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Sediment and phosphorus dynamics behind weirs in agricultural drainage ditches

By

Elizabeth Louise Usborne

A Thesis Submitted to the Faculty of Mississippi State University

in Partial Fulfillment of the Requirements for the Degree of Master of Science

in Wildlife and Fisheries Science in the Department of Wildlife, Fisheries and Aquaculture

Mississippi State, Mississippi

August 2012


Copyright 2012

By



Sediment and phosphorus dynamics behind weirs in agricultural drainage ditches

By


Approved: _________________________________ _________________________________ Robert Kröger Michael Cox Assistant Professor of Wildlife, Fisheries Professor of Plant and Soil Sciences & Aquaculture (Committee Member) (Director of Thesis) _________________________________ _________________________________ Matthew Moore Christopher Boyd Research Ecologist Assistant Extension Professor of USDA Agricultural Research Service Environmental Sciences (Committee Member) Coastal Research & Extension Center (Committee Member) _________________________________ _________________________________ Eric Dibble Bruce Leopold Professor of Wildlife, Fisheries Professor of Wildlife, Fisheries & Aquaculture & Aquaculture (Graduate Coordinator) (Department Head) _________________________________ George Hopper Dean of College of Forest Resources


Name: Elizabeth Louise Usborne Date of Degree: August 11, 2012 Institution: Mississippi State University Major Field: Wildlife and Fisheries Science Major Professor: Dr. Robert Kröger Title of Study: Sediment and phosphorus dynamics behind weirs in agricultural drainage

ditches Pages in Study: 98 Candidate for Degree of Master of Science

Low grade rip rap weirs installed in agricultural surface drainage ditches manage

downstream eutrophication by slowing water flow, allowing sediments time to settle out

of the water column and phosphorus (P) to sorb to soil. A laboratory experiment was

conducted in microcosm chambers to simulate increased hydraulic residence time caused

by weirs and two field studies were conducted to compare experimental data with field

data and determine sediment deposition rates. One field study monitored weirs monthly

after installation and the other measured weirs of varying ages. Weirs retained

significantly more water and sediment than controls. Longer inundation times led to

abiotic factors known to release P during hydrologic flux, but did not translate to reduced

P storage. By converting intermittently inundated sediments into more consistently

saturated sediments, weirs function as a viable conservation practice for about a year until

temporary P retention mechanisms and sediment retention capacities are reached.

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DEDICATION

I dedicate this thesis to my parents, whose love and support carried me through

this process. Especially my dad, whose enthusiasm for the environment and natural

history inspired my career since childhood.

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ACKNOWLEDGEMENTS

I am well aware it took a lot of effort from a lot of people for this research project

to reach completion and their contributions never for a moment went unnoticed. To all of

them I offer heartfelt thanks. I would like to acknowledge and thank the Departments of

Wildlife, Fisheries and Aquaculture and Plant and Soil Sciences at Mississippi State

University, and the Mississippi Agricultural and Forestry Experiment Station (MAFES)

South Farm Facility for their support and use of their facilities. I would also like to thank

my funding agency, the United States Department of Agriculture, Agriculture and Food

Research Initiative (USDA AFRI). A huge debt of gratitude is owed, of course, to

Robbie Kröger and my entire graduate committee for your advice, patience, and guidance

through this experience. I know that the skills and lessons you’ve bestowed to me are

still to be fully realized and will serve me well into the future. Thank you to the

undergraduate student workers Thomas Arington, Thomas Morgan, Alex Blake, and

Jonathan Stoll who will never fully understand how indispensable and appreciated their

long hours were. I would like to thank Samuel Pierce for help in the lab, field, and,

especially, statistics. Thank you, Dan Prevost and Tim Huggins from Delta Wildlife for

your collaboration and help in the field. Thank you to Sam Testa and Melinda Josey for

not only tolerating an outsider in their laboratories, but extending generous hospitality.

To Keith Crouse, at Mississippi State University Soil Testing Extension Service, thank

you for your patience with my questions and my quantity of samples. Thank you to my

colleagues and fellow graduate students Alex Littlejohn, Jason Brandt, Andrew

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McDonell, Tyler Stubbs, Dan Goetz, Beth Poganski, Corry Shoemaker, Corrin Flora, and

Jeanie Barlow for your companionship and hard work. Thanks to Dr. Jeff Hatten and

Janet Dewey of the Forestry Department for their expertise, insight, and use of their soil

core probe. Thank you, especially, to all of my neighbors and friends for all of the

support needed outside of work. Without the strong network they comprise, I’m

convinced this journey would have been impossible. Finally, thank you to Ursa, my dog,

for taking me on walks to clear my head. Always quick with a soft furry head to pet, her

moral support is boundless.

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TABLE OF CONTENTS

Page

DEDICATION .................................................................................................................... ii

ACKNOWLEDGEMENTS ............................................................................................... iii

LIST OF TABLES ............................................................................................................ vii

LIST OF FIGURES ......................................................................................................... viii

CHAPTER

I. INTRODUCTION .............................................................................................1

1.1 References ..............................................................................................5

II. EFFECTS OF INUNDATION DURATION ON PHOSPHORUS RETENTION BY LOWER MISSISSIPPI RIVER ALLUVIAL VALLEY AGRICULTURAL DRAINAGE DITCH SEDIMENTS ................7

2.1 Introduction ............................................................................................7 2.2 Methods..................................................................................................9

2.2.1 Study sites and sediment collection .................................................9 2.2.2 Experimental set up..........................................................................9 2.2.3 Sediment analysis...........................................................................10 2.2.4 Statistical analysis ..........................................................................11

2.3 Results ..................................................................................................12 2.3.1 Abiotic controls on P .....................................................................12 2.3.2 Phosphorus .....................................................................................13

2.4 Discussion ............................................................................................14 2.5 References ............................................................................................26

III. PRELIMINARY EVIDENCE OF SEDIMENT AND PHOSPHORUS DYNAMICS BEHIND NEWLY INSTALLED LOW GRADE WEIRS IN AGRICULTURAL DRAINAGE DITCHES ...............................29

3.1 Introduction ..........................................................................................29 3.2 Methods................................................................................................31

3.2.1 Study sites and sediment collection ...............................................31

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3.2.2 Sediment accumulation and water depth .......................................32 3.2.3 Sediment analysis...........................................................................32 3.2.4 Statistical analysis ..........................................................................34

3.3 Results ..................................................................................................34 3.3.1 Sediment accumulation and water depth .......................................34 3.3.2 Abiotic controls on P .....................................................................35 3.3.3 Total P and P fractions ...................................................................35 3.3.4 Correlations ....................................................................................35


IV. PRELIMINARY EVIDENCE OF SEDIMENT AND PHOSPHROUS DYNAMICS BEHIND LOW GRADE WEIRS OF VARYING AGES INSTALLED IN AGRICULTURAL DRAINAGE DITCHES .....................58

4.1 Introduction ..........................................................................................58 4.2 Methods................................................................................................60

4.2.1 Study sites and sediment collection ...............................................60 4.2.2 Sediment accumulation and water depth .......................................61 4.2.3 Sediment analysis...........................................................................62 4.2.4 Statistical analysis ..........................................................................64

4.3 Results ..................................................................................................65 4.3.1 Sediment accumulation and water depth .......................................65 4.3.2 Abiotic controls on P .....................................................................65 4.3.3 Total P and P fractions ...................................................................66


V. SYNTHESIS ....................................................................................................95

5.1 References ............................................................................................98

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LIST OF TABLES

TABLE Page

2.1 Crops and soil texture at sample sites throughout the Lower Mississippi River Alluvial Valley (LMAV), spring, 2010. ...................................................17

4.1 Calculated average volume of sediment accumulated behind weirs of various ages at different sites in Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) in August 2010. .........................................................71

4.2 Soil texture at study sites associated with Bee and Wolf Lake in Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) in August 2010. ......................................................................................................71

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LIST OF FIGURES

FIGURE Page

2.1 Location of sampling sites along the Lower Mississippi River Alluvial Valley (LMAV) and schematic showing sampling design and field replication. Samples collected March 13-15, 2010. ..........................................18

2.2 Microcosm chambers with collected Lower Mississippi River Alluvial Valley (LMAV) ditch sediments for laboratory experiment at South Farm Aquaculture Facility, Mississippi State University, 2010. .......................19

2.3 Average percentage of organic matter in drainage ditch sediments collected from 16 sites (n = 48) throughout the Lower Mississippi River Alluvial Valley (LMAV) March 13-15, 2010, and subjected to four inundation treatments in microcosm chambers through summer of 2010 (none = no inundation). Similar letters denote no significant difference (F = 0.56, P = 0.64). Standard deviation bars shown. .......................................20

2.4 Trends of sediment pH and overlying water pH through time for 3 day inundation treatments (A), 3 week inundation treatments(B), and 3 month inundation treatments (C) of drainage ditch sediments from 16 sites (n = 48) throughout the Lower Mississippi River Alluvial Valley (LMAV) in 2010. ...............................................................................................21

2.5 Average final pH of overlying water column (χ2 = 77.33, P < 0.01) (A) and sediment interface (χ2 = 66.33, P < 0.01) (B) of drainage ditch sediments collected from 16 sites (n=48) throughout the Lower Mississippi River Alluvial Valley (LMAV) and placed within microcosm chambers for four inundation treatments in 2010 (none = no inundation). Different letters denote significant difference. Standard deviation bars shown. .........................................................................................22

2.6 Average Eh (mV) of drainage ditch sediments collected from 16 sites (n = 48) throughout the Lower Mississippi River Alluvial Valley (LMAV) and placed within microcosm chambers for four inundation treatments in 2010 (none = no inundation). Different letters denote significant difference (χ2 = 147.19, P < 0.01). Standard deviation bars shown. .................23

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2.7 Average TP concentrations (µg/g) of drainage ditch sediments collected from 16 sites (n = 48) throughout the Lower Mississippi River Alluvial Valley (LMAV) and placed within microcosm chambers for four inundation treatments in 2010 (none = no inundation). Different letters denote significant difference (χ2 = 10.29, P = 0.02). Standard deviation bars shown. .........................................................................................................24

2.8 Average concentrations (µg/g) of soluble and loosely bound P (χ2 = 32.49, P < 0.01) (A), AlPO4 (χ

2 = 44.45, P < 0.01) (B), FePO4 (χ2 =

20.31, P < 0.01) (C), and reductant-soluble P (χ2 = 19.05, P < 0.01) (D) of drainage ditch sediments collected from 16 sites (n = 48) throughout the Lower Mississippi River Alluvial Valley (LMAV) and placed within microcosm chambers for four inundation treatments in 2010 (none = no inundation). Different letters denote significant difference. Standard deviation bars shown. .........................................................................................25

3.1 Location of study site along within an agricultural drainage ditch associated with Harris Bayou, Mississippi. ........................................................40

3.2 Initial limestone rip rap weir installation in two stage ditch design at East ditch, Harris Bayou, Mississippi, December 2010. ....................................41

3.3 Schematic showing how the distance between locations was chosen so as to avoid influences from other weirs. As depicted, the position of another study location occurs past the point where weir height intersects slope. ..................................................................................................................41

3.4 Schematic showing sediment accumulation and water depth measurement methods in Harris Bayou ditch site, December 2010 through November 2011. ....................................................................................42

3.5 Average sediment deposition at four sites within the east ditch of Stovall Farms in Harris Bayou, Mississippi after one year (December 2010 – November 2011) of weir installation (no weir located at inflow location). Similar letters denote no significant differences (χ2 = 2.42, P = 0.49). Standard deviation bars shown. ...........................................................43

3.6 Average water depth at four sites within the east ditch of Stovall Farms in Harris Bayou, Mississippi after one year (December 2010 – November 2011) of weir installation (no weir located at inflow location). Different letters denote significant differences (χ2 = 7.67, P = 0.05). Standard deviation bars shown. ..............................................................44

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3.7 Average percentage organic matter at four sites within the east ditch of Stovall Farms in Harris Bayou, Mississippi after six months (January 2011 – June 2011) of weir installation (no weir located at inflow location). Different letters denote significant differences (χ2 = 6.41, P = 0.09). Standard deviation bars shown. ..............................................................45

3.8 Average pH of sediment at four sites within the east ditch of Stovall Farms in Harris Bayou, Mississippi after six months (January 2011 – June 2011) of weir installation (no weir located at inflow location). Different letters denote significant differences (F = 25.05, P < 0.01). Standard deviation bars shown. ..........................................................................46

3.9 Average TP of sediment at four sites within the east ditch of Stovall Farms in Harris Bayou, Mississippi after six months (January 2011 – June 2011) of weir installation (no weir located at inflow location). Similar letters denote no significant differences (χ2 = 3.22, P = 0.36). Standard deviation bars shown. ..........................................................................47

3.10 Average soluble and loosely bound P (χ2 = 5.89, P = 0.12) (A), AlPO4 (χ2 = 6.45, P = 0.09) (B), FePO4 (χ

2 = 6.70, P = 0.08) (C), and reductant-soluble P (χ2 = 2.24, P = 0.52) (D) concentrations in sediment sampled at four sites within the east ditch of Stovall Farms in Harris Bayou, Mississippi after six months (January 2011 – June 2011) of weir installation (no weir located at inflow location). Different letters denote significant differences. Standard deviation bars shown. ...................................48

3.11 Average bioavailability ratios of sediment at four sites within the east ditch of Stovall Farms in Harris Bayou, Mississippi after six months (January 2011 – June 2011) of weir installation (no weir located at inflow location). Similar letters denote no significant differences (χ2 = 2.09, P = 0.55). Standard deviation bars shown. ...............................................49

3.12 Correlations between water depth (F = 5.86, P = 0.02, R2 = 0.13) (A), organic matter percentage (F = 5.23, P = 0.03, R2 = 0.19) (B), pH (F = 18.12, P < 0.01, R2 = 0.45) (C), and TP (F = 3.57, P = 0.07, R2 = 0.14) (D) and location within the east ditch of Stovall Farms in Harris Bayou, Mississippi. Location 0 denotes inflow, 1 denotes weir 1, 2 denotes weir 2, and 3 denotes weir 3. ..............................................................................50

3.13 Correlations between water depth (F = 3.99, P = 0.08, R2 = 0.33) (A), pH (F = 12.93, P = 0.02, R2 = 0.76) (B), and bioavailability ratio (F = 44.14, P < 0.01, R2 = 0.92) (C) with time at the inflow location within the east ditch of Stovall Farms in Harris Bayou, Mississippi. Time 0 denotes weir installation in December 2010, 1 denotes one month after installation in January 2011, 2 denotes two months after installation in February 2011, etc. .............................................................................................51

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3.14 Organic matter percentage and pH correlation (F = 11.80, P = 0.03, R2 = 0.75) at inflow location (A), and TP and pH correlation (F = 6.74, P = 0.06, R2 = 0.63) at weir 3 location (B) within the east ditch of Stovall Farms in Harris Bayou, Mississippi. ..................................................................52

3.15 Correlations between bioavailability ratio and pH (F = 6.50, P = 0.06, R2 = 0.62) at the inflow (A), bioavailability ratio and sediment depth (F = 9.14, P = 0.04, R2 = 0.70) at weir 1 (B), and bioavailability ratio and organic matter (F = 23.10, P = 0.01, R2 = 0.85) at weir 3 (C) within the east ditch of Stovall Farms in Harris Bayou, Mississippi from January 2011 to June 2011. .............................................................................................53

4.1 Location of study sites associated with two oxbow lakes, Bee Lake and Wolf Lake, in Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) region. Nine drainage ditches containing 14 weirs of various ages at seven sites were surveyed and sampled in August 2010. .......................72

4.2 Schematic depicting control site location decision. Controls were located within the same drainage ditch as their respective weirs so as to eliminate undesired variables, but also located far enough away so as to be outside the influence of the weir. The distance of weir influence ended where height of the weir intercepted slope of the ditch. ..........................73

4.3 Example of a soil core taken from an agricultural drainage ditch to determine visually and tactilely where depositional sediment layer is different from original ditch bottom soil, and therefore sediment depth, Yazoo City, July 2010. .......................................................................................74

4.4 A technique to determine sediment depth behind a weir. A laser is leveled on the top of the weir and shot to a marker held above and below the weir. Height from the top of the sediment to the laser mark is recorded. Sediment depth is determined by the difference between heights. ...............................................................................................................75

4.5 Schematic showing measurements used to calculate estimated volume of sediment (m3) retained behind a weir. Volume of two triangular based pyramids are calculated and subtracted from each other to determine accumulated sediment volume. Volume of ditch is calculated using height and width of the exposed weir for the area of the base. Length is determined by ditch slope. The second pyramid has measured sediment depth added to the base height. ...........................................................76

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4.6 Average accumulation depth of sediments (F = 3.55, P = 0.08) (A) and water depth (S = 50.50, P = 0.02) (B) at 7 sites within Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) in August 2010. Different letters denote significant differences. Standard error bars shown. ................................................................................................................77

4.7 Averages (χ2 = 7.58, P = 0.02) (A) and regression analysis (F = 5.25, P = 0.05) (B) for difference in sediment depth between weirs and their corresponding controls in 7 drainage ditch sites within Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) in August 2010 for 1, 2, and 4 year old weirs. Different letters denote a significant difference. Standard error bars shown. .................................................................................78

4.8 Averages (F = 3.18, P = 0.13) (A) and regression analysis (F = 6.46, P = 0.04) (B) for difference in water depth between weirs and their corresponding controls in 7 drainage ditch sites within Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) in August 2010 for 1, 2, and 4 year old weirs. Similar letters denote no significant difference. Standard error bars shown. .................................................................................79

4.9 Average pH of drainage ditch sediments at 7 sites within Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) in August 2010. Similar letters denote no significant difference (F = 1.44, P = 0.24). Standard error bars shown. .................................................................................80

4.10 Averages (F = 2.74, P = 0.17) (A) and regression analysis (F = 3.32, P = 0.10) (B) for difference in sediment pH between weirs and their corresponding controls in 7 drainage ditch sites within Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) in August 2010 for 1, 2, and 4 year old weirs. Negative numbers indicate pH at control is greater than weir. Similar letters denote no significant difference. Standard error bars shown. .................................................................................81

4.11 Average organic matter percentage of drainage ditch sediments at 7 sites within Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) in August 2010. Similar letters denote no significant difference (F = 0.30, P = 0.59). Standard error bars shown. ......................................................82

4.12 Averages (F = 1.03, P = 0.39) (A) and regression analysis (F = 0.26, P = 0.62) (B) for difference in sediment organic matter percentage between weirs and their corresponding controls in 7 drainage ditch sites within Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) in August 2010 for 1, 2, and 4 year old weirs. Negative numbers indicate organic matter percentage greater at control than weir. Similar letters denote no significant difference. Standard error bars shown. ................83

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4.13 Average TP of drainage ditch sediments at 7 sites within Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) in August 2010. Similar letters denote no significant difference (S = 238.00, P = 0.76). Standard error bars shown. .................................................................................84

4.14 Averages (F = 1.14, P = 0.38) (A) and regression analysis (F = 0.99, P = 0.33) (B) for difference in sediment TP between weirs and their corresponding controls in 7 drainage ditch sites within Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) in August 2010 for 1, 2, and 4 year old weirs. Negative numbers indicate TP greater at control than weir. Similar letters denote no significant difference. Standard error bars shown. ................................................................................................85

4.15 Average soluble and loosely bound P (S = 247.00, P = 0.71) (A), AlPO4 (S = 262.00, P = 0.96) (B), FePO4 (S = 248.00, P = 0.74) (C), and reductant-soluble P (S = 327.00, P = 0.05) (D) concentrations of drainage ditch sediments at 7 sites within Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) in August 2010. Different letters denote significant differences. Standard error bars shown. ....................86

4.16 Average bioavailability ratio of drainage ditch sediments at 7 sites within Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) in August 2010. Similar letters denote no significant difference (S = 236.00, P = 0.48). Standard error bars shown. ..................................................87

4.17 Average difference in soluble and loosely bound P (χ2 = 3.22, P = 0. 20) (A), AlPO4 (χ

2 = 3.50, P = 0.17) (B), FePO4 (F = 3.30, P = 0.06) (C), and reductant-soluble P (F = 3.20, P = 0.06) (D) concentrations between weirs and their corresponding controls in 7 drainage ditch sites within Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) in August 2010 for 1, 2, and 4 year old weirs. Negative numbers indicate concentrations greater at control than weir. Different letters denote significant differences. Standard error bars shown. ..........................................88

4.18 Average difference in sediment bioavailability ratios between weirs and their corresponding controls in 7 drainage ditch sites within Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) in August 2010 for 1, 2, and 4 year old weirs. Different letters denote significant differences (χ2 = 4.77, P = 0.09). Standard error bars shown. .........89

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4.19 Least squares regression analysis of the difference in soluble and loosely bound P (F = 4.16, P = 0.05) (A), AlPO4 (F = 4.21, P = 0.05) (B), FePO4 linear (F < 0.01, P = 0.99) and polynomial (2 degrees, F = 3.30, P = 0.06) (C), and reductant-soluble P (F = 3.12, P = 0.09) (D) concentrations between weirs and their corresponding controls in 7 drainage ditch sites within Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) in August 2010 for 1, 2, and 4 year old weirs. Negative numbers indicate concentrations greater at control than weir. ...........90

4.20 Least squares regression analysis of the difference in sediment bioavailability ratios between weirs and their corresponding controls in 7 drainage ditch sites within Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) in August 2010 for 1, 2, and 4 year old weirs (F = 0.03, P = 0.86). Negative numbers indicate bioavailability ratio greater at control than weir. ................................................................................91

1

CHAPTER I

INTRODUCTION

The Mississippi River watershed drains half the continental United States

(USACE 2011). Its rich river valley soils make for very productive agriculture. Excess

nutrient inputs from industrial agriculture travel through this vast watershed until they are

finally deposited in the Gulf of Mexico where they contribute to a hypoxic “dead zone”

the size of Connecticut (UNESCO 2011). Not only are these nutrients problematic on

such a large, continental scale, but also locally and regionally in headwaters and smaller

drainage basins as they make their way downstream to the coast. Excessive phosphorus

(P) is one of the leading causes of eutrophication (Reddy and DeLaune 2008). As a

natural function of ecosystem succession, eutrophication is a slow process but can be

accelerated by excess nutrients, like P, from upland rural and urban land use. Primary

producers take advantage of large inputs of P, a normally limiting resource in aquatic

systems, leading to an increase in biomass. Through decomposition of this biomass by

respiring microorganisms, oxygen levels in water are depleted, leading to aquatic species

migration or die off. In addition to the ecological ramifications of nutrient loading,

recreation and fisheries economies are impacted by eutrophication (Pretty et al. 2002,

Rabalais et al. 2002).

Finding solutions to mitigating cultural eutrophication requires an understanding

of one of the main contributors. The challenge of managing P is that it has no significant

gaseous phase in its cycle. Once it enters a system, the only way it leaves is through

2

further runoff or erosion to another system downstream. Plant and microorganism

assimilation is a temporary solution, as biological uptake only sequesters P for the life of

the tissue. After death or senescence, P is released back into the system. Adsorption to

mineral surfaces or precipitation with ions in soils and sediment burial are the only long

term solutions to P storage.

Although most P is bound to soil solids, complex interactions can cause P to

move into solution. Most P in the overlying water column is particulate P which settles

out of the water column and deposits onto sediments. Consequently, overall net flux of P

is out of the water column and into soils. Due to greater dissolved P concentrations in

pore water, P has a positive diffusive flux, meaning movement from soil or sediment into

the overlying water column (Reddy and DeLaune 2008). Because P is such a limiting

resource in freshwater systems, the small amount that dissolves out of sediments can

greatly impact aquatic ecosystems. Therefore, increased P deposition leads to increased

dissolved P concentration in pore water, which leads to more P diffusing out of soils and

becoming bioavailable, encouraging eutrophication.

The interplay in equilibrium processes of P movement into and out of the water

column determines whether sediments behave like P sources or sinks. Ability of soils and

sediments to act as P sinks depends indirectly on abiotic regulated factors, such as 1)

redox potential, 2) pH, 3) clay and organic matter content, 4) temperature, and 5) time

(Reddy and D'Angelo 1997, Reddy and DeLaune 2008, Wang and Pant 2010).

Anaerobic soil conditions reduce chemical species in sediment. Specifically, reduction of

ferric iron in FePO4 to ferrous iron causes PO43- to become soluble (Patrick and Khalid

1974). Phosphorus solubility is pH dependent because of the effect pH has on iron and

aluminum phosphate precipitation at low pH and calcium phosphate precipitation at high

3

pH. In acidic conditions, P becomes insoluble by specific adsorption through a process

known as ligand exchange, whereby oxygen ions on iron and aluminum hydroxides are

replaced by phosphate anions. Electrostatic attraction, defined as nonspecific adsorption,

of phosphate anions is also regulated by pH, as well as clay content and organic matter,

which provide positive binding surfaces. Increased temperature accelerates rates of

adsorption and precipitation. Finally, over time, P can be incorporated into precipitates,

eventually assuming crystalline forms, leading to long term P storage.

Manipulation of hydraulic residence time affects many of the above mentioned

variables and therefore long term P retention by sediments. Controlled drainage with

slotted or drop pipes, flash board risers, and vegetated drainage ditches are some common

examples that use hydrological manipulation to effectively manage nutrients (Kröger et

al. 2008a, Kröger et al. 2008b). Low grade weirs, an in-ditch tool for controlled surface

drainage, are also now being implemented. As opposed to a dam, “low grade” refers to

the low profile of the weir, which allows drainage through the ditch during storm events.

Installed in agricultural drainage ditches, low grade weirs (here on referred to simply as

weirs) offer a potential solution that combine benefits of controlled drainage with wetland

ecosystem services as a means to reduce nutrient runoff into primary aquatic systems

(Kröger et al. 2008b). By altering drainage ditch hydrology, several things could happen.

Flow velocities within the ditch channel are reduced, promoting sediment deposition and

reducing bank erosion (Kuhnle et al. 1998). This is a physical way to limit transport of

particulate P, but rate of sediment accumulation behind a weir is unknown. Hydraulic

residence time is increased allowing more time for dissolved P in solution to interact with

bed sediments, which decreases bioavailability of P (Fisher and Acreman 2004, Kröger et

al. 2012). However, longer inundation times also alter sediment redox and pH, possibly

4

creating conditions that increase phosphate release from sediments in a system that

accumulates particulate P. The following research addresses the question of weirs’

potential to manipulate physicochemical factors that determine sediment ability to retain

P with the following objectives.

Objective 1: Determine if weirs affect abiotic factors (e.g., pH, Eh) relating to

changes in sediment P.

H01 = There are no differences in abiotic factors in ditch sediments influenced by

weirs.

H02 = P changes in sediment behind weirs does not differ from P changes in

sediment not influenced by weirs.

Objective 2: Determine if weirs affect sediment accumulation in agricultural

drainage ditches.

H0 = There is no difference in sediment accumulation in ditches with or without

weirs.

Objective 3: Estimate at what age physical and chemical processes for P retention

behind weirs reach saturation.

H01=Sediment accumulation rates in ditches behind weirs are the same as

sediment accumulation rates in ditches not influenced by weirs.

H02= P retention by sediment behind weirs reaches saturation at the same rate as P

retention by sediment not influenced by weirs.

H03=There is no difference in the saturation rates of sediment accumulation or P

retention by sediments behind a weir.

5

1.1 References Fisher, J. and M. C. Acreman. 2004. Wetland nutrient removal: a review of the evidence.

Hydrology and Earth System Sciences 8:673-685.

Kröger, R., C. M. Cooper, and M. T. Moore. 2008a. A preliminary study of an alternative controlled drainage strategy in surface drainage ditches: low-grade weirs. Agricultural Water Management 95:678-684.

Kröger, R., M. M. Holland, M. T. Moore, and C. M. Cooper. 2008b. Agricultural drainage ditches mitigate phosphorus loads as a function of hydrological variability. Journal of Environmental Management 37:107-113.

Kröger, R., R. E. Lizotte, Jr., F. D. Shields, Jr., and E. Usborne. 2012. Inundation influences on bioavailability of phosphorus in managed wetland sediments in agricultural landscapes. Journal of Environmental Management 41:1-11.

Kuhnle, R. A., R. L. Bingner, G. R. Foster, and E. H. Grissinger. 1998. Land use changes and sediment transport in Goodwin Creek. Pages 279-292 in P. C. Klingeman, R. L. Beschta, P. D. Komar, and J. B. Bradley, editors. Gravel-bed rivers in the environment. Water Resources Publications, LLC, Highlands Ranch, CO.

Patrick, W. H. J. and R. A. Khalid. 1974. Phosphate release and sorption by soils and sediments: effect of aerobic and anaerobic conditions. Science 186:53-55.

Pretty, J. N., C. F. Mason, D. B. Nedwell, R. E. Hine, S. Leaf, and R. Dils. 2002. Environmental costs of freshwater eutrophication in England and Wales. Environmental Science and Technology 37:201-208.

Rabalais, N. N., R. E. Turner, and D. Scavia. 2002. Beyond science into policy: Gulf of Mexico hypoxia and the Mississippi River. BioScience 52:129-142.

Reddy, K. R. and E. M. D'Angelo. 1997. Biogeochemical indicators to evaluate pollutant removal efficiency in constructed wetlands. Water Science and Technology 35:1-10.

Reddy, K. R. and R. D. DeLaune. 2008. Biogeochemistry of wetlands: science and applications. 774 Pages. CRC Press, Boca Raton, FL.

UNESCO. 2011. Sediment issues and sediment management in large river basins interim case study synthesis report CN/2011/SC/IHP/PI/2. International Sediment Initiative. http://www.irtces.org/isi/isi_document/2011/ISI_Synthesis_Report2011.pdf. Accessed May 25, 2012.

6

USACE. 2011. The Mississippi Drainage Basin. Army Corps of Engineers New Orleans District. http://www.mvn.usace.army.mil/bcarre/missdrainage.asp. Accessed Feb. 13, 2012.

Wang, J. and H. K. Pant. 2010. Enzymatic hydrolysis of organic phosphorus in river bed sediments. Ecological Engineering 36:963-968.

7

CHAPTER II

EFFECTS OF INUNDATION DURATION ON PHOSPHORUS RETENTION BY

LOWER MISSISSIPPI RIVER ALLUVIAL VALLEY AGRICULTURAL

DRAINAGE DITCH SEDIMENTS

2.1 Introduction

Excess phosphorus (P) inputs from agricultural non-point source pollution

contribute to accelerated eutrophication in the Mississippi River drainage basin and Gulf

of Mexico (Lund 1972, Foy and Withers 1995, Rabalais et al. 2002, Khan and Ansari

2005, USACE 2011). Plant and microorganism assimilation is a temporary solution for P

management, as biological uptake only sequesters P for the life of the tissue, and death or

senescence releases P back into the system. Adsorption to mineral surfaces, precipitation

with ions, and sediment burial are long-term solutions to P sequestration (Nguyen and

Sukias 2002, Smith et al. 2005, Kröger et al. 2012). Known mechanisms that regulate P

solubility in sediments are reduced (anaerobic) conditions, pH, clay and organic matter

content, time, and temperature (Patrick and Khalid 1974, Reddy and D'Angelo 1997,

Reddy and DeLaune 2008, Lu et al. 2009, Wang and Pant 2010).

All these factors, except clay content and time, are regulated directly or indirectly

as a function of hydrology. Controlled drainage with slotted or drop pipes, flash board

risers, and vegetated drainage ditches use hydrologic manipulation to manage nutrients

(Kröger et al. 2008a, Kröger et al. 2008b). Weirs are now being implemented as in-ditch

tools for controlled surface drainage. Installed to reduce water velocity, weirs slow

8

transport of runoff, thereby allowing sediments and P time to settle and sorb to soils

while maintaining the original drainage function of the ditch during storm events (Reed

1992, Jordan et al. 2003). Due to the relatively new application of weirs in Lower

Mississippi River Alluvial Valley (LMAV) drainage ditches, weirs’ impacts on drainage

ditch systems over time are unknown. Traditionally, agricultural drainage ditches are

ephemeral systems. In ditches equipped with weirs, after the rain stops and the fields

drain, this added physical barrier pools water behind it, keeping ditch sediments

inundated longer.

The increase in hydraulic residence time potentially alters physicochemical

factors that regulate P retention, which may cause sediments to change from a P sink to

source (Richardson 1985, Schramm et al. 2009). Wright et al. (2001) predicted

inundation would increase P availability due to flooding-induced solubilization of

phosphate minerals. They conducted a microcosm experiment of floodplain soils that

were continuously inundated, flooded and drained once, repeatedly flooded and drained,

or never inundated soils that acted as a control. Results showed P availability was

significantly greater in all flooded versus control soils, but attributed it to microbial

activity, as they did not observe any differences in iron and aluminum fractions, and

suggest the mechanism of release from iron and aluminum phosphates to be more

applicable in wetland soils. Olila et al. (1997) investigated influence of flooding and

draining on bioavailable P in wetland soil. They, too, found that inundating previously

drained soil rapidly released P to the overlying water column. However, fluctuations in P

were significantly less in soil that was continuously inundated compared to soil that

alternated between drained and re-flooded. Their results suggest that P fractions in newly

9

accreted soil are more susceptible to release under alternating flooding and drying

conditions.

Drainage ditch sediments can be described as continuously accreting and

subjected to wet/dry oscillations. Therefore, converting ephemerally inundated ditch

sediments to continuously inundated sediments, based on results by Olila et al. (1997),

could be a viable P management strategy in agricultural land use areas. This study aims

to understand, at a broad spatial scale, effects of inundation on physicochemical factors

and P fractions of sediments within agricultural surface drainage systems. To simulate

increased hydraulic residence time created by weirs, a laboratory experiment was

conducted in microcosm chambers on samples collected throughout the LMAV.

2.2 Methods

2.2.1 Study sites and sediment collection

Sixteen agricultural sites without influence of controlled drainage within the

LMAV were sampled March 13-15, 2010 (Figure 2.1). Three agricultural drainage

ditches were sampled from each site (total n = 48). Sediment was collected from three

locations along each ditch (Figure 2.1) and homogenized within a zip seal plastic bag,

then placed on ice for transport back to the laboratory. A sample site photo was taken

and drainage ditch dimensions, sampling time, vegetation (if present), location

coordinates, water presence, and crop cover were recorded.

2.2.2 Experimental set up

Forty-eight sediment samples were distributed into chambers and subjected to 3

day, 3 week, or 3 month inundation treatments (Figure 2.2). Sub samples of sediments

were withheld from any inundation and regarded as “no inundation treatment” for

10

comparison. All sediments were subjected to all treatments. Deionized water was used

to initially inundate sediments and replenish loss from evaporation. Sediments were then

kept inundated during the treatment. pH of the overlying water column and sediment-

water interface was recorded with a Mettler Toledo S47 SevenMulti™ pH meter. pH

measurements were taken hourly the first four to seven hours after inundation. pH

measurements tapered off to daily and then every other day for longer inundation

treatments as pH began to equilibrate. At the end of the appropriate treatment duration, a

platinum electrode that had been in place for the length of the experiment recorded redox

potential (Eh) of sediments to determine anaerobic severity, and approximately 25 g of

sediment was collected for further analysis.

2.2.3 Sediment analysis

Soil particle size, organic matter percentage, total phosphorus (TP), and P

fractions were determined on sediments collected after each inundation treatment. Moist

sediments were held at 4˚C to minimize microbial activity until preservation by oven

drying and crushing for long term storage before analysis.

Soil particle size analysis was performed at the United States Department of

Agriculture (USDA) National Sedimentation Laboratory in Oxford, Mississippi. Moist,

well mixed, uncrushed sediment samples were run through a Horiba Scientific LA-910

particle size analyzer. Percentage sand (>50 to <1020µm), silt (>2 to <50µm), and clay

(<2µm) were determined then translated to soil texture using the soil pyramid (Sylvia et

al. 2004). Organic matter content was determined by the Mississippi State University

Soil Testing Extension Service with wet combustion with colorimetry (DeBolt 1974).

11

Total P was extracted using the sulfuric acid/hydrogen peroxide/hydrofluoric acid

digestion method determined by Bowman (1988). In a fluoropolymer beaker, sulfuric

acid, hydrogen peroxide, and hydrofluoric acid were added in succession to 0.5 g of dried

and finely ground soil then heated on a hot plate to approximately 150oC for 12 minutes.

Cooled, filtered samples were then brought to 50 mL volume and pH was adjusted to fall

within the range of 3.8-6.0 using concentrated sodium hydroxide.

Phosphorus fractionation was performed to determine concentrations of four

forms of P in each soil sample using the modified method proposed by Chang and

Jackson (1957). Sequentially, soluble and loosely bound P was first extracted from soil

with ammonium chloride. Next, aluminum phosphate (AlPO4) was extracted with

ammonium fluoride. Then iron phosphate (FePO4) was extracted with sodium hydroxide,

followed finally by reductant-soluble P which was extracted with sodium citrate, sodium

bicarbonate, and sodium dithionate in a hot water bath.

Total P and P fraction concentrations were then determined on filtered samples

using the procedure of Watanabe and Olsen (1965). Absorbance was measured at 880

nm on a Hach® DR/4000 U Spectrophotometer. A standard curve was made using

standards of 0.0, 0.01, 0.05, 0.10, 0.20, 0.40, 0.60, 0.80, 1.00, 1.50, 2.00, 3.00, and 5.00

µg/mL. P concentration of the original sample was then determined using the formula: μ / μ /

(2.1)

2.2.4 Statistical analysis

Statistical analysis was run on JMP® (Copyright © 2008 SAS Institute Inc.)

software. Data sets were tested for normality with a Shapiro-Wilk test. If data were

distributed normally, a Levene Homogeneity of Variance test was run. A one-way

12

analysis of variance (ANOVA) was run on data with equal variance. A Welch ANOVA

was used in the case of significantly different variances. A Tukey- Kramer HSD

(honestly significant difference) means comparisons test was used to determine

significantly different means. If data were not distributed normally, a Kruskal-Wallis

rank sums test was used to determine if data were distributed centrally. A Wilcoxon

pairwise comparison test was then used to determine which treatment means differed.

For all data analysis, α = 0.05.

2.3 Results

2.3.1 Abiotic controls on P

Soil texture of study sites was dominated by silt loam, which was expected from

agricultural land use areas located in the LMAV (Table 2.1). Organic matter percentage

averages ranged from 1.83% to 2.15%, and did not differ significantly among treatments

(F = 0.56, P = 0.64, Figure 2.3). Figure 2.4 describes rate of change of water and

sediment pH through time for three inundation treatments. Overlying water column pH

tends to be greater than sediment pH in all three treatments through time. Sediment pH

in 3 week inundation treatments and water and sediment pH in 3 month inundation

treatments significantly increase through time. Comparison of final average water

column pH demonstrates that with longer inundation pH increases (Figure 2.5A). Water

column pH in the two longest inundation treatments (7.07 for 3 week and 7.23 for 3

month) were significantly greater (χ2 = 77.33, P < 0.01) than the two briefest (6.36 for no

treatment and 6.61 for 3 day). Sediment pH, similarly, shows longer inundation led to an

increase in pH (Figure 2.5B). Sediment pH rose by 0.33 from no inundation to 3 days of

inundation and another 0.43 from 3 days to 3 weeks of inundation before leveling off (χ2

13

= 66.33, P < 0.01). Longer inundation time also led to significantly lesser redox potential

(χ2 = 147.19, P < 0.01, Figure 2.6). A significantly more oxidized environment under the

no inundation treatment (552.38 mV ± 101.90) moved to moderately reduced conditions

after 3 days (155.98 mV ± 170.69) of inundation, and was further significantly reduced

after 3 weeks (-101.34 mV ± 108.70) and 3 months (-147.76 mV ± 56.25) of inundation.

2.3.2 Phosphorus

Total P concentration in sediments inundated for 3 days were significantly greater

than the longest inundated sediments (3 months) and no inundation (χ2 = 10.29, P = 0.02,

Figure 2.7). Total P from sediments that were inundated for the intermediate amount of

time (3 weeks) did not differ significantly from TP in any other treatment (χ2 = 10.29, P =

0.02).

There was a significant incremental increase in average soluble and loosely bound

P (χ2 = 32.49, P < 0.01) with non-inundated sediments having the least (9.58 µg/g) and 3

week inundated sediments having the most (65.1 8 µg/g) (Figure 2.8A). However,

sediments inundated for 3 months did not differ significantly from 3 weeks or 3 days (χ2

= 32.49, P < 0.01, Figure 2.8A). Sediments treated with no inundation had significantly

more AlPO4 than other inundation treatments (χ2 = 44.45, P < 0.01, Figure 2.8B).

Average AlPO4 concentrations did not differ significantly among 3 day, 3 week, or 3

month inundation treatments (χ2 = 44.45, P < 0.01, Figure 2.8B). Three day inundation

treatment sediments had the least concentration of FePO4 (χ2 = 20.31, P < 0.01, Figure

2.8C). There was no significant difference in concentrations of FePO4 among no

inundation, 3 week, and 3 month inundation treatments (χ2 = 20.31, P < 0.01, Figure

2.8C). Reductant-soluble P concentrations were greatest in no inundation and 3 month

14

inundation treatment sediments (χ2 = 19.05, P < 0.01, Figure 2.8D). Reductant-soluble P

concentrations were least in 3 day and 3 week inundation treatment sediments.

2.4 Discussion

Increase in inundation time did alter physicochemical factors that regulate P

retention. Longer inundation times had expected effects on pH and Eh and some

unexpected outcomes in effects on P fractions. As expected, brief inundation meant

lower pH. Availability of P depends indirectly on pH, and at lesser pH forms non-

bioavailable compounds with iron and aluminum oxides and hydroxides through specific

adsorption (Bohn et al. 1979). Conversely, P is most bioavailable in a more neutral pH

range of 6 to 7 (Lindsay and Moreno 1960, Bohn et al. 1979, Lindsay 1979). Redox

potential, as well as pH, indirectly dictates bioavailability of P, and like pH, had an

expected response to inundation. Longer inundation led to a decrease in Eh. In aerobic

soils, phosphate precipitates as iron(III) and aluminum phosphates (Ponnamperuma 1972,

Bohn et al. 1979). Under inundated, anaerobic conditions (ranging from 300-100 mV),

ferric iron in FePO4 is reduced to soluble ferrous iron causing PO43- to become soluble

(Patrick and Khalid 1974, Bohn et al. 1979).

In the no inundation treatments, average sediment pH was below the previously

mentioned neutral range and Eh was above 500 mV which reflected a smaller amount of

soluble and loosely bound P and a greater amount of AlPO4 and FePO4 (refer to Figure

2.5, Figure 2.6, and Figure 2.8, respectively). As expected, after a brief inundation of

aerobic and relatively acidic sediments, average sediment pH rose above 6 and redox

dropped below 200 mV. This change corresponded with a rise in soluble P, a drop in

AlPO4 and FePO4, and low reductant-soluble P. Average sediment pH was even greater

15

and Eh least in 3 month inundated sediments; however, FePO4 and reductant-soluble P

concentrations were as high as no inundation sediments. In fact, FePO4 concentrations

were back to no inundation levels after only three weeks of inundation.

This apparent discrepancy can be explained by two stage phosphorus-layer

silicate adsorption reactions which consists of an initial rapid stage followed by a slower

stage (Bohn et al. 1979). The initial rapid stage is completed in about one day and is a

combination of nonspecific adsorption and ligand exchange on mineral edges (Bohn et al.

1979). Once all potential sorption sites are occupied, P adsorption reaches saturation,

after which, dissolution and precipitation reactions continue to retain P (Reddy and

DeLaune 2008). These more complex dissolution and precipitation reactions compose

the slower stage which continues for several weeks or longer (Bohn et al. 1979).

Several previous studies have demonstrated the importance of time for P retention

by soil. Low and Black (1950) showed that P retention increases with time by digesting

kaolinite in phosphate solutions and then analyzing the clay for fixed P. Lu et al. (2009)

found a significant positive correlation between hydraulic residence time and overall P

removal in the field by measuring TP removal in a surface water wetland receiving

agricultural runoff. Kröger et al (2012) went a step further, showing sediments exposed

to briefer hydraulic residence times had greater concentrations of bioavailable P

compared to sediments exposed to longer hydraulic residence times.

If sediments have the ability to alternate between being P sources and sinks, and

hydraulic residence time determines the outcome, then installing controlled drainage

structures such as weirs could improve retention of P leaving agricultural landscapes in

runoff (Kröger et al. 2008b). In this study, under conditions where drainage ditch

sediments were quickly inundated and drained, sediments displayed characteristics of

16

behaving like P sources and would be detrimental in the pursuit for P reduction to

downstream waters. In ditches equipped with weirs, sediments would be inundated longer

and more consistently. With weirs holding a steady hydraulic residence time, they could

influence sediments to act as P sinks even if they alter environmental conditions known

to release P and can therefore be utilized as an effective nutrient management tool.

Ideally, based on these results, weirs should be designed to inundate sediments for three

months or longer. These results support using weirs to manipulate agricultural drainage

ditch hydrology to create longer inundation of sediments to lessen P contribution to

eutrophication of receiving waters.

17

Tab

le 2

.1

Cro

ps a

nd s

oil t

extu

re a

t sam

ple

site

s th

roug

hout

the

Low

er M

issi

ssip

pi R

iver

All

uvia

l Val

ley

(LM

AV

), s

prin

g, 2

010.

Sam

ple

Site

Ave

rage

%

San

d

(>50

to

< 10

20 µ

m)

Ave

rage

%

Silt

(>

2 to

<

50 µ

m)

Ave

rage

%

Cla

y (<

2 µm

) So

il T

extu

re

Cro

p

Por

tage

vill

e, M

O

0.05

73

.36

26.5

9 S

ilt L

oam

C

otto

n, C

orn

Suc

cess

, AR

0.

66

73.8

8 25

.46

Sil

t Loa

m/S

ilty

Cla

y L

oam

R

ice

Judd

Hil

l, A

R

4.81

82

.55

12.6

4 S

ilt L

oam

/Sil

t C

otto

n T

unic

a, M

S 4.

30

77.1

0 18

.59

Sil

t Loa

m

Cot

ton

Har

ris

Bay

ou, M

S

0.56

76

.63

22.8

1 S

ilt L

oam

/Sil

ty C

lay

Loa

m

Fal

low

Tal

laha

chee

Cou

nty,

MS

1.

36

73.1

8 25

.46

Sil

t Loa

m/S

ilty

Cla

y L

oam

C

orn,

Soy

bean

s P

orte

rs B

ayou

, MS

0.

18

74.1

5 25

.67

Sil

t Loa

m/S

ilty

Cla

y L

oam

R

ice,

Soy

bean

s S

tone

vill

e A

quac

ultu

re, M

S

0.00

65

.95

34.0

6 S

ilty

Cla

y L

oam

C

atfi

sh

Bea

sley

, MS

1.24

78

.60

20.1

5 S

ilt L

oam

/Sil

ty C

lay

Loa

m

Soy

bean

s B

ee L

ake,

MS

3.

70

84.8

6 11

.44

Sil

t Loa

m/S

ilt

Cor

n W

ashi

ngto

n C

ount

y, M

S

0.13

75

.89

23.9

8 S

ilt L

oam

C

orn

Wol

f L

ake,

MS

3.

61

75.9

6 20

.44

Sil

t Loa

m/S

ilty

Cla

y L

oam

C

otto

n, C

orn,

Soy

bean

s

18

Figure 2.1 Location of sampling sites along the Lower Mississippi River Alluvial Valley (LMAV) and schematic showing sampling design and field replication. Samples collected March 13-15, 2010.

19

Figure 2.2 Microcosm chambers with collected Lower Mississippi River Alluvial Valley (LMAV) ditch sediments for laboratory experiment at South Farm Aquaculture Facility, Mississippi State University, 2010.

20

Figure 2.3 Average percentage of organic matter in drainage ditch sediments collected from 16 sites (n = 48) throughout the Lower Mississippi River Alluvial Valley (LMAV) March 13-15, 2010, and subjected to four inundation treatments in microcosm chambers through summer of 2010 (none = no inundation). Similar letters denote no significant difference (F = 0.56, P = 0.64). Standard deviation bars shown.

21

Figure 2.4 Trends of sediment pH and overlying water pH through time for 3 day inundation treatments (A), 3 week inundation treatments(B), and 3 month inundation treatments (C) of drainage ditch sediments from 16 sites (n = 48) throughout the Lower Mississippi River Alluvial Valley (LMAV) in 2010.

22

Figure 2.5 Average final pH of overlying water column (χ2 = 77.33, P < 0.01) (A) and sediment interface (χ2 = 66.33, P < 0.01) (B) of drainage ditch sediments collected from 16 sites (n=48) throughout the Lower Mississippi River Alluvial Valley (LMAV) and placed within microcosm chambers for four inundation treatments in 2010 (none = no inundation). Different letters denote significant difference. Standard deviation bars shown.

23

Figure 2.6 Average Eh (mV) of drainage ditch sediments collected from 16 sites (n = 48) throughout the Lower Mississippi River Alluvial Valley (LMAV) and placed within microcosm chambers for four inundation treatments in 2010 (none = no inundation). Different letters denote significant difference (χ2 = 147.19, P < 0.01). Standard deviation bars shown.

24

Figure 2.7 Average TP concentrations (µg/g) of drainage ditch sediments collected from 16 sites (n = 48) throughout the Lower Mississippi River Alluvial Valley (LMAV) and placed within microcosm chambers for four inundation treatments in 2010 (none = no inundation). Different letters denote significant difference (χ2 = 10.29, P = 0.02). Standard deviation bars shown.

25

Figure 2.8 Average concentrations (µg/g) of soluble and loosely bound P (χ2 = 32.49, P < 0.01) (A), AlPO4 (χ

2 = 44.45, P < 0.01) (B), FePO4 (χ2 = 20.31, P <

0.01) (C), and reductant-soluble P (χ2 = 19.05, P < 0.01) (D) of drainage ditch sediments collected from 16 sites (n = 48) throughout the Lower Mississippi River Alluvial Valley (LMAV) and placed within microcosm chambers for four inundation treatments in 2010 (none = no inundation). Different letters denote significant difference. Standard deviation bars shown.

26

2.5 References Bohn, H. L., B. L. McNeal, and G. A. O'Connor. 1979. Soil chemistry. 329 Pages. John

Wiley and Sons, Inc., New York, NY.

Bowman, R. A. 1988. A rapid method to determine total phosphorus in soils. Soil Science Society of America Journal 52:1301-1304.

Chang, S. C. and M. L. Jackson. 1957. Fractionation of soil phosphorus. Soil Science 84:133-144.

DeBolt, D. C. 1974. A high sample volume procedure for the colorimetric determination of soil organic matter. Communications in Soil Science and Plant Analysis 5:131-137.

Foy, R. H. and P. J. A. Withers. 1995. The contribution of agricultural phosphorus to eutrophication. The Fertilizer Society 365:1-32.

Jordan, T. E., D. F. Whigham, K. H. Hofmockel, and M. A. Pittek. 2003. Nutrient and sediment removal by a restored wetland receiving agricultural runoff. Journal of Environmental Quality 32:1534 - 1547.

Khan, F. A. and A. A. Ansari. 2005. Eutrophication: an ecological vision. The Botanical Review 71:449-482.

Kröger, R., C. M. Cooper, and M. T. Moore. 2008a. A preliminary study of an alternative controlled drainage strategy in surface drainage ditches: low-grade weirs. Agricultural Water Management 95:678-684.

Kröger, R., M. M. Holland, M. T. Moore, and C. M. Cooper. 2008b. Agricultural drainage ditches mitigate phosphorus loads as a function of hydrological variability. Journal of Environmental Management 37:107-113.


Lindsay, W. L. 1979. Chemical equilibria in soils. 449 Pages. Wiley, New York, NY.

Lindsay, W. L. and E. C. Moreno. 1960. Phosphate phase equilibria in soils. Soil Science Society of America Journal 24:177-182.

Low, P. F. and C. A. Black. 1950. Reactions of phosphate with kaolinite. Soil Science 70:273-290.

27

Lu, S. Y., F. C. Wu, Y. F. Lu, C. S. Xiang, P. Y. Zhang, and C. X. Jin. 2009. Phosphorus removal from agricultural runoff by constructed wetland. Ecological Engineering 35:402-409.

Lund, J. W. G. 1972. Eutrophication. Proceedings of the Royal Society of London 180:371-382.

Nguyen, L. and J. Sukias. 2002. Phosphorus fractions and retention in drainage ditch sediments receiving surface runoff and subsurface drainage from agricultural catchments in the North Island, New Zealand. Agriculture, Ecosystems and Environment 92:49-69.

Olila, O. G., K. R. Reddy, and D. L. Stites. 1997. Influence of draining on soil phosphorus forms and distribution in a constructed wetland. Ecological Engineering 9:157-169.

Patrick, W. H. J. and R. A. Khalid. 1974. Phosphate release and sorption by soils and sediments: effect of aerobic and anaerobic conditions. Science 186:53-55.

Ponnamperuma, F. N. 1972. The chemistry of submerged soils. Pages 29-96 in N. C. Brady, editor. Advances in Agronomy. Academic Press, Inc., New York, NY.

Rabalais, N. N., R. E. Turner, and D. Scavia. 2002. Beyond science into policy: Gulf of Mexico hypoxia and the Mississippi River. BioScience 52:129-142.

Reddy, K. R. and E. M. D'Angelo. 1997. Biogeochemical indicators to evaluate pollutant removal efficiency in constructed wetlands. Water Science and Technology 35:1-10.


Reed, D. J. 1992. Effect of weirs on sediment deposition in Louisiana coastal marshes. Environmental Management 6:55-65.

Richardson, C. J. 1985. Mechanisms controlling phosphorus retention capacity in freshwater wetlands. Science 228:1424-1427.

Schramm, H. L., M. S. Cox, T. E. Tietjen, and A. W. Ezell. 2009. Nutrient dynamics in the Lower Mississippi River Floodplain: comparing present and historic hydrologic conditions. Wetlands 29:476-487.

Smith, D. R., B. E. Haggard, E. A. Warnemuende, and C. Huang. 2005. Sediment phosphorus dynamics for three tile fed drainage ditches in Northeast Indiana. Agricultural Water Management 71:19-32.

28

Sylvia, D. M., J. J. Fuhrmann, P. G. Hartel, and D. A. Zuberer. 2004. Principles and applications of soil microbiology. 672 Pages 2nd edition. Prentice Hall, Upper Saddle River, NJ.

USACE. 2011. The Mississippi Drainage Basin. Army Corps of Engineers New Orleans District. http://www.mvn.usace.army.mil/bcarre/missdrainage.asp. Accessed Feb. 13, 2012.

Wang, J. and H. K. Pant. 2010. Enzymatic hydrolysis of organic phosphorus in river bed sediments. Ecological Engineering 36:963-968.

Watanabe, F. S. and S. R. Olsen. 1965. Test of an ascorbic acid method for determining phosphorus in water and NaHCO3 extracts from soil. Soil Science Society of America Journal 29:677-678.

Wright, R. B., B. G. Lockaby, and M. R. Walbridge. 2001. Phosphorus availability in an artificially flooded southeastern floodplain forest soil. Soil Science Society of America Journal 65:1293-1302.

29

CHAPTER III

PRELIMINARY EVIDENCE OF SEDIMENT AND PHOSPHORUS DYNAMICS

BEHIND NEWLY INSTALLED LOW GRADE WEIRS IN

AGRICULTURAL DRAINAGE DITCHES

3.1 Introduction

Agricultural drainage ditches are conduits of nutrient effluent. Phosphorus (P)

carried by surface runoff through ditches, contributes to downstream eutrophication

(Dunne et al. 2007, Gordon et al. 2008, Blann et al. 2009). Sediment is also a major non-

point source pollutant that contributes to habitat degradation and acts as a transporter of

particulate P (Sharpley 1980, House 1995, Waters 1995, Uusitalo et al. 2001, Zaimes et

al. 2004, Blann et al. 2009). In fact, Uusitalo et al. (2001) report that particulate P

comprises 92% of P found in runoff. To address this problem in the Mississippi River

Basin, the United States Department of Agriculture (USDA) Natural Resources

Conservation Service (NRCS) created the Mississippi River Basin Healthy Watersheds

Initiative (MRBI) to help implement practices that avoid, control, and trap nutrient runoff

(MRBI 2012). Weirs implemented in agricultural drainage ditches are seen as a potential

tool to help achieve this goal. The intention is for weirs to reduce sediments and

nutrients like P from entering primary aquatic systems by altering ditch hydrology.

Slowing water flow through drainage ditches allows sediments time to settle out of the

water column and P in solution to sorb to sediments and ditch soil (Bowmer et al. 1994,

Jiang et al. 2007).

30

P retention by soils through sorption, precipitation, and sedimentation, is a long

term P storage mechanism (Craft 1996, Hoffmann et al. 2009). P retention by soil is

regulated by iron and aluminum oxides, which are affected by pH which in turn is

affected by hydraulic residence time (Bohn et al. 1979, Boström et al. 1988, Dunne et al.

2007, Reddy and DeLaune 2008). Straight and deep channel designs of surface drainage

ditches leads to shortened hydraulic residence time and also affects channel morphology,

sediment dynamics, and nutrient cycling (Blann et al. 2009).

Presence of a weir can counteract effects of channelized ditches by increasing

hydraulic residence time and slowing down velocity of runoff. Slowing flow velocities

within ditch channels promotes sediment deposition and reduces bank erosion (Kuhnle et

al. 1998). This is a physical way to limit transport of sediment and particulate P, but rates

of sediment accumulation behind weirs in agricultural drainage ditches are unknown.

Increasing hydraulic residence time also allows more time for dissolved P in solution to

interact with bed sediments, which has been shown to decrease bioavailability of P

(Fisher and Acreman 2004, Kröger et al. 2012). However, inundation also alters

sediment pH, possibly creating conditions that increase phosphate release from

sediments in a system designed to accumulate particulate P (Reddy and DeLaune 2008,

Schramm et al. 2009).

Due to recent novel use of weirs in drainage ditches, there are many uncertainties

regarding not only what ways weirs change the ditch system and how that translates to P

and sediment responses, but also how quickly those changes occur after installation.

Weirs have the potential to convert ephemerally inundated sediments into perennially

saturated sediments. Because it is unclear how quickly weirs begin to affect their

surroundings, this study tracked changes in water and sediment depth, as well as P

31

fractions behind weirs starting at time of installation. This will show whether weirs

influence hydrology, sediment accumulation, and P fractions at all, and whether those

changes occur immediately or there is a lag period.

3.2 Methods


The study site was located in the east ditch of Stovall Farms associated with

Harris Bayou in Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) region

(Figure 3.1). Installation of weirs at three locations within the ditch was funded through

the Environmental Protection Agency’s (EPA) Clean Water Act Section 319 Nonpoint

Source Management Program, administered in Mississippi by the Mississippi Department

of Environmental Quality (MDEQ). Additional funding came from the MRBI. Weirs

were built with limestone rip rap placed at intervals along a two stage ditch (Figure 3.2).

Precipitation data was recorded at a United States Geological Survey (USGS) monitoring

station within the study site and was accessed through the USGS Water Resources web

site (USGS 2012). Starting at time of installation and then once a month for six months,

drainage ditch sediment samples were collected directly upstream of each of three weir

sites and at an inflow site located upstream from all three weirs. Each location was

spaced at a distance considered uninfluenced by other weirs. This was determined based

on ditch slope and weir height (Figure 3.3). A spade was used to extract the first 15-

30cm of surface sediment. Sediments were transported in 50 mL centrifuge tubes on ice

and stored at 4˚C until preservation by oven drying and crushing for long term storage in

Whirl-Paks® before analysis. Sediments were collected for analysis from January 2011,

one month after weirs were installed, through June 2011.

32

3.2.2 Sediment accumulation and water depth

Sediment accumulation was measured using a stick of rebar driven into the center

of the ditch upstream of the weir as a permanent reference marker. This was done by

holding a ruler at the top of the sediment and measuring to the top of the rebar (Figure

3.4). Water depth, as documentation of weirs altering ditch system hydrology, was

measured using a similar permanent reference marker technique (Figure 3.4). Sediment

accumulation and water depth were recorded from December 2010, when weirs were

installed, until November 2011, one year later.


Field, laboratory, and instrument duplicates were run on all analyses for quality

assurance and control. Soil particle size analysis was performed on moist, well mixed,

uncrushed sediment samples at the United States Department of Agriculture (USDA)

National Sedimentation Laboratory in Oxford, Mississippi on a Horiba Scientific LA-910

particle size analyzer. Percentage sand (>50 to <1020µm), silt (>2 to <50µm), and clay

(<2µm) were determined then translated to soil texture using the soil pyramid (Sylvia et

al. 2004). The Mississippi State University Soil Testing Extension Service determined

sediment sample pH and used the DeBolt (1974) procedure to determine percentage

organic matter.

The sulfuric acid/hydrogen peroxide/hydrofluoric acid digestion method

established by Bowman (1988) was used to determine total phosphorus (TP). Sulfuric

acid, hydrogen peroxide, and hydrofluoric acid were added in succession to 0.5 g of dried

and finely ground soil within a fluoropolymer beaker. Contents were then heated on a

hot plate to approximately 150oC for 12 minutes, after which, samples were cooled and

33

filtered. Samples were then brought to 50 mL volume and using concentrated sodium

hydroxide pH was adjusted to fall within the range of 3.8-6.0.

A modified P fractionation method first proposed by Chang and Jackson (1957)

was performed to determine concentrations of four forms of P in each soil sample.

Soluble and loosely bound P was extracted from soil first with ammonium chloride,

followed by aluminum phosphate (AlPO4) which was extracted with ammonium fluoride.

Next, iron phosphate (FePO4) was extracted with sodium hydroxide. Reductant-soluble P

was extracted last with sodium citrate, sodium bicarbonate, and sodium dithionate in a

hot water bath.

Total P and P fraction concentrations were then determined on filtered samples

using the procedure of Watanabe and Olsen (1965). Absorbance was measured at 880

nm on a Hach® DR/4000 U Spectrophotometer using a standard curve made from

standards of 0.00, 0.01, 0.05, 0.10, 0.20, 0.40, 0.60, 0.80, 1.00, 1.50, 2.00, 3.00, and 5.00

µg/mL. P concentration of the original sample was then determined using the formula: μ / μ /

(3.1)

For purposes of comparison, P fractions were classified as potentially bioavailable

(soluble and loosely bound P, and FePO4) and non-bioavailable (AlPO4, and reductant-

soluble P). Potentially bioavailable and non-bioavailable concentrations were then

translated into bioavailability ratios with large ratio numbers representing more

potentially bioavailable P and less non-bioavailable P. Conversely, small ratio numbers

represent less potentially bioavailable P and more non-bioavailable P. Therefore, in the

interest of nutrient reduction management, smaller ratio numbers are preferred in

drainage ditch sediments.

34


Statistical analysis was run on JMP® (Copyright © 2008 SAS Institute Inc.)

software with α = 0.10. To test overall differences among sites based spatially along the

ditch, the data set was organized by site and tested for normality with a Shapiro-Wilk

test. Normally distributed data was tested for homogeneity of variance with a Levene

test. A oneway analysis of variance (ANOVA) was run on data with equal variance, and

a Welch ANOVA was used for data with unequal variance. A Tukey-Kramer HSD all

pairs means comparison test was used to determine and rank significantly different

means. If data were not distributed normally, a Kruskal-Wallis rank sums test

determined if means differed, if they did, a Wilcoxon pairwise comparison test was then

used to determine and rank significantly different means. Spatial and temporal

correlations for precipitation, sediment depth, water depth, organic matter percentage,

pH, TP, and bioavailability ratio variables for each site were determined with bivariate

least squares linear regression. To find correlations among variables by location,

pairwise multivariate analysis was run on precipitation, sediment depth, water depth,

organic matter percentage, pH, TP, and bioavailability ratio.

3.3 Results


Average sediment deposition did not differ significantly among sites (χ2 = 2.42, P

= 0.49, Figure 3.5). Weir sites had significantly greater average water depths, and had

between three to five times greater water depth than the inflow site (χ2 = 7.67, P = 0.05,

Figure 3.6).

35


Soil texture at the site is silt loam (Sand 0.48%; Silt 81.99%; Clay 17.53%).

Average organic matter percentages ranged between 1% and 2%. Inflow site had greatest

average organic matter percentage (1.67% ± 0.16) and weirs 2 and 3 had the least (χ2 =

6.14, P = 0.09, Figure 3.7). Overall average pH increased significantly from inflow (5.50

± 0.51) to weir 1 (6.13 ± 0.19) to weir 2 (6.80 ± 0.11). Overall average pH at weir 3

(6.48 ± 0.56) did not differ significantly from weirs 1 and 2 (χ2 = 25.05, P < 0.01, Figure

3.8).

3.3.3 Total P and P fractions

Average TP ranged from 1147.4 µg/g ± 517.37 to 1815.1 µg/g ± 716.72 and did

not differ significantly among sites (χ2 = 3.22, P = 0.36, Figure 3.9). There also were no

significant differences in average soluble and loosely bound P (χ2 = 5.89, P = 0.12) and

reductant-soluble P (χ2 = 2.24, P = 0.52) concentrations among sites, however, there were

significant differences in AlPO4 (χ2 = 6.45, P = 0.09) and FePO4 (χ

2 = 6.70, P = 0.08)

concentrations among sites (Figure 3.10). Weir 2 had greatest average concentration of

AlPO4 whereas weir 3 and inflow had the least (Figure 3.10B). Average concentration of

FePO4 was greatest at inflow and least at weir 1 (Figure 3.10C). Average bioavailability

ratios ranged from 1.53 ± 0.90 to 4.73 ± 8.75 but did not differ significantly among sites

(χ2 = 2.09, P = 0.55, Figure 3.11).

3.3.4 Correlations

Four variables were correlated with location in the ditch. Water depth (F = 5.86,

P = 0.02, R2 = 0.13), pH (F = 18.12, P < 0.01, R2 = 0.45), and TP (F = 3.57, P = 0.07, R2

= 0.14) increased moving downstream from inflow to weir 3. Organic matter percentage

36

(F = 5.23, P = 0.03, R2 = 0.19) decreased moving downstream from inflow to weir 3

(Figure 3.12). The only variables correlated significantly with time were found at the

inflow site. Water depth (F = 3.99, P = 0.08, R2 =0.33) decreased over time (Figure

3.13). pH (F = 12.93, P = 0.02, R2 = 0.76) and bioavailability ratio (F = 44.14, P < 0.01,

R2 = 0.92) increased over time (Figure 3.13). Organic matter percentage and pH were

correlated negatively at inflow (F = 11.80, P = 0.03, R2 = 0.75, Figure 3.13A). At weir 3

TP and pH were correlated positively (F = 6.74, P = 0.06, R2 = 0.63, Figure 3.14B).

Bioavailability ratio was correlated positively with pH at inflow (F = 6.50, P = 0.06, R2 =

0.62, Figure 3.15A). At weir 1, bioavailability was correlated negatively with sediment

depth (F = 9.14, P = 0.04, R2 = 0.70, Figure 3.15B). At weir 3, bioavailability was

correlated negatively with organic matter percentage (F = 23.10, P = 0.01, R2 = 0.85,

Figure 3.15C).

3.4 Discussion

Several studies have suggested there is seasonality to P and sediment loss relating

to a wet/dry cycle (Zhang et al. 2004, Puustinen et al. 2007, Surridge et al. 2007).

Sediments subjected to hydrologic fluctuation can act as substantial P sources. Loeb et

al. (2008) showed that inundating floodplain soils increases concentrations of P in soil

pore water 2-90 times. Baldwin and Mitchell (2000) also report that re-wetting

desiccated soils and sediments will result in an initial flush of available P. Alternatively,

they say the act of drying previously inundated sediments will reduce P availability.

Schramm et al. (2009) stated the mechanism for this is that under aerobic conditions

FePO4 complexes are non-bioavailable, but become reduced to ferrous iron under

anaerobic conditions created through inundation which releases bioavailable P quickly

37

and persistently throughout the period of inundation. This information suggests that due

to their very nature of rapid and frequent wetting and drying cycles, drainage ditch

sediments pose a detriment to P management. This study demonstrated that weirs sustain

greater water depths, compared to the inflow location which had lower water depth on

average, and declining water depth over time. Although presence of a weir did lead to

longer inundation of sediments, it did not translate into a rise in bioavailability ratios. By

converting inconsistently inundated sediments into more reliably saturated sediments,

weirs had no variables correlated with time. Comparatively, the inflow which remained

in traditional state of flux, revealed a rising bioavailability ratio in the first six months of

observations.

There are other studies that reflect this apparent contradiction to the

aforementioned literature. Chacon et al. (2006) acknowledge that reduction of iron

oxides contributes to seasonally flooded sediments releasing P, but also takes into

consideration secondary reactions involving ferrous iron which can lead to decreases in

soluble P concentrations. Additionally, Lucassen et al. (2005) says that in seepage zone

sediment, where reduced iron levels are high, phosphate is effectively immobilized under

oxidized and reduced conditions. Watts (2000) even recommends maintaining high water

levels in reservoirs to prevent desiccation of littoral sediments for P retention.

Not only was the outcome of inundation effects on P unexpected, but so was the

outcome of pH effects on P. A rise in pH is expected to release bioavailable P due to

displacement of inorganic P from oxides and hydroxides of iron and aluminum (Andersen

1975, Koski-Vähälä et al. 2001, Reddy and DeLaune 2008). However, average

bioavailability ratios behind all three weirs were no different than at the inflow even

though sediment pH behind all three weirs was significantly greater than at the inflow.

38

Averages of water depth and sediment pH were greater behind weirs. These

variables also increased spatially moving down the ditch. Kröger et al. (2011) suggest

one of the benefits of using weirs is that they can be placed at multiple intervals along the

stretch of ditch in a stepwise fashion, allowing for sequential accumulation going

downstream. With soil P found at significant spatial variations in drainage ditches by

Vaughan at al. (2007), placing weirs at multiple locations could also increase the

probability of intercepting non point source pollutants leaving the landscape in irregular

patches. Although there was no difference in overall averages, as there was with water

depth and sediment pH, TP was another variable that increased spatially from inflow to

weir 3. Of the total TP measured along the ditch throughout the study, 20% was found

behind weir 1, 25% behind weir 2, and 33% behind weir 3.

While it appears weirs are an effective way to manage P, sediment was the other

non-point source pollutant examined in this study. Sediment accumulation did not

correlate with increased water depth behind weirs, nor did it vary among locations or

time. There are some reasons why sediment accumulation was not detected in this study.

The channel itself was formed as a two-stage ditch, opposed to the more common one-

stage trench design. Also, weirs were comprised of stacked limestone rip rap, unlike

other weirs constructed with a solid inner core, like an earthen berm or concrete. This

study assessed sediment depths from one permanent reference marker positioned in the

center of the channel at each location. Because of the differences with this study,

sediment could be deposited within recesses of the rip rap or on the more densely

vegetated second stage shoulders instead of at the reference marker (Steiger et al. 2001,

Powell et al. 2007, Usborne et al. 2012). Future studies should incorporate additional

39

sediment measurement reference points at varying distances from the weir and ditch

banks to create a more accurate picture of weir contribution to sediment dynamics.

Despite not observing differences in sediment accumulation, it is clear that weirs

affect hydrology and P dynamics quickly after installation. Although the application of

weirs in agricultural drainage ditches is new, this study observed differences in as soon as

six months. Weirs temper the traditional ephemeral nature of agricultural drainage ditch

sediments, prolonging inundation times and allowing P to interact with sediments under

equilibrated conditions (Petry et al. 2002, Kröger et al. 2008, Kröger et al. 2012). With

the exploratory observational data collected from this study, weirs are observed to be

effective tools at managing P. Recommended future investigation should be given to

long term effects of weirs and mechanisms by which they operate.

40

Figure 3.1 Location of study site along within an agricultural drainage ditch associated with Harris Bayou, Mississippi.

41

Figure 3.2 Initial limestone rip rap weir installation in two stage ditch design at East ditch, Harris Bayou, Mississippi, December 2010.

Figure 3.3 Schematic showing how the distance between locations was chosen so as to avoid influences from other weirs. As depicted, the position of another study location occurs past the point where weir height intersects slope.

42

Figure 3.4 Schematic showing sediment accumulation and water depth measurement methods in Harris Bayou ditch site, December 2010 through November 2011.

43

Figure 3.5 Average sediment deposition at four sites within the east ditch of Stovall Farms in Harris Bayou, Mississippi after one year (December 2010 – November 2011) of weir installation (no weir located at inflow location). Similar letters denote no significant differences (χ2 = 2.42, P = 0.49). Standard deviation bars shown.

44

Figure 3.6 Average water depth at four sites within the east ditch of Stovall Farms in Harris Bayou, Mississippi after one year (December 2010 – November 2011) of weir installation (no weir located at inflow location). Different letters denote significant differences (χ2 = 7.67, P = 0.05). Standard deviation bars shown.

45

Figure 3.7 Average percentage organic matter at four sites within the east ditch of Stovall Farms in Harris Bayou, Mississippi after six months (January 2011 – June 2011) of weir installation (no weir located at inflow location). Different letters denote significant differences (χ2 = 6.41, P = 0.09). Standard deviation bars shown.

46

Figure 3.8 Average pH of sediment at four sites within the east ditch of Stovall Farms in Harris Bayou, Mississippi after six months (January 2011 – June 2011) of weir installation (no weir located at inflow location). Different letters denote significant differences (F = 25.05, P < 0.01). Standard deviation bars shown.

47

Figure 3.9 Average TP of sediment at four sites within the east ditch of Stovall Farms in Harris Bayou, Mississippi after six months (January 2011 – June 2011) of weir installation (no weir located at inflow location). Similar letters denote no significant differences (χ2 = 3.22, P = 0.36). Standard deviation bars shown.

48

Figure 3.10 Average soluble and loosely bound P (χ2 = 5.89, P = 0.12) (A), AlPO4 (χ2 =

6.45, P = 0.09) (B), FePO4 (χ2 = 6.70, P = 0.08) (C), and reductant-soluble

P (χ2 = 2.24, P = 0.52) (D) concentrations in sediment sampled at four sites within the east ditch of Stovall Farms in Harris Bayou, Mississippi after six months (January 2011 – June 2011) of weir installation (no weir located at inflow location). Different letters denote significant differences. Standard deviation bars shown.

49

Figure 3.11 Average bioavailability ratios of sediment at four sites within the east ditch of Stovall Farms in Harris Bayou, Mississippi after six months (January 2011 – June 2011) of weir installation (no weir located at inflow location). Similar letters denote no significant differences (χ2 = 2.09, P = 0.55). Standard deviation bars shown.

50

Figure 3.12 Correlations between water depth (F = 5.86, P = 0.02, R2 = 0.13) (A), organic matter percentage (F = 5.23, P = 0.03, R2 = 0.19) (B), pH (F = 18.12, P < 0.01, R2 = 0.45) (C), and TP (F = 3.57, P = 0.07, R2 = 0.14) (D) and location within the east ditch of Stovall Farms in Harris Bayou, Mississippi. Location 0 denotes inflow, 1 denotes weir 1, 2 denotes weir 2, and 3 denotes weir 3.

51

Figure 3.13 Correlations between water depth (F = 3.99, P = 0.08, R2 = 0.33) (A), pH (F = 12.93, P = 0.02, R2 = 0.76) (B), and bioavailability ratio (F = 44.14, P < 0.01, R2 = 0.92) (C) with time at the inflow location within the east ditch of Stovall Farms in Harris Bayou, Mississippi. Time 0 denotes weir installation in December 2010, 1 denotes one month after installation in January 2011, 2 denotes two months after installation in February 2011, etc.

52

Figure 3.14 Organic matter percentage and pH correlation (F = 11.80, P = 0.03, R2 = 0.75) at inflow location (A), and TP and pH correlation (F = 6.74, P = 0.06, R2 = 0.63) at weir 3 location (B) within the east ditch of Stovall Farms in Harris Bayou, Mississippi.

53

Figure 3.15 Correlations between bioavailability ratio and pH (F = 6.50, P = 0.06, R2 = 0.62) at the inflow (A), bioavailability ratio and sediment depth (F = 9.14, P = 0.04, R2 = 0.70) at weir 1 (B), and bioavailability ratio and organic matter (F = 23.10, P = 0.01, R2 = 0.85) at weir 3 (C) within the east ditch of Stovall Farms in Harris Bayou, Mississippi from January 2011 to June 2011.

54

3.5 References Andersen, J. M. 1975. Influence of pH on release of phosphorus from lake sediments.

Archiv für Hydrobiologie 76:411-419.

Baldwin, D. S. and A. M. Mitchell. 2000. The effects of drying and re-flooding on the sediment and soil nutrient dynamics of lowland river–floodplain systems: a synthesis. Regulated Rivers: Research and Management 16:457-467.

Blann, K. L., J. L. Anderson, G. R. Sands, and B. Vondracek. 2009. Effects of agricultural drainage on aquatic ecosystems: a review. Critical Reviews in Environmental Science and Technology 39:909-1001.

Bohn, H. L., B. L. McNeal, and G. A. O'Connor. 1979. Soil chemistry. 329 Pages. John Wiley and Sons, Inc., New York, NY.

Boström, B., J. M. Andersen, S. Fleischer, and M. Jansson. 1988. Exchange of phosphorus across the sediment-water interface. Hydrobiologia 170:229-244.


Bowmer, K. H., M. Bales, and J. Roberts. 1994. Potential use of irrigation drains as wetlands. Water Science and Technology 29:151-158.

Chacon, N., S. Flores, and A. Gonzalez. 2006. Implications of iron solubilization on soil phosphorus release in seasonally flooded forests of the lower Orinoco River, Venezuela. Soil Biology and Biochemistry 38:1494-1499.


Craft, C. B. 1996. Dynamics of nitrogen and phosphorus retention during wetland ecosystem succession. Wetlands Ecology and Management 4:177-187.


Dunne, E. J., K. A. McKee, M. W. Clark, S. Grunwald, and K. R. Reddy. 2007. Phosphorus in agricultural ditch soil and potential implications for water quality. Journal of Soil and Water Conservation 62:244-252.

Fisher, J. and M. C. Acreman. 2004. Wetland nutrient removal: a review of the evidence. Hydrology and Earth System Sciences 8:673-685.

Gordon, L. J., G. D. Peterson, and E. M. Bennett. 2008. Agricultural modifications of hydrological flows create ecological surprises. Trends in Ecology and Evolution 23:211-219.

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Hoffmann, C. C., C. Kjaergaard, J. Uusi-Kämppä, H. C. B. Hansen, and B. Kronvang. 2009. Phosphorus retention in riparian buffers: review of their efficiency. Journal of Environmental Quality 38:1942-1955.

House, W. A., F. H. Denison, and P. D. Armitage. 1995. Camparison of the uptake of inorganic phosphorus to a suspended and stream bed-sediment. Water Resources 29:767-779.

Jiang, C., X. Fan, G. Cui, and Y. Zhang. 2007. Removal of agricultural non-point source pollutants by ditch wetlands: implications for lake eutrophication control. Hydrobiologia 581:319-327.

Koski-Vähälä, J., H. Hartikainen, and P. Tallberg. 2001. Phosphorus mobilization from various sediment pools in response to increased pH and silicate concentration. Journal of Environmental Quality 30:546-552.

Kröger, R., M. M. Holland, M. T. Moore, and C. M. Cooper. 2008. Agricultural drainage ditches mitigate phosphorus loads as a function of hydrological variability. Journal of Environmental Management 37:107-113.

Kröger, R., M. Moore, J. Farris, and M. Gopalan. 2011. Evidence for the use of low-grade weirs in drainage ditches to improve nutrient reductions from agriculture. Water, Air, and Soil Pollution 221:223-234.


Kuhnle, R. A., R. L. Bingner, G. R. Foster, and E. H. Grissinger. 1998. Land use changes and sediment transport in Goodwin Creek. Pages 279-292 in P. C. Klingeman, R. L. Beschta, P. D. Komar, and J. B. Bradley, editors. Gravel-bed rivers in the environment. Water Resources Publications, LLC, Highlands Ranch, CO.

Loeb, R., L. P. M. Lamers, and J. G. M. Roelofs. 2008. Prediction of phosphorus mobilisation in inundated floodplain soils. Environmental Pollution 156:325-331.

Lucassen, E. C. H. E. T., A. J. P. Smolders, and J. G. M. Roelofs. 2005. Effects of temporary desiccation on the mobility of phosphorus and metals in sulphur-rich fens: differential responses of sediments and consequences for water table management. Wetlands Ecology and Management 13:135-148.

MRBI. 2012. Mississippi River Basin Healthy Watersheds Initiative fact sheet. United States Department of Agriculture Natural Resources Conservation Service. http://www.nrcs.usda.gov/Internet/FSE_DOCUMENTS//nrcs143_008142.pdf. Accessed May 23, 2012.

Petry, J., C. Soulsby, I. A. Malcolm, and A. F. Youngson. 2002. Hydrological controls on nutrient concentrations and fluxes in agricultural catchments. Science of The Total Environment 294:95-110.

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Powell, G. E., A. D. Ward, D. E. Mecklenburg, and A. D. Jayakaran. 2007. Two-stage channel systems: Part 1, a practical approach for sizing agricultural ditches. Journal of Soil and Water Conservation 62:277-286.

Puustinen, M., S. Tattari, J. Koskiaho, and J. Linjama. 2007. Influence of seasonal and annual hydrological variations on erosion and phosphorus transport from arable areas in Finland. Soil and Tillage Research 93:44-55.


Schramm, H. L., M. S. Cox, T. E. Tietjen, and A. W. Ezell. 2009. Nutrient dynamics in the Lower Mississippi River Floodplain: comparing present and historic hydrologic conditions. Wetlands 29:476-487.

Sharpley, A. N. 1980. The enrichment of soil phosphorus in runoff sediments. Journal of Environmental Quality 9:521-526.

Steiger, J., A. M. Gurnell, and G. E. Petts. 2001. Sediment deposition along the channel margins of a reach of the middle River Severn, UK. Regulated Rivers: Research and Management 17:443-460.

Surridge, B. W. J., A. L. Heathwaite, and A. J. Baird. 2007. The release of phosphorus to porewater and surface water from river riparian sediments. Journal of Environmental Quality 36:1534-1544.


Usborne, E., K. A. Littlejohn, S. C. Pierce, and R. Kröger. 2012. Nutrient reduction capabilities of agricultural drainage ditch wetlands: creation and policy implications. Pages 185-202 Wetlands: Ecology, Management and Conservation. Nova, New York, NY.

USGS. 2012. Water-resources data for the United States, water year 2011: United States Geological Survey water-data report WDR-US-2011, site 341404090385600. http://nwis.waterdata.usgs.gov/nwis/uv?cb_00045=on&format=html&begin_date=2010-12-01&end_date=2011-11-30&site_no=341404090385600. Accessed Mar. 9, 2012.

Uusitalo, R., E. Turtola, T. Kauppila, and T. Lilja. 2001. Particulate phosphorus and sediment in surface runoff and drainflow from clayey soils. Journal of Environmental Quality 30:589-595.

Vaughan, R. E., B. A. Needelman, P. J. A. Kleinman, and A. L. Allen. 2007. Spatial Variation of Soil Phosphorus within a Drainage Ditch Network. J. Environ. Qual. 36:1096-1104.


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Waters, T. 1995. Sediment in streams: sources, biological effects, and control. 251 Pages. American Fisheries Society, Bethesda, MD.

Watts, C. J. 2000. Seasonal phosphorus release from exposed, re-inundated littoral sediments of two Australian reservoirs. Hydrobiologia 431:27-39.

Zaimes, G. N., R. C. Schultz, and T. M. Isenhart. 2004. Stream bank erosion adjacent to riparian forest buffers, row-crop fields, and continuously-grazed pastures. Journal of Soil and Water Conservation 59:19-27.

Zhang, M., Z. He, D. V. Calvert, and P. J. Stoffella. 2004. Spatial and temporal variations of water quality in drainage ditches within vegetable farms and citrus groves. Agricultural Water Management 65:39-57.

58

CHAPTER IV

PRELIMINARY EVIDENCE OF SEDIMENT AND PHOSPHROUS DYNAMICS

BEHIND LOW GRADE WEIRS OF VARYING AGES INSTALLED

IN AGRICULTURAL DRAINAGE DITCHES

4.1 Introduction

Non-point source pollutants like sediment and P are a concern for watershed

management. Excess P triggers the process of eutrophication. Sediment causes habitat

degradation as well as transports P (Braskerud 2002, UNESCO 2011). These pollutants

leave agricultural landscapes and travel through drainage ditches in runoff. Therefore,

through alteration of the hydrology of drainage ditches, these pollutants can be managed

(Kröger et al. 2012).

Slowing flow of runoff allows sediments time to settle out (Bowmer et al. 1994,

Jiang et al. 2007). Based on this principle, weirs are being installed in drainage ditches as

a controlled surface drainage strategy to manage sediment and P leaving agricultural

landscapes in runoff. It has been shown that relatively small channels are not only close

to sediment sources, but their morphology and hydraulics also are significantly

influenced by large physical obstructions (Hassan et al. 2005). Drainage ditches can be

seen as mimicking headwater streams in agriculturally impacted watersheds, in which

case weirs, acting as an influential large physical obstruction, could have a substantial

impact on sedimentation. As a physical barrier, weirs also cause pooling of water (see

Chapter III, Figure 3Error! Reference source not found..6). An increase in hydraulic

59

residence time would allow P time to interact with ditch sediments (Bowmer et al. 1994,

Jiang et al. 2007).

Sediments have usually been regarded as long term means to sequester P. Mitsch

et al. (1995) and Braskerud (2002) showed sedimentation of particle bound P was the

main mechanism for P retention. But there also are chemical processes of specific

adsorption with iron and aluminum oxides and hydroxides that limit P bioavailability.

Iron and aluminum oxides and hydroxides are affected by pH which in turn is affected by

hydraulic residence time (Boström et al. 1988). Weirs have the potential to dramatically

alter typical hydraulic residence times to which ditch sediments are exposed. As long as

there are available binding sites for P, sediments can still act as a P sink (Reynolds and

Davies 2001). This implies there are a finite number of P binding sites, and those sites

can be reduced as pH approaches neutral. Therefore sediments can become P saturated

when there are no longer any P binding sites available (Maguire and Sims 2002). In fact,

Richardson (1985) found that systems became P saturated in only a few years. However,

after saturation of sorption sites there also are more complex precipitation reactions that

occur at greater concentrations of P and require more time (Bohn et al. 1979, Reddy and

DeLaune 2008). Low and Black (1950), for example, showed that P retention by soil

increased with time and P concentration.

Sharpley et al. (2007) in an indoor fluvarium, Lu et al. (2009) and Kröger et al.

(2012) in wetlands, and Kröger et al. (2011) in drainage ditches at an agricultural

experiment station all found that increasing retention time promotes sediment

accumulation and increases P retention. Although promoting sediment and P retention

through increased retention time is the premise under which weirs are installed, due to

their nascent application within agricultural drainage ditch systems, there are several

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unknowns. First, it is unsure if the findings of the abovementioned studies apply to

agricultural drainage ditch weirs in a functioning field setting, and also if any potential

sediment accumulation or P retention benefits apply to extended use of weirs over many

years.

The rate of sediment accumulation, changes in abiotic factors, and P fraction

behavior in sediment behind weirs in operational agricultural drainage ditches over the

course of several years is unknown. If weirs are expected to become an established

conservation practice, it is important to understand how they perform years after they are

installed. This study first aims to observe if there are differences in sediment

accumulation, abiotic factors such as pH and organic matter percentage, and P fraction

composition between agricultural drainage ditch sediments with or without weirs.

Furthermore, this study evaluates if there are any differences in sediment and P variables

among weirs of increasing age.

4.2 Methods


Nine drainage ditches containing 14 weirs at seven sites, associated with Bee

Lake and Wolf Lake in Mississippi’s Lower Mississippi River Alluvial Valley (LMAV)

region, were evaluated (Figure 4.1). Weirs built with limestone rip rap had been installed

with funds allocated to the Mississippi Department of Environmental Quality as part of

the United States Environmental Protection Agency’s (EPA) Clean Water Act Section

319 Nonpoint Source Management Program. The oldest weirs were in ditches that

emptied into Bee Lake and had been established for 4 years (n = 3). The youngest weirs

were 1 year old and most abundant (n = 9). Intermediate aged weirs were 2 years old and

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were most scarce (n = 2). One and two year old weirs were found in ditches that emptied

into Wolf Lake. Each weir was paired with a control (Figure 4.2). Control sites were

located within an area of the same agricultural drainage ditch uninfluenced by the weir

(influence was determined based on ditch slope and weir height). There was no

indication of any maintenance occurring after weir installation at any of the ditch sites.

Field measurements were recorded and samples were collected directly up stream

of weir sites and control sites. All data was collected during non-storm event conditions

in August 2010 to minimize potential climate variables. A spade was used to extract the

first 15-30 cm of surface sediment. All sediment samples were collected in the center of

the ditch above each weir structure (when present) and placed within plastic 50 mL

centrifuge tubes. Samples were transported on ice and stored at 4˚C until preservation by

oven drying and crushing for long term storage in Whirl-Paks® before analysis.


Water depth, as documentation of weirs altering ditch system hydrology, was

measured using a meter stick. Sediment accumulation was measured using three

techniques. Deposition layer in soil cores was determined with visual and tactile

differences from original ditch bottom soil (Figure 4.3). Differences between sediment

height above and below the weir were determined using a laser held at the top of the weir

and the height measured from the sediment surface (Figure 4.4). A meter stick was

pushed through the deposition layer until reaching compacted original ditch bottom and

depth was recorded. This was the most reliable technique due to vegetation and water

depth constraints and was able to be used at all sites. Where conditions allowed, one of

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the more traditional methods such as soil cores and taking the difference of sediment

height above and below the weir were used to ensure accuracy of the meter stick method.

Weir and ditch dimensions were measured using Topcon Positioning Hiper Ga

Real Time Kinematic GPS equipment (horizontal accuracy of 10 mm and vertical

accuracy of 15 mm) with American GPS satellites as well as Russian GLONASS. The

geometry of the ditch extending behind the influence of the weir was determined with

Topcon software (Copyright Topcon © 2002-2010). Total volume of sediment

accumulated was calculated using dimension measurements and equation for the volume

of a triangular base pyramid (Figure 4.5):

width height (4.1)


Quality assurance and control was ensured by running field, laboratory and

instrument duplicates on all analyses. Soil particle size analysis was performed on the

Horiba Scientific LA-910 particle size analyzer located at the United States Department

of Agriculture (USDA) National Sedimentation Laboratory in Oxford, Mississippi. Soil

texture was determined using results from soil particle size divided into percentage sand

(>50 to <1020 µm), silt (>2 to <50 µm), and clay (<2 µm) and translated onto the soil

pyramid (Sylvia et al. 2004). Analysis of sediment sample pH and percentage organic

matter was performed by The Mississippi State University Soil Testing Extension

Service. The DeBolt (1974) procedure was used to determine percentage organic matter.

Total phosphorus (TP) was extracted with the method proposed by Bowman

(1988). A sample of 0.5 g of dried and finely ground soil was measured and placed

within a fluoropolymer beaker. Sulfuric acid, hydrogen peroxide, and hydrofluoric acid

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were then added in succession. The beaker was then placed on a hot plate heated to

approximately 150oC for 12 minutes. Samples were cooled and filtered. Volume of the

sample was brought to 50 mL with deionized water. Then using concentrated sodium

hydroxide, pH was adjusted to fall within the range of 3.8-6.0. Concentrations of four

forms of P in each soil sample were determined by using a P fractionation procedure

modified from the method first proposed by Chang and Jackson (1957). Each

fractionation was extracted from a sediment sample by adding reagents in succession.

Soluble and loosely bound P was extracted from soil first with ammonium chloride.

Next, aluminum phosphate (AlPO4) was extracted with ammonium fluoride. Then iron

phosphate (FePO4) was extracted with sodium hydroxide, followed finally by reductant-

soluble P which was extracted with sodium citrate, sodium bicarbonate, and sodium

dithionate in a hot water bath. The procedure of Watanabe and Olsen (1965) was then

used to determine concentrations of TP and P frations in filtered samples. Absorbance

was measured at 880 nm on a Hach® DR/4000 U Spectrophotometer. Standards of 0.00,

0.01, 0.05, 0.10, 0.20, 0.40, 0.60, 0.80, 1.00, 1.50, 2.00, 3.00, and 5.00 µg/mL were

prepared and used to make a standard curve. P concentration of the original sample was

then determined using the formula:

μ / μ /

(4.2)

P fractionations were classified as potentially bioavailable (soluble and loosely

bound P, and FePO4) and non-bioavailable (AlPO4, and reductant-soluble P). A

bioavailability ratio was then calculated using determined potentially bioavailable and

non-bioavailable concentrations. A large ratio indicates a greater proportion of

potentially bioavailable P to non-bioavailable P. In contrast, a small ratio indicates a

64

smaller proportion of non-bioavailable P to potentially bioavailable P. From a nutrient

management standpoint it is better to have less potentially bioavailable phosphorus and

that is reflected in a smaller ratio number.


Statistical analyses were run on JMP® (Copyright © 2008 SAS Institute Inc.)

software, with all α = 0.10. Data set distributions were tested for normality with a

Shapiro-Wilk W test. If normal, a Levene Homogeneity of Variance test was run. A

one-way analysis of variance (ANOVA) was run on data with equal variance. A Welch

ANOVA was used in the case of significantly different variances. Tukey-Kramer HSD

all pairs means comparison test was then used to determine and rank significantly

different means. If data were not normal, a Kruskal-Wallis rank sums test determined if

means differed, if they did, a Wilcoxon pairwise comparison test was then used to

determine and rank significantly different means.

Due to differences in weir heights, ditch lengths and widths, and acreages drained,

comparison of weirs and controls across systems cannot be done. This study seeks to

compare weirs of different ages; however, age was only one variable. Weirs were located

in different ditches on different farms at different locations, and built to different

specifications. To eliminate these other variables, each weir was paired with a respective

control (see Study sites and sample collection, section 4.2.1). Results from the control

were subtracted from results from the weir. This relativized variables allowing for

comparisons to be made among weirs based solely on age.

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4.3 Results


There was no significant difference between sediment depths detected using the

meter stick method and the soil core method (F = 0.02, P = 0.89) nor between the meter

stick method and the laser measuring method (S = 55.50, P = 0.75). Weirs had

significantly greater sediment (F = 3.55, P = 0.08) and water depths (S = 50.50, P = 0.02)

than controls (Figure 4.6). One year old weirs had significantly greater average sediment

depths than 2 and 4 year old weirs (χ2 = 7.58, P = 0.02, Figure 4.7A). Sediment depth

was correlated negatively with weir age (F = 5.25, P = 0.05, R2 = 0.37, Figure 4.7B).

Average sediment volume accumulated behind weirs was 73.09 m3 ± 77.14 and ranged

from 2.49 m3 to 217 m3 (Table 4.1). There were no significant differences in average

water depth among weir ages (F = 3.18, P = 0.13, Figure 4.8A); however, water depth

significantly decreases with weir age (F = 6.46, P = 0.04, R2 = 0.52, Figure 4.8B).


Soil texture of sites ranged from silt to silty clay loam, but was mostly silt loam

with 16.75% average clay content (Table 4.2). There were no significant differences in

average sediment pH between weirs and controls (F = 1.44, P = 0.24, Figure 4.9), nor

were there any significant differences in average sediment pH among weir ages (F = 2.74,

P = 0.17, Figure 4.10A) or correlation between sediment pH and weir age (F = 3.32, P =

0.10, Figure 4.10B). There were no significant differences in average organic matter

percentage between weirs and controls (F = 0.30, P = 0.59, Figure 4.11). Similarly there

were no significant differences in average organic matter percentage among weir ages (F

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= 1.03, P = 0.39, Figure 4.12A) or correlation between organic matter percentage and

weir age (F = 0.26, P = 0.62, Figure 4.12B).

4.3.3 Total P and P fractions

There were no significant differences in average TP between weirs and controls

(S = 238.00, P = 0.76, Figure 4.13). There were no significant differences in average TP

among weir ages (F = 1.14, P = 0.38, Figure 4.14A) and no correlation between TP and

weir age (F = 0.99, P = 0.33, Figure 4.14B). The only P fraction that differed

significantly between weirs and controls was reductant-soluble P (S = 327.00, P = 0.05),

with controls having greater concentrations than weirs (Figure 4.15D). There were no

significant differences between weir and control bioavailability ratios (S = 236.00, P =

0.48, Figure 4.16).

Among weir ages, fractions that differed significantly were FePO4 (F = 3.30, P =

0.06, Figure 4.17C) and reductant-soluble P (F = 3.20, P = 0.06, Figure 4.17D). Two

year old weirs had the greatest concentration of FePO4 and the least concentration of

reductant-soluble P. One year old weirs had the least concentration of FePO4, whereas 4

year old weirs had the greatest concentration of reducant-soluble P. This translated to 1

year old weirs having the least bioavailability ratios and 2 year old weirs having the

greatest bioavailability ratios (χ2 = 4.77, P = 0.09, Figure 4.18).

Soluble and loosely bound P was correlated positively with weir age (F = 4.16, P

= 0.05, R2 = 0.17, Figure 4.19A), as was reductant-soluble P concentrations (F = 3.12, P=

0.09, R2 = 0.14, Figure 4.19D). AlPO4 was correlated negatively with weir age (F = 4.21,

P = 0.05, R2 = 0.17, Figure 4.19B). While FePO4 was not linearly correlated with weir

age, it was correlated polynomially (F = 3.30, P = 0.06, R2 = 0.26). FePO4 concentrations

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increased from 1 year old to 2 year old weirs and then decreased from 2 year old to 4 year

old weirs, (Figure 4.19C). There was no correlation between bioavailability ratios and

weir age (F = 0.03, P = 0.86, R2 < .01, Figure 4.20).

Organic matter percentage was correlated positively with bioavailability ratio (F =

10.26, P = 0.02, R2 = 0.63). A larger organic matter percentage corresponded with a

larger ratio. No other factors in controls were correlated. Similar to controls, organic

matter percentage was correlated positively with bioavailability ratio in weirs (F = 5.27, P

= 0.04, R2 = 0.31). When split among ages, 1 year old weir bioavailability ratios are

negatively correlated with sediment pH (F = 4.25, P = 0.09, R2 = 0.46) and correlated

positively with organic matter percentage (F = 4.42, P = 0.09, R2 = 0.47). Two and 4

year old weir bioavailability ratios were not correlated with any abiotic variables

measured in this study.

4.4 Discussion

Weirs installed in agricultural drainage ditches are intended to mimic ecosystem

services of sediment and P retention provided by wetlands. Using wetlands as a

conservation practice has removed effectively sediment and P from agricultural runoff

(Chescheir et al. 1991a). This conservation practice is more effective with slower flow

velocities and longer hydraulic residence time (Chescheir et al. 1991b, Mander et al.

1991, Richardson and Craft 1993, Fisher and Acreman 2004). Weirs implemented in

agricultural ditches are intended to use hydrologic manipulation to mimic wetland

ecosystem services of sediment and P retention. In this study, weirs are clearly effective

at retaining water better than controls, with weirs collectively having an average water

depth 8 times greater than controls (Figure 4.6B). Because ditch sediments more closely

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resemble upland soils rather than wetland soils, controlled drainage can be used to

prolong inundation of ditch sediments, lessening P loss (Dunne et al. 2007).

There are limits to the ability for soil to retain P. As wetlands age, nutrient

reduction efficiencies decrease due to binding site saturation (Richardson 1985). Yoo et

al. (2006) states the study marsh in Korea designed for P removal is estimated to act as a

P sink for 6-13 years. Comparing the Yoo et al. study to this study highlights the

difference between using wetlands and weirs as conservation strategies. In this study, the

least average bioavailability ratio among weir ages was with one year old weirs,

suggesting that drainage ditch sediment P retention efficiency should be expected to be

much shorter than that of constructed wetlands. Initial uptake of P by sediments can be

quite rapid and effective in the short-term but has a fixed storage capacity and should not

be confused with long-term retention (Richardson and Craft 1993). In fact, Fisher and

Acreman (2004) compared various investigations on wetland P reduction efficiency and

report that studies conducted for a year or longer indicated wetlands actually began

contributing to P loss. Burial and soil formation are the only true long-term mechanisms

that assure P is not released back to solution (Richardson and Craft 1993, Needelman et

al. 2007).

White et al. (1999) states that sedimentation is the main mechanism for P

retention in their study, but also observe that sediments near the inflow have reduced

sorption ability after five years. This occurs because the physical volume capable of

containing accumulating sediment is rapidly depleted and the slower, more permanent

mechanism of soil production cannot keep pace providing new P binding sites (Gehrels

and Mulamoottil 1989). In this study, agricultural drainage ditches equipped with weirs

had even more rapid sediment accumulation than the five years observed by White et al.,

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with physical containment limits appearing to be reached after one year. Certainly, the

slow process of soil production cannot be considered a reasonable P retention mechanism

with rates of accumulation exhausting sediment retention capacities in one year.

The rapid pace of sediment accumulation observed underscores the critical need

for weirs in drainage ditches to intercept the quantity of sediment headed for downstream

waters from the agricultural landscape. In other systems, such as the slough studied by

Williams (1995), no sediment was trapped behind the constructed weir. However, weirs

in this study retain sediment more than twice as much as controls. Because volume of

sediment depends on dimensions of the ditch and weir, comparisons among ages can’t be

made, however, substantial volumes of sediment are quantitatively being withheld behind

weirs and preventing impacts downstream (Table 4.1).

For this conservation practice to remain effective, maintenance is required.

Sánchez-Carrillo et al. (2001) observed sedimentation rates of 1 to 4 cm per year within

their study wetland, and predicted physical capacity reached by the end of the century. In

comparison to this study, average sediment depths behind one year old weirs were over

30 cm. This would imply drainage ditches are dynamic systems and can be expected to

reach physical sediment retention capacity much sooner when a weir is installed. Most

sedimentation occurred behind one year old weirs and drastically declines behind two

year old weirs (Figure 4.7). This suggests yearly maintenance would be required.

Maintenance timing can be based on physical or chemical results, and in this study they

appear to coincide. Bioavailability ratios were least behind one year old weirs, again

suggesting yearly maintenance requirements. The mechanism behind P retention by ditch

sediments at weir locations was not investigated in this study, and it should be noted that

successful retention of P by one year old weir sediments should not be confused with

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long term removal. But because long term soil formation is not a realistic expectation for

intensively managed and beleaguered drainage ditch systems, minimum yearly

maintenance at precise weir locations would reset the system for optimal sediment and P

reduction.

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Table 4.1 Calculated average volume of sediment accumulated behind weirs of various ages at different sites in Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) in August 2010.

Age (Years) Site Volume of Sediment Accumulated (m3)

1 Coleman 14.68 1 HS Weir 1 55.13 1 HS Weir 2 191.30 1 Straughter 30.44 1 Terrace Weir 1 217.00 1 Terrace Weir 2 117.05 2 RC Weir 1 2.49 2 RC Weir 2 9.19 4 Bee Lake Weir 1 118.86 4 Bee Lake Weir 2 15.83 4 Bee Lake Weir 3 31.99

Table 4.2 Soil texture at study sites associated with Bee and Wolf Lake in Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) in August 2010.

Sample Site

Average % Sand (>50 to

< 1020 µm)

Average % Silt (>2 to

< 50 µm)

Average % Clay (<2 µm)

Soil Texture

Terrace 2.62 69.18 28.19 Silty Clay Loam Straughter 5.838 80.77 13.392 Silt Loam Coleman 4.682 80.051 15.265 Silt Loam

DP 10.33 79.25 10.41 Silt/Silt Loam HS 8.99 77.29 13.73 Silt Loam RC 2.43 78.49 19.08 Silt Loam

Bee Lake 3.27 79.59 17.15 Silt Loam

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Figure 4.1 Location of study sites associated with two oxbow lakes, Bee Lake and Wolf Lake, in Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) region. Nine drainage ditches containing 14 weirs of various ages at seven sites were surveyed and sampled in August 2010.

73

Figure 4.2 Schematic depicting control site location decision. Controls were located within the same drainage ditch as their respective weirs so as to eliminate undesired variables, but also located far enough away so as to be outside the influence of the weir. The distance of weir influence ended where height of the weir intercepted slope of the ditch.

74

Figure 4.3 Example of a soil core taken from an agricultural drainage ditch to determine visually and tactilely where depositional sediment layer is different from original ditch bottom soil, and therefore sediment depth, Yazoo City, July 2010.

75

Figure 4.4 A technique to determine sediment depth behind a weir. A laser is leveled on the top of the weir and shot to a marker held above and below the weir. Height from the top of the sediment to the laser mark is recorded. Sediment depth is determined by the difference between heights.

76

Figure 4.5 Schematic showing measurements used to calculate estimated volume of sediment (m3) retained behind a weir. Volume of two triangular based pyramids are calculated and subtracted from each other to determine accumulated sediment volume. Volume of ditch is calculated using height and width of the exposed weir for the area of the base. Length is determined by ditch slope. The second pyramid has measured sediment depth added to the base height.

77

Figure 4.6 Average accumulation depth of sediments (F = 3.55, P = 0.08) (A) and water depth (S = 50.50, P = 0.02) (B) at 7 sites within Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) in August 2010. Different letters denote significant differences. Standard error bars shown.

78

Figure 4.7 Averages (χ2 = 7.58, P = 0.02) (A) and regression analysis (F = 5.25, P = 0.05) (B) for difference in sediment depth between weirs and their corresponding controls in 7 drainage ditch sites within Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) in August 2010 for 1, 2, and 4 year old weirs. Different letters denote a significant difference. Standard error bars shown.

79

Figure 4.8 Averages (F = 3.18, P = 0.13) (A) and regression analysis (F = 6.46, P = 0.04) (B) for difference in water depth between weirs and their corresponding controls in 7 drainage ditch sites within Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) in August 2010 for 1, 2, and 4 year old weirs. Similar letters denote no significant difference. Standard error bars shown.

80

Figure 4.9 Average pH of drainage ditch sediments at 7 sites within Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) in August 2010. Similar letters denote no significant difference (F = 1.44, P = 0.24). Standard error bars shown.

81

Figure 4.10 Averages (F = 2.74, P = 0.17) (A) and regression analysis (F = 3.32, P = 0.10) (B) for difference in sediment pH between weirs and their corresponding controls in 7 drainage ditch sites within Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) in August 2010 for 1, 2, and 4 year old weirs. Negative numbers indicate pH at control is greater than weir. Similar letters denote no significant difference. Standard error bars shown.

82

Figure 4.11 Average organic matter percentage of drainage ditch sediments at 7 sites within Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) in August 2010. Similar letters denote no significant difference (F = 0.30, P = 0.59). Standard error bars shown.

83

Figure 4.12 Averages (F = 1.03, P = 0.39) (A) and regression analysis (F = 0.26, P = 0.62) (B) for difference in sediment organic matter percentage between weirs and their corresponding controls in 7 drainage ditch sites within Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) in August 2010 for 1, 2, and 4 year old weirs. Negative numbers indicate organic matter percentage greater at control than weir. Similar letters denote no significant difference. Standard error bars shown.

84

Figure 4.13 Average TP of drainage ditch sediments at 7 sites within Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) in August 2010. Similar letters denote no significant difference (S = 238.00, P = 0.76). Standard error bars shown.

85

Figure 4.14 Averages (F = 1.14, P = 0.38) (A) and regression analysis (F = 0.99, P = 0.33) (B) for difference in sediment TP between weirs and their corresponding controls in 7 drainage ditch sites within Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) in August 2010 for 1, 2, and 4 year old weirs. Negative numbers indicate TP greater at control than weir. Similar letters denote no significant difference. Standard error bars shown.

86

Figure 4.15 Average soluble and loosely bound P (S = 247.00, P = 0.71) (A), AlPO4 (S = 262.00, P = 0.96) (B), FePO4 (S = 248.00, P = 0.74) (C), and reductant-soluble P (S = 327.00, P = 0.05) (D) concentrations of drainage ditch sediments at 7 sites within Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) in August 2010. Different letters denote significant differences. Standard error bars shown.

87

Figure 4.16 Average bioavailability ratio of drainage ditch sediments at 7 sites within Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) in August 2010. Similar letters denote no significant difference (S = 236.00, P = 0.48). Standard error bars shown.

88

Figure 4.17 Average difference in soluble and loosely bound P (χ2 = 3.22, P = 0. 20) (A), AlPO4 (χ

2 = 3.50, P = 0.17) (B), FePO4 (F = 3.30, P = 0.06) (C), and reductant-soluble P (F = 3.20, P = 0.06) (D) concentrations between weirs and their corresponding controls in 7 drainage ditch sites within Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) in August 2010 for 1, 2, and 4 year old weirs. Negative numbers indicate concentrations greater at control than weir. Different letters denote significant differences. Standard error bars shown.

89

Figure 4.18 Average difference in sediment bioavailability ratios between weirs and their corresponding controls in 7 drainage ditch sites within Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) in August 2010 for 1, 2, and 4 year old weirs. Different letters denote significant differences (χ2 = 4.77, P = 0.09). Standard error bars shown.

90

Figure 4.19 Least squares regression analysis of the difference in soluble and loosely bound P (F = 4.16, P = 0.05) (A), AlPO4 (F = 4.21, P = 0.05) (B), FePO4 linear (F < 0.01, P = 0.99) and polynomial (2 degrees, F = 3.30, P = 0.06) (C), and reductant-soluble P (F = 3.12, P = 0.09) (D) concentrations between weirs and their corresponding controls in 7 drainage ditch sites within Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) in August 2010 for 1, 2, and 4 year old weirs. Negative numbers indicate concentrations greater at control than weir.

91

Figure 4.20 Least squares regression analysis of the difference in sediment bioavailability ratios between weirs and their corresponding controls in 7 drainage ditch sites within Mississippi’s Lower Mississippi River Alluvial Valley (LMAV) in August 2010 for 1, 2, and 4 year old weirs (F = 0.03, P = 0.86). Negative numbers indicate bioavailability ratio greater at control than weir.

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Richardson, C. J. and C. B. Craft. 1993. Effective phosphous retention in wetlands: fact or fiction? Pages 271-282 in G. A. Moshiri, editor. Constructed wetlands for water quality improvement. CRC Press, Inc, Boca Raton, FL.

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CHAPTER V

SYNTHESIS

Literature suggests that P release may occur from either inundation of previously

dry sediments, or desiccation of previously wet sediments (Qiu and McComb 1994,

Baldwin and Mitchell 2000, Surridge et al. 2007, Loeb et al. 2008). Due to the

ephemeral nature of many agricultural drainage ditches, either of these scenarios could

exist throughout the year. Results from the current research concur that unstable

fluctuation between drying and wetting leads to P release. Suggesting that keeping

sediments inundated as long as possible without intermittent drying will prevent

excessive P solubility (Watts 2000, Kröger et al. 2012). Without weirs in place

encouraging prolonged inundation, ditch sediments are more likely to act as P sources.

Creating a state of persistent inundation in agricultural drainage ditches also can

affect vegetative and microbial communities. A large portion of the P cycle not explored

by this research involves biological processes, which could explain more complex

mechanisms regarding weir influence on P fractions in the field. Although biological

uptake is considered a temporary means to sequester P, weirs can be expected to perform

efficiently for one year before requiring maintenance; therefore biological uptake could

be considered a relevant portion of weir effectiveness. Peat accretion and litter burial are

the only long term biotic means of P storage, and comprise a small amount of the P

returned to the system after death or senescence. But because capacities for sediment

retention and advantageous bioavailability ratios maximize behind a weir after one year,

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additional P could be withdrawn throughout one growing season, eliminating P

reintroduction because of annual maintenance already scheduled based on sediment

accumulation rates.

Annual maintenance may seem frequent compared to other conservation

practices. However, this research demonstrates not only value of weirs, but also the

challenge of comparing them to other known conservation practices. Agricultural surface

drainage ditch weirs can be viewed as sharing characteristics of constructed wetlands,

headwater streams, vegetated buffer strips, and floodplains, but should be considered

their own distinct system. Arguably, one of the largest distinctions is the accelerated

process of sediment accumulation. This also highlights the heightened need for weirs.

Of the calculated sediment volumes from the weirs in this study, a collective total of

803.96 m3 of sediment was retained.

With sediment loss on a scale that considerable, a rapid solution is not only

required, but also appropriate. Relying on short-term, temporary mechanisms of P

adsorption by sediments is justifiable under an annual maintenance plan. This mitigation

strategy also could be compatible with agriculture management needs. Yearly

maintenance may be viewed by the farmer as labor intensive; however, efforts would be

restricted to weir locations. Additionally, excavated sediment from the drainage ditch

can be transported upland to supplement areas on the farmer’s property that is

experiencing soil loss. This confers further benefit by transporting associated P with

previously inundated sediment to a terrestrial system, where through drying and rewetting

from rain or irrigation, P becomes available for crop production, decreasing input costs.

With the continued depletion of finite minable P reserves, inorganic P fertilizer input

costs will steadily rise.

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For weirs to be implemented broadly and to become a suggested agriculture best

management practice (BMP), they must include worthwhile incentives to those using

them, as well as demonstrate environmental benefits. This research not only

demonstrates weirs’ environmental applicability, but it reveals several potential

advantages over other established BMPs. Drainage ditches are a necessary, preexisting,

and widely found feature on the agricultural landscape. Consequently weirs installed in

drainage ditches do not require land to be taken out of production. They can be

constructed out of a range of building materials with various costs. Multiple small weirs

can be arranged within a drainage system to allow for a compounding positive water

quality effect while allowing a flexible design. Future studies may wish to examine

differences with building materials and configurations to refine the sediment retention

ability of weirs. For weirs to become adopted commonly, they must display such

practical and economical attributes. Therefore, it is recommended that those aspects not

be overlooked in future research into this burgeoning conservation asset.

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5.1 References Baldwin, D. S. and A. M. Mitchell. 2000. The effects of drying and re-flooding on the

sediment and soil nutrient dynamics of lowland river–floodplain systems: a synthesis. Regulated Rivers: Research and Management 16:457-467.

Kröger, R., R. E. Lizotte, Jr., F. D. Shields, J., and E. Usborne. 2012. Inundation influences on bioavailability of phosphorus in managed wetland sediments in agricultural landscapes. Journal of Environmental Management 41:1-11.

Loeb, R., L. P. M. Lamers, and J. G. M. Roelofs. 2008. Prediction of phosphorus mobilisation in inundated floodplain soils. Environmental Pollution 156:325-331.

Qiu, S. and A. McComb. 1994. Effects of oxygen concentration on phosphorus release from reflooded air-dried wetland sediments. Australian Journal of Marine and Freshwater Research 45:1319-1328.

Surridge, B. W. J., A. L. Heathwaite, and A. J. Baird. 2007. The release of phosphorus to porewater and surface water from river riparian sediments. Journal of Environmental Quality 36:1534-1544.

Watts, C. J. 2000. Seasonal phosphorus release from exposed, re-inundated littoral sediments of two Australian reservoirs. Hydrobiologia 431:27-39.

Final

Documents

weirs function

fisheries professor

department of wildlife

phosphorus p

james nail

aquaculture mississippi

phosphorus dynamics

field study