, indicating that Little Bean Marsh has a large 

unused capacity to remove NO3 from the water column. The 

Nitrogen Cycling and Denitrification 49 

rerouting of Short Creek around the marsh deprives the creek of 

this nutrient removal function. Thus, the bypassing of wetlands 

with drainage ditches in the flood plain may increase NO3 loads 

to the Missouri River. The marsh’s large capacity for denitrification also explains the low concentrations of NO3 in the marsh. 

In other types of wetlands, nitrification has been determined to 

be the rate-limiting step in N removal. However, concentrations 

of dissolved NH4 in Little Bean Marsh also were usually at or 

near the detection limit of 0.005 mg/L (fig. 19), indicating that 

NH4 in the water column is rapidly nitrified or taken up by 

plants. Because about two-thirds of the marsh is covered by 

emergent macrophytes that derive most of their N from the soils 

and shade out algae that derive N from the water column, 

decomposition of organic N may be a more likely rate-limiting 

step than nitrification in the removal of N. Anoxic and acidic 

conditions in the bottom sediments probably limit both nitrification and decomposition rates (Chen and others, 1972). 

The lack of denitrification in December when marsh temperatures averaged 6 °C verifies the findings of Broderick and 

others (1988) that winter denitrification rates approach 0 at 

approximately 5 °C. One might suspect that winter concentrations of NO3 might increase as denitrification ceases. As previously described, ice formation also may serve to increase NO3 

concentrations in the underlying unfrozen water. However, only 

small increases in NO3 occurred during the winters of 1997 and 

1998 (fig. 20) with the exception of two increases in February 

and April 1997 that are associated with the two largest rainstorms. The seasonal pattern of inorganic N concentrations in 

Little Bean Marsh does not closely conform with the pattern of 

low concentrations in the growing season and higher concentrations in the fall and early spring suggested by Kibby (1978). The 

minimal increases in winter NO3 concentrations may be 

explained by the fact that processes that generate NO3, such as 

nitrification, decomposition, and surface-water loadings also 

slow down in winter. Although Reddy and others (1980) determined a laboratory Q10 value (the ratio of two denitrification 

rate constants 10 degrees apart) of 1.5 between 18 and 28 °C, a 

relation between temperature and potential denitrification rates 

in Little Bean Marsh at temperatures between 19 and 27 °C was 

not evident (fig. 27) indicating other factors control the process 

after a threshold temperature is reached. The decrease in N 

removal rates in winter can cause wetland management for 

water quality to be in conflict with management for waterfowl. 

Water levels commonly are raised in the late fall to attract 

migrating waterfowl and allowed to stay up through the spring 

migration, coinciding with the period of minimal N removal. 

Wetlands commonly are drained in the summer to allow vegetation growth for waterfowl, which is the period of maximum N 

removal. Without the overlying water, reducing conditions in 

the sediments are unlikely, and denitrification will cease. However, short periods of drying may increase rates of decomposition, and thus increase the NO3 available for denitrification 

upon reinundation. 

Dissolved-oxygen measurements made at the beginning of 

incubation ranged from 3.0 to 8.5 mg/L, and also had no apparent relation to potential denitrification rates (fig. 28). A dis­
-- -- -- -- -- --

-- -- 

-- -- -- -- -- --

-- -- 

-- -- -- -- -- --

-- -- 

-- -- 

-- -- -- 

-- -- -- -- 

-- --

-- -- -- -- 

-- -- -- -- -- 

Table 10. Amounts of denitrified nitrogen generated in in-situ experiments. [mg/L/d, milligrams per liter per day; --, no data] 

Ambient, Nitrate added, Ambient, Nitrate added, Ambient, 

Time in mg/L/d in mg/L/d Time in mg/L/d in mg/L/d Time in mg/L/d 

Date Site M1 Site M2 Site M3 

06/24/1997 1445 Experiment 06/25/1997 1645 Experiment initiated 1600 Experiment initiated 1100 0.00 06/26/1997 1615 0.00 0.25 1500 0.00 0.00 1030 .00 06/27/1997 1110 .00 1.36 1050 .00 .75 07/28/1997 1145 Experiment 07/29/1997 1400 Experiment initiated 1530 Experiment initiated 07/30/1997 1610 0.00 0.36 1550 0.23 1.19 1410 0.00 07/31/1997 1615 .68 .21 1550 .04 1.48 08/01/1997 1200 .97 1.01 .00 .00 09/08/1997 Dry 1335 Experiment initiated 09/09/1997 Dry 1315 0.00 1.31 09/10/1997 Dry 1400 .00 4.61 12/01/1997 1315 Experiment 12/02/1997 1030 Experiment initiated 1110 Experiment initiated 1315 0.00 12/03/1997 1130 0.00 0.00 1145 0.00 0.00 1030 .00 12/04/1997 1300 .00 .00 1130 .00 .00 1330 .00 04/21/1998 1450 Experiment initiated 04/22/1998 1230 0.00 0.17 1000 Experiment initiated 1200 Experiment 04/23/1998 1445 .00 .38 1400 0.00 0.41 1500 0.00 04/24/1998 1320 .00 .66 1340 .00 -.50 1130 .00 
-- -- -- -- -- --

-- -- -- -- -- -- -- -- 

-- -- -- -- -- -- --

-- -- 

-- -- -- -- 

Table 10. Amounts of denitrified nitrogen generated in in-situ experiments.—Continued 

[mg/L/d, milligrams per liter per day; --, no data] 

Ambient, Nitrate added, Ambient, Nitrate added, Ambient, 

Time in mg/L/d in mg/L/d Time in mg/L/d in mg/L/d Time in mg/L/d 

Date Site M4 Site M5 Site M6 

06/24/1997 1420 Experiment initiated 1500 Experiment initiated 06/25/1997 1045 -0.04 -0.07 1115 0.00 0.00 1615 Experiment 06/26/1997 1210 .59 .85 1315 .00 .80 1515 0.00 06/27/1997 1040 .00 07/28/1997 1345 Experiment initiated 1400 Experiment initiated 1615 Experiment 

07/29/1997 07/30/1997 1430 0.00 0.53 1445 0.00 0.53 1540 0.02 07/31/1997 1215 .00 2.75 1340 .00 2.16 1530 .00 08/01/1997 .00 09/09/1997 1120 Experiment initiated 1210 Experiment initiated 1345 Experiment 

09/09/1997 1100 0.00 1.40 1200 0.00 0.43 1330 0.00 09/10/1997 1250 .00 2.30 1310 .00 5.03 1420 .00 12/01/1997 1235 Experiment initiated 1235 Experiment initiated 12/02/1997 1345 0.00 -0.03 1345 0.00 0.00 1145 Experiment 12/03/1997 1045 .00 -.07 1045 .00 .05 1200 0.00 12/04/1997 1400 .00 .24 1400 .08 .09 1000 .00 04/21/1997 1200 Experiment initiated 1345 Experiment initiated 04/22/1998 1300 0.00 0.99 1330 0.00 0.12 1100 Experiment 04/23/1998 1515 .00 .19 1545 .00 .23 1400 0.00 04/24/1998 1200 .00 1.62 1220 .00 .28 1400 .00 
52 Hydrology and Cycling of Nitrogen and Phosphorus in Little Bean Marsh: A Remnant Missouri-River Wetland 



















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solved-oxygen concentration of 0.3 mg/L is sufficiently oxidizing to limit denitrification in the water column (Broderick and 

others, 1988). Therefore, a relation with potential denitrification 

rates was not expected. 

Although evidence for rapid denitrification in other types 

of wetlands is common (Brinson and others, 1984), it is not universal. Therefore, identification and characterization of denitrification in Little Bean Marsh is important for the management 

of water quality, wetland restoration, and flood plains in the 

Missouri River riparian zone. 

Phosphorus Cycling 

Compartments and processes involving P in wetlands are 

numerous and complex (fig. 29). However, sediment deposition 

in wetlands with mineral soils can be a primary mechanism for 

removing nutrients from surface water. Measurements of nutrient sedimentation fluxes in 11 wetland studies surveyed by 

Johnston (1991) averaged 15 grams N per square meter per year 

and 1.5 grams P per square meter per year. Mitsch (1978) determined that 15.7 grams P per square meter were removed in 1 

year in a cypress swamp in southern Illinois, and that 3.6 grams 

P per square meter were from a single flood event. Less than 10 

percent of the P input from all sources was released to the adjacent Cache River. However, in subsequent years, Dortch (1996) 

reported much lower removal rates for the Cache River wetland. 

Sorption and desorption of P has been reported to be substantial 

in wetlands with mineral soils (Dierberg and Brezonik, 1985). 

Phosphorus adsorption potential is best predicted by the presence of amorphous aluminum and iron minerals in the soil 

(Nichols, 1983; Richardson, 1985). The presence of these compounds results in more PO4 being sorbed where dissolved PO4 

concentrations are high, and more PO4 being solubilized where 

dissolved PO4 concentrations are low under anaerobic conditions (Patrick and Khalid, 1974; Holford and Patrick, 1981). 

The abundant supply of dissolved iron available to Little Bean 

Marsh from ground water (Ziegler and others, 1993) should 

provide a large capacity for sorption and desorption of P. 

Indeed, coatings of iron on the marsh bottom commonly were 

observed on the north side of the marsh where ground-water 

seepage was most prevalent. 

Several mechanisms, both aerobic and anoxic, can resuspend or redissolve P into the water column. Wind is one of these 

mechanisms. However, Carper and Bachmann (1984) determined that in a 104-ha Iowa lake (Little Wall Lake) with an 

average depth of 1.3 m, wind was rarely strong enough to resuspend mineral sediments from the bottom. Winds of 30 km/h 

(kilometers per hour) were required to produce waves that “felt” 

the lake bottom. Although the open-water depths in Little Bean 

Marsh (approximately 0.75 m) are less than Little Wall Lake, 

the fetch, and thus the wave heights, also were much less. The 

lack of fetch primarily is caused by macrophytes that break up 

open water and protect large areas from wind. Therefore, a wetland would trap sediment more efficiently when open-water 

Phosphorus Cycling 53 

areas are small, and macrophytes dominate the area near the 

surface outflow point of the wetland. The outflow area in Little 

Bean Marsh was an area of open water. 

Macrophyte leaching is another source of aerobic P 

release. Klopatek (1978) reported that 58 percent of P taken up 

by Scirpus fluviatilis was leached back to the water. 

Correlations 

The P-trapping efficiency of a wetland usually is determined by P release from the sediments to the water column 

where P can then escape the wetland in surface outflow. As previously described, several aerobic and anaerobic processes can 

release P. Therefore, concentrations of P were correlated with 

wind speed, dissolved oxygen, pH, and turbidity to identify the 

primary mechanisms for P release from sediments and subsequent escape from the marsh. 

Whereas total P and PO4 concentrations were not significantly correlated with wind speed or pH at any site, P concentrations were significantly correlated with mean daily dissolved 

oxygen at open-water sites M3 and M5 (p<0.05) (table 11). 

Total P at site M2 also was significantly correlated with mean 

daily dissolved oxygen at site M3 (p<0.05). The significance of 

these correlations was somewhat unexpected because mean 

daily dissolved oxygen was measured only at site M3. Apparently, low dissolved oxygen values at site M3 are indicative of 

widespread anoxia or near anoxia over large parts of the marsh. 

These correlations indicate that anoxic conditions have a larger 

effect on P concentrations than other variables such as wind 

speed and pH. However, anoxic conditions usually were limited 

to low-water periods in 1997 (figs. 8 and 12) that prevented 

large quantities of P from escaping the marsh in surface outflows. Annual climatic patterns in the Midwest likely make this 

a common scenario in Missouri wetlands. Although surface outflows during periods of P release could cause eutrophication 

problems downstream, high P concentrations in wetlands are 

most likely during low-water periods, when wetlands are not 

discharging surface water and, therefore, P can be retained by 

the wetlands. 

Total P was significantly correlated (p<0.1) with chlorophyll concentrations at partially open-water site M2 and openwater site M5 only (table 12), indicating that in open-water 

areas, most P is tied up in algae similar to most reservoirs in the 

region (Jones and Knowlton, 1993). Conversely, most P is not 

tied up in algae where emergent plants shade the water. The 

highest concentrations of total P occurred during a September 

algal bloom at open-water sites (table 3). This bloom occurred 

during macrophyte senescence and soon after an extended 

period of low dissolved oxygen and low water levels in the 

marsh (figs. 8 and 12). High summer temperatures likely accelerated decomposition that consumed oxygen and caused large 

areas of anoxic conditions in the marsh, which released P into 

the water that contributed to the algal bloom. The bloom and the 

period of low dissolved oxygen ended in early October with a 

sudden decrease in temperature (figs. 14 and 15) and more than 
RUNOFF $ECOMPOSITION $ISSOLVEDORTHOPHOSPHORUS

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Phosphorus Cycling 55 

Table 11. Spearman’s rho values and levels of significance (p) for correlation of phosphorus with mean daily dissolved oxygen at five 

sites in Little Bean Marsh. Mean-daily dissolved oxygen was measured at site M3. 

Site 

Total phosphorus Soluble reactive phosphorus (PO4) 

(fig. 1) rho p rho p 

M1 -0.38 0.15 +0.19 0.68 

M4 -.06 .42 -.38 .11 

M2 -.72 .0005 -.14 .26 

M3 -.54 .0027 -.35 .034 

M5 -.68 .0002 -.44 .012 

Table 12. Spearman’s rho values and levels of significance (p) for correlation of phosphorus with chlorophyll concentrations at five 

sites in Little Bean Marsh. 

Site 

Total phosphorus Soluble reactive phosphorus (PO4) 

(fig. 1) rho p rho p 

M4 -0.39 0.84 -0.25 0.75 

M6 .71 .063 -.03 .55 

M2 .71 .018 -.40 .89 

M3 .44 .095 .35 .15 

M5 .76 .002 .13 .35 

7.5 cm of rain. Soluble reactive P had no significant correlations 

(p>0.1) with chlorophyll at any site (table 12). 

Sinks and Sources of Phosphorus 

Approximately 53 kg of total P left Little Bean Marsh in 

surface outflows in 1997 (table 8). Dividing this value by the 

952 ha drainage area gives an areal yield of 0.056 kg/ha, which 

is less than one-half of the 0.12 kg/ha areal yield of the Missouri 

River at St. Joseph, Missouri, in 1997 (Hauck and others, 1998) 

and less than one-third of the 1980 to 1996 average P yield of 

0.19 for the entire Missouri River Watershed (Goolsby and others, 1999). Most P in Little Bean Marsh was in the particulate 

phase (table 3); the majority of the dissolved P usually was PO4, 

although nonreactive (organic) P was a substantial part. The 

largest P loads leaving the marsh occurred in response to the 

two largest storms in February and April 1997 (table 8). Even 

though Little Bean Marsh removed a large percentage of the P 

it received, concentrations of available soluble reactive P in the 

outflows were large enough to stimulate phytoplankton growth 

in a hypothetical P-limited water body downstream (Horne and 

Goldman, 1994). 

Concentrations of P in rainfall generally are less than those 

present in the Missouri River (Hauck and others, 1999) or in 

runoff from its tributaries in Missouri and Kansas. However, 

because precipitation constituted 41 percent of the inflow to Little Bean Marsh and other sources of P were relatively low, precipitation was an important source of P to Little Bean Marsh. P 

data are not collected by the National Atmospheric Data Program and, consequently, wet deposition data for P are difficult 

to obtain. At 28 nonurban locations in the contiguous 48 states, 

atmospheric deposition of P ranged from 0.05 to 1.1 kg/ha/yr 

(kilograms per hectare per year), and averaged 0.4 kg/ha/yr 

(Johnston, 1991). Orthophosphorus (PO4) concentrations in 

precipitation measured at Centralia, Missouri, by the Agricultural Research Service (Alberts and others, written commun., 

1993) averaged 0.007 mg/L, or about 0.07 kg/ha, in 1997. 

Because particulate and organic P concentrations are expected 

to be small in precipitation, PO4 concentrations are likely good 

estimates of total P in precipitation. A rate 0.07 kg/ha from 

direct precipitation into Little Bean Marsh would result in about 

3 kg of P input into Little Bean Marsh in 1997, which is less 

than 6 percent of the P leaving the marsh in surface outflow. 

The amount of dry deposition is not known, but when added to 

wet deposition, would be expected to substantially increase the 

amount of total atmospheric deposition. 

Storm loads were measured from four storms at three sites 

during the spring of 1997. Storm loads were measured from two 

storms at site I1 (fig. 1). These storms contributed 64,000 m3

 of 

surface runoff to Little Bean Marsh (table 2), which is approximately 12 percent of the total surface runoff into the marsh during 1997. These four storm events also contributed approximately 136 kg of P to Little Bean Marsh. Eighty-four percent of 
56 Hydrology and Cycling of Nitrogen and Phosphorus in Little Bean Marsh: A Remnant Missouri-River Wetland 

the 136 kg of P was attached to suspended sediment. If one prorates the loading of total P from these four storms to the total 

amount of surface runoff (516,000 m3) during 1997, it is apparent that surface runoff contributes more than one order of magnitude more P than was released in surface outflow from the 

marsh. Little Bean Marsh evidently traps P-laden sediments 

very efficiently as the long hydraulic residence time would indicate (Moustafa, 1997). 

Ground-water seepage usually is a small source of nutrients to freshwater marshes (Mitsch and Gosselink, 1993). The 

median concentration of total dissolved P in 21 ground-water 

samples collected from five monitoring wells near Little Bean 

Marsh (fig. 1) was less than 0.02 mg/L (table 1). Nearly all of 

the P in ground water was in PO4 form and is, therefore, readily 

available to phytoplankton. Samples from well W4 periodically 

had abnormally high concentrations of dissolved P. Using a net 

ground-water seepage of 157,000 m3

 in 1997 (fig. 7) and 

assuming total P concentrations of 0.02 mg/L, ground-water 

seepage would have contributed less than 8 percent of the total 

P leaving Little Bean Marsh in surface outflow and less than 1 

percent of the amount of P received in storm flows. Therefore, 

ground water probably is a minor source of P to wetlands along 

the Missouri River that receive at least minimal amounts of 

storm runoff. 

The mean concentration of iron in samples from the Missouri River alluvial aquifer in northwest Missouri was 8 mg/L 

(Ziegler and others, 1993). Iron in ground-water samples collected from nine wells at Little Bean Marsh at approximately 

monthly intervals between September 27, 1996, and April 13, 

1997, ranged from 1.0 to 18 mg/L (Walter, 1998). At 8 mg/L, 

the net 157,000 m3 of ground water received by Little Bean 

Marsh in 1997 provided approximately 1,200 kg of iron that 

usually would be oxidized to form a large and relatively constant source of sorption sites for P, metals, and other hydrophobic compounds. The P immobilization capacity of Little Bean 

Marsh would be expected to be substantial as long as the marsh 

was oxic. Therefore, the source of wetland water can have a 

substantial effect on P retention. However, the frequent lack of 

detectable NO3 in Little Bean Marsh means that Fe+3 is the next 

likely electron acceptor during anoxic conditions. Decomposition and other oxygen-consuming processes can reduce ferric to 

ferrous iron and allow release of bound P, regardless of the 

amount of iron available. 

Soils contain, by far, the largest stocks of P of any wetland 

storage component (Johnston, 1991). The mean concentration 

of total P in the top 15 cm of bottom sediments collected from 

four locations (M1, M2, M3, and M5) in Little Bean Marsh is 

590 parts per million (table 9), which is near the low end of the 

range of total P concentrations in the mineral soils of six wetlands in Wisconsin, North Dakota, and North Carolina (ranging 

from 220 to 1,500 mg/kg, Johnston, 1991). However, assuming 

a bulk bottom-sediment density of 1.5 g/cm3 (grams per cubic 

centimeter) and a mean concentration of total N in bottom sediments of 590 parts per million (table 9), the top 15 cm of Little 

Bean Marsh contained 1,500 kg of P. This amount of P is more 

than one order of magnitude greater than either the amount of P 

entering the marsh in surface runoff or the amount of P in surface outflow during 1997. 

Annual leaching fluxes of P from Scirpus fluviatilis in a 

Wisconsin wetland were 0.37 g/m2

/yr and litterfall rates from 

Typha latifolia were 2.5 g/m2

/yr (Klopatek, 1978). Annual P litterfall fluxes in an Iowa pothole wetland were 0.63 g/m2/yr 

(Davis and Van der Valk, 1978). Whereas these leaching rates 

would provide less than 2 percent of the total amount of P lost 

from the marsh in surface outflow, the highest concentrations of 

P in the marsh occurred during the September/early-October 

period of senescence (fig. 21). Scirpus fluviatilis is slow to 

decay (Whitley and others, 1990) and in the Iowa pothole, Scirpus fluviatilis retained more than 50 percent of its dry weight 

after 525 days of decomposition. Klopatek (1978) determined 

that 38 percent of the P uptake by Scirpus fluviatilis was 

retained over the winter months, with the rest lost to leaching 

and litterfall. Prentki and others (1978) reported 30 percent 

overwinter retention of P in Typha latifolia. Consequently, 

much of the P tied up in Scirpus fluviatilis is prone to burial. 

Even though most (95 to 99 percent) of the nutrients used by 

emergent macrophytes are taken in from the soil rather than the 

overlying water (Wetzel, 1983), the amount of P returned into 

the water column by emergent vegetation would likely be a tiny 

fraction of the amount in the soil, and, therefore, burial of P 

probably is a substantial sink of P in Little Bean Marsh. 

Comparisons with the Missouri River, 

nearby Wetlands, and Missouri Reservoirs 

Whereas few studies of wetland water quality have been 

undertaken in Missouri or along the Missouri River, reports by 

Knowlton and Jones (1997) and Richards (1995) provide some 

data for comparison. Average total N concentrations (ranging 

from 0.6 mg/L at site M1 to 1.7 mg/L at site M6) (fig. 11) in Little Bean Marsh, are substantially less than four Missouri River 

oxbows sampled by Knowlton and Jones (1997) (table 7), 

roughly one-half of mean concentrations of total organic, NH3, 

and NO3-N in the Missouri River reported by Hauck and others 

(1999) (table 4) and Ford (1982), but somewhat greater than the 

average total N concentrations reported for reservoirs of the glaciated plains of Missouri (0.65 mg/L) by Jones and Knowlton 

(1993). Low N concentrations in Little Bean Marsh primarily 

result from the near absence of inorganic N (figs. 19 and 20). 

Others also have reported low concentrations of inorganic N in 

riparian wetlands (Kibby, 1978; Mitsch, 1978). Apparently, 

denitrification and biotic uptake keep dissolved inorganic N 

concentrations low. This finding is similar to N speciation 

found in Missouri River oxbows by Knowlton and Jones 

(1997), who reported that dissolved organic N composed more 

than 95 percent of the total dissolved N, and more than 50 percent of the total N. Most N lost from Little Bean Marsh in surface outflow was organic N, and most of that was dissolved. 

This N speciation contrasts with the receiving Missouri River, 
Comparisons with the Missouri River, nearby Wetlands, and Missouri Reservoirs 57 

where the dominant N species is NO3 (Knowlton and Jones, 

1997) (table 4). 

Richards (1995, 1999) collected samples approximately 

every 3 months on an unnamed flood-plain water body near the 

Eagle Bluffs of the Missouri River in central Missouri beginning in 1992. This water body was created by flood scour, and 

is several feet deeper than Little Bean Marsh. A similar setting 

in the Missouri River flood plain makes comparisons with Little 

Bean Marsh useful. Specific conductance in the scour was similar to that in alluvial ground water. Although the scour had 

somewhat greater concentrations of all N and P species except 

organic nitrogen than Little Bean Marsh, both water bodies had 

low concentrations of NO3, indicating N limitation caused by 

denitrification in the scour. The greater percentage of organic 

nitrogen in Little Bean Marsh may have been related to the large 

quantity of emergent vegetation compared to the deeper, openwater scour. The greater concentrations of P and inorganic N in 

the scour may have been because of lesser rates of vegetative 

uptake in the deeper scour, because much of the water in the 

scour was likely below the photic zone. The pH values in the 

scour were well within the range measured at Little Bean 

Marsh. 

An artificially-created and intensively-managed wetland 

near the Eagle Bluffs scour provides another useful comparison. 

This water source for this artificial wetland primarily is outflow 

from municipal sewage-treatment wetlands, supplemented with 

Missouri River water. Consequently, the Eagle Bluffs wetland 

receives a much greater loading of N and P than Little Bean 

Marsh. The Eagle Bluffs wetland was filled in 1996 and was 

sampled at the outflow from 1996 to 1999 (Richards, 1999). As 

might be expected, N and P concentrations were much larger 

that those present in Little Bean Marsh, indicating that the created wetland may not have the nutrient removal capacity of Little Bean Marsh, a remnant native wetland. However, the cause 

of this difference in removal capacity is not apparent, because 

in addition to greater loading rates, the Eagle Bluff wetland may 

not yet have a thick layer of organic matter on the marsh bottom, and may have more algal growth in the abundant open 

water. All N and P species concentrations also were greater in 

the Eagle Bluffs Wetland than in the Missouri River. Therefore, 

outflow from the Eagle Bluffs wetland actually increases concentrations of nutrients in the Missouri River compared with 

Little Bean Marsh outflow that dilutes nutrient concentrations 

in the river. 

Comparison of Little Bean Marsh water quality with Missouri River water quality is useful as, historically, much local 

runoff from small streams passed through riparian wetlands, 

such as Little Bean Marsh, before entering the river. Because 

there are very few remnant riparian wetlands left in the Missouri 

River flood plain, a comparison of the water quality in Little 

Bean Marsh and the Missouri River can help determine the 

effects that restored wetlands might have on water quality in the 

Missouri River, and document the effect that riparian wetlands 

may have had on Missouri River water quality before floodplain development. 

Annual means and ranges of selected constituents and 

properties of sites where at least 12 samples were collected over 

a 12-month period are listed in table 4. A Kruskal-Wallis ranksum test was done on data collected monthly from the Missouri 

River at St. Joseph in 1997 and data collected biweekly at site 

M5 near the outflow of Little Bean Marsh during the same year 

to determine the statistical significance of the differences 

between the means of several constituents and properties. Test 

results and observation of data in table 4 indicate that concentrations of NO3 were significantly less at Little Bean Marsh 

than concentrations in the Missouri River (p<0.0001). Nitrate 

concentrations in Little Bean Marsh usually were less than 

0.005 mg/L, whereas concentrations in the Missouri River consistently were between 0.3 and 1.7, and averaged 1.1 mg/L. This 

difference in NO3 concentration was likely the result of the 

rapid removal of NO3 in the marsh by denitrification, whereas 

most water in the river is derived from ditches that convey water 

across the flood plain with little opportunity for denitrification. 

Large rivers, such as the Missouri, typically have low amounts 

of primary production because of high turbidity, large depths, 

and constant mixing. Consequently, inorganic nitrogen is not 

rapidly removed, can build up to substantial concentrations in 

large rivers such as the Missouri and Mississippi, and cause 

large zones of hypoxia in the Gulf of Mexico, which usually is 

N-limited (Goolsby and others, 1999). However, there was no 

significant difference between Little Bean Marsh and the Missouri River in the annual means of total NH4 and NH4 plus 

organic N (p = 0.65 and p = 0.99). Consequently, wetlands such 

as Little Bean Marsh may not reduce the amount of organic 

nitrogen in receiving rivers. However, suspended organic material in rivers often is refractory and not readily available to 

biota. Nevertheless, these data indicate that the vast extent of 

riparian marshes along the Missouri and Mississippi River may 

have had a substantial role in limiting NO3 loads to the Gulf of 

Mexico before agricultural development of flood plains, and 

that drainage and removal of riparian marshes may be a major 

cause of the increased NO3 loads to the Gulf of Mexico. 

Annual mean concentrations of total and soluble reactive P 

were less in Little Bean Marsh than the Missouri River. However, whereas the differences between the means for soluble 

reactive P were statistically significant (p<0.0001), the difference between the means for total P were not significant at the 95 

percent confidence level (p = 0.063). Nevertheless, it is likely 

that the water contributed from wetlands like Little Bean Marsh 

would decrease the concentration of P in the Missouri River, 

despite the potential for occasional P release during periods of 

wetland anoxia. 

The annual mean of specific conductance at site M5 in Little Bean Marsh (500 μS/cm) was significantly less (p = 0.0014) 

than the annual mean conductance of the Missouri River (718 

μS/cm). This probably reflects the large contribution of direct 

precipitation relative to runoff in the marsh compared to the 

river. 
58 Hydrology and Cycling of Nitrogen and Phosphorus in Little Bean Marsh: A Remnant Missouri-River Wetland 

Summary and Conclusions 

Water-quality benefits have been assumed to accrue from 

wetlands in Missouri and other parts of the Midwest despite the 

lack of concurrent water-quality and hydrologic data. Consequently, this study was undertaken to determine the magnitude 

and character of selected water-quality benefits that can occur 

in a remnant riparian wetland along the Missouri River, and to 

identify critical processes that can be managed in remnant or 

restored riparian wetlands for water-quality improvement. Little Bean Marsh was selected for this study because it is one of 

only a few remaining riparian wetlands along the Missouri 

River where nutrient cycling processes were expected to be relatively undisturbed. Nutrient cycling processes discovered in 

Little Bean Marsh may predict the effects that large-scale wetland restorations along the Missouri River will have on the 

water quality of the Missouri River. 

In 1997, 41 percent of the water received by Little Bean 

Marsh was from direct precipitation, 14 percent from groundwater seepage, 30 percent from watershed runoff, and 15 percent from backflows from Bean Lake. Like many flood-plain 

water bodies, more than 90 percent of the marsh’s drainage area 

is flood plain. The nearly flat slopes, sandy soils, and large shallow depressions keep runoff coefficients in the flood plain low. 

No runoff was observed from several storms that had more than 

4 cm (centimeters) of rain; consequently, surface-water contributions from the watershed were small. River levees and floodplain drainage have converted Little Bean Marsh and many 

other Missouri River wetlands from open to closed systems. 

Open wetlands typically are large sinks for alloctanous sediment and nutrients, whereas closed wetlands rely primarily on 

precipitation for nutrients. Consequently, the hydrology, water 

quality, and ecological functions of remnant wetlands located 

along today’s Missouri River usually are substantially different 

from the predevelopment condition. 

Ground-water levels near Little Bean Marsh did not 

closely track Missouri River stages primarily because of the 1.6 

km (kilometers) separating them. However, ground-water levels closely tracked marsh water levels indicating a substantial 

hydraulic connection between ground water and the marsh. 

Although Little Bean Marsh was both a recharge and discharge 

area, ground water entering the marsh was three times the 

recharge to ground water. Ground water typically recharged the 

marsh on the north side, and marsh water seeped down through 

the marsh bottom on the south side. The close interaction with 

shallow ground water indicates lowering of the water table by 

as little as 1 m (meter) by pumping would likely make Little 

Bean Marsh much drier, and could substantially affect the 

marsh ecosystem. Reduced surface runoff and ground-water 

availability is a stabilizing influence on marsh hydrology, and 

probably leads to the persistence of emergent vegetation. The 

ground-water connection of Little Bean Marsh indicates that the 

hydrologic regime of most wetlands along the lower Missouri 

River is largely a function of marsh depth relative to the water 

table. 

More water was lost from the marsh through evapotranspiration (59 percent) than all other pathways combined. This characteristic partially is because of abundant macrophytes that can 

increase evapotranspiration above that lost from open-water 

surfaces. Surface outflow accounted for 36 percent and groundwater seepage accounted for only 5 percent of the losses. Long 

residence times and large evaporative losses allow Little Bean 

Marsh to perform the hydrologic functions of peak-flow attenuation and surface-runoff reduction. Routing of streams around 

marshes obviously precludes these benefits. Large residence 

times also allows the marsh to greatly affect water quality 

before water escapes as ground-water recharge or surface outflow. 

The greater percentage of precipitation as an inflow source 

has diluting effects on all water-quality constituents, and the 

substantial difference between specific conductance of ground 

water and precipitation/surface runoff make specific conductance a useful indicator of the primary source of water in the 

marsh at a given time. However, because of the shallowness of 

the marsh, ion exclusion during ice formation in the winter 

caused the greatest specific conductances during the entire 

monitoring period. This substantial increase in the concentration of all dissolved solutes could make wetlands more vulnerable to toxic contaminants than other surface-water bodies. 

Dissolved-oxygen concentrations were less than 5 mg/L 

(milligrams per liter) for 3 to 4 months in 1996 and 1997, and 

were near zero for about 1 month during the summers of both 

years. These periods of anoxia produce widespread reducing 

conditions that inhibit nitrification, enhance denitrification, and 

cause iron reduction and phosphorus (P) release. The largest 

diel changes were about 5 mg/L that often occurred during periods of low dissolved-oxygen concentrations. Dissolved-oxygen 

concentrations were significantly greater at open-water sites 

than at sites with emergent vegetation because of shading and 

decomposition of a large amount of detritus. These extended 

periods of anoxia during the summer undoubtedly limit the habitat for certain fish and other fauna that have not adapted to low 

oxygen conditions and have no refuge in a shallow marsh. 

Water temperatures peaked between 25 and 30 °C (degrees 

Celsius) for almost 2 months each year. Despite depths of less 

than 1 m, vertical temperature differences of 3 to 9 °C persisted 

for more than 3 months during the summers of 1996 and 1997. 

The period of least dissolved-oxygen concentration corresponded with the period of maximum stratification. The ability 

of open-water marshes to stratify increases the area and periods 

of time with reducing conditions that inhibit nitrification and 

potentially cause denitrification and P release from iron reduction. Chlorophyll data indicate that much of Little Bean Marsh 

underwent a major algal bloom in September 1997, shortly after 

summer stratification abruptly ended. The bloom may have 

been triggered by mixing of the water column that made ammonia nitrogen (N) from the hypolimnion available to a N-limited 

population of algae. Shallow depths and extended periods of 

anoxia in the marsh would limit the ability of some organisms 

to escape high-temperature stress. 
Most of the N in Little Bean Marsh was in organic forms. 

Denitrification and biotic uptake kept concentrations of inorganic N extremely small. This situation contrasts with N in the 

Missouri River and many of its tributaries, where nitrate (NO3) 

is the dominant species of N. Consequently, artificial floodplain drainage that bypasses wetlands can deliver substantially 

more biotically available inorganic N to receiving streams than 

surface water that has been routed through a remnant wetland. 

Water clarity in Little Bean Marsh usually was high for 

several reasons. Sediment loadings from the largely flood-plain 

drainage were low, emergent vegetation shade out algae and 

shield the water from wind, and high concentrations of bivalent 

cations increase flocculation rates of inorganic suspended material. The high divalent cation concentration in the marsh probably was a result of ground water received by the marsh. Openwater sites had significantly greater turbidities and suspended 

sediment concentrations than sites covered with emergent vegetation. This difference was likely caused by shading that limited algal growth in vegetated sites. Consequently, marshes 

along the lower Missouri River are likely to have better water 

clarity if they are incised deep enough to receive substantial 

amounts of ground water and grow abundant emergent macrophytes. 

Little Bean Marsh usually was either eutrophic or hypereutrophic. Vegetated sites had less persistent periods of eutrophy. The most intense period of hypereutrophy occurred in September, and coincided with both an algal bloom and senescence 

of Scirpus fluviatilis and Typha latifolia. The rapid leaching of 

N that occurs soon after macrophyte senescence, combined with 

a recent destratification of the marsh, probably provided N to 

the N-limited open-water areas and triggered a phytoplankton 

bloom. This type of bloom could be a common phenomenon in 

Missouri River wetlands with thick stands of macrophytes. N 

was better correlated with chlorophyll than was P at open-water 

sites, indicating N is more limiting than P on phytoplankton 

growth. The lack of any correlation of chlorophyll with N or P 

at some vegetated sites is consistent with light limitation. 

Despite the rarity of runoff events, surface runoff, combined with atmospheric deposition, contributed more than 

seven times the 530 kg (kilograms) of N that escaped Little 

Bean Marsh in surface outflow in 1997. Atmospheric deposition alone was more than the 530 kg. Ground water contributed 

less than 1.5 percent of the N leaving the marsh in surface outflow. Undoubtedly, much more N is internally recycled through 

plants in Little Bean Marsh than is externally gained or lost 

annually. The slow decay rate of Scirpus fluviatilis and reducing 

conditions in bottom sediments potentially make burial of 

organic N a substantial sink of N. Annual peak concentrations 

of nearly all N and P species occurred during the fall period of 

senescence. 

Results of denitrification experiments indicate that denitrification rates were limited by NO3 in the water column. Consequently, decomposition and nitrification of ammonium (NH4) 

and organic N are the rate-limiting steps of N removal in Little 

Bean Marsh. Conditions for high rates of denitrification were 

almost ideal in the bottom sediments of the marsh. The domiSummary and Conclusions 59 

nant macrophytes in the marsh were Scirpus fluviatilis and 

Typha latifolia; both have oxidizing rhizospheres around their 

roots. These rhizospheres greatly increase rates of the rate-limiting decomposition and nitrification processes over rates 

occurring in open-water areas where all NO3 must diffuse from 

the water column through the sediment/water interface, as 

opposed to being created in the sediments. The NO3 limitation 

on rates of denitrification also indicate that Little Bean Marsh 

has a large unused capacity for N removal. 

Only the largest storm had a measurable effect on P concentrations in Little Bean Marsh. Most dissolved P usually was 

in the orthophosphate (PO4) form, but most total P was on particulates. Correlations of wind speed, pH, turbidity, and dissolved oxygen with concentrations of total P indicate that periods of anoxia have much larger effects on P release than the 

other variables. Periods of most extreme anoxia occurred in late 

summer and corresponded with high rates of decomposition and 

senescing macrophytes to produce the highest concentrations of 

both N and P and a large algal bloom in Little Bean Marsh. Low 

water levels prevented the escape of P in surface outflow during 

the periods of highest P concentrations. Annual climatic patterns in the Midwestern United States probably make the correspondence of these conditions common in marshes along the 

Missouri River. 

Little Bean Marsh removed the majority of the P it 

received. The amount of P in surface inflows to the marsh were 

more than one order of magnitude greater than that escaping in 

surface outflows. P contributions from atmospheric deposition 

to the marsh are small compared to the 53 kg of P in surface outflows in 1997. Low concentrations of P in ground water caused 

ground water to be a small source of P to Little Bean Marsh as 

well. The long hydraulic residence time of the marsh and large 

contributions of iron from ground water (that provide many 

sorption sites for P) make the marsh an effective sediment and 

P trap. These characteristics, plus the large long-term retention 

of P by the macrophytes after senescence, make burial of P a 

substantial sink of P in Little Bean Marsh. 

N and P concentrations were much larger in an artificiallycreated sewage-treatment wetland along the Missouri River in 

central Missouri than those present in Little Bean Marsh. Much 

greater loadings of N and P, a lack of organic material in wetland bottom sediments, and more abundant phytoplankton in 

the sewage-treatment wetland may have all contributed to this 

phenomenon. 

Concentrations of total P in Little Bean Marsh were less 

than those in other Missouri River wetlands. This difference is 

likely related to the amount of open water that is more conducive to algal growth. Average total N concentrations in Little 

Bean Marsh were substantially less than other Missouri River 

wetlands; roughly one-half of mean concentrations in the Missouri River, but roughly twice the mean N concentrations found 

in reservoirs of the glaciated plains of Missouri. Low concentrations of N primarily result from the near absence of inorganic N 

in Little Bean Marsh. Apparently, denitrification and biotic 

uptake keep dissolved inorganic N concentrations low. Dissolved organic N was the dominant N species in Little Bean 
60 Hydrology and Cycling of Nitrogen and Phosphorus in Little Bean Marsh: A Remnant Missouri-River Wetland 

Marsh and other Missouri River wetlands. This situation contrasts with the Missouri River, in which NO3 dominates. 

Concentrations of NO3 in Little Bean Marsh usually were 

less than 0.005 mg/L and were significantly less than concentrations in the Missouri River (p<0.0001), which averaged 1.1 mg/ 

L. Large rivers, such as the Missouri, typically have low 

amounts of primary production because of turbidity, large 

depths, and constant mixing. Consequently, inorganic nitrogen 

is not rapidly removed and can build up to substantial concentrations in large rivers, such as the Missouri and the Mississippi, 

and can cause large zones of oceanic hypoxia, such as in the 

Gulf of Mexico, which typically are N limited. Concentrations 

of organic N in Little Bean Marsh and the Missouri River, however, were not significantly different. Consequently, wetlands 

such as Little Bean may not reduce the amount of organic nitrogen in receiving rivers. Suspended organic material in rivers 

often is refractory, however, and not readily available to biota. 

Nevertheless, these data indicate that the vast historical extent 

of riparian marshes along the Missouri and Mississippi Rivers 

may have had a substantial role in limiting NO3 loads to the 

Gulf of Mexico before agricultural development of flood plains, 

and that drainage and removal of riparian marshes may be a primary cause of the increased NO3 loads to the Gulf of Mexico 

today. Annual mean concentrations of total and soluble reactive 

P were both less in Little Bean Marsh than the Missouri River; 

therefore, the water contributed from Little Bean Marsh and 

marshes like it would likely decrease the concentration of P in 

the Missouri River, despite the potential for occasional P release 

during periods of wetland anoxia. 

Whereas the data indicate that Little Bean Marsh has the 

capacity to remove much more N and P than is presently 

removed, large quantities of additional nutrients could alter species composition of wetland vegetation and cause other hydrochemical changes. For example, rerouting of Short Creek 

through Little Bean Marsh might allow the removal of additional N from runoff through denitrification, but also might 

cause the marsh to fill with sediment. Because remnant riparian 

wetlands along the lower Missouri River are so rare, preservation of these wetlands may be more important than using them 

to improve the quality of downstream and downgradient waters 

such as the Missouri River. 

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alluvium, northwestern Missouri--July 1988 and June-July 

1989: U.S. Geological Survey Open-File Report 93–101, 

30 p. 
TABLE


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-- -- -- -- ----

-- --

--

--

-- -- -- -- -- -- -- -- 

-- -- -- --

-- -- -- --

--

--

-- --

-- --

--

-- -- --

-- -- -- -- -- -- -- -- 

-- --

-- -- --

-- --

-- --

-- --

-- -- -- -- -- -- -- -- 

--

--

-- --

--

--

--

--

--

-- -- 

--

--

--

--

Table 3. Analyses of surface-water samples collected from Little Bean Marsh. [ft, feet; mg/L, milligrams per liter; °C, degrees Celsius; μS/cm, microsiemens per centimeter at 25 degrees Celsius; FTU, formazine turbidity units; Eh, oxidation-reduction pot

--, no data; <, less than] 

Site Gage Water 

number height, depth, 

(fig. 1) Date Time in ft in ft 

Dissolved Dissolved Water Water 

oxygen oxygen temperature temperature pH, 

(top), (bottom), (top), (bottom), in standard co

in mg/L in mg/L in °C in °C units 

M1 M1 M1M1 M1 M1 M1 M1 M1 M1M1M1 

M2 M2 M2 M2 M2M2 M2M2M2M2M2M2M2 M2M2M2 M2M2

03/17/1997 03/31/1997 04/17/1997 05/05/1997 

05/21/1997 

06/10/1997 

06/25/1997 

07/08/1997 

07/29/1997 

08/12/1997

12/02/1997 

12/19/1997 

11/15/1996 

12/02/1996 12/19/1996 

01/03/1997 

01/20/1997 02/05/1997 

02/17/1997 

02/24/1997 03/17/1997 03/31/1997 

04/17/199705/05/1997

05/21/1997

06/10/1997

06/25/1997 

07/08/1997 

07/29/1997 08/12/1997

1550 1200 1430 

1330 

1500 

1345 

1420 1200 

1600 

13451230 

1500 

1545 1430 

1500 

1450 

1230 1530 

1400 

1100 

1430 

1425 

28.25 

28.05 29.27 

29.59 

29.56 

28.98 

28.65 

28.37 

29.16 29.12 

28.27 28.27 

28.25 

28.05 

29.27 

28.98 

28.65 

28.37 

1 

.8 

.6 

.5 

1.4

2 

2

1.5 

.04 

1 

.4 

2 

1.2 

1 

.7 

8.3 

11.6 

6.4 

4.2 

2.8 

1.6 

3 

1.6 

7.1 

12.9 

13.6

13.7

3.7 

15.6 

4.6 

10.3 

13 

7.8 

6.6 

9.3 

1 

3.4 

5 

1.4 

6.4 

4.3 

2.2 

1.2 1.4 

12.2 

3.0 

7.2 

6.5 

3.3 

.4 

2.4 

1 

.7 

5.3 

13.3 

23.1 

23.3

20.8 

24.9 

25.2 

22.8 

5.8 

.5 

6.6 

4.1 

2.3 

3 

.2 

1 

1.5 

8.3 

13.5 

13.3 

21.5 

25.5 

21 

27.1 

31 

22.3 

22.3 

20.6 

24.6 

24.9 

22.8 

6.6 

11.4 

20 

18.9 

19.8 

24.9 

27 

22.3 

7.5 

7.86

7.92 

7.66 

7.77 

7.46 

7.28 

6.74 

7.06 

7.15 

7.3 

7.34 

7.69 

7.66 

6.67 

7.13 

7.02 

7.09 

7.64 

7.31 

7.55 

8.1 

7.6 

7.4 

7.21 

6.76 

7.06 
--

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--

--

-- -- -- -- -- -- -- -- 

--

-- -- 

-- -- --

-- 

-- -- --

--

-- -- -- -- -- -- -- -- 

--

-- --

--

-- 

-- ----

--

-- --

-- --

--

--

--

--

--

--

---- --

--

Table 3. Analyses of surface-water samples collected from Little Bean Marsh.—Continued [ft, feet; mg/L, milligrams per liter; °C, degrees Celsius; μS/cm, microsiemens per centimeter at 25 degrees Celsius; FTU, formazine turbidity units; Eh, oxidation-reduction pot

--, no data; <, less than] 

Dissolved Dissolved Water Water 

Site Gage Water oxygen oxygen temperature temperature pH, 

number height, depth, (top), (bottom), (top), (bottom), in standard co

(fig. 1) Date Time in ft in ft in mg/L in mg/L in °C in °C units 

M2 08/27/1997 1330 28 0.3 3 27.7 7.47 M2 09/10/1997 1315 27.54 8.17 M2 09/24/1997 1300 27.49 .4 2.1 16.8 7.46 M2 10/14/1997 1330 27.68 .5 10.8 8.5 15.9 14.9 7.78 M2 10/28/1997 1145 27.94 .5 9.4 8.5 7.42 M2 11/19/1997 1430 27.86 .6 9.6 9.2 7.4 M2 11/19/1997 M2 12/02/1997 28.3 4.8 6.1 7.29 M2 12/19/1997 1230 1.5 5.2 7.23 M3 11/15/1996 1415 11 7.1 7.13 M3 11/25/1996 1400 29.09 2.3 12.0 1.6 8.36 M3 12/18/1996 1145 13.3 3.5 8.25 M3 01/03/1997 1117 28.85 2.1 16.1 5.1 7.32 M3 01/20/1997 1145 28.27 1.5 22.9 24.1 2.4 2.7 7.32 M3 02/05/1997 1100 28.27 1.5 12.8 1.3 2.8 6.8 M3 02/17/1997 1030 27.94 1.1 13 11.1 1.3 3.8 M3 02/24/1997 M3 03/17/1997 0945 28.25 1.5 11.7 7.1 8.19 M3 03/31/1997 1130 28.05 1.3 11.8 14.3 8.19 M3 04/17/1997 1310 29.27 2.5 11.1 11.1 15.7 15.7 8.28 M3 05/05/1997 1110 8.2 8.2 18.3 18.3 8.15 M3 05/21/1997 1120 29.59 2 14.2 14.5 21.7 214 8.44 M3 06/09/1997 29.42 2.6 5.8 5.5 24 24 7.68 M3 06/25/1997 1050 29.2 2.4 8.6 8.2 27.5 7.92 M3 07/08/1997 1315 28.98 2 9.8 7.5 28.5 28.3 7.6 M3 07/29/1997 1215 28.65 1.8 10 5 29 27.5 7.62 M3 08/12/1997 1325 28.37 1.6 5.6 5.6 23 23 7.85 M3 08/27/1997 1200 28 1.5 7 2.5 29.1 28.1 8.27 M3 09/10/1997 1115 27.54 1 6.4 6.3 22.7 22.8 8.13 M3 09/24/1997 1125 27.49 .8 2.9 2.8 16.8 16.8 7.88 M3 10/14/1997 1130 27.68 1 11.1 11 12.3 12.4 8.31 
--

--

-- -- 

--

-- --

-- -- -- -- -- -- -- -- 

--

--

--

-- -- 

-- --

-- -- -- -- -- -- -- 

-- -- --

-- ---- --

-- --

-- -- --

-- -- 

-- -- -- -- -- -- -- -- 

-- -- -- -- -- -- -- -- 

-- -- --

-- -- -- -- -- -- -- 

-- -- -- -- -- -- -- -- 

-- -- -- -- -- -- -- -- 

-- -- -- -- -- -- -- -- 

-- --

--

--

-- -- --

-- --

-- --

--

--

----

--

--

--

--

Table 3. Analyses of surface-water samples collected from Little Bean Marsh.—Continued [ft, feet; mg/L, milligrams per liter; °C, degrees Celsius; μS/cm, microsiemens per centimeter at 25 degrees Celsius; FTU, formazine turbidity units; Eh, oxidation-reduction pot

--, no data; <, less than] 

Dissolved Dissolved Water Water 

Site Gage Water oxygen oxygen temperature temperature pH, 

number height, depth, (top), (bottom), (top), (bottom), in standard co

(fig. 1) Date Time in ft in ft in mg/L in mg/L in °C in °C units 

M3 10/28/1997 1045 27.94 1.1 13.9 13.9 7.5 7.67 M3 11/19/1997 1315 27.86 1.3 19 19.5 6.2 6.3 7.5 M3 12/02/1997 28.3 1.3 11.4 11.6 6.5 6.5 7.75 M3 12/19/1997 1115 2 4.3 8.37 M4 05/21/1997 1055 1.5 6.7 5.3 19.4 18.8 7.45 M4 06/10/1997 1 .5 .5 20.3 7.6 M4 06/25/1997 M4 07/08/1997 1245 28.98 1.25 4.6 .2 26.8 24.2 7.75 M4 07/29/1997 1145 28.65 .9 8.5 4.5 26.9 24.8 6.83 M4 08/12/1997 1145 28.37 .92 1.4 1.6 21.5 21.5 7.01 M4 08/27/1997 1120 28 .33 3.8 .1 23 7.3 M4 10/28/1997 0945 27.94 .25 12.3 5.9 6.71 M4 11/19/1997 1130 27.86 .17 13.8 .6 M4 12/02/1997 1345 28.3 8.4 5.9 7.48 M4 12/19/1997 1100 .83 1.5 7.78 M5 11/15/1996 1415 2.5 11 7.1 7.13 M5 11/25/1996 29.09 M5 12/02/1996 1210 12.7 3.9 7.52 M5 12/19/1996 1100 29.12 15.7 2.1 8.25 M5 01/03/1997 1315 28.85 13.1 2.7 7.48 M5 01/20/1997 1400 28.27 13 2.6 7.49 M5 02/05/1997 28.27 16.3 3.4 6.98 M5 02/17/1997 1210 27.94 9.2 3.7 7.29 M5 02/21/1997 M5 02/24/1997 M5 03/17/1997 28.25 12.4 6.3 8.37 M5 03/31/1997 28.05 M5 04/05/1997 M5 04/10/1997 M5 04/11/1997 
--

-- -- -- 

--

-- -- --

--

--

--

--

-- --

-- --

-- -- -- -- -- --

-- --

--

-- -- --

-- -- 

-- --

-- -- -- -- -- -- 

-- --

-- --

-- -- 

-- --

--

--

-- --

Table 3. Analyses of surface-water samples collected from Little Bean Marsh.—Continued [ft, feet; mg/L, milligrams per liter; °C, degrees Celsius; μS/cm, microsiemens per centimeter at 25 degrees Celsius; FTU, formazine turbidity units; Eh, oxidation-reduction pot

--, no data; <, less than] 

Dissolved Dissolved Water Water 

Site Gage Water oxygen oxygen temperature temperature pH, 

number height, depth, (top), (bottom), (top), (bottom), in standard co

(fig. 1) Date Time in ft in ft in mg/L in mg/L in °C in °C units 

M5 04/17/1997 1400 29.27 12.5 16.8 8.42 M5 05/05/1997 8.4 8.1 19.9 20.3 8.22 M5 05/21/1997 1015 29.59 1.4 1.7 19.4 19.1 7.59 M5 06/10/1997 5.3 4.2 21.3 21.2 7.85 M5 06/25/1997 29.2 M5 07/08/1997 1215 28.98 1.6 6.3 5.8 27.4 27.4 8.36 M5 07/29/1997 1120 28.65 2.2 3 3 27.4 26.8 7.18 M5 08/12/1997 1240 28.37 2.3 3.6 3.4 22.8 22.8 7.57 M5 08/27/1997 1140 28 2 7 .1 28.3 27.3 8.09 M5 09/10/1997 1105 27.54 2 7.6 9.8 22.4 22.2 7.7 M5 09/24/1997 1100 27.49 2 1.9 1.6 17.5 17.5 7.54 M5 10/14/1997 1100 27.68 7.2 12 7.93 M5 10/28/1997 1015 27.94 1.5 15.4 15.4 6 7.25 M5 11/19/1997 1215 27.86 18 17.5 5.5 5.7 7.62 M5 12/02/1997 28.3 10.2 6.3 7.45 M5 12/19/1997 1045 2 4 7.4 M6 09/10/1997 1330 8.5 M6 09/24/1997 1.5 3.9 3.5 17 17 7.6 M6 10/14/1997 1355 27.68 .7 16.2 12.7 8.08 M6 11/19/1997 1445 27.86 1.2 14.3 9 7.96 M6 12/02/1997 1115 8 5.9 6.98 M6 12/19/1997 1300 2 4.4 7.25 
-- -- -- -- -- -- -- -- 

--

-- 

--

--

--

--

-- -- -- -- -- -- 

--

--

--

-- -- -- -- -- 

-- --

--

Table 3. Analyses of surface-water samples collected from Little Bean Marsh.—Continued [ft, feet; mg/L, milligrams per liter; °C, degrees Celsius; μS/cm, microsiemens per centimeter at 25 degrees Celsius; FTU, formazine turbidity units; Eh, oxidation-reduction pot

--, no data; <, less than] 

Total Total Total Dissolved Suspended 

Site Total dissolved suspended ammonium ammonium ammonium N

number Eh, nitrogen, nitrogen, nitrogen, nitrogen, nitrogen, nitrogen, 

(fig. 1) Date Color in millivolts in mg/L in mg/L in mg/L in mg/L in mg/L in mg/L 

M1 03/17/1997 M1 03/31/1997 10 364 M1 04/17/1997 2 354 0.53 0.35 0.18 0.015 0.005 0.01 M1 05/05/1997 7 .77 .42 .35 .031 <.005 .029 M1 05/21/1997 7 322 .75 .48 .27 .10 <.005 .10 M1 06/10/1997 0 365 .43 .3 .13 .015 .015 0 M1 06/25/1997 0 .42 .4 .02 .0025 <.005 0 M1 07/08/1997 0 .41 .29 .12 .02 .017 .003 M1 07/29/1997 2 378 .7 .62 .08 .011 .007 .004 M1 08/12/1997 0 308 .58 .3 .28 .03 .028 .002 M1 12/02/1997 17 .65 .49 .16 .046 .005 .041 M1 12/19/1997 0 .067 0 M2 11/15/1996 22 390 .9 .5 .4 .18 <.005 .18 M2 12/02/1996 20 377 .67 .39 .28 .11 .017 .096 M2 12/19/1996 33 458 .52 .42 .1 .039 <.005 .037 M2 01/03/1997 14 371 .35 .24 .11 .048 <.005 .046 M2 01/20/1997 375 .92 0 .211 0 M2 02/05/1997 6 390 .61 .48 .13 .052 .014 .038 M2 02/17/1997 2 410 .51 .38 .13 .006 .022 0 M2 02/24/1997 15 2.93 .92 2.01 .16 .011 .15 M2 03/17/1997 11 .88 .25 .63 <.005 <.005 0 M2 03/31/1997 13 385 .91 .5 .41 <.005 <.005 0 M2 04/17/1997 85 357 1.34 .54 .8 .15 .006 .15 M2 05/05/1997 5 1.01 .44 .57 .12 <.005 .12 M2 05/21/1997 9 355 1.32 .83 .49 .32 .17 .15 M2 06/10/1997 15 368 1.08 .8 .28 .012 .008 .004 M2 06/25/1997 13 .94 .83 .11 <.005 <.005 0 M2 07/08/1997 16 .51 .84 0 .031 .011 .02 M2 07/29/1997 16 368 1.5 .86 .64 .047 .035 .012 M2 08/12/1997 12 308 2.12 1.1 1.02 .038 .072 0 
-- -- -- -- -- -- -- -- 

--

--

--

-- -- 

--

--

--

----

--

--

--

Table 3. Analyses of surface-water samples collected from Little Bean Marsh.—Continued [ft, feet; mg/L, milligrams per liter; °C, degrees Celsius; μS/cm, microsiemens per centimeter at 25 degrees Celsius; FTU, formazine turbidity units; Eh, oxidation-reduction pot

--, no data; <, less than] 

Total Total Total Dissolved Suspended 

Site Total dissolved suspended ammonium ammonium ammonium N

number Eh, nitrogen, nitrogen, nitrogen, nitrogen, nitrogen, nitrogen, 

(fig. 1) Date Color in millivolts in mg/L in mg/L in mg/L in mg/L in mg/L in mg/L 

M2 08/27/1997 20 401 2.77 1.69 1.08 0.28 0.14 0.13 M2 09/10/1997 0 357 1.84 1.1 .74 .012 <.005 .01 M2 09/24/1997 26 387 1.85 1.14 .71 .24 .17 .071 M2 10/14/1997 24 390 .62 .76 0 .01 .16 0 M2 10/28/1997 19 365 .68 .54 .14 .023 .017 .006 M2 11/19/1997 8 395 .74 .65 .09 .038 <.005 .036 M2 11/19/1997 M2 12/02/1997 35 .85 .63 .22 .20 <.005 .20 M2 12/19/1997 13 .6 .41 .19 .017 <.005 .015 M3 11/15/1996 22 360 .66 .46 .2 .045 <.005 .043 M3 11/25/1996 380 .44 .39 .05 .037 <.005 .035 M3 12/18/1996 26 410 .52 .48 .04 .021 M3 01/03/1997 20 377 .67 .59 .08 .022 <.005 .02 M3 01/20/1997 42 378 1.28 1.26 .02 .13 <.005 .13 M3 02/05/1997 9 421 .73 .5 .23 .066 .02 .046 M3 02/17/1997 18 414 .63 .39 .24 .037 <.005 .035 M3 02/24/1997 19 .69 .43 .26 .16 .009 .15 M3 03/17/1997 2 .69 .32 .37 .028 <.005 .026 M3 03/31/1997 7 .85 .5 .35 .022 <.005 .02 M3 04/17/1997 75 370 .91 .74 .17 .13 <.005 .12 M3 05/05/1997 5 .94 .53 .41 <.005 <.005 .023 M3 05/21/1997 10 .96 .66 .3 .028 <.005 .026 M3 06/09/1997 8 370 1.06 .76 .3 .009 <.005 .007 M3 06/25/1997 9 .94 .74 .2 .01 <.005 .008 M3 07/08/1997 11 .99 .67 .32 .022 .009 .013 M3 07/29/1997 6 352 1.35 .81 .54 .027 <.005 .025 M3 08/12/1997 5 382 1.11 .89 .22 .045 .012 .033 M3 08/27/1997 0 1.51 .95 .56 .14 <.005 .14 M3 09/10/1997 20 383 1.68 .99 .69 .13 .012 .11 M3 09/24/1997 24 368 1.9 1.34 .56 .41 .31 .098 M3 10/14/1997 18 360 1.11 .82 .29 .12 .089 .028 
--

----

-- -- -- -- -- -- -- -- 

--

-- ---- -- --

-- -- --

-- --

--

--

--

--

--

--

Table 3. Analyses of surface-water samples collected from Little Bean Marsh.—Continued [ft, feet; mg/L, milligrams per liter; °C, degrees Celsius; μS/cm, microsiemens per centimeter at 25 degrees Celsius; FTU, formazine turbidity units; Eh, oxidation-reduction pot

--, no data; <, less than] 

Total Total Total Dissolved Suspended 

Site Total dissolved suspended ammonium ammonium ammonium N

number Eh, nitrogen, nitrogen, nitrogen, nitrogen, nitrogen, nitrogen, 

(fig. 1) Date Color in millivolts in mg/L in mg/L in mg/L in mg/L in mg/L in mg/L 

M3 10/28/1997 17 352 0.9 0.74 0.16 0.095 <0.005 0.093 M3 11/19/1997 8 389 .74 .77 0 .13 .014 .11 M3 12/02/1997 33 1.09 .95 .14 .11 .007 .10 M3 12/19/1997 8 .85 .55 .3 .026 .019 .007 M4 05/21/1997 11 401 .89 .68 .21 .024 <.005 .022 M4 06/10/1997 15 365 .98 .82 .16 .018 .006 .012 M4 06/25/1997 23 1.24 .94 .3 .11 .065 .041 M4 07/08/1997 10 379 .97 .77 .2 .03 .014 .016 M4 07/29/1997 16 374 1.06 .7 .36 .033 .016 .017 M4 08/12/1997 15 387 1.18 .74 .44 .043 .029 .014 M4 08/27/1997 15 381 1.21 .96 .25 .079 .05 .029 M4 10/28/1997 20 365 .95 .77 .18 .071 <.005 .069 M4 11/19/1997 8 406 .67 .72 0 .063 <.005 .061 M4 12/02/1997 10 1.29 1.16 .13 .13 .011 .12 M4 12/19/1997 6 .88 .67 .21 .087 .011 .076 M5 11/15/1996 25 360 1.4 1 .4 .34 .2 .14 M5 11/25/1996 M5 12/02/1996 22 399 .84 .49 .35 .22 <.005 .22 M5 12/19/1996 43 473 .92 .74 .18 .11 <.005 .11 M5 01/03/1997 14 367 .95 .45 .5 .37 <.005 .37 M5 01/20/1997 377 .81 .55 .26 .0025 <.005 0 M5 02/05/1997 6 410 .46 .32 .14 .043 <.005 .041 M5 02/17/1997 10 411 .64 .42 .22 .06 <.005 .058 M5 02/21/1997 7 .59 .31 .28 .059 <.005 .057 M5 02/24/1997 9 .96 .36 .6 .087 <.005 .085 M5 03/17/1997 7 1.17 .42 .75 .019 <.005 .017 M5 03/31/1997 1.75 .78 .97 .078 <.005 .076 M5 04/05/1997 1.35 .62 .73 .008 0 M5 04/10/1997 1.22 .81 .41 .12 0 M5 04/11/1997 .79 .67 .12 .051 0 
-- --

--

--

--

----

Table 3. Analyses of surface-water samples collected from Little Bean Marsh.—Continued [ft, feet; mg/L, milligrams per liter; °C, degrees Celsius; μS/cm, microsiemens per centimeter at 25 degrees Celsius; FTU, formazine turbidity units; Eh, oxidation-reduction pot

--, no data; <, less than] 

Total Total Total Dissolved Suspended 

Site Total dissolved suspended ammonium ammonium ammonium N

number Eh, nitrogen, nitrogen, nitrogen, nitrogen, nitrogen, nitrogen, 

(fig. 1) Date Color in millivolts in mg/L in mg/L in mg/L in mg/L in mg/L in mg/L 

M5 04/17/1997 11 348 1.04 0.34 0.7 0.081 0.028 0.053 M5 05/05/1997 1.61 .47 1.14 .031 <.005 .029 M5 05/21/1997 1 461 .97 .68 .29 .026 <.005 .024 M5 06/10/1997 0 365 .88 .68 .2 .017 .008 .009 M5 06/25/1997 3 .79 .73 .06 .1 <.005 .098 M5 07/08/1997 3 376 .48 .84 0 .017 .009 .008 M5 07/29/1997 0 352 1.05 .82 .23 .04 .016 .024 M5 08/12/1997 5 379 1.21 1.07 .14 .12 .084 .034 M5 08/27/1997 0 346 1.9 .92 .98 .11 <.005 .11 M5 09/10/1997 2 375 3 .85 2.15 .69 <.005 .69 M5 09/24/1997 22 340 2.99 1.92 1.07 1.32 .93 .39 M5 10/14/1997 25 370 1.35 1.36 0 .59 .54 .046 M5 10/28/1997 13 356 .84 .86 0 .13 .071 .058 M5 11/19/1997 10 389 .86 .72 .14 .15 <.005 .14 M5 12/02/1997 27 .96 .67 .29 .02 .011 .009 M5 12/19/1997 3 .92 .66 .26 .028 <.005 .026 M6 09/10/1997 2 342 2.2 1.11 1.09 .19 <.005 .192 M6 09/24/1997 24 389 2.21 1.08 1.13 .30 .18 .124 M6 10/14/1997 23 377 .7 .85 0 .009 .016 0 M6 11/19/1997 8 390 .8 .77 .03 .11 <.005 .108 M6 12/02/1997 35 1.34 .62 .72 .049 <.005 .047 M6 12/19/1997 18 .58 .46 .12 .018 <.005 .016 
-- -- -- -- --

-- -- -- -- --

--

--

--

---- 

-- -- --

-- -- 

-- -- 

-- -- 

-- -- 

-- -- --

-- -- 

-- -- 

-- -- 

-- -- 

--

--

-- 

--

-- 

-- -- 

-- -- 

---- 

-- 

-- -- 

-- -- 

--

Table 3. Analyses of surface-water samples collected from Little Bean Marsh.—Continued [ft, feet; mg/L, milligrams per liter; °C, degrees Celsius; μS/cm, microsiemens per centimeter at 25 degrees Celsius; FTU, formazine turbidity units; Eh, oxidation-reduction pot

--, no data; <, less than] 

Dissolved Soluable Total 

Site Total organic organic Total Dissolved reactive Suspended Chlorophyll suspended s

number nitrogen, nitrogen, phosphorus, phosphorus, phosphorus, phosphorus, mean, solids, 

(fig. 1) Date in mg/L in mg/L in mg/L in mg/L in mg/L in mg/L in μg/L in mg/L 

M1 03/17/1997 0 M1 03/31/1997 0 M1 04/17/1997 0.408 0.238 0.074 0.038 0.033 .036 5.2 M1 05/05/1997 .737 .415 .147 .042 .031 .105 16 M1 05/21/1997 .639 .47 .118 .066 .057 .052 12.9 M1 06/10/1997 .413 .283 .247 .057 .043 .19 40.8 M1 06/25/1997 .394 .374 .138 .055 .046 .083 10 M1 07/08/1997 .388 .271 .125 .065 .061 .06 3.2 M1 07/29/1997 .687 .611 .119 .045 .036 .074 2.7 M1 08/12/1997 .548 .27 .211 .036 .032 .175 30.6 10.5 M1 12/02/1997 .359 .24 .739 .587 .608 .152 8.2 23 M1 12/19/1997 .106 .054 0 11.7 M2 11/15/1996 .706 .488 .070 .019 .015 .051 M2 12/02/1996 .550 .366 .061 .018 .006 .043 M2 12/19/1996 .479 .415 .034 .018 .011 .016 M2 01/03/1997 .300 .235 .101 .060 .071 .041 M2 01/20/1997 .020 .009 0 M2 02/05/1997 .504 .412 .097 .025 .017 .072 M2 02/17/1997 .502 .356 .055 .015 .012 .04 M2 02/24/1997 2.374 .516 .366 .052 .046 .314 M2 03/17/1997 .875 .245 .116 .012 .006 .104 M2 03/31/1997 .905 .495 .108 .029 .020 .079 M2 40/17/1997 .999 .346 .189 .014 .008 .175 56.6 M2 05/05/1997 .887 .435 .159 .019 .005 .14 70 M2 05/21/1997 .999 .658 .147 .035 .022 .112 59.6 M2 06/10/1997 1.07 .79 .328 .111 .074 .217 350 M2 06/25/1997 .922 .812 .150 .095 .072 .055 14 M2 07/08/1997 .477 .827 .151 .04 .026 .111 16.4 M2 07/29/1997 1.45 .823 .224 .049 .030 .175 38.5 21 M2 08/12/1997 2.08 1.03 .274 .024 .012 .25 83.4 19 
-- -- -- -- --

-- -- 

-- -- -- 

-- -- --

-- -- -- -- -- -- -- -- 

-- -- 

-- -- 

-- -- 

--

--

--

--

--

--

-- -- 

-- -- 

Table 3. Analyses of surface-water samples collected from Little Bean Marsh.—Continued [ft, feet; mg/L, milligrams per liter; °C, degrees Celsius; μS/cm, microsiemens per centimeter at 25 degrees Celsius; FTU, formazine turbidity units; Eh, oxidation-reduction pot

--, no data; <, less than] 

Dissolved Soluable Total 

Site Total organic organic Total Dissolved reactive Suspended Chlorophyll suspended s

number nitrogen, nitrogen, phosphorus, phosphorus, phosphorus, phosphorus, mean, solids, 

(fig. 1) Date in mg/L in mg/L in mg/L in mg/L in mg/L in mg/L in μg/L in mg/L 

M2 08/27/1997 2.49 1.54 0.316 0.028 0.007 0.288 85.5 66.8 M2 09/10/1997 1.83 1.10 .216 .034 .015 .182 76 14.5 M2 09/24/1997 1.60 .965 .156 .061 .035 .095 72.7 19.5 M2 10/14/1997 .608 .597 .128 .044 .034 .084 5.4 10.2 M2 10/28/1997 .645 .511 .098 .03 .022 .068 5.3 11 M2 11/19/1997 .700 .645 .057 .019 .008 .038 10.2 10.4 M2 11/19/1997 0 M2 12/02/1997 .645 .622 .373 .254 .239 .119 18.2 16 M2 12/19/1997 .581 .405 .142 .066 .030 .076 10.1 20 M3 11/15/1996 .607 .45 .036 .014 .014 .022 M3 11/25/1996 .394 .379 .030 .015 .015 M3 12/18/1996 .499 .480 0 M3 01/03/1997 .646 .585 .043 .017 .011 .026 M3 01/20/1997 1.11 1.22 .084 .046 .063 .038 M3 02/05/1997 .656 .472 .062 .039 .028 .023 M3 02/17/1997 .591 .385 .070 .021 .010 .049 M3 02/24/1997 .492 .38 .111 .015 .007 .096 M3 03/17/1997 .66 .315 .062 .014 .005 .048 M3 03/31/1997 .826 .495 .071 .02 .006 .051 M3 04/17/1997 .782 .735 .134 .012 .004 .122 28.8 M3 05/05/1997 .913 .525 .157 .030 .017 .127 45 M3 05/21/1997 .93 .655 .149 .030 .016 .119 85.2 M3 06/09/1997 1.05 .755 .142 .051 .021 .091 26 M3 06/25/1997 .928 .735 .088 .043 .026 .045 6.3 M3 07/08/1997 .966 .659 .092 .030 .018 .062 22.5 M3 07/29/1997 1.32 .805 .214 .047 .033 .167 30.1 95 M3 08/12/1997 1.06 .876 .134 .034 .024 .1 47 26 M3 08/27/1997 1.37 .945 .165 .043 .026 .122 42.7 24.4 M3 09/10/1997 1.55 .976 .232 .058 .033 .174 67.5 36 M3 09/24/1997 1.43 .972 .208 .089 .036 .119 84.4 39.5 M3 10/14/1997 .943 .681 .126 .030 .019 .096 5.3 31.3 
--

-- 

--

-- -- 

-- -- 

-- -- -- -- -- -- -- -- 

-- -- 

-- -- 

-- -- 

-- -- 

-- -- 

-- -- 

-- -- 

-- -- 

-- -- 

-- 

-- 

-- -- 

-- -- 

Table 3. Analyses of surface-water samples collected from Little Bean Marsh.—Continued [ft, feet; mg/L, milligrams per liter; °C, degrees Celsius; μS/cm, microsiemens per centimeter at 25 degrees Celsius; FTU, formazine turbidity units; Eh, oxidation-reduction pot

--, no data; <, less than] 

Dissolved Soluable Total 

Site Total organic organic Total Dissolved reactive Suspended Chlorophyll suspended s

number nitrogen, nitrogen, phosphorus, phosphorus, phosphorus, phosphorus, mean, solids, 

(fig. 1) Date in mg/L in mg/L in mg/L in mg/L in mg/L in mg/L in μg/L in mg/L 

M3 10/28/1997 0.803 0.735 0.100 0.022 0.009 0.078 16.4 17.5 M3 11/19/1997 .549 .692 .066 .020 .005 .046 17.2 8.4 M3 12/02/1997 .963 .925 .211 .169 .155 .042 8.4 4.8 M3 12/19/1997 .822 .529 .092 .047 .019 .045 8.6 9.6 M4 05/21/1997 .864 .675 .084 .034 .02 .05 12 M4 06/10/1997 .96 .812 .271 .164 .128 .107 13.7 M4 06/25/1997 1.13 .873 .164 .102 .070 .062 43.3 M4 07/08/1997 .938 .754 .069 .030 .017 .039 M4 07/29/1997 1.02 .682 .117 .043 .027 .074 32.4 5.7 M4 08/12/1997 1.13 .709 .104 .026 .015 .078 22.2 14.7 M4 08/27/1997 1.13 .908 .159 .034 .023 .125 9.4 3.6 M4 10/28/1997 .877 .765 .142 .031 .016 .111 21.6 27 M4 11/19/1997 .605 .715 .076 .020 .007 .056 18.8 17.6 M4 12/02/1997 1.15 1.15 .300 .238 .229 .062 10 11 M4 12/19/1997 .791 .657 .093 .043 .016 .05 11.7 13.5 M5 11/15/1996 .921 .665 .114 .018 .018 .096 M5 11/25/1996 M5 12/02/1996 .62 .485 .072 .021 .007 .051 M5 12/19/1996 .803 .735 .069 .017 .002 .052 M5 01/03/1997 .578 .445 .037 .015 .008 .022 M5 01/20/1997 .805 .545 .070 .020 .008 .05 M5 02/05/1997 .373 .274 .053 .018 .010 .035 M5 02/17/1997 .578 .415 .090 .013 .006 .077 M5 02/21/1997 .505 .282 .130 .017 .014 .113 M5 02/24/1997 .812 .297 .087 .016 .011 .071 M5 03/17/1997 1.15 .415 .141 .170 .004 0 M5 03/31/1997 1.67 .775 .480 .030 .010 .454 M5 04/05/1997 1.35 .61 .235 .031 .013 .204 M5 04/10/1997 1.2 .67 .150 .038 .020 .112 38.6 M5 04/11/1997 .713 .542 .069 .034 .019 .035 4 
--

-- 

--

--

-- 

-- 

Table 3. Analyses of surface-water samples collected from Little Bean Marsh.—Continued [ft, feet; mg/L, milligrams per liter; °C, degrees Celsius; μS/cm, microsiemens per centimeter at 25 degrees Celsius; FTU, formazine turbidity units; Eh, oxidation-reduction pot

--, no data; <, less than] 

Dissolved Soluable Total 

Site Total organic organic Total Dissolved reactive Suspended Chlorophyll suspended s

number nitrogen, nitrogen, phosphorus, phosphorus, phosphorus, phosphorus, mean, solids, 

(fig. 1) Date in mg/L in mg/L in mg/L in mg/L in mg/L in mg/L in μg/L in mg/L 

M5 04/17/1997 0.761 0.114 0.016 0.016 0.005 0 50.4 M5 05/05/1997 1.58 .465 .268 .028 .012 .24 62 M5 05/21/1997 .942 .675 .100 .036 .021 .064 15 M5 06/10/1997 .861 .67 .059 .052 .028 .007 25.4 M5 06/25/1997 .688 .725 .100 .044 .025 .056 22.3 M5 07/08/1997 .461 .829 .154 .031 .019 .123 54.5 M5 07/29/1997 1.01 .802 .144 .036 .022 .108 33.8 29 M5 08/12/1997 1.09 .984 .16 .033 .022 .127 40.1 28 M5 08/27/1997 1.78 .915 .237 .033 .017 .204 104.5 44 M5 09/10/1997 2.31 .845 .604 .023 .011 .581 885.6 112 M5 09/24/1997 1.66 .979 .409 .205 .058 .204 86.5 88 M5 10/14/1997 .716 .772 .160 .031 .016 .129 13.7 35.2 M5 10/28/1997 .670 .748 .111 .023 .009 .088 12.7 16 M5 11/19/1997 .654 .658 .068 .02 .005 .048 21.5 9.2 M5 12/02/1997 .930 .649 .170 .086 .072 .084 14.2 14 M5 12/19/1997 .890 .655 .110 .05 .02 .06 14.7 10.9 M6 09/10/1997 2.00 1.10 .220 .045 .018 .175 45.2 26 M6 09/24/1997 1.90 .897 .278 .113 .033 .165 139 50 M6 10/14/1997 .689 .832 .084 .039 .024 .045 3.1 3 M6 11/19/1997 .502 .58 .092 .031 .01 .061 20.6 28 M6 12/02/1997 1.27 .594 .247 .163 .152 .084 17 11 M6 12/19/1997 .56 .455 .127 .068 .032 .059 7 14 

1Depth below ice. 
For additional information contact: 

U.S. Geological Survey 

Water Resources Discipline 

1400 Independence Road 

Mail Stop 100 

Rolla, MO 65401 

(573) 308-3664 or 

http://mo.water.usgs.gov 
Hydrology and Cycling of Nitrogen and Phosphorus in Little Bean Marsh: —Scientific Investigations Report 2004–5171

A Remnant Riparian Wetland along the Missouri River in Platte County, Missouri, 1996–97Dale W. Blevins—