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Treatment Potential of Wastewater Drainage Ditches in a Rural Community of the Andean Amazon

University of Florida Institutional Repository Howard T. Odum Center for Wetlands
Permanent Link: http://ufdc.ufl.edu/UFE0021048/00001

Material Information

Title: Treatment Potential of Wastewater Drainage Ditches in a Rural Community of the Andean Amazon
Physical Description: 1 online resource (225 p.)
Language: english
Creator: Saunders, Lynn V
Publisher: University of Florida
Place of Publication: Gainesville, Fla.
Publication Date: 2007

Subjects

Subjects / Keywords: appropriate, canal, cbnrm, country, desagues, developing, drain, ecological, ecology, engineering, humedales, management, natural, peru, phosphorus, sanitation, sewage, systems, technology, wetland
Environmental Engineering Sciences -- Dissertations, Academic -- UF
Genre: Environmental Engineering Sciences thesis, Ph.D.
bibliography   ( marcgt )
theses   ( marcgt )
government publication (state, provincial, terriorial, dependent)   ( marcgt )
born-digital   ( sobekcm )
Electronic Thesis or Dissertation

Notes

Abstract: Ditches are ubiquitous features in altered landscapes. Although ditches are designed to be efficient at moving water across the landscape, there is growing evidence that ditches provide services beyond basic water transport. The position of ditches in the landscape lends them great importance for controlling the timing and magnitude of terrestrially-derived contaminant exports to downstream water bodies. My study examined the treatment function of ditches receiving domestic sewage effluents. The use of ditches for discharging wastewaters is common in regions where treatment systems are often non-existent. My study investigated the occurrence of in-stream contaminant retention in wastewater ditches, identified mechanisms responsible for contaminant retention, and implemented experimental modifications to the design of existing ditches to test whether treatment performance was improved. The setting for the study was the town of Oxapampa, Peru where approximately two-thirds of wastewater generated is discharged to the Chorobamba River via earthen, vegetated ditches while the remainder is routed to the river via underground pipes. Dilution-corrected concentrations of E. coli, biochemical oxygen demand, total suspended solids and soluble reactive phosphorus exported from ditches were found to be significantly lower than the same effluents discharged via underground pipes, indicating that transport in ditches is not conservative. Conservative solute tracer experiments revealed the important influence of ditch vegetation on transport characteristics such as residence time and transient storage, which have positive implications for improved contaminant retention and decreased export of sediment and E. coli. Deposition of silts and clays and organic matter accumulation were shown to be important drivers of improved phosphate retention by promoting sorption with benthic sediments. Channel modifications to ditches were performed within a participatory framework that directly involved community members in planning, implementation and management. Two different modification approaches were tested: an open water flow design and an alternating subsurface flow/open water design. The latter design proved to be effective at sediment and E. coli removal and shows promise for treatment of water to irrigation standards. Reuse of treated ditch water should be promoted to prevent the continued eutrophication of the river.
General Note: In the series University of Florida Digital Collections.
General Note: Includes vita.
Bibliography: Includes bibliographical references.
Source of Description: Description based on online resource; title from PDF title page.
Source of Description: This bibliographic record is available under the Creative Commons CC0 public domain dedication. The University of Florida Libraries, as creator of this bibliographic record, has waived all rights to it worldwide under copyright law, including all related and neighboring rights, to the extent allowed by law.
Statement of Responsibility: by Lynn V Saunders.
Thesis: Thesis (Ph.D.)--University of Florida, 2007.
Local: Adviser: Brown, Mark T.

Record Information

Source Institution: UFRGP
Rights Management: Applicable rights reserved.
Classification: lcc - LD1780 2007
System ID: UFE0021048:00001

Permanent Link: http://ufdc.ufl.edu/UFE0021048/00001

Material Information

Title: Treatment Potential of Wastewater Drainage Ditches in a Rural Community of the Andean Amazon
Physical Description: 1 online resource (225 p.)
Language: english
Creator: Saunders, Lynn V
Publisher: University of Florida
Place of Publication: Gainesville, Fla.
Publication Date: 2007

Subjects

Subjects / Keywords: appropriate, canal, cbnrm, country, desagues, developing, drain, ecological, ecology, engineering, humedales, management, natural, peru, phosphorus, sanitation, sewage, systems, technology, wetland
Environmental Engineering Sciences -- Dissertations, Academic -- UF
Genre: Environmental Engineering Sciences thesis, Ph.D.
bibliography   ( marcgt )
theses   ( marcgt )
government publication (state, provincial, terriorial, dependent)   ( marcgt )
born-digital   ( sobekcm )
Electronic Thesis or Dissertation

Notes

Abstract: Ditches are ubiquitous features in altered landscapes. Although ditches are designed to be efficient at moving water across the landscape, there is growing evidence that ditches provide services beyond basic water transport. The position of ditches in the landscape lends them great importance for controlling the timing and magnitude of terrestrially-derived contaminant exports to downstream water bodies. My study examined the treatment function of ditches receiving domestic sewage effluents. The use of ditches for discharging wastewaters is common in regions where treatment systems are often non-existent. My study investigated the occurrence of in-stream contaminant retention in wastewater ditches, identified mechanisms responsible for contaminant retention, and implemented experimental modifications to the design of existing ditches to test whether treatment performance was improved. The setting for the study was the town of Oxapampa, Peru where approximately two-thirds of wastewater generated is discharged to the Chorobamba River via earthen, vegetated ditches while the remainder is routed to the river via underground pipes. Dilution-corrected concentrations of E. coli, biochemical oxygen demand, total suspended solids and soluble reactive phosphorus exported from ditches were found to be significantly lower than the same effluents discharged via underground pipes, indicating that transport in ditches is not conservative. Conservative solute tracer experiments revealed the important influence of ditch vegetation on transport characteristics such as residence time and transient storage, which have positive implications for improved contaminant retention and decreased export of sediment and E. coli. Deposition of silts and clays and organic matter accumulation were shown to be important drivers of improved phosphate retention by promoting sorption with benthic sediments. Channel modifications to ditches were performed within a participatory framework that directly involved community members in planning, implementation and management. Two different modification approaches were tested: an open water flow design and an alternating subsurface flow/open water design. The latter design proved to be effective at sediment and E. coli removal and shows promise for treatment of water to irrigation standards. Reuse of treated ditch water should be promoted to prevent the continued eutrophication of the river.
General Note: In the series University of Florida Digital Collections.
General Note: Includes vita.
Bibliography: Includes bibliographical references.
Source of Description: Description based on online resource; title from PDF title page.
Source of Description: This bibliographic record is available under the Creative Commons CC0 public domain dedication. The University of Florida Libraries, as creator of this bibliographic record, has waived all rights to it worldwide under copyright law, including all related and neighboring rights, to the extent allowed by law.
Statement of Responsibility: by Lynn V Saunders.
Thesis: Thesis (Ph.D.)--University of Florida, 2007.
Local: Adviser: Brown, Mark T.

Record Information

Source Institution: UFRGP
Rights Management: Applicable rights reserved.
Classification: lcc - LD1780 2007
System ID: UFE0021048:00001


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TREATMENT POTENTIAL OF WASTEWATER DRAINAGE DITCHES IN A RURAL
COMMUNITY OF THE ANDEAN AMAZON




















By

LYNN VELISHA SAUNDERS


A DISSERTATION PRESENTED TO THE GRADUATE SCHOOL
OF THE UNIVERSITY OF FLORIDA IN PARTIAL FULFILLMENT
OF THE REQUIREMENTS FOR THE DEGREE OF
DOCTOR OF PHILOSOPHY

UNIVERSITY OF FLORIDA

2007

































O 2007 Lynn Velisha Saunders





























To my parents and to my namesake, Dr. Mary Velisha McIndoo









ACKNOWLEDGMENTS

I am grateful to the many special people in my life that have directly or indirectly led me to

this point in time and space. Endless gratitude is deserved to my parents, Bill and Brenda

McIndoo, who have always been incredible beacons of love and support, even throughout those

rough adolescent years. I thank Andrew McIndoo for always being an awesome and caring

brother, not to mention a great travel companion. I also thank Kori Jacobs, a wonderful friend

who's kept me laughing over these last five years during the Bartow days, field work in Peru,

and the many nights just hanging out over dinner and wine.

I express my great appreciation to my mentors at Humboldt State University: Bob

Gearheart, Beth Eschenbach and Brad Finney. It was with their guidance, support and inspiration

that I chose to continue with my formal education.

I thank my current advisor, Mark Brown for allowing and trusting me to spread my wings

and develop my own research proj ect; for providing me with financial support for conferences

and equipment; and most importantly, for introducing me to the world of systems ecology. I

thank Joe Delfino for his support by loaning me much-needed laboratory equipment. I also

express my appreciation to Michael McClain for his inspiration and support as well as for taking

the time out of his busy schedule to come up to Gainesville from Miami to attend my qualifying

exams. I also thank Jim Jawitz for helpful suggestions on my research and for his wry humor that

made subsurface contaminant hydrology lectures entertaining. Thanks to Mark Clark for his

support and insightful inquiries into this research. Thanks to Ed Dunne for his thoughtful review

of my chapter related to phosphorus sorption.

My sincere gratitude goes to Joaquin Arteaga, former director of public works related to

water in Oxapampa and later an assistant for this research, who provided me with important

background data that were crucial for this research. I thank my hard-working and wastewater









ditch-tolerant field assistants, Jaime Guerovich and Gino Arteaga Koo. I also thank to Percy

Summers, director of ProPachitea at the Instituto de Bien Comun for allowing me to use their

research facilities in Oxapampa. Special thanks goes to all my friends in Oxapampa, Peru who

supported and worked with me on this study and made my experiences there so treasured.

And finally, and most of all, to my husband Tom, the greatest husband, friend, mentor,

collaborator and lifelong partner that I could ever imagine.












TABLE OF CONTENTS


page

ACKNOWLEDGMENTS .............. ...............4.....


LIST OF TABLES ................. ...............9..._. .....


LIST OF FIGURES ........._._ ...... .__ ...............11...


AB S TRAC T ............._. .......... ..............._ 18...


CHAPTER


1 INTRODUCTION ................. ...............20.......... ......


Statement of the Problem ................. ...............20................

Study Overview and Obj ectives ................. ...............21.......... ....
Study Setting ................... ..... ...... .... ... ...............22....
Water Quality in the Chorobamba River............... ...............23.
Water Quality in Wastewater Ditches and Pipes............... ...............24.
Research Approach ................. ...............25.................


2 LITERATURE REVIEW ................. ...............34................


Introduction Ditches in the Landscape ............... ...............34....
Conceptual Basis for Pollutant Retention in Ditches .............. ...............37....
Characteristics of Ditch Ecosystems ................. ...............37................
Contaminant Exposure and Retention Efficiency .............. ...............38....
Effect of Ditch Structure on Contaminant Retention ................. ..............................40

Vegetation in Ditches ................. ...... ..... ... .. .............4
Effect of Vegetation on Velocity Attenuation and Transient Storage ................... ..........41
Influence of Hydrologic Retention on Water Quality ................. ....__. ...............43
Sedimentation ............... ... .. ...._.__.. .. ........ ...... ............4
Retention processes influenced by interactions with the benthos ................... .........46
Synthesis and research needs ...._... ................. ........._.. ........5
Conclusions............... ..............5


3 INFLUENCE OF MACROPHYTES ON TRANSPORT CHARACTERISTICS AND
NUTRIENT ST ORAGE S............... ...............66


Introducti on ................. ...............66.................
M ethod s .............. ...............67....
Site Descriptions............... ..............6
San M artin ................. ...............67.................
Gustayson ........._..... ..... ._. ...............68....
K oell .............. ...............68....

Frey............... ...............68..












Field Procedures for Water Chemistry, Flow and Channel Characterization .................69
Laboratory Analysis of Water Quality ....___. ................ .........___. ....... 6
Transport Characteristics of Ditches ....___. ................ ...............70. ....
Trac er exp eri ment fi eld procedure s ...._.. ...._._._.. ........__._ .........7
Analysi s of tracer data ................. ...............71........... ...
Vegetation Sampling .............. ...... ..... .............7
Collection of plants at sampling locations .................. .......... .... .......... .........73
Comparison of biomass and nutrient content of plants grown out of wastewater ...74
Statistical Analyses............... ...............74
Re sults................ .. ...........__ ... .......___.. ...........7
Water Quality of Wastewater Ditches ...._ ......_____ .......___ ............7
General characteristics .............. .. ...............74.
Water quality during tracer experiments ...._ ......_____ ...... ....__.........7
Transport Characteristics............... ............7
Plant Biomass and Nutrient Storages .............. ...............78....
Discussion ................. ...............80........ ......
Conclusions............... ..............8


4 PHOSPHORUS RETENTION BY SORPTION WITH BENTHIC SEDIMENTS.............100


Introducti on ................. ...............100................
Methods .............. .... ...............101..

Experimental Sites ................. ...............101................
Field Procedures ................. ...............102......... ......
Laboratory Procedures............... ... .. ...............10
Water chemistry and sediment characterization............... ...........10
Phosphorus sorption index ................. ......... ...............102 ....
Equilibrium phosphorus concentration .............. ...............103....
Phosphorus fractionation............... .............10
Statistical Analyses............... ...............10
Re sults ................ ...............106................
D discussion ............... .... .. .. ....... ....... ..... ... .. .... .... .. ......... 0
Effect of Seasonal Trends in Particle Size and Organic Matter Content on P
Buffering Capacity............... ...............10
Phosphorus Sorption Index ................. ............. ...............110 .....
Sediment Equilibrium Phosphorus Concentrations ................. .......... ................11 3
Phosphorus Fractionation to Examine P Status of Ditch Sediments ................... ..........1 14
Management Implications ................. ...............117......... ......

5 EVALUATION OF CHANNEL MODIFICATIONS TO EXISTING DITCHES FOR
IMPROVED WATER QUALITY AND AESTHETICS ........................... ...............131

Introducti on ................. ...............131................
M ethods .............. .. ......... ...... .. .. .........13
Government and Community Involvement ................ ...............132...............
M odification Designs .............. ...............133....
Open-water flow design ................ ...............134................












S ub surface fl ow d esi gn .............. ............... 13 5...
Monitoring and Statistical Analyses............... ...............13
R e sults................... .... ........ ... ......... .. ... ................... 3
Effects of Modifications on Residence Times and Water Quality .............. .... ............137
Management Requirements ............ _...... ._ ...............139...
Government and Community Response ....__ ......_____ .......___ ...........14
Discussion ............ ..... .._ ...............142...
Conclusions............... ..............14


6 SYNTHESIS AND FUTURE WORK .............. ...............164....

APPENDIX


A ENERGY CIRCUIT LANGUAGE ..........._ .....___ ...............167.

B DITCH CHANNEL CROSS-SECTIONAL AND LONGITUDINAL PROFILES ............168


C QA/QC RESULTS.................... ...........17

D OTIS-P SIMULATION RESULTS ................. ...............179...............


E PHO SPHORUS FRACTIONATION DATA ................. ...............181........... ...


F PHOSPHORUS SORPTION INDEX (PSI) DATA ................. ..............................193


G EQUILIBRIUM PHO SPHORUS CONCENTRATION (EPCO) DATA ................... .........200

H INTERPRETIVE MATERIALS USED FOR COMMUNITY PARTICIPATION............_.209


LIST OF REFERENCES ................. ...............211...___ ......


BIOGRAPHICAL SKETCH .............. ...............225....










LIST OF TABLES


Table page

2-1 Summary of environmental conditions and findings from published studies
examining sediment, nutrient, BOD and pesticide retention in drainage ditches..........._...60

2-2 Summary of the retention effectiveness of ditches based on the studies reported in
Table 2-1 .............. ...............64....

2-3 Possible reactants and conditions that limit the retention capacity in drainage ditches ....65

2-4 Dominant macrophytes reported by the 15 studies out of 22 in Table 2-1 that
provided plant species information............... ..............6

3-1 Dominant macrophyte species common to the study sites. ............. .....................8

3-2 Description of tracer experiments conducted at each site ................. ................ ...._.90

3-3 Channel characteristics and mean water quality parameters for wastewater drainage
ditch sites .............. ...............91....

3-4 Biomass and hydrologic conditions of each tracer experiment and resulting transient
storage zone parameter values .............. ...............96....

3-5 Estimated P and N uptake rates based on measured above and belowground biomass
and plant tissue nutrient storage at each study site ................. .............................97

3-6 Average P and N loads and estimated gross nutrient retention by aboveground
biomass assimilation at each study site ................. ...............97...............

3-7 Mean molar ratios of carbon, nitrogen and phosphorus of aboveground (AG) and
belowground (BG) plant tissues between sites .............. ...............98....

4-1 Site averages of particle size distributions and sediment pH for the wastewater
ditches in October-November 2005 and June-August 2006 ................ .....................122

4-2 Comparison of P sorption measures determined in each site in years 2005 and 2006 ....122

4-3 Phosphorus sorption index (PSI) values for wastewater ditches calculated using three
different sets of units to compare with values reported for ditches, streams and
w wetlands .............. ...............129....

5-1 Activities performed to promote local support and participation in channel
m odifications............... .............14

5-2 Primary design elements used in open-water flow design and their intended functions
for improving water quality and aesthetics ................. ...............150..............











5-3 Primary design elements used in the alternating subsurface flow/open water design
and their intended functions for improving water quality and aesthetics ................... .....153

5-4 Comparison of median travel times in Koell and Frey before and after channel
m odifications............... .............15

Cl Summary of QA/QC results for nutrient analyses .............. ...............178....

E-1 Phosphorus fractionation data for Koell Site 1 .............. ...............181....

E-2 Phosphorus fractionation data for Koell Site 2 ......___ ..... .._._. ....._.._........18

E-3 Phosphorus fractionation data for Koell Site 3 .............. ...............183....

E-4 Phosphorus fractionation data for Koell Site 4 ...._.._.._ ..... .._._. ....._.._........18

E-5 Phosphorus fractionation data for San Martin Site 1 .................... ..............18

E-6 Phosphorus fractionation data for San Martin Site 2 ....._____ .......___ ................1 86

E-7 Phosphorus fractionation data for San Martin Site 3 .............. ...............187....

E-8 Phosphorus fractionation data for San Martin Site 4 ....._____ .......___ ................1 88

E-9 Phosphorus fractionation data for Gustayson Site 1 .............. ...............189....

E-10 Phosphorus fractionation data for Gustayson Site 2 ................ .......... ................1 90

E-11 Phosphorus fractionation data for Gustayson Site 3 .............. ...............191....

E-12 Phosphorus fractionation data for Gustayson Site 4 ................ .......... ................1 92

F-1 Year 2005 PSI data for Koell sites............... ...............193.

F-2 Year 2005 PSI data for Frey sites .............. ...............194....

F-3 Year 2005 PSI data for Gustayson sites............... ...............195.

F-4 Year 2005 PSI data for San Martin sites............... ...............196.

F-5 Year 2006 PSI data for Koell sites............... ...............197.

F-6 Year 2006 PSI data for Gustayson sites............... ...............198.

F-7 Year 2006 PSI data for San Martin sites............... ...............199.










LIST OF FIGURES


Figure page

1-1 Map of Peru showing the location of Oxapampa ........__............_. ........._.._....2

1-2 View of the Chorobamba River valley and the city of Oxapampa ........._.._... ........._.....27

1-3 The urban center of Oxapampa and the Chorobamba River............._ .........___.......27

1-4 Simplified map of the urban center of Oxapampa, indicating the locations of
wastewater ditches (solid lines) and underground pipes (dashed lines) all of which
drain to the Chorobamba River............... ...............28.

1-5 Daily discharge in the Chorobamba River between December 2001 and January 2006...29

1-6 Summary of E coli concentrations in the Chorobamba River (2005-2006) .....................29

1-7 Summary of SRP concentrations in the Chorobamba River (2005-2006) ................... ......30

1-8 Summary of dissolved inorganic nitrogen (sum of nitrate and ammonium)
concentrations in the Chorobamba River (2005-2006)............... ..............3

1-9 Dilution-corrected E. coli concentrations in A) ditch effluents and B) pipe effluents......31

1-10 Dilution-corrected carbonaceous biochemical oxygen demand (CBODS)
concentrations in A) ditch effluents and B) pipe effluents ................ .......__ ..........31

1-11 Dilution-corrected total suspended solids (TSS) concentrations in A) ditch effluents
and B) pipe effluents ........... __... ......... ...............32...

1-12 Dilution-corrected soluble reactive phosphorus (SRP) concentrations in A) ditch
effluents and B) pipe effluents ................. ...............32........... ...

1-13 Dilution-corrected ammonium (NH4-N) concentrations in A) ditch effluents and B)
pipe effluents............... ...............33

2-1 Hypothetical relationships for A) contaminant processing efficiency versus discharge
proposed by Meyer and Likens (1979) and B) contaminant processing efficiency
versus contaminant loading, as proposed by Odum et al. (1979) ........._.._. ........._.......58

2-2 Simplified systems diagram of the interactions between macrophyte growth, kinetic
energy ("KE," a proxy for water velocity), sediment transport and contaminant
retention in a generalized ditch system. .............. ...............59....

3-1 Systems diagram illustrating the influence of macrophyte harvesting on water quality
in wastewater ditches. ............ .............85......










3-2 Map of Oxapampa showing the locations of the sampling sites in each of the study
ditches .............. ...............86....

3-3 Ditch site San Martin looking upstream from sampling location 4 in A) November
2005 and B) June 2006 .............. ...............87....

3-4 Gustayson site looking downstream at times of solute transport experiments in
November 2005 at A) three weeks after plants were harvested and B) lower reach,
location 4 before plant removal; C) in June 2006; and D) in July 2006, one week
after vegetation removal and herbicide application. .............. ...............88....

3-5 Koell site looking downstream from sampling location 1 at the time of tracer
experiments A) in November 2005 and B) July 2006 .............. ..... ............... 8

3-6 Frey site at the marsh location at the time of biomass sampling in 2005 ........._.._.............89

3-7 View of the Mariotte siphon used for delivering the conservative tracer at a constant
flow rate .............. ...............90....

3-8 Box plots of San Martin and Koell at upstream and downstream sampling locations
for A) E. coli and B) TS S............... ...............92..

3-9 Box plots of San Martin and Koell at upstream and downstream sampling locations
for A) TP, B) TN, C) SRP and D) NH4-N .............. ...............93....

3-10 Longitudinal water quality trends of E coli, TSS, SRP and NH4-N during tracer
experiments performed in sites San Martin, Koell and Gustayson under both high
and low biomass conditions. .............. ...............94....

3-11 Comparison of conservative tracer breakthrough curves in Gustayson before and
after removal of ditch vegetation in 2006 ................ ...............95........... ..

3-12 Comparison of conservative tracer breakthrough curves in San Martin with low and
high ditch vegetation biomass............... ...............95

3-13 Comparison of breakthrough curves in Koell under different biomass conditions and
after trash was removed from the stream ...._.._.._ .... ... ..._. ...._.._ ..........9

3-14 Comparison of leaf, stem and root/rhizome tissue N and P content of macrophyte
species found in wastewater ditches .............. ...............98....

3-15 Comparison of common species in Oxapampa that are found growing in
(wastewater) and out (control) of wastewater ................. ...............99...............

4-1 Looking upstream from sampling site 4 in San Martin in A) November 2005 and B)
July 2006 ................. ...............118................










4-2 Looking downstream from sampling site 1 in Gustayson in A) November 2005 and
B) July 2006 ................. ...............119................

4-3 Looking downstream from sampling site 1 in Koell in A) November 2005 and B)
July 2006 ................. ...............119................

4-4 Sampling sites in Frey............... ...............120.

4-5 Example of the approach used to determine the sediment EPCo value and K, a
measure of P buffering capacity, from P sorption isotherms. ..........._... ............... 120

4-6 Rainfall in Oxapampa from September 2005 September 2006 .............. ..................121

4-7 Examination of longitudinal variation in 2006 PSI values for each ditch from
upstream (1) to downstream (4) sampling locations ................. ...........................123

4-8 Longitudinal trends of percent organic matter in 2006 for each ditch from upstream
("1") to downstream ("4") sampling locations. ................ ...............................123

4-9 Relationships between K (a measure of P buffer capacity determined from EPCo
experiments) and percentage of sand, silt, clay and organic matter for all sites. .............124

4-10 Contributions of biotic and abiotic sorption for phosphorus sorption indices (PSI)
determined for sediments in 2005 ....._._................. ...............124 ....

4-11 Comparison of ditch water SRP concentrations with sediment equilibrium
phosphorus concentration (EPCo) experiments............... ..............12

4-12 Sediment EPCo values as a function of water column SRP concentrations grouped
by whether sampling site acted as a P sink or source. ............. ...... ............... 12

4-13 Relationship between P sorption index and sediment EPCo indicates that highest
EPCo values (and thus greater potential to act as P sources) tend to be associated
with lowest P buffering capacity............... ...............12

4-14 Percentages of sediment pools of sorbed phosphorus for each site ............... ... ............127

4-15 Average total sediment P for San Martin, Gustayson and Koell calculated from the
sum of the P fractions for each ditch. ........._.._ ...._.. ......_.. ..........2

4-16 Average sediment TP for each sampling location at San Martin, Gustayson and
Koell............... ...............128.

4-17 Gustayson sediments collected on filters after the NaOH extraction step indicate the
presence of iron oxides. ............. .....................130

5-1 Neighborhood meetings were attended to assess whether the community supported
implementation of modified ditches ........._.._. ....._.. ...._.. ......._.. ........148










5-2 Residents participated in all aspects of the channel modification activities ................... .149

5-3 Ornamental plants such calla lily were transplanted from other sewage ditches to the
modified ditches ................. ...............150................

5-4 Schematic of the sedimentation basin design used in channel modifications .................15 1

5-5 Views of the sedimentation basin in Koell ................ ................ ......... ....... 151

5-6 Sections of cascades and riffles created to improve oxygenation and increase pockets
of transient storage ........... __..... ._ ...............152...

5-7 Creation of gravel bars to increase channel sinuosity and filtration in Koell ..................1 52

5-8 Schematic of the alternating subsurface flow/open-water design implemented in Frey .153

5-9 Views of the alternating subsurface flow/open water design used in Frey....................153

5-10 Cross-sectional view of the subsurface flow design ................. ....__ .................154

5-11 Comparison of breakthrough curves before and after channel modifications in Koell ...154

5-12 Comparison of breakthrough curves before and after channel modifications in Frey.....155

5-13 Comparison of E coli concentrations at upstream, middle and downstream sampling
locations in Koell before and after ditch modifications ................. ......._.._........._..155

5-14 Comparison of TSS concentrations at upstream, middle and downstream sampling
locations in Koell before and after ditch modifications ......___ ........__ ..............156

5-15 Comparison of dissolved oxygen (DO) concentrations at upstream, middle and
downstream sampling locations in Koell before and after ditch modifications...............156

5-16 Comparison of NH4-N concentrations at upstream, middle and downstream sampling
locations in Koell before and after ditch modifications ................. ........................157

5-17 Comparison of NO3-N concentrations at upstream, middle and downstream sampling
locations in Koell before and after ditch modifications ................. ........................157

5-18 Comparison of SRP concentrations at upstream, middle and downstream sampling
locations in Koell before and after ditch modifications ......____ ........._ .............158

5-19 Comparison of E coli concentrations at upstream, middle and downstream sampling
locations in Frey before and after ditch modifications .............. .....................5

5-20 Comparison of TSS concentrations at upstream, middle and downstream sampling
locations in Frey before and after ditch modifications .............. .....................5










5-21 Comparison of dissolved oxygen (DO) concentrations at upstream, middle and
downstream sampling locations in Frey before and after ditch modifications ................159

5-22 Comparison of NO3-N concentrations at upstream, middle and downstream sampling
locations in Frey before and after ditch modifications .............. .....................6

5-23 Comparison of NH4-N concentrations at upstream, middle and downstream sampling
locations in Frey before and after ditch modifications .............. .....................6

5-24 Comparison of SRP concentrations at upstream, middle and downstream sampling
locations in Frey before and after ditch modifications .............. .....................6

5-25 Simple tool consisting of a combined net and brush was provided to residents to
facilitate ditch maintenance .............. ...............161....

5-26 Images of the open water flow design in Koell ditch four months after modifications...162

5-27 Images of the alternating subsurface flow/open water design in Frey four months
after modifications ........... ......__ ...............162..

5-28 Residents from Frey held a celebration to inaugurate the ditch and new bridge............. 163

5-29 Other residents began attempts at modifying their ditches ................. ......................163

A-1 Description of the symbols used in energy circuit diagrams (from Odum 1994)...........167

B-1 San Martin cross-sectional profiles for sections 1 through 4............... ....................6

B-2 San Martin cross-sectional profiles for sections 5 through 8............... ...................16

B-3 San Martin longitudinal profile............... ...............170

B-4 Gustayson cross-sectional profiles for sections along Block 1.............. ....................7

B-5 Gustayson cross-sectional profiles for sections along Block 2 ........._.._.. ........_........171

B-6 Gustayson cross-sectional profiles for sections along Block 3 ........._.._.. ........_........172

B-7 Gustayson cross-sectional profiles for sections along Block 4 ........._.._.. ........_........173

B-8 Gustayson longitudinal profiles for Blocks 1-4 ................. ...............174........._..

B-9 Frey cross-sectional profiles .............. ...............175....

B-10 Frey longitudinal profile ........._. ...... .... ...............175..

B-11 Koell cross-sectional profiles............... ...............17

B-12 Koell longitudinal profile............... ...............177












D-1 OTIS-P modeling results for 2005 tracer experiment in San Martin .............. ..............179

D-2 OTIS-P modeling results for 2006 tracer experiment in San Martin .............. ..............179


D-3 OTIS-P modeling results for 2005 tracer experiment in Gustayson ............... .... ........._..179

D-4 OTIS-P modeling results for 2006 tracer experiment in Gustayson ............... .... ........._..180


D-5 OTIS-P modeling results for 2005 tracer experiment in Koell ................. ................ ..180

D-6 OTIS-P modeling results for 2006 tracer experiment in Koell ................. ................ ..180

G-1 2005 EPCO data for Koell site 1............... ...............200...

G-2 2005 EPCO data for Koell site 2............... ...............200...


G-3 2005 EPCO data for Koell site 3............... ...............200...

G-4 2005 EPCO data for Koell site 4............... ...............201...


G-5 2005 EPCO data for Gustayson site 1............... ...............201...

G-6 2005 EPCO data for Gustayson site 2............... ...............201...


G-7 2005 EPCO data for Gustayson site 3............... ...............202...

G-8 2005 EPCO data for Gustayson site 4............... ...............202...


G-9 2005 EPCO data for San Martin site 2............... ...............202...

G-10 2005 EPCO data for San Martin site 4............... ...............203...


G-11 2005 EPCO data for Frey site 1 .............. ...............203....

G-12 2005 EPCO data for Frey site 2 .............. ...............203....


G-13 2005 EPCO data for Frey site 3 .............. ...............204....

G-14 2006 EPCO data for Koell site 1............... ...............204...


G-15 2006 EPCO data for Koell site 2............... ...............204...

G-16 2006 EPCO data for Koell site 3............... ...............205...


G-17 2006 EPCO data for Koell site 4............... ...............205...

G-18 2006 EPCO data for Gustayson site 1............... ...............205...


G-19 2006 EPCO data for Gustayson site 2............... ...............206...











G-20 2006 EPCO data for Gustayson site 3............... ...............206...

G-21 2006 EPCO data for Gustayson site 4............... ...............206...

G-22 2006 EPCO data for San Martin site 1............... ...............207...

G-23 2006 EPCO data for San Martin site 2............... ...............207...

G-24 2006 EPCO data for San Martin site 3............... ...............207...

G-25 2006 EPCO data for San Martin site 4............... ...............208...


H-1 Pamphlet distributed to citizens to inform and receive their input on the
implementation of modified ditches ........__......... ........._ ....__ .......209

H-2 Image of the sign painted and installed at each of the modified ditches, translated as
"Please do not throw trash: Pilot proj ect for community participation of wastewater
treatment in ditches." .......... ..............210......











Abstract of Dissertation Presented to the Graduate School
of the University of Florida in Partial Fulfillment of the
Requirements for the Degree of Doctor of Philosophy

TREATMENT POTENTIAL OF WASTEWATER DRAINAGE DITCHES IN A RURAL
COMMUNITY OF THE ANDEAN AMAZON

Lynn Velisha Saunders

August 2007

Chair: Mark T. Brown
Major: Environmental Engineering Sciences

Ditches are ubiquitous features in altered landscapes. Although ditches are designed to be

efficient at moving water across the landscape, there is growing evidence that ditches provide

services beyond basic water transport. The position of ditches in the landscape lends them great

importance for controlling the timing and magnitude of terrestrially-derived contaminant exports

to downstream water bodies.

My study examined the treatment function of ditches receiving domestic sewage effluents.

The use of ditches for discharging wastewaters is common in regions where treatment systems

are often non-existent. My study investigated the occurrence of in-stream contaminant retention

in wastewater ditches, identified mechanisms responsible for contaminant retention, and

implemented experimental modifications to the design of existing ditches to test whether

treatment performance was improved.

The setting for the study was the town of Oxapampa, Peru where approximately two-thirds

of wastewater generated is discharged to the Chorobamba River via earthen, vegetated ditches

while the remainder is routed to the river via underground pipes. Dilution-corrected

concentrations of E coli, biochemical oxygen demand, total suspended solids and soluble

reactive phosphorus exported from ditches were found to be significantly lower than the same










effluents discharged via underground pipes, indicating that transport in ditches is not

conservative.

Conservative solute tracer experiments revealed the important influence of ditch vegetation

on transport characteristics such as residence time and transient storage, which have positive

implications for improved contaminant retention and decreased export of sediment and E. coli.

Deposition of silts and clays and organic matter accumulation were shown to be important

drivers of improved phosphate retention by promoting sorption with benthic sediments.

Channel modifications to ditches were performed within a participatory framework that

directly involved community members in planning, implementation and management. Two

different modification approaches were tested: an open water flow design and an alternating

subsurface flow/open water design. The latter design proved to be effective at sediment and E.

coli removal and shows promise for treatment of water to irrigation standards. Reuse of treated

ditch water should be promoted to prevent the continued eutrophication of the river.









CHAPTER 1
INTTRODUCTION

Statement of the Problem

The widespread deficiency of sanitation services in developing countries poses

increasingly grave threats to human and ecosystem health. From the 49% of rural communities in

Latin America that have access to sanitation services, only 14% provide some level of

wastewater treatment (WHO and UNICEF 2000). Barriers to providing adequate sanitation

services to developing countries include insufficient resources (both structural and financial),

inappropriate technological approaches, and lack of public awareness (WHO and UNICEF

2000). Innovative strategies that couple the use of existing landscape features and local materials

with public education will yield more appropriate and sustainable wastewater management

solutions that are better tailored to a particular region.

In the Andean Amazon region of Peru, untreated wastewaters are commonly discharged to

open ditches that drain to nearby streams. The physical, chemical and biological features of

these wastewater drainage ditches suggest that they may provide some level of pollutant

attenuation prior to discharge to the river. Like small headwater streams, these unlined and

vegetated ditches often possess high surface-to-volume ratios as a result of their shallow water

depths thus increasing contact with reactive benthic substrates. These interactions with the

benthos enhance biogeochemical reactions resulting in greater rates of nitrification-

denitrification, sorption to sediments and nutrient assimilation by plant and algal communities

(Peterson et al. 2001). Plant communities commonly found growing in or along the banks of

wastewater drainage ditches in Peru include Eichhornia cra~ssipes (water hyacinth), Hydrocotyle

spp. (pennywort), Lemna spp. (duckweed), Polygonum spp. (smartweed), M~yriophyllum

aquatica (parrot feather), Cynadon dactyon (bermudagrass), Hedychium coronarium (ginger










lily), and numerous others (personal observation) that have been recognized for their treatment

potential in constructed wetlands (e.g. DeBusk et al. 1995, Brix 1997, Greenway 1997, Hume et

al. 2002). This combination of physical, chemical and biological features thus provide

compelling evidence that material transport in wastewater ditches is not conservative.

Study Overview and Objectives

Drainage ditches are common interfaces between terrestrial and aquatic ecosystems in

altered landscapes throughout the world. Yet despite this important position of ditches in the

landscape, few studies have characterized the function of ditches to influence the fate and

transport of potential contaminants such as pathogens, nutrients and sediments. Existing studies

of water quality in ditches have taken place predominantly in agricultural settings (e.g. Cooper et

al. 2002, Moore et al. 2005, Dunne et al. Dr Press). My study evaluated the role of ditches in a

different context than previous studies by examining the capacity of ditches to attenuate levels of

domestic sewage contamination in a rural Peruvian community. The primary obj ectives of my

study were the following:

* Investigate the occurrence of in-stream contaminant retention in wastewater ditches

* Identify the mechanisms responsible for contaminant retention

* Implement design modifications to existing ditches to enhance treatment performance.

The end-goal of this research was to examine the feasibility of using modified ditches to

mitigate water contamination, thereby offering rural communities a simple and cost-effective

approach for improving wastewater management in a region that currently lacks the resources for

more conventional treatment approaches. Studying the effects on water quality of the physical,

chemical and biological features in wastewater drainage ditches both before and after

modifications may yield design and management implications that are relevant not only for









dealing with wastewater discharges but also for other point and non-point sources of pollution

draining into ditch systems.

Study Setting

The setting for this study was Oxapampa, Peru (100 35'S, 750 24' W) located at an

elevation of 1,814 m above sea level in the Andean Amazon (Figures 1-1 and 1-2). The

population of Oxapampa is approximately 13,400 with 58% of inhabitants residing in the urban

center (Figure 1-3) and the remaining 42% in the rural periphery of town (CEPID 2002).

Oxapampa experiences minimal seasonal temperature fluctuations; average temperatures range

between 15 and 29 OC (Municipalidad Provincial de Oxapampa 2003). Total annual precipitation

is 1,400 mm per year, 80% of which falls between November and April (EPS Selva Central S.A.

2000).

The economy of Oxapampa is predominately agrarian. Important agricultural products

include coffee, chili peppers, avocados, plantains and squash. Oxapampa is a major contributor

of livestock production with over 4,910 head of cattle (CEPID 2002). These practices influence

water quality as storm events flush sediments, animal wastes, fertilizers and pesticides into

drainage ditches that are ultimately discharged to the Chorobamba River. A local slaughterhouse

facility is located on the river bank and discharges its wastes directly without treatment.

The Chorobamba River is used extensively by local inhabitants for bathing, laundry,

washing vehicles, recreation, fishing, and in some cases human consumption. In 2000, 55% of

Oxapampa residents had access to running water from the San Alberto watershed. Water

treatment consists of a settling tank and chlorination. In 2004 this number increased to 67%

(9,000 people). In 2004, approximately 1,170 homes had running water and a sewage

connection, 33 1 houses had running water only, and 200 families had neither of these services

(Arteaga, personal communication). Water usage rates are high (approximately 500 L per capital









per day) as the result of wasteful consumption practices (e.g. leaving taps open) and leaking

pipes (Arteaga, personal communication).

The city of Oxapampa currently has no wastewater treatment facility or household septic

tanks. Sewage disposal consists of discharging both grey and black waters (including solids) into

open, earthen and vegetated ditches. Approximately two-thirds of the wastewater generated in

Oxapampa is discharged to open ditches that drain to the Chorobamba River. The remainder of

the wastewaters is discharged directly to the Chorobamba River via networks of underground

pipes. Both pipes and ditches function as combined sewers during storm events. The linear

distance of open wastewater ditches and underground pipes is 18 km and 8 km, respectively

(Arteaga, personal communication, Figure 1-4). The pipes and ditches discharge an average of

150 L sl to the Chorobamba River, representing 1 to 2.5% of the total river discharge during the

dry season (Figure 1-5).

While these percent contributions of wastewater to the river are small, they nevertheless

have resulted in changes to the ecological integrity of the river, notably reductions in fish

populations and increasing growths of filamentous algae in shallow channels, presumably from

excess loads of phosphorus (Municipalidad Provincial de Oxapampa 2003).

The impact of untreated wastes on the health of citizens is reflected by morbidity reports.

Over 60% of children under the age of five years suffer from either acute or chronic diarrhea.

Gastrointestinal diseases and parasites account for 37% of all hospital visits (CEPID 2002). Of

the total population, approximately 40% are under the age of 15 and are thus at greater risk of

waterborne diseases. Oxapampa has the only hospital facility within 1.5 hours.

Water Quality in the Chorobamba River

To examine the effect of wastewater discharges on water quality, the Chorobamba River

was sampled on multiple occasions in years 2005 and 2006 at one site upstream of the city and at









three consecutive sites downstream of Oxapampa. Grab samples of water were collected at each

site at several locations along a channel cross-section and composite. Water samples were

analyzed for E. cobi concentrations, soluble reactive phosphorus (SRP), ammonium (NH4-N) and

nitrate (NO3-N) following the procedures detailed in Chapter 3. Results of these analyses

indicated that wastewater discharges from the city negatively impact water quality in the

Chorobamba River by elevating pathogen concentrations (Figure 1-6) and nutrient levels

(Figures 1-7 and 1-8) downstream of the city. The health threat posed by fecal contamination is

illustrated by E. cobi concentrations that consistently exceeded limits established by the USEPA

for recreational waters.

Water Quality in Wastewater Ditches and Pipes

To examine water quality characteristics of wastewater effluents entering the river, water

samples were collected from discharge points of three underground pipes (Grau, Nueva Berna

and Santa Domingo) and three ditches (Gustayson, Koell and San Martin). Samples were

analyzed for E. cohi, total suspended solids (TSS), 5-day carbonaceous biochemical oxygen

demand (CBODS), SRP, NH4-N and NO3-N following analytical procedures described in

Chapter 3. Parameter concentrations in ditches were corrected for dilution by accounting for the

percentage of total flow that originated from natural streams (approximately 45%, 37%, 20% and

5% in San Martin, Gustayson, Koell and Frey, respectively). This correction was performed to

distinguish possible in-stream treatment processes from dilution to provide a fair comparison

against pipe water quality. Differences in water quality between pipes and ditches were

determined using t-tests on In-transformed data. Concentrations of dilution-corrected E. cohi,

TSS, CBODS and SRP were found to be significantly lower (p < 0.01) in ditch effluents

compared to pipe effluents (Figures 1-9 through 1-13).










There is little reason to believe that waste-streams or water usage differs substantially in

the town leading to the differences in water quality observed between ditches and pipes (Arteaga,

personal communication). These findings thus imply that in-stream retention mechanisms

influence concentrations ofE. coli, sediments, oxygen demand and nutrients during transport in

ditches. This central hypothesis forms the basis for the remainder of this study as described

below.

Research Approach

A general assessment was made of the potential for treatment to occur in drainage ditches

by synthesizing findings from the literature. The conceptual framework for understanding

potential treatment mechanisms occurring in ditches was then applied to several wastewater

drainage ditches in Oxapampa, Peru. The potential for treatment in wastewater ditches was first

examined with respect to the function of vegetation for influencing transport characteristics,

sedimentation processes and nutrient uptake. The significant differences in SRP concentrations

observed between ditches and pipes prompted an evaluation of the role of benthic sediments for

mitigating phosphorus exports via sorption processes. Finally, utilizing insights gained from

studying treatment influences of plants and sediments, channel modifications were implemented

in an effort to amplify treatment processes.

























~ifi


i,


Ioca tion:
Oxap~ampla,
SPerui


r!

1
F_



















N

W P"


I_


11I
500


1111
1,000 Kilometers


Figure 1-1. Map of Peru showing the location of Oxapampa, situated in the Andean Amazon at
an elevation of 1,814 masl. Map created using ArcGIS 9. 1 using ESRI' s publicly
available Shaded World Relief metadata (ESRI Data & Maps, 2004)







































gure I-z. view or tnle L~norocamoa Klver valley ano tnle city or uxapampa


gure 1-3. The urban center of Oxapampa and the Chorobamba River












Oxapampa, Pleru
Scale 1:5000


J,


I
~-t-r+'
)I
------- '~~


JM'" 11..U.~i
..


Figure 1-4. Simplified map of the urban center of Oxapampa, indicating the locations of
wastewater ditches (solid lines) and underground pipes (dashed lines) all of which
drain to the Chorobamba River


i
~
~















150



S100





50


18-Dec-01 18-Dec-02 18-Dec-03 18-Dec-04 18-Dec-05
Date

Figure 1-5. Daily discharge in the Chorobamba River between December 2001 and January 2006
(Note: missing data between September and December 2004; data source:
AARS/ProPachitea)




1200



800


8 5 / 235 CFU 100 mL-1
= 400 Water quality
oO
a standard applied to
L~i-1 recreational waters
(EPA 2003)

1 2 3 4

Sampling site
Figure 1-6. Summary of E coli concentrations in the Chorobamba River (2005-2006), indicating
that sampling sites downstream of Oxapampa (sampling sites 2-4) exceed maximum
allowable E. coli concentrations for recreational waters. Points are average values
(n = 6) and bars are +1 SD.














0.15


0.1


0.05


Sampling site
Summary of dissolved inorganic nitrogen (sum of nitrate and ammonium)
concentrations in the Chorobamba River (2005-2006), showing that sites downstream
of Oxapampa (sampling sites 2-4) have elevated concentrations relative to upstream
of the town (Site 1). Points are average values (n=6) and bars are 11 SD.


Figure 1-8.


Sampling site
Figure 1-7. Summary of SRP concentrations in the Chorobamba River (2005-2006), showing
that sites downstream of Oxapampa (sampling sites 2-4) have elevated SRP
concentrations relative to upstream of the town (Site 1). Points are average values
(n=6) and bars are 11 SD.



0.5 1






















Ccl I,


B to


FF ~tl


Pipe sites


50000

40000

30000

20000

10000

0


nE


Ditch sites


Pipe sites


Figure 1-9. Dilution-corrected E. coli concentrations in ditch effluents (A) compared to E. coli in
pipe effluents (B) where filled circles represent median values(n=12), open circles are
outliers, boxes delineate inter-quartile range (IQR), and whiskers are 1.5* IQR


250

j 200

L150

S100


Ditch sites


Figure 1-10. Dilution-corrected carbonaceous biochemical oxygen demand (CBODS)
concentrations in ditch effluents (A) compared to CBODS in pipe effluents (B) where
filled circles represent median values (n=5), open circles are outliers, boxes delineate
inter-quartile range (IQR), and whiskers are 1.5* IQR





TI


250

,200

E 150
100
50oo r

0o


t




n


Pipe sites


Ditch sites


Figure 1-11. Dilution-corrected total suspended solids (TSS) concentrations in ditch effluents (A)
compared to TSS in pipe effluents (B) where filled circles represent median values
(n=12), open circles are outliers, boxes delineate inter-quartile range (IQR), and
whiskers are 1.5* IQR


6


E~4


2

0


Ditch sites


Pipe sites


Figure 1-12. Dilution-corrected soluble reactive phosphorus (SRP) concentrations in ditch
effluents (A) compared to SRP in pipe effluents (B) where filled circles represent
median values (n=12), open circles are outliers, boxes delineate inter-quartile range
(IQR), and whiskers are 1.5* IQR
















;





O'~lo~


Pipe sites


Ditch sites


Figure 1-13. Dilution-corrected ammonium (NH4-N) concentrations in ditch effluents (A)
compared to NH4-N in pipe effluents (B) where filled circles represent median values
(n=12), open circles are outliers, boxes delineate inter-quartile range (IQR), and
whiskers are 1.5* IQR









CHAPTER 2
LITERATURE REVIEW

Introduction Ditches in the Landscape

Ditches are ubiquitous features in altered landscapes. Their presence grows in proportion

to the expansion of anthropogenic land use and as a consequence, ditches have become an

inevitable component of ecosystems around the world. In developed nations, their primary

purpose is to route water for irrigation or to improve drainage from agricultural fields and

roadways. Drainage ditches are equally pervasive in many parts of the developing world;

however, in the absence of wastewater treatment facilities, drainage ditches often serve as the

primary transport conduits for untreated human sewage.

Ditches are designed and managed to be efficient at moving water across the landscape and

as a consequence, are commonly perceived as little more than conduits for routing water and

materials. Until recently, little interest had been given to understanding and promoting potential

ancillary functions of drainage ditches. However, there is growing evidence that ditches provide

services beyond basic water transport. Studies have shown that some ditches serve as important

habitat for maintaining species richness (Armitage et al. 2003), function as wildlife corridors

(Mauritzen et al. 1999, Mazerolle 2005) and provide recharge of surficial aquifers (Fernald and

Guldan 2006).

The position of ditches in the landscape suggests their unique importance as interface

ecosystems (Odum and Odum 2003), connecting terrestrial and aquatic environments and

controlling the timing and magnitude of terrestrially-derived contaminant exports to downstream

water bodies. Surprisingly little attention has been given to investigating the treatment function

of drainage ditches despite their downstream connection with receiving water bodies. While

there are studies that point to discharges from drainage ditches as maj or sources of water quality









contamination (e.g. Hunt et al. 1999, Fletcher et al. 2004, Cahoon et al. 2006) there is also

growing evidence that some ditches mitigate negative impacts to downstream water quality.

Ditch networks may be viewed as low-order stream systems that are often responsible for

draining a disproportionate area of many watersheds, especially in agricultural landscapes

(Alexander et al. 2000). The association between land-use practices, diffuse pollution and the

continued deterioration of water quality is well recognized and has been discussed in great detail

in the literature (see Howarth et al. 1996). Despite the large number of studies addressing runoff

issues, mitigating the effects of diffuse pollution remains a somewhat elusive goal in practice

(Sharpley and Tunney 2000, McDowell et al. 2004). A number of best management practices

(BMPs) such as buffer strips have been recommended to reduce runoffloads to ditches. It makes

good sense to couple these load-reducing practices with ditch design and management

approaches that amplify the abilities of drainage systems to retain pollutants (Cooper et al. 2004,

Dabney et al. 2006).

The obj ective of this review is to provide a general, conceptual framework for

understanding potential retention mechanisms in ditches and to synthesize findings from the

literature that have reported on the treatment capabilities of drainage ditch systems. The purpose

of this exercise is to gain an understanding of the degree to which ditches have been shown to

mitigate contaminant exports to receiving bodies and to examine the conditions under which

retention has been shown to occur. The term retention is used here to mean the difference

between imports and exports in a system resulting from transformation and storage processes

which ultimately affect the timing and magnitude of downstream exports. This review also

highlights where future research efforts should be focused to expand upon our current

understanding of the treatment function of drainage ditches.









The studies covered in this review are strongly biased towards agricultural drainage ditches

because this is the setting in which the vast majority of the studies focusing on contaminant

processing in ditches have been conducted. Regardless of this bias towards agricultural ditches,

the lessons learned from examining the physical and biogeochemical processes taking place in

agricultural systems may be transferred and applied to other settings such as roadside ditches and

ditches associated with aquaculture, forestry, industrial practices and even with routing of

wastewater effluents. The existence of open channels for routing untreated sewage is poorly

documented. However, in regions such as the Andean Amazon of Peru, these ditches are

currently the norm for handling wastewaters. Studies of the treatment function of ditches thus

yield important implications for improved water management across diverse land-use types and

across the globe.

For purposes of clarification, ditches will be referred to as long, linear water conveyance

features that are earthen, unlined and vegetated (unless managed otherwise). Of the 22 studies

identified in this review that examined contaminant retention in ditches, six employed

approaches developed originally to understand material cycling in stream ecosystems and

referred to their study systems as streams rather than ditches (Macrae et al. 2003, Royer et al.

2004, Schaller et al. 2004, Bernot et al. 2006, Ensign et al. 2006, Gucker and Pusch 2006). Ditch

characteristics are as diverse as the settings in which they are found (Moore et al. 2005), thus

complicating their strict definition. However, if the systems are channelized and convey drainage

waters from altered landscapes, such as agricultural lands, they are referred to here as ditches.

The take-home message from this review is that the consideration of drainage ditches in the

landscape should not be limited strictly to their conveyance function of water and materials but










should more broadly recognize their position in the landscape as important interface ecosystems

between the terrestrial environment and downstream water bodies.

Conceptual Basis for Pollutant Retention in Ditches

Characteristics of Ditch Ecosystems

It is not surprising that some ditches possess the capacity for significantly reducing

concentrations of contaminants such as pesticides, nutrients and pathogens. The treatment

function of ditches arises from their physical, biological and chemical characteristics, many of

which are shared with other systems recognized for their importance in pollutant retention.

Many earthen and vegetated ditches share important physical characteristics with

headwater streams. For example, many ditches have relatively shallow water depths and thus

tend to have high surface area to volume ratios, which may enhance biogeochemical reactions

(Alexander et al. 2000) resulting in increased rates of denitrification, sorption to reactive

substrates and assimilation by plant and algal communities. High nutrient loads and colonization

by emergent and submerged aquatic plants often yield a system whose biological features are

reminiscent of nutrient-rich treatment wetlands. Accumulation of plant litter and deposition of

fine organic-rich materials creates a benthic environment with redox conditions that are more

similar to a wetland than to a stream system. However, ditches are lotic systems and experience

pulses of water and materials that may frequently disturb ecosystem structure thus re-setting or

maintaining the system in early stages of succession. This blend of physical, biological, and

chemical attributes creates a unique ecosystem consisting of hydrogeomorphic features common

to many headwater streams and chemical and biological characteristics reminiscent of treatment

wetlands.









Contaminant Exposure and Retention Efficiency

The physical, chemical and biological characteristics expressed by a ditch reflect the

conditions to which it is exposed. Ditch exposure to contaminants is sometimes direct such as

during the application of herbicides to control vegetation. Other times, it is incidental, as in the

case of drift during aerial applications of fertilizers or pesticides. Most often, however, exposure

in ditches occurs when there is water movement, particularly during storm events that generate

runoff draining to ditches or from subsurface tile drains that remove excess water to improve

agricultural field conditions. For wastewater drainage ditches, exposure is often greatest during

hours of peak water use. The patterns of water flow within a ditch thus dictate the regularity and

the intensity of the contaminant exposure.

The exposure regime of potential contaminants in a ditch governs its treatment efficiency.

The exposure regime in a ditch consists of two important factors that function as controls on

pollutant retention: discharge and contaminant loading. Discharge governs the residence time

within the ditch. Meyer and Likens (1979) proposed that discharge controls whether a stream

system is operating in a "processing mode" or a "throughput mode." In the processing mode, the

system is efficient at storing, utilizing and transforming potential environmental pollutants. This

mode tends to occur at lower flows when contact time is greatest, thereby promoting more

physical and biogeochemical interactions during transport. For higher rates of flow, the system

switches to a throughput mode, whereby potential contaminant retention mechanisms are

bypassed due to low residence times in the channel.

Contaminant loading is the second important control on the efficiency of pollutant

retention. The effect of loading on treatment efficiency can be conceptualized using the subsidy-

stress gradient theory of Odum et al. (1979). A ditch ecosystem may initially become stimulated

by the availability of macronutrients or organic materials until a certain threshold is reached at









which the loading exceeds the processing capacity of the system (see Figure 2-1). At this point,

loadings no longer serve to subsidize system metabolism but instead become a stressor to system

and reduce its processing efficiency. In this sense, contaminant loading vis-a-vis the subsidy-

stress gradient theory may embody a switch between a processing and throughput mode similar

to that hypothesized for discharge by Meyer and Likens (1979).

The subsidy-stress gradient of Odum et al. (1979) also considers inputs of toxic

substances (not depicted in Figure 2-1), which in ditch ecosystems could represent toxicity limits

of pesticides and metals in plants and other organisms. In the case of either a usable or a toxic

input to the system, once the loading has exceeded the processing capability of the system, the

system is no longer efficient at pollutant retention and instead switches to a throughput mode,

thus exporting environmental pollutants to downstream water-bodies (Haggard et al. 2001, Marti

et al. 2004, Gucker and Pusch 2006).

The obvious implication behind these two hypotheses is that ditches will be of limited use

for reducing pollutant loads to downstream water-bodies if they are overwhelmed by high

discharges and contaminant loadings. It would be insightful to develop a general understanding

of the discharge and loading levels at which this "switch" between operating in a processing

versus a throughput mode tends to occur in order to predict the treatment reliability of a system

under different conditions. However, such thresholds may be very site-specific and thus less

appropriate for extrapolating to other drainage ditch systems. Perhaps a more widely-reaching

contribution by scientists would be to investigate approaches to better modify and manage

ditches with the obj ective of maintaining or extending the effective operating range of the

processing mode under elevated discharge or loading scenarios. Modifications and management










affect and re-organize system structure. Therefore, for such approaches to be effective, it is

imperative to understand how the structure of ditches influences its treatment function.

Effect of Ditch Structure on Contaminant Retention

In a very broad sense, ditch ecosystems can be viewed as consisting of two dominant

physical/biological components: vegetation and benthic substrates. These two components affect

and are affected by flows of water, materials and energy thereby creating a dynamic three-way

relationship between water flow, sediment transport and plant growth (Figure 2-3, Clarke 2002).

It is within this dynamic relationship that the potential for contaminant retention exists via

transformations and storage governed by plants and sediments that control contaminant fate and

transport in ditch ecosystems.

In many ditch ecosystems, vegetation and often to a lesser degree, sediments are removed

to improve drainage and prevent the risk of flooding. This review emphasizes the role of

vegetation in ditches, not only because this is the component most easily controlled by managers,

but as depicted in Fig. 2-2, the removal of vegetation typically results in the subsequent export of

sediments .

Vegetation in Ditches

Ditches are often unshaded, tend to be relatively shallow and often experience low water

velocities except perhaps during pulse events such as storms. These characteristics, often in

conjunction with high concentrations of bioavailable nutrients, support abundant growth of in-

stream vegetation. In-stream vegetation is a distinctive feature of many drainage ditches;

therefore, understanding how vegetation influences the fate and transport of contaminants is

particularly germane in ditch ecosystems.

While much attention has been paid to the treatment function of vegetation in constructed

wetland ecosystems (Brix 1997, Richardson and Qian 1999, Dierberg et al. 2002, Karathanasis et









al. 2003, Collins et al. 2005), littoral and riparian zones (Tabacchi et al. 1998, Spruill 2000, Jia et

al. 2006), and other wetland ecosystems (Odum et al. 2000), relatively few studies have

specifically examined this function within lotic ecosystems (Clarke 2002). Perhaps this is

derived from a perception of limited plant-water contact time in streams and ditches compared to

very low-velocity systems. Another explanation that vegetation is often overlooked for uptake of

organic contaminants such as pesticides in ditches is that partitioning is often assumed to take

place only between the water and sediments (Cooper et al. 2004). However, increasing numbers

of studies are examining the effects of vegetation in flowing-water ecosystems and reporting that

the presence of vegetation yields significant implications for contaminant fate and transport (e.g.

Wilcock et al. 1999, Champion and Tanner 2000, Clarke 2002, Salehin et al. 2003, Cooper et al.

2004, Scholz 2005, Dabney et al. 2006).

The potential treatment processes promoted by the presence of ditch vegetation are varied

and interrelated. Many studies have shown the importance of direct uptake by plants for removal

of potential environmental contaminants such as nutrients (e.g. Brix and Schierup 1989, DeBusk

et al. 1995, DeBusk et al. 2001). However, release of assimilated nutrients during plant

decomposition often makes this retention process ineffective unless plants are periodically

harvested (Reddy et al. 1995, DeBusk et al. 2001). Often, the role of vegetation for contaminant

retention in aquatic systems is less by their direct uptake and more by the conditions they create

which in turn promote retention (Brix 1997, Sobolewski 1999, Kim and Geary 2001, Schulz et

al. 2003). Arguably the most important contribution of in-stream vegetation for retention is the

attenuation of water velocity.

Effect of Vegetation on Velocity Attenuation and Transient Storage

Stands of emergent and submerged vegetation increase channel roughness (Wilcock et al.

1999, Champion and Tanner 2000, Madsen et al. 2001) and provide structural complexity










(Clarke 2002, Salehin et al. 2003, Grimm et al. 2005) that serve to attenuate water velocity and

increase solute residence time. Flow resistance by transport through and around stands of

vegetation results in flow becoming generally deeper and slower. Patches of vegetation increase

channel heterogeneity, creating eddies and pockets of slower moving water that impede solute

transport. This delay of downstream advective transport (or hydrologic retention) due to time

spent in stream channel quiescent zones is termed in-channel transient storage (Bencala and

Walters 1983).

The impedance of flow due to stands of in-stream vegetation may also promote the

subsurface component of transient storage, called hyporheic exchange. Hyporheic exchange is

the exchange of surface water with shallow groundwater (Harvey and Wagner 2000) due to

increased hydraulic pressure which drives water into and through porous channel beds, a process

termed advective pumping by Woirman et al. (2002). However, the importance of hyporheic

exchange in many drainage ditches may be severely limited by the presence of fine sediments

that clog interstices and impede the transfer of water across the hyporheic zone (Grimaldi and

Chaplot 2000, Hancock 2002). Hyporheic zones may also have reduced importance in ditches as

a result of channel straightening which disconnects the stream channel from parafluvial zones,

reduces channel complexity and increases water velocities (Hancock 2002).

The capacity of in-stream vegetation to attenuate flow velocities and promote transient

storage is predicated on macrophyte morphology and growth characteristics (Clarke 2002, Dodds

and Biggs 2002, Schulz et al. 2003). For example, submerged macrophytes with finely divided

leaves that tend to grow in dense stands will exert greater influence over flow and sedimentation

patterns than those in more open stands with broader or ribbon-like leaves (Sand-Jensen and

Mebus 1996, Clarke 2002). While there is no single variable that can be used to determine









velocity attenuation of different community types under varying hydrologic conditions, Dodds

and Biggs (2002) found that the mass density of macrophytes and periphyton served as an

adequate proxy, explaining 81% of the variance in velocity attenuation in their study. However,

it is necessary to note that the presence of macrophytes in lotic systems does not always ensure

greater velocity attenuation. The velocity attenuation capacity of macrophytes is linked to the

spatial distribution of their stands as illustrated by Wilcock et al. (2004) who found that emergent

plants colonizing stream margins reduced the channel cross-sectional area, resulting in increased

water velocities. In contrast, submerged plants growing within the channel served to impede

flow, thereby reducing velocities and increasing water depths.

Influence of Hydrologic Retention on Water Quality

Hydrologic retention, the average time spent by solutes in transient storage zones per unit

downstream distance (Harvey et al. 2003), increases solute contact (in time and space) with

biogeochemically reactive substrates and has been linked to significant changes of downstream

water chemistry (Mulholland et al. 1997, Harvey and Fuller 1998, Hill et al. 1998). The

importance of in-stream obstructions such as vegetation and woody-debris dams for increasing

channel residence time and promoting contaminant retention has been documented in recent

years (Valett et al. 2002, Wilcock et al. 2004, Ensign et al. 2006). Wilcock et al. (2004) showed

that ammonium retention in four agricultural streams with varying coverage of macrophytes

could be approximated by the stream length (X) and velocity (v) such that % retention X v.

The link between transient storage and nutrient retention in streams is commonly made

using nutrient spiraling methods (Newbold et al. 1981, Stream Solute Workshop 1990). In this

approach, solute inj section experiments are performed to fit a 1-D transport model to observed

solute breakthrough curves (BTCs). Several transport models have been used in this approach

and all incorporate transient storage parameters that relate exchange rates and size of storage









zones (see Stream Solute Workshop 1990, Runkel 1998). Conservative solute BTCs allow the

estimation of transient storage and other hydraulic parameters while changes in the dilution-

corrected concentrations of reactive solutes with downstream distance provide evidence of

retention processes occurring during transport. Spiraling metrics, such as uptake length, uptake

rate and uptake velocity, have been used as proxies for the "self-purifying capacity" of stream

ecosystems (Marti et al. 2004) and are used to compare nutrient retention over time, space and

between different systems.

Nutrient spiraling theory has its roots in understanding and comparing nutrient dynamics in

pristine stream systems. However, the methodology has been applied in recent years to eutrophic

systems as well (Marti et al. 2004, Royer et al. 2004, Haggard et al. 2005, Bernot et al. 2006,

Ensign et al. 2006, Gucker and Pusch 2006). The nutrient spiraling concept by itself does not

reveal the mechanisms responsible for retention but instead provides a "whole-reach" measure of

retention with respect to the solute in question. This approach has allowed nutrient retention in

drainage ditch systems to be compared with that in less-altered stream systems.

Ensign et al. (2006) used this approach to compare spiraling metrics on various dates to

create an empirical model to predict uptake in four drainage systems under various discharge

regimes. The authors reported spiraling metrics that fell within the same ranges reported for

headwater streams (see Ensign and Doyle 2006) and predicted relatively high retention (65 and

37% for ammonium and phosphate, respectively) during storm flows when retention tends to be

lowest. These encouraging results contrast with results from other ditch spiraling studies which

concluded that despite high nutrient uptake rates and uptake velocities, overall nutrient retention

in ditches tended to be low relative to high contaminant loads (Table 2-1, Macrae et al. 2003,

Royer et al. 2004, Gucker and Pusch 2006).









Low nutrient retention in ditches is often hypothesized to be the result of the lack of

channel structural complexity (e.g. Grimm et al. 2005, Bernot et al. 2006) that limits retention

times in transient storage zones and availability of reactive surface areas. To explicitly

investigate the relationship between channel complexity, transient storage and nutrient retention,

Ensign and Doyle (2005) performed channel manipulations in an agricultural drainage canal and

a channelized blackwater stream in North Carolina, USA. Transient storage was measured under

3 different channel conditions: unaltered, removal of all vegetation and coarse woody debris

(CWD) and the addition of PVC baffles. Removal of vegetation and CWD reduced transient

storage zones by 61% and 43%, while baffles increased storage zones by 227% and 119%.

Phosphate and ammonium retention were quantified using the uptake velocity (V7) spiraling

metric, which represents the relative demand for the solute by the benthos (Bernot et al. 2006).

After vegetation removal, Vfvalues for phosphate and ammonium decreased by 38% and 88%

and with baffles in place they increased by approximately 3000% (from -1.5 to 53 mm min )

and 143%. The Ensign and Doyle (2005) study supports the importance of maintaining structural

complexity for improved drainage ditch treatment function. The following sections examine

other study results to provide an overview of the effects of sedimentation and contact with

reactive substrates in ditches for contaminant processing and retention.

Sedimentation

Suspended sediments and BOD removal. Sedimentation is an important retention

mechanism in ditches as many potential contaminants are in particulate form or are associated

with sediment surfaces. Lecce et al. (2006) reported that vegetated ditches in North Carolina

functioned as efficient sediment traps, reducing sediment exports from the basin by 2,978 Mg yrl

or 3.83 Mg ha- yr- Hargreaves (2005) reported that approximately 97%, 67% and 69% of high

loads of TSS, TN and TP were removed within a 150-200 m distance along a vegetated ditch in









Mississippi, USA. Scholtz and Trepel (2004) reported decreases in BOD from approximately 15

mg L-1 to less than 1 mg L^1 within 90 m. However, between 90 and 150 m BOD values tended

to fluctuate with final concentrations around 6.0 mg L^1 on average. Unless particulate matter is

buried and incorporated into the sediment (Saunders and Kalff 2001), the process of

sedimentation unfortunately does not ensure permanent contaminant removal. Erosion can re-

suspend and transport particulate and sediment-bound contaminants to downstream water-bodies.

The presence of in-stream vegetation reduces the potential for re-suspension (Madsen et al.

2001) but during senescence this function decreases markedly (Schulz et al. 2003).

Besides removing suspended materials from the water column, sedimentation can promote

other retention processes by allowing deposition and accumulation of organic matter within and

downstream of plant stands (Schulz et al. 2003). The accumulation of organic matter in addition

to the presence of root systems increases the porosity and permeability of benthic sediments and

therefore can affect sediment pore water fluxes (Salehin et al. 2003). Sedimentation in the

presence of vegetation thus results in significant changes to the chemistry of benthic sediments

leading to important implications for contaminant retention.

Retention processes influenced by interactions with the benthos

Sedimentation and deposition of seston and plant litter occurring in the presence of

macrophytes or woody debris dams create benthic environments that set in motion important

processes such as denitrification and sorption.

Nitrate removal. Denitrification is often the primary mechanism for nitrogen retention in

aquatic systems (Howarth et al. 1996, Alexander et al. 2000) and is of particular importance

because it represents a permanent loss from the system. Saunders and Kalff (2001) in their

review of nitrogen retention of wetlands, lakes and rivers found that from the sites studied, rivers

provided the least amount of N retention and wetlands the greatest. Hence, rates of retention due









to denitrification in vegetated ditches are likely to fall somewhere in the middle. Denitrification

rates in streams are often limited by contact time, zones of anoxia and the availability of organic

carbon or nitrate (Hedin et al. 1998). Macrophyte presence in ditches provides extensive surface

areas for microbial growth, increases residence time and promotes the accumulation of organic

matter thus providing the necessary carbon source. The importance of promoting this retention

mechanism is highlighted in agricultural ditches because TN loading to ditches is often

dominated by nitrate loads that have bypassed riparian controls due to the use of subsurface tiles.

Reports concluding on the effectiveness of denitrification in ditches are varied and appear

in some cases to be largely a function of discharge. Scholtz and Trepel (2004) reported high

losses of nitrate (~79%), presumably due to denitrification, along a 150 m reach in Germany.

Royer et al. (2004) and Schaller et al. (2004) examined denitrification under several discharge

regimes in Illinois and concluded that although denitrification rates tended to be high, retention

was insignificant since the highest nitrate loads coincided with highest flowrates Nitrate

concentrations for all 3 studies were within the same ranges (13-18 mg L^)~; however flowrates

reported by Scholtz and Trepel (2004) were ~0.5 L s^l whereas those for Royer et al. (2004) and

Schaller et al. (2004) were as high as 1,840 and 1,370 L s^l. High nitrate loads in conjunction

with high flows switches the system to a throughput mode (as noted by Royer et al. 2004)

rendering in-stream processes incapable of mitigating downstream exports. Bernot et al. (2006)

also reported that nitrate demand in six ditches in Indiana and Michigan appeared to be saturated;

however no significant correlations were found between retention and nitrate concentration or

discharge. The results from these studies suggest that while the potential for high nitrate

retention exists in ditches, it is likely to be severely compromised under high flowrates that limit

residence time, especially in systems with extremely limited channel complexity.










Hyporheic exchange is often considered to be of minor significance for improving water

quality in ditches due to channel straightening and clogging ofinterstices by fine sediments.

However, it has been shown that nitrate can be depleted by denitrification within the first few

centimeters (Hill et al. 1998, Revsbech et al. 2005), making even small vertical fluxes potentially

significant. No studies were identified that examined the role of hyporheic exchange specifically

in drainage ditches; however, LeFebvre et al. (2004) reported high rates of denitrification in a

straightened river reach in France. Denitrification rates were correlated to the presence of high

total organic matter and fine sediments resulting from direct sediment exports from surrounding

fields.

Ammonium retention. High concentrations of ammonium are common in ditches

receiving waste inputs of manure or with high rates of decaying plant material (e.g. Dukes and

Evans 2006). Retention of ammonium is an important goal in many ditch systems because of

ammonia toxicity to aquatic organisms. Macrophytes translocate oxygen to their root systems

providing aerobic microsites that promote mineralization of organic N and nitrification of

ammonium thus setting the stage for coupled nitrification-denitrification interactions (Reddy et

al. 1989). Ammonium may also be utilized by algae and heterotrophic microbes that make up

reactive biofilms on sediments and other benthic substrates. Physical adsorption of ammonium to

sediment surfaces is most important in pristine systems with low background ammonium

concentrations (Triska et al. 1994).

For ditches, the more likely pathway for ammonium retention is nitrification-denitrification

and temporary storage by macrophytes and biofilms. Findings from Gucker and Pusch (2006)

support this hypothesis by reporting that nitrification represented the dominant ammonium

retention mechanism in channelized, eutrophic systems in Germany. Scholtz and Trepel (2004)










reported no net change in ammonium concentration along ditches in northern Germany despite

increasing dissolved oxygen levels with downstream distance. One possible explanation is that

decay of plant material contributed ammonium to the water column; this hypothesis is supported

by fluctuations in BOD. Similar to the results reported for nitrate enriched systems by Royer et

al. (2004), Schaller et al. (2004), Inwood et al. (2005) and Grimm et al. (2005), Gucker and

Pusch (2006) found that while absolute uptake rates of ammonium and nitrate were higher in the

eutrophic systems than in pristine streams, the percent retention was minimal due to

comparatively much greater N loads.

Phosphorus retention. Retention of phosphorus (P) presents another formidable challenge

in drainage ditch systems. While biological assimilation is an important retention mechanism, it

is short-term as P stored in tissues eventually becomes re-mineralized and re-released to the

water column (DeBusk et al. 2001). Dierberg et al. (2002) reported that high water column pH

occurring during photosynthesis by submerged aquatic vegetation resulted in co-precipitation of

calcium and phosphorus that removed 50-79% of P in constructed wetlands for agricultural

runoff. However, in most stream and wetland systems, sediment sorption is the dominant

retention process (Reddy et al. 1995).

The capacity of benthic sediments to temporarily retain dissolved P from the water column

by sorption is governed by physiochemical properties of the sediment such as soil composition

(mineralogy and texture), pH, organic matter content and redox conditions (Reddy et al. 1995,

Sallade and Sims 1997a, Reddy et al. 1999, Nguyen and Sukias 2002). Biotic processes

occurring on sediment surfaces may enhance sorption capacity (Haggard et al. 1999). Nguyen

and Sukias (2002) determined fractions of sorbed P in 26 agricultural drainage ditches in New

Zealand to determine the extent of and mechanisms responsible for sediment P sorption.









Although 64-68% of sediment surfaces were found to be saturated, the ditches still demonstrated

high P retention capacity by removing 44-84% of the 5,000 mg P kgl added during P buffer

capacity experiments. The authors speculated that large presence of fine particles (63-67%) and

high organic carbon concentrations (3.8-14. 1%) likely resulted in sorption of Al and Fe which

could have formed humic-Fe/Al complexes that are effective at promoting even greater P

sorption. From the P fractionation analysis, Ngyugen and Sukias (2002) found that

approximately half of P sorbed represented a temporary storage as those fractions were loosely

bound and hence could be re-released to the water column. However, some of the ditches were

also long-term sinks as 6% to 39% of total sorbed P was stored as residual P.

The concentration gradient between water column soluble reactive P (SRP) and sorbed

sediment P controls whether sediments release or remove phosphorus from the water column.

The aqueous concentration at which P is neither released nor sorbed from sediments is termed

the equilibrium phosphorus concentration (EPCo). This value is used as a benchmark to

determine whether sediments tend to function as a sink or a source of SRP to the water column.

If SRP is greater than EPCo, sediments act as a temporary P sink and vice-versa. Smith et al.

(2005) reported that the addition of alum to drainage ditches reduced EPCo values (but not the P

buffering capacity) due to phosphorus precipitation of with aluminum. This precipitate is

generally more robust than that of iron because under reducing condition ferric iron is reduced

and releases P to the water column. In a separate study Smith et al. (2006) found that alum

treatments were not effective at reducing exchangeable P and EPCo and hypothesized that a

storm event mobilized alum-treated sediments and deposited new sediments. While the addition

of alum is common to water treatment facilities and in controlling P concentrations in lakes, the

practice may not prove economically viable to many farmers or land managers.









Pesticides and heavy metals retention. While the organic matter content of sediments has

been shown to promote greater sorption of P, it is also responsible for promoting retention of

heavy metals (Gambrell 1994, Debusk et al. 1996, Hammer and Keller 2002) and pesticides

(Margoum et al. 2006). Retention of metals and pesticides have also been linked to direct uptake

by macrophytes and sorption to macrophyte-associated substrates (e.g. Moore et al. 2001, Collins

et al. 2005). While no studies on heavy metal retention in vegetated ditches were found, grassed

swales are nevertheless common BMPs for retaining stormwater and have been shown to be

moderately effective at reducing sediment-bound heavy metals (Backstrom 2003, Zanders 2005).

Few studies have explicitly studied pesticide retention in drainage ditches; however, the

studies that do exist illustrate that vegetated ditches show great promise as viable BMPs for

mitigation of pesticide exports. Bennett et al. (2005) found that ditch lengths of 120 and 280 m

were sufficient to reduce two pyrethroid pesticides (bifenthrin and lamda-cyhalothrin)

concentrations to 1.0% and 0.1% of the initial concentrations, respectively, with macrophytes,

rather than sediments, functioning as the dominant pesticide sink. In similar experiments

performed by Cooper et al. (2002), 96-97% and 99% of added lamda-cyhalothrin and bifentrhrin

was associated with macrophytes within only 3 hours of the exposure event and concentrations

for both pesticides were below ecotoxicological limits within a downstream distance of 200 m.

Cooper et al. (2004) reported that a ditch length of 510 meters would be sufficient to remove

99.9% of initial exposure concentration of the pesticide esfenvalerate. Using microcosm

experiments, Bouldin et al. (2005) found that Juncus effuses was more resilient to exposure of

atrazine and lamda-cyhalothrin compared to Luakigia peploides. However, L. peploides tended

to accumulate more pesticides in its tissues (up to 2,461 Cpg kgl for atrazine).









While no studies of metals retention were identified specifically in drainage ditch

ecosystems, studies from wetlands show that the ability of macrophytes to remove heavy metals

also differs significantly by species (e.g. Debusk et al. 1996, Miretzky et al. 2004, Maine et al.

2006), with some species exhibiting internal detoxification mechanisms that allow them to

accumulate metals beyond accepted plant toxicity limits (Deng et al. 2004). It is evident that

some plant species may be more effective for pesticides and metals retention based on their

accumulation capacity and toxicity tolerance.

Synthesis and research needs

The studies reviewed here paint varying pictures of the ability of ditch systems to mitigate

exports of potential environmental contaminants such as sediments, nutrients and pesticides

(Tables 2-1 and 2-2). In general, current research supports that vegetated ditches show high

potential for retention of sediments and pesticides. Dissolved nutrients present a greater

challenge due to inadequate residence times and high nutrient loads relative to available reactive

surface areas. Unfortunately, the timing of high concentration events is often in conjunction with

the greatest levels of discharge, thereby exacerbating the ineffectiveness of in-stream processing.

Increasing residence times in ditches pose a risk for flooding during high flow events thus

creating an incompatibility between promoting both in-stream treatment and efficient water

transport.

To reconcile these conflicting goals, several researchers have reported on diversified and

integrated buffer measures that serve as multiple checks for dampening runoff rates and

concentrations (Bouldin et al. 2004, Wang et al. 2005, Dabney et al. 2006). Front-end mitigation

measures in agricultural settings include BMPs such as buffer strips, conservation tillage and

winter cover crops (Cooper et al. 2004). Dukes et al. (2003) reported that controlled drainage

was shown to dampen the flashiness of storm events, which could have important implications









for improved processing within ditches. Magnitude of loading is also a function of the timing of

fertilizer and pesticide applications. Jia et al. (2006) found that scheduling of irrigation played a

more important role in controlling losses of nutrients from fields than the BMPs in place. For

urban systems, front-end mitigation measures include the reduction of impervious surfaces

contributing to drains (Walsh et al. 2005).

Continued research on load-reduction should then be coupled with innovative

experimentation to seek approaches that amplify the efficiency of in-stream processing in

ditches. An appropriate theoretical template is that of "hot spots and hot moments" presented by

McClain et al. (2003). Mechanisms for retention, particularly those governed by rates of

biogeochemical reactions, are controlled by one or more limiting reactants in space or time

(Table 2-3). Identifying limiting factors as a starting point, ecological engineers can then begin

the creative design process of implementing those "missing" reactants into ditch systems to

enhance overall retention.

Hedin et al. (1998) provided a conceptual example of this approach by offering the advice

that managers interested in promoting denitrification should consider promoting natural inputs of

labile carbon to the near-stream region. Groffman et al. (2005) found that local inputs of labile

carbon could be provided by promoting the presence of organic debris jams which functioned as

denitrification hot spots in urban steams. Taking the idea of providing local carbon sources one

step further, Cooke et al. (2001) presented the idea of a denitrification bioreactor consisting of

gravel, sawdust and corncobs for the outfalls of subsurface tiles. Where anoxic conditions are

limited, water levels or hydraulic gradients in some cases may be adjusted as proposed by

Grimaldi and Chaplot (2000). Sorption sites for P retention may be enhanced by using soil

amendments such as calcium carbonate, hematite, vermiculate among others (Dunne et al. In









Press). Ann et al. (2000) suggested aluminum or calcium carbonate amendments that are less

influenced by redox conditions. Recycled concrete gravel or other sources of limestone could be

added to the benthic substrate to create riffle sections for improved oxygenation, mixing and

possible co-precipitation of calcium and phosphorus.

Other thought-provoking design ideas can be gained from innovative stormwater

infiltration trenches such as the "Ecology Ditch" described by Barber et al. (2003). This design

consists of vegetation that is underlain by a layer of compost, followed by sand and finally by a

perforated drainage pipe surrounded by gravel. Vegetation traps pollutants and organic matter

provides adsorption and filtration sites. Infiltration to the sand and gravel layers provides further

filtration and dampens peak discharges. The pipe provides drainage when the storage capacity of

the sands and gravels is exceeded.

In most ditches, efficient retention is primarily constrained by lack of channel complexity

and low residence times. Creating or restoring channel meanders may be possible in some cases

to increase channel heterogeneity and lengthen transport time. While this approach would require

additional land, the total area required may still be less than that needed for a constructed

wetland. For ditches that have sufficient hydraulic capacity, baffles or other structural

modifications might be installed in place of creating meanders. Materials such as straw bales

(except where grazing animals have access) might be useful as permeable baffle structures that

also provide a source of labile carbon and reactive surface areas.

Although some findings contradict the contaminant retention benefits of vegetation in

ditches (e.g. Barlow et al. 2003, Dukes and Evans 2006), there appears to be a general consensus

that plants are key components due to their direct uptake and sorption of potential contaminants

and especially by the hydraulic and microbial conditions they promote. Despite their importance,









relatively few studies have specifically addressed macrophyte management strategies in ditches

and other macrophyte-rich lotic systems (van Strien et al. 1991, Milsom et al. 2004, Wilcock et

al. 2004, Vereecken et al. 2006). Fifteen of the 22 studies presented in Table 2-1 identified

dominant macrophyte species present at the study sites. Ten species were prevalent in more than

one study and are summarized in Table 2-4. Future ditch studies should determine which species

are best suited for mitigating target contaminants given the geomorphic and climatic conditions

present. While some macrophyte species have higher uptake rates and tolerances to toxic

substances, studies have shown that polycultures nevertheless tend to be more effective than

monocultures for removal of nutrients (Kadlec and Knight 1996, Picard et al. 2005) and

pesticides (Bouldin et al. 2005). Future studies should also examine at what biomass densities

macrophytes are most effective at promoting the treatment function of ditches without

compromising the risk of flooding during large storm events. Improved understanding of which

species to promote and how often to harvest will enable managers to develop more informed

strategies for management of ditches as potential BMPs.

Some ditches may not support in-stream macrophyte growth due to their geomorphology.

These ditches could be widened to reduce their water depth and velocity or where land is limited,

one option for introducing macrophytes to the system is the use of floating platforms that support

hydroponic growth of macrophytes. Constructed floating islands are frequently used in lagoons

to support hydroponic growth of fast-growing species with extensive root systems such as

Vetiveria zizanioides for improved water quality (Lavania et al. 2004). For ditches with perennial

water flow, floating platforms could be made to freely move vertically with changing water

levels yet be fixed in horizontal space.









Examination of Table 2-2 also indicates where research in ditches is most lacking.

Despite concerns about water quality impairment due to pathogen contamination (Rodgers et al.

2003, Kay et al. 2007), few studies have addressed BMPs for pathogen reduction in agriculture

(Oliver et al. 2005) and surprisingly, no studies were found that specifically addressed the ability

of vegetated ditches to mitigate pathogens. Similarly, no studies were found that examined heavy

metal retention during transport in vegetated ditches. Another important research area for ditch

management is the impact of sediment dredging on retention mechanisms. Macrae et al. (2003)

found no difference between phosphorus retention before and after dredging; however others

have reported decreases in retention due to the loss of accumulated organic matter (Sallade and

Sims 1997a, Nguyen and Sukias 2002). Furthermore, there is the question of what should be

done with the extracted sediments. Van Strien et al. (1991) recommended utilizing organic and

nutrient-rich sludge by re-applying it to surrounding Hields.

Experimental approaches for management and modifications to existing channels may

not always be feasible due to possible interference with channel operations. To provide insights

into the effects of differing vegetation and sediment management strategies on ditch performance

under controlled yet realistic Hield conditions, experimental drainage ditch research facilities

have been constructed (Drent and Kersting 1993, Strock et al. 2005). Whether research is

conducted in experimental or actual field sites, there is clearly the need to achieve a greater

understanding of how ditches can be better designed and managed to reduce impacts to receiving

water bodies.

Conclusions

This review presented an overview of the potential treatment mechanisms occurring in

ditches and provided a synopsis of the current literature that has reported on the use of ditches

for reducing loads to downstream water bodies. Findings from 22 individual studies suggest that









as stand-alone systems, vegetated ditches are likely to be unreliable as effective mitigation

measures, especially for reducing loads of dissolved nutrients during high discharge events.

However, if coupled with a load-reduction scheme, vegetated ditches may serve as a cost-

effective and efficient BMP for mitigating contaminant exports to receiving water bodies.

Recognition of their potential role in mitigating contaminant loads may help to realign design

obj ectives for ditches to achieve goals beyond that of water transport. Experimental modification

of drainage ditch structure for amplifying the treatment function of ditches represents a virtually

untapped and much-needed realm of research that could have significant implications for

improved downstream water quality in urban and rural environments around the world.











Processing |
mode |


Processing |
mode |


Throughput
mode


Throughput
mode


0 1
load


discharge


Figure 2-1. Hypothetical relationships for A) contaminant processing efficiency versus discharge
proposed by Meyer and Likens (1979) and B) contaminant processing efficiency
versus contaminant loading, as proposed by Odum et al. (1979)























/ DOWNSTREAM
x EXPEX ORT
N, P, PEST
METALS






IF CAPACITY


SUNI
PLANT HARVESR~E
BIOMASS

MACROPHYTES ~ PERMAN ENT
REMOVAL


N, P, PEST
METALS

BENTHIC
SEDIMENTS

DITCH SYSTEM




Figure 2-2. Simplified systems diagram of the interactions between macrophyte growth, kinetic
energy ("KE," a proxy for water velocity), sediment transport and contaminant
retention in a generalized ditch system: Rain events or water usage deliver
environmental contaminants to the system ("N"=nitrogen, "P"=phosphorus,
"Pest"=pesticides, metals and sediments). Greater plant biomass leads to lower water
KE and more sediment accumulation. Lower KE promotes contaminant retention,
represented here by storage in benthic sediments and uptake and temporary storage
by macrophytes. When the uptake capacity of nutrients is reached, they no longer
contribute to plant biomass. For simplicity, toxicity effects of metals and pesticides
are not represented here. When plant biomass reaches the maximum density threshold
(as determined by the land manager) the plants are harvested, providing a permanent
sink of accumulated contaminants but temporarily diminishing their KE attenuation
function. Higher KE results in reduced retention, represented here as re-suspension
from and the lack of deposition to benthic sediment storage.












Table 2-1. Summary of environmental conditions and findings from published studies examining sediment, nutrient, BOD and
pesticide retention in drainage ditches (Note: NP = data not provided)


Contaminant Ditch type
(source of water)

Sediment 4 agricultural ditches,
Coastal Plain, North
Carolina, USA


Discharge Concentration(s)
(Ls -) (mg L- )

804-8940 TSS = 107-1345


Substrate


Vegetation


Represent
significant
effect?
Yes


Citation


(Lecce et al.
2006)


Sandy


Variable cover
depending on
season


50-75% cover
Polygonum,
Ludwigia,
Sagittaria, Typha,
Eleocharis


Sediment TN,
TP


Receiving pond water
from catfish farm,
Mississippi, USA



Agricultural and
pastoral ditches, Florida,
USA


7 agricultural ditches,
Indiana, USA


2 dairy farm irrigation
ditches, Australia


95-127 TSS=740-4549
TN=13.5-32.5
TP=0.42-3.89


(Hargreaves et al.
2005)


sandy,siliceous
OM=11-22%



silt+clay = 5-
93%
TC= 0.9-11.4%


Not clear (Dunne et al. In
Press)


SRP=0.02-0.09



TP=0.1-5.7
SRP=0.07-5.23


Not clear


(Smith et al.
2005)


(Barlow et al.
2003)


(Macrae et al.
2003)


TP, PO4


silt+clay=73- 5-85% cover
89% Lolium, Trifolium
Paspalum, Cyperus


Agricultural ditch,
Canada


1.5-7.3 SRP=0.7-2.1 sands and silts


Tall grasses and
periphyton


TP, PO4


12 pastoral farms, New
Zealand


TP = 0.05-0.40
SRP= 0.015-0.03


silt+clay >60%
TC=4.7-14.1%


Potamogeton,
Polygonum,
Glyceria, Nitella,
Lemna,
Ranunculus,
Bromus


Treatment
effects)

Ditches were sinks
for loads from fields
Sediment storage =
8.6-130.1 kg m 'yr

Avg.% removal
(n=3)
TSS = 96.5
TP = 68.8
TN =67.1

Degree of P
saturation > 25%
indicating potential
of P release

Some sites acted as P
sinks and others as P
sources

Estimated 14%
decrease in P exports
using a bare drain

P areal uptake rates
in same range
pristine systems but
represented only 5-
10% retention

Sediments acted as a
long-term P sink
(residual P 6-39%)


(Nguyen and
Sukias 2002)












Table 2-1. Continued
Contaminant Ditch type
(source of water)

PO4 17 agricultural
ditches, Delaware,
USA




P (forms Wet meadow ditches,
not The Netherlands
specified)
TN, TP 5 irrigation ditches.
NO3, PO4 Australia


Wetted
width
(m)


Discharge Concentration(s) Substrate
(L s- ) (mg L- )


Vegetation


NP


Treatment
effects)

Sediments were
enriched with
bioavailable P but
still had high buffer
capacity

9()-95% removal
capacity of P

Reductions in all
parameters in 4 of 5
ditches


Represent
significant
effect'?
Not clear






Yes


Yes


Citation


(Sallade and Sims
1997a)





(Meuleman and
Beltman 1993)

(Bowmer et al.
1994)


NP NP TP = ().21-6.14
SRP = .()4-().74




NP NP NP


silt+clay =24%
OM = 8%


Heavy cover
Fontinalis


-2-4 2-9()


TP=().11-1.)2
SRP = .()36-).53
TN= 1.56-3.47
NO3-N=).()2-2.14


Variable cover
Paspalunt, Tipha,
Potaniogeton,
Schoenoplectus,
Ludwigia,
Vallisnen'a,
Elodea, Sagitaria

>5()% flow
obstruction
Phalaris, Carex



Presence of algae
and macrophytes


4 groundwater-fed
agricultural ditches,
Germany




6 agricultural
ditches, Michigan
and Indiana, USA



2 agricultural row
crops ditches and 2
canals, North
Carolina, USA


~79% and 6)%
reduction in NO3-N
and BOD in 15() m,
no net change in
NH4-N

Uptake velocities
within same range as
those reported in
relatively pristine
systems

Uptake velocities
within ranges
reported for pristine
systems
Estimated 46-75%
and 13-66% NH4 and
PO4 removed during
high flow periods


(Scholz and
Trepel 2()(4)


NH4
NO3
BOD


5 1.5 > ().5 NH4-N=
().13+().)5
NO3-N= 13.1+2.()
BOD= 14.8


NP 1.5 575 NH4-N=).()12-
().144
NO3-N=().2-5.1
SRP=).()(62-().262


().9 and 12 and NH4 ).()86-().648
2.6 18 SRP ).()39-().6)4


NO3
NH4
PO4


Not clear (Bemot et al.
2()(6)


Coarse to
medium-
grained sand,
accumulations
of silt. No
CWD


Ditches Ludwigia
Canals
Potaniogeton


(Ensign et al.
2()(6)












Table 2-1. Continued
Contaminant Ditch type
(source of water)

PO4 2 agricultural
NH4 ditches, receiving
NO3 additional point
source inputs of
septic and WWTP
effluents, Germany


Wetted
width
(m)
NP


Discharge Concentration(s)
(L s ) (mg L- )

22-23 NO3-N= 0.8-
164-511 16.4
NH4-N= 0.03-
0.3
SRP = 0.01 -
0.27


Substrate


Vegetation


Treatment
effects)

Uptake rates were
high compared to
pristmne systems but
retention was low
due to high loads


Denitrification rates
ranged from~-0 to 16
mgN m-2 hr-1


Denitrification rates
ranged from <0.1 to
15 mgN/hr/m2



Pyrethroid
COncentrations
decreased to 0.1% of
initial concentration
within 280 m

Pyrethroid
concentrations
decreased to 0.1% of
initial concentration
within 510 m

94-98% insecticide
added was removed
from water column
by day 3


Represent
significant
effect?
No


Citation


(Gucker and
Pusch 2006)


fine sands Potamogeton,
OM = 5.9% Sparganium,
Phalaris, Glyceria


Agricultural tile-
drained row crops,
Illinois, USA


5 agricultural tile-
drained ditches,
Illinois, USA


7.4 20 -1370 NO3-N= 3-17.7


Gravel, sand,
some organic
sediments.


Gravel and sand
>95%


27-86% cover
Cladophora,
Potamogeton


Open canopy in 4
of 5 sites


(Schaller et al.
2004)



(Royer et al.
2004)


NO3


NP 10-1840 NO3-N= 0.6-
15.8


Pesticides Row crops,
1Vississippi Delta,
USA



Pesticides Row crops,
1Vississippi Delta,
USA



Pesticides 3 experimental
ditches, The
Netherlands


2.8 29.4* pyrethroid = 3.74
and 6.66


~88% cover (up to
6000 g m-2),
Ludwigia, Lemna
Polygonum


Ludwigia,
Polygonum Leersia


(Bennett et al.
2005)




(Cooper et al.
2004)


2.8 16.2 pyrethroid =0.15 NP


Yes


1.6 NP lamba-
cyhalothrin
0.0025


Silty clay loam


Variable densities
Myriphyllum,
Elodea, Sagittaria


(Leistra et al.
2004)












Table 2-1. Continued
Contaminant Ditch type
(source of water)

Pesticides 2 ditches, row crops,
Mississippi Delta,
USA



Pesticides 2 agricultural ditches
draining irrigation
runoff, Canada


Wetted
width
(m)
2.8


Discharge Concentration(s) Substrate
(L s- ) (mg L- )


Vegetation


Polygonum,Leersia
, Lemna


Treatment
effects)

lamba-cyhalothrin
and bifenthrin
reduced to below
ecotoxicological
levels within 200 m

2 herbicides
occasionally
exceeded aquatic life
guidelines
1.9 and 2.2% N and P
applied to fields
exited ditches

Atrazine and
pyrethroid
concentrations were
decreased to "no-
effects" level within
50 m

40-60% of applied
linuron was
discharged from the
40-m ditches


Represent
significant
effect?
Yes


Citation


(Cooper et al.
2002)


1.0 lamba-cyhalothrin
= 0.06
bifentrhin = 0.11



0 to No
2000 initial/upstream
concentrations
provided


Loams to fine
sandy loams


Not clear (Cessna et al.
2001)


16-881 g m2
Polygonum,
Leersia Sporobolus


Pesticides Row crops,
Mississippi Delta,
USA


NP 61.3 pyrethroid = 0.46
atrazine = 28.9


(Moore et al.
2001)


sandy loam 20-110g m2
OM = 26% Myriophyllum,


Pesticides 8 experimental
ditches, The
Netherlands


3.4 NP linuron= 0.0005-
0.5


Not clear (Crum et al.
1998)


*estimated from channel and velocity information










Table 2-2. Summary of the retention effectiveness of ditches based on the studies reported in Table 2-1
Contaminant Number of Number Number Inconclusive
cases* concluding concluding
ditches to be ditches to be
effective ineffective
Sediments 2 2 0 0
Pesticides 7 5 0 2
BOD 1 1 0 0
Total N 2 2 0 0
Total P 5 4 0 1
Phosphate 10 4 2 4
Nitrate 6 2 3 1
Ammonium 5 1 3 1
Metals 0 0 0 0
Pathogens 0 0 0 0
38 55% 21% 24%
* Individual studies from Table 2-1 were counted more than once if they considered multiple contaminants









Table 2-3. Possible reactants and conditions that limit the retention capacity in drainage ditches
Contaminant Retention limited by
Nitrate Labile carbon
Anoxic zones
Assimilation capacity

Ammonium Oxygen for nitrification
Assimilation capacity

Phosphate Sorption sites
Appropriate redox conditions
Assimilation capacity

Sediments Quiescent zones and channel friction

Pathogens Sorption sites
Quiescent zones

Pesticides Sorption sites
Assimilation capacity

Table 2-4. Dominant macrophytes reported by the 15 studies out of 22 in Table 2-1 that provided
plant species information
Scientific name Common name Occurrence frequency Plant type
Polygonunt spp. Knotweed 6 Emergent
Potanzttt~~~~~ttttt~~~ogeo spp. Pondweed 5 Submerged or floating leaves
Luakuigia spp. Primrose 4 Emergent
Leersia spp. Cutgrass 3 Emergent
Lenana spp. Duckweed 3 Floating
Sagittaria spp. Arrowhead 3 Emergent
Elodea spp. Pondweed 2 Submerged
Glyceria spp. Manna grass 2 Emergent
M~yriophyllunt spp. Water milfoil 2 Submerged
Typha spp. Cattail 2 Emergent









CHAPTER 3
INFLUENCE OF MACROPHYTES ON TRANSPORT CHARACTERISTICS AND
NUTRIENT STORAGE

Introduction

Macrophytes are important features affecting the transport and fate of potential

environmental contaminants in running water ecosystems. Plants provide temporary retention of

potential contaminants such as nutrients directly through uptake and assimilation into tissues.

They also indirectly influence in-stream retention mechanisms by providing structural

complexity that creates zones of in-stream transient storage. These zones of slowly-moving

water increase channel residence time, promote sedimentation of suspended materials, and

amplify biogeochemical reactions in both time and space due to greater contact with reactive

substrates (Figure 3-1).

Vegetation is likely to influence the transport and fate of pathogens, sediments and

nutrients in wastewater drainage ditches in Oxapampa. The mild climate coupled with high

nutrient conditions characteristic of these ditches results in year-round abundant growth of

vegetation. Due to the risk of flooding, potential mosquito and other pest problems and the local

negative perception of "weedy" plants species, vegetation in ditches is periodically removed. The

current management practice by the city of Oxapampa and other municipalities in the region

consists of harvesting aboveground ditch vegetation with machetes followed by application of a

non-selective glyphosate herbicide (Roundup). For wastewater ditches with medium to high

daily water flows (>0.5 liters/s), these procedures are carried out two to four times per year

(Arteaga, personal communication).

Few studies have explicitly linked the presence of plants with transient storage in lotic

ecosystems (Harvey et al. 2003). This study examines how plant management in wastewater

ditches influences properties of water transport by comparing transport times and transient









storage parameters under different biomass conditions. It also quantifies nutrient storage in

macrophyte tissues to examine whether plant uptake serves as an important retention mechanism.

The specific hypotheses tested in my study are: (1) transport in vegetated ditches results in

decreased concentrations of suspended sediments, E. coli bacteria and nutrients; (2) the presence

of ditch vegetation significantly increases transient storage and channel residence time; and (3)

aboveground biomass harvesting provides an important nutrient sink.

Methods

Site Descriptions

Four wastewater ditches were chosen as study sites for examination of water quality and

plant biomass characteristics: San Martin, Gustayson, Koell and Frey. At the times of sampling,

the lower reach of Frey was ponded with no flow exiting the site for discharge to the

Chorobamba River. As a result, transport studies to examine transient storage were limited to the

San Martin, Gustayson and Koell sites.

The site names are derived from the names of the streets along which the ditches are

located. Wastewater ditches within the city run parallel to roads (Figure 3-2) and consequently

are straight and typically incised. Most sites were dominated primarily by Cynadon dactylon and

Hedychium coronarium with the exception of Frey, which consisted of a larger variety of both

floating and emergent macrophytes due to the presence of a marsh environment in the lower

reach (Table 3-1).

San Martin

Four locations were sampled in San Martin, the first of which was located 30 m below the

junction between a natural stream and the inflow of wastewaters. The second sampling location

receives occasional inputs from a small ditch fed by wastewater from a single house. There are

no wastewater inputs between locations 2 and 4, the most downstream location, which is










approximately 6 m upstream from the discharge point to the river (Figure 3-3). All of the San

Martin sampling sites are surrounded by heavily grazed pasture.

Gustayson

The Gustayson ditch exists along four city blocks and is surrounded by homes that

discharge wastewaters along the total length of the ditch (Figure 3-4). Samples were collected

along each block for a total of four sampling sites before discharge to the river. Additional

wastewaters enter Gustayson from two perpendicular streets. Hence, discharge in Gustayson

tends to increase at each consecutive sampling site.

Koell

Koell is located along a densely-inhabited city block and then passes through an

abandoned pasture before discharging to the river (Figure 3-5). The first two sampling locations

are at the top and bottom of the city block where homes contribute wastewater that results in

increased flowrates between the two locations. The channel along this section is deeply incised

and subj ect to severe slumping. The third and fourth sampling locations are within a pasture

where the channel gradient is lower and the channel bed is wider and shallower. The Koell ditch

site has a history of trash dumping into the mid-reach of the ditch due to the lack of a trash

collection service for residents living at the end of the block. Therefore between sampling sites 2

and 3 there existed an accumulation of predominantly plastic items such as bags and bottles as

well as miscellaneous items such as tires, pieces of scrap wood and metal.

Frey

Frey has the lowest flows of the four study ditches and on occasion dries out. The first

sampling location in Frey is in the upper portion of ditch where the channel is surrounded by

homes that discharge wastes. At the end of the city block, the channel abruptly drops 2 m and

drains into a marsh covered where the second sampling site is located (Figure 3-6). The city









occasionally attempts to ditch the marsh to promote flow towards the river to avoid mosquito

problems. However, the site's low topography results in ponding despite the city's efforts.

During high flows, a portion of the water in Frey flows out of the marsh and j oins with the Koell

site. Just before this confluence is the Einal sampling location in Frey.

Field Procedures for Water Chemistry, Flow and Channel Characterization

Grab samples of ditch water for each site were collected periodically between May and

November 2005 and March through July 2006 to examine water quality over distance traveled in

each ditch (hereafter longitudinal samplings) and through time during the dry season. Water

samples were collected in acid-washed, amber 500 mL Nalgene bottles. Samples were placed on

ice in a cooler and taken to the research station in Oxapampa for analysis. Water temperature and

dissolved oxygen (DO) were measured in situ using a handheld DO meter (Model 55, YSI

Environmental, Yellow Springs, OH). Water pH and specific conductivity were also determined

in situ using a combined pH and conductivity meter (ExStik EC500, Extech Instruments,

Waltham, MA). Water velocity was measured at each sampling location using an impeller

flowmeter (Model 2030, General Oceanics, Miami, FL) and used with channel width and depth

measurements for calculating discharge. Sites were surveyed to evaluate channel characteristics

such as length, slope and cross-sectional areas.

Laboratory Analysis of Water Quality

After collection, water samples were immediately filtered through 0.45 Cpm membrane

filters and analyzed for soluble reactive phosphorus (SRP), nitrate (NO3-N) and ammonium

(NH4-N) with a portable Hach DR/890 colorimeter and powder pillow reagents (Hach Company,

Loveland, CO). SRP was analyzed with the PhosVer3 ascorbic acid method (equivalent to

USEPA Method 365.2 and Standard Method 4500-PE; detection limit 0.05 mg L^1 as PO4,

precision + 0.03 mg PO4 L^)~. NO3-N was analyzed using the cadmium reduction method









(detection limit of 0.01 mg L^1 as NO3, preCISion 10.03 mg NO3 L^)~. Liquid wastes from the NO3

analysis were collected and disposed of through Environmental Health and Safety at the

University of Florida. NH4-N was analyzed with the salicylate method (detection limit 0.02 mg

L^ as N, precision & 0.02 mg N L^)~. Unfiltered samples were analyzed for total phosphorus (TP)

using the Hach acid persulfate digestion method followed by analysis of SRP as described above.

Total nitrogen (TN) was measured using Hach Test 'N Tube digestion method (detection limit of

2 mg N L^1, precision & 0.5 mg N L^)~. Proper operator and equipment performance was tested

for each set of samples through the use of known standards (purchased as ampules from Hach

Company), blanks and replicates.

Sediments collected on 0.45 Cpm membrane fi1ters were used for gravimetric determination

of total suspended sediments. Concentrations of E coli bacteria were measured using Coliscan

Easy Gel kits (Micrology Laboratories, Goshen, INT). Carbonaceous biochemical oxygen demand

(CBODS) was analyzed following Standard Method 5210 B (APHA 1992) and measured with a

YSI 58 DO meter fitted with a YSI 5010 BOD probe (YSI Environmental, Yellow Springs, OH).

Transport Characteristics of Ditches

Transport characteristics of each ditch were determined through the analysis of

breakthrough curves of conservative solute tracer experiments (described below, Table 3-2).

Tracer experiments were performed on three occasions in Gustayson. The first experiment took

place approximately 3 weeks following the removal of plants from the upper 3 reaches of the

ditch and before plant removal in the lower reach. Therefore, the analysis of the first tracer

experiment in Gustayson compares transport properties between the cleared and vegetated

reaches. Tracer experiments were performed twice in San Martin and Koell; once in November

2005 and again in June 2006 to compare transport characteristics under low and high biomass









conditions, respectively. In Koell, the second experiment also coincided with the removal of

approximately 0.5 metric tons of trash from the stream channel.

Tracer experiment field procedures

Constant rate inj sections of NaCl were conducted following experimental design protocols

detailed by Wagner and Harvey (1997). Five and ten gallon Mariotte siphons (Figure 3-7) were

constructed and used to deliver the conservative tracer at a constant flowrate such that the

resulting breakthrough curve (BTC) reached a plateau concentration at the downstream

monitoring location. For most experiments a constant rate inj section of 0.9 L min' (accuracy

+0.02 L min- ) was used. However, flows could be adjusted if desired by the changing the head

differential between the vent tube and the outlet from the Mariotte siphon. Quantities of the

conservative tracer (between 20-40 g NaCl L^) were pre-weighed at the laboratory and dissolved

at the site using ditch water that had been sieved to remove any solids. The solution was then

added to the Mariotte siphon. Tracer breakthrough curves were determined by measurements of

specific conductivity (YSI 556 MPS, YSI Environmental, Yellow Springs, OH). Grab samples

of water were collected at 3 to 4 points during tracer experiments to analyze water quality

following procedures described above.

Analysis of tracer data

Transport and transient storage parameters were estimated from the resulting tracer

breakthrough curve data using the One-dimensional Transport with Inflow and Storage with

Parameter Estimation (OTIS-P) model (Runkle 1998). The OTIS-P model conceptualizes

flowing water systems as having two distinct hydrologic regimes (Wagner and Harvey 1997):

surface water flow in the main channel and immobile or slowly moving water in surface

quiescent zones and in subsurface flowpaths. The mathematical model is a 1-D advective-

dispersive transport model that accounts for advection, dispersion, lateral inflow, and transient









storage. Transient storage is represented as a first-order mass transfer that yields two parameters.

The first parameter represents the cross-sectional area of the transient storage zone (As) and the

second is a first-order mass transfer coefficient (a) that describes the rate of hydrologic exchange

between storage zones and the main channel. The governing equations are:

aC Q BC 1a 8 C\ q, c, c+~, c 3
=-- +---- AD- I+ C-C+ C-C(31
at A Dx A ax \ x A
aC, A
-a- (C- C)(3 -2)
at A "

where t and x are time and distance along the stream; C, Cs, and CL are the solute concentrations

in the stream, storage zones and groundwater (mg L^)~, Q is the stream volumetric flow rate (m3

s- ), A and As are the cross-sectional areas of the stream channel and storage zone (m2), D is the

dispersion coefficient (m2 S-1), qL is the lateral inflow rate (m3 S-1 m-1), and a is the storage

exchange coefficient (s- ).

The OTIS-P model uses a least squares algorithm to adjust parameter values of D, A,

a, and A, to fit the observed tracer BTCs. Confident interpretation of model results is predicated
on the reliability parameter estimations. Wagner andHrve (T~P~1997) sulggest reporting of the

Damkohler number (Dal) as an indicator of parameter uncertainty where values on the order of

0.1 to 1.0 indicate high parameter reliability. Dal is calculated as:


Dal = a1+ A)L(3-3)


where L is the reach length (m) and v is mean water velocity (ms )~.

Parameter values and metrics that are commonly reported to facilitate cross-study

interpretations of transient storage studies include As, As/A, a, Tmed, Fmed and Fmed200. The ratio

of As/A shows the relative importance of storage zone and main channel cross-sectional areas

(Harvey et al. 2003). Tmed represents the median transport time, the time at which 50% of the









plateau curve is realized (Runkel 2002). The fraction of transport time spent in transient storage

zones is denoted as Fined and may be approximated by (Runkel 2002):

Fmed-[ 1e-L~a u) [A,/A+A,] (3-4)

Fined is alSo calculated using a standardized distance of L = 200 m (Fined200) to allow comparison

with other systems (Runkel 2002).

Vegetation Sampling

Collection of plants at sampling locations

Above and below-ground biomass in ditches were sampled in November 2005 and June

2006 in San Martin, Gustayson and Koell in conjunction with tracer experiments as described

above. Biomass sampling in Frey took place only in November 2005. Representative biomass

samples were collected using a 0.25m x 0.25m quadrat at three random locations within af l0m

distance of each water quality sampling location in each study ditch for a total of 12 biomass

samples per ditch. Only plants growing within the channel or with plant parts exposed to the

water were collected. If the plant had no contact with the water or channel it was not included in

the sample. This sampling approach was chosen to only include plants that might have an effect

on transport characteristics or contaminant retention.

At the research station, the plant tissues were carefully rinsed to remove trapped sediments,

air dried for an hour and then separated into leaves, stems and roots/rhizomes and weighed. The

plant tissues were then dried at 1050C for 48 hours and reweighed. Moisture content was

determined from the difference between the wet and dry weights. Biomass was calculated as the

tissue dry weight multiplied by the total sampled area as determined by use of the quadrat.

Dried plant tissue was ground to a 40 mesh size using a Wiley mill for analysis of total

carbon, nitrogen and phosphorus. TN and TC content were determined using approximately 15










mg of dry sediments. Samples were analyzed by dry combustion using a CNS analyzer (Carlo

Erba Model NA-1500). Total P content was determined on 0.5 g of dry sample that was

combusted at 5500C in a muffle furnace for four hours. The remaining ash was then dissolved in

6 M~HCI (Andersen 1976) and the digestate was analyzed for P on a Technicon autoanalyzer III

using an automated ascorbic acid method (Method 365.1, USEPA 1993). Plant tissue nutrient

content was used to estimate above and belowground nutrient storage at each study site.

Comparison of biomass and nutrient content of plants grown out of wastewater

Biomass samples of dominant macrophyte species found in wastewater ditches in

Oxapampa were collected and compared with the same species found growing out of wastewater

following the protocols given above. The purpose of this comparison was to examine whether

nutrient content and allocation of nutrients differed in plants found growing in wastewater.

Statistical Analyses

Comparisons of upstream and downstream concentrations of water quality parameters were

made using paired one-tail t-tests of data that were In-transformed to meet normality

requirements. Plant tissue N and P of macrophyte species growing in wastewater ditches were

compared with the same species growing out of wastewater using the same statistical approach.

Results

Water Quality of Wastewater Ditches

General characteristics

The ditch sites differed from one another with respect to their channel and flow

characteristics as well as their water chemistry (Table 3-3). San Martin experiences the highest

water flows and has the lowest nutrient and suspended sediment concentrations. The lower

temperature and specific conductivity (17.60C, 155 CLS cm-') in San Martin compared to that of

the other sites (18.7-20.90C, 243-255 CLS cm-') further supports that San Martin experiences more









dilution than the other sampling sites. DO values for all sites were low (average DO values for

the sites varied between 1 and 2.5 mg L-') and CBODS values ranged from 4.6 mg L^1 in San

Martin to 18 mg L^1 in Gustayson. These low oxygen environments limit nitrification, as

evidenced from the high ammonium and organic N concentrations and low nitrate

concentrations.

Longitudinal water quality samplings indicated that concentrations of nutrients, TSS, and

E. coli tended to increase with downstream distance in Gustayson due to wastewater inputs along

the length of the ditch. Similar results were seen in the upper reach of Frey. In the marsh of Frey,

water was ponded on all sampling occasions, precluding water quality parameter measurements

after passage through the marsh.

The lower reaches of the San Martin and Koell sites (reach lengths 255 m and 73 m,

respectively) have no additional wastewater inputs and therefore allow observation of potential

retention of TSS, E. coli and nutrient exports to the Chorobamba River. However, there were

only 8 and 10 sampling occasions for San Martin and Koell, respectively, for which discharge

measurements in the lower reaches did not vary significantly (i.e. the change in flowrate between

sampling sites was less than the flowmeter accuracy of 10 cm s- ). Seven of the 8 sampling dates

in San Martin were characterized by high biomass conditions. Eight of the 10 sampling dates in

Koell had high biomass as well. These small sample sizes of low biomass conditions prevents a

statistical comparison to made between upstream and downstream trends in water quality based

on biomass conditions.

Attempts were made to sample the same parcel of water at each sampling location by

measuring mean water velocity to estimate arrival times to downstream sites. It is recognized

that this approach does not ensure that the same water was sampled; therefore, these









measurements were grouped to examine overall trends in upstream and downstream water

quality to seek evidence of in-stream E. coli, TSS and nutrient retention in San Martin and Koell

during transport (Figures 3-8 and 3-9).

Paired t-tests indicated a significant decrease in E. coli concentrations between upstream

and downstream sampling sites for both Koell (t = 3.38, df = 6, p<0.01) and San Martin (t =

2.29, df = 6, p-value < 0.05). Downward trends in TSS were observed for both sites but were not

significant at p < 0.05. TN and TP concentrations tended to decrease at both sites as well but

only TN in Koell was significantly lower at the downstream location (t = 3.80, df = 3, p < 0.05).

In San Martin, downstream SRP concentrations tended to be significantly lower than the

upstream sampling location (t = 4.24, df = 7, p<0.01) while downstream SRP values at Koell

tended increase slightly compared to the upstream sampling location. Ammonium and nitrate

concentrations did not change significantly between upstream and downstream sampling

locations (NO3-N values not shown).

Water quality during tracer experiments

Water samples collected during tracer experiments have the advantage of allowing the

same parcel of water to be sampled through space. However, samples collected during the

experiments revealed no consistent patterns that clearly distinguished the effects of high biomss

from low biomass conditions on contaminant retention (Figure 3-10). TSS was variable in all

sites but tended to be lower under high biomass conditions in Koell and San Martin. E. coli

concentrations consistently decreased with distance downstream in all cases except during the

low biomass condition experiment in Gustayson. Nitrate values for all sites on all tracer

experiment dates were less than 0.05 mg N L^1 and did not vary more than + 0.03 mg N L^1 and

therefore were not included in Figure 3-10. Phosphate and ammonium concentrations were









variable or increased with downstream distance for all sites. The only exception was phosphate

which decreased from 0.9 to 0.55 mg P L^ in San Martin under high biomass conditions.

Transport Characteristics

Comparison of breakthrough curves for sites under different biomass conditions revealed

the importance of macrophytes for providing structural complexity that served to extend travel

times and increase sizes and exchanges with transient storage zones (Figures 3-11, 3-12 and 3-13

and Table 3-4). Median transport times (Tmed) in San Martin and Gustayson increased by 89 and

143% under high biomass conditions. The difference in Tmed between tracer experiments in

Gustayson was likely exaggerated by the higher flowrates during the low biomass experiment

(17.9 versus 11.8 L s^)~. However, the opposite was true for the experiments in San Martin (32

versus 38 L s for the low and high biomass experiments, respectively) indicating that the

calculated percent increase in Fmed WaS likely underestimated for this site.

Estimated parameter values revealed notable differences in transient storage between the

low and high biomass tracer experiments (Table 3-4). The ratio of the transient storage and

channel cross-sectional areas (As/A) in the 2005 Gustayson experiment increased from 3.0 to

1 1.6 between the low biomass and high biomass reaches, respectively. The fraction of median

transport time spent in transient storage zones (Fmed) alSO increased from 18 to 62%. Similar

results were observed in Gustayson the following year in the second set of tracer experiments.

The As/A ratio increased from 0.008 to 0.27 in Gustayson and 0.36 to 1.33 in San Martin

for the low and high biomass tracer experiments. Fmed inCTreSed from zero to 21% in Gustayson

and 5.5 to 56% in San Martin. Exchange rates between the stream and storage zones (a) also

increased under high biomass conditions from 6.0 E-8 to 27.9 E-4 s^l in Gustayson and from 3.7

E-4 to 30.5 E-4 s^l in San Martin. Damkohler numbers were on the order of 0. 1 to 1.0 for all









tracer experiments except for the second high biomass experiment in Gustayson (Dal = 47)

suggesting high reliability of the parameter estimations (Wagner and Harvey 1997).

The tracer studies performed in Koell revealed an additional source of structural

complexity common to wastewater drainage ditches in the study area: garbage. Before removal

of approximately 0.5 metric tons of refuse from the channel, the conservative tracer BTC

displayed a long tail implying greater contact with storage zones compared to the second

experiment which took place shortly after trash removal but during high biomass conditions

(Figure 3-13). Values of As/A and Fmed after garbage removal decreased from 15.1 to 0.07 and

from 52% to 0.9%. As seen in Table 3-4, median transport times are approximately the same

between the two experiments in Koell despite higher transient storage during the low biomass

conditions. The majority of the refuse had accumulated in the channel between reach lengths of

93 and 121 m. The channel above and below this section was relatively clear of other flow

obstructions. Therefore, transport times were likely to be distributed between rapid transport in

the upper and lower reaches with slow transport due to mixing with storage zones in the section

with trash accumulation.

Plant Biomass and Nutrient Storages

At the times of plant sampling in 2005, mean total biomass (sum of above and

belowground biomasses, Table 3-5) was highest in Gustayson (2,980 g m-2), followed by San

Martin (1,137 g m-2), Koell (596 g m-2) and Frey (487 g m-2). In 2006, the sites showed a more

even distribution in biomass with mean total biomass values of 2,574 g m-2 in Gustayson, 2,725

g m-2 in San Martin and 2,826 g m-2 in Koell.

Vegetation in 2005 had been harvested by the city from Gustayson and San Martin

approximately 3 months and 2 months prior to biomass collection and from Koell and Frey

approximately one month prior (Arteaga, personal communication). Using these approximate









time periods of 90, 60 and 30 days together with biomass areal estimates and measured values of

biomass and plant tissue nutrient content, daily N and P uptake by plants was estimated (Table

3-5). These calculations assumed a zeroth order growth model due to a lack of intermediary

biomass data. Belowground biomass growth rates could not be calculated due to an unknown

quantity of roots and rhizomes remaining even after herbicide applications. Therefore the

contribution of belowground biomass was not considered in daily uptake calculations. Estimated

nutrient uptake rates for aboveground biomass were used with average dissolved inorganic N and

P loads to each of the ditches to arrive at a rough approximation of the potential gross N and P

retention in ditches by plant uptake (Table 3-6). The approximations made in Table 3-6 suggests

that aboveground biomass of ditch vegetation assimilates the equivalent of up to 9.6 and 1 1.7%

of daily N and P loads in ditches. While these estimates do not include uptake contributions by

belowground biomass, nutrient content of roots and rhizomes tended be minimal compared that

stored in aboveground tissues. Nevertheless, the overall plant assimilations rates are therefore

underestimates. However, N and P retention values do not account for nutrient releases from

decaying plant material.

Mean molar ratios of above and belowground C, N and P indicate high P enrichment at all

sites (Table 3-7), particularly in Gustayson (molar N:P of 10. 1 and 3.6 in above and

belowground plant tissues, respectively). Nutrient content per species (Figure 3-14) indicated

that Cynadon had the highest foliar P content (7.0 mg g- ) while highest P in roots was

M~yriphyllum (9.4 mg g )~. Hydrocotyle had the highest foliar N content (47.4 mg g- ) followed

by Cynadon (37. 1 mg g )~. Eichornia and Myriophyllum contained the highest N in roots (42.3

and 40.1 mg g- ). Polygonum and Hydrocotyle had the highest P in stems (8.2 and 8.0 mg g- )

while Equisetum contained the highest N in stems (32.2 mg g- ).










Comparisons of whole plant nutrient content of species growing in and out of wastewater

show that plants in wastewater ditch to have elevated N and P tissue content (Figure 3-15). Both

N and P in Hydrocotyle tissues were higher (p < 0.01) in wastewater. P content was significantly

higher (p < 0.01) in M~yriophyllum while N content was significantly higher for Hedychium.

However, caution must be noted that the nutrient status of the soils of the control species was not

determined.

Discussion

Waste streams discharged to wastewater drainage ditches are highly variable through time

and space reflecting punctuated and diverse uses of water by residents. The variability of inputs

to a system leads to high uncertainty during longitudinal sampling campaigns that the same

parcel of water is being captured over space in order to fairly assess in-stream retention. A

decrease in downstream concentration may easily be misinterpreted as retention when instead a

non-representative slug of water had been sampled. The problem of teasing apart within-stream

variability from evidence of retention is addressed to some degree by lumping the data to

evaluate changes in upstream and downstream concentrations. The resulting set of box plots

illustrate that though there are trends in the data suggesting concentration decreases with distance

downstream, these differences may often be due to chance alone given the wide spread of the

data. Notwithstanding, results do strongly imply potential for retention in the study ditches by

sedimentation of E coli, TSS, TN and TP. The San Martin site also presents compelling

evidence of SRP retention. Similar sampling campaigns were not carried out during the wet

season when flows tend to be highest. Therefore, the results of the dry season conditions

presented here are likely to overestimate overall system performance.

While a growing number of studies are reporting on the effects of in-stream vegetation for

reducing velocities and promoting sedimentation (Wilcock et al. 1999, Champion and Tanner










2000, Clarke 2002, Schulz et al. 2003, Hargreaves et al. 2005, Lecce et al. 2006), only a few to

date have explicitly quantified the importance of macrophytes for creating zones of transient

storage (Harvey et al. 2003, Salehin et al. 2003, Ensign and Doyle 2005). Analysis of BTCs

using OTIS-P does not distinguish between surface and subsurface storage zones. Historically,

the maj ority of OTIS-P interpretations assumed that exchanges with storage zones took place

between the main channel and subsurface flowpaths and often negated the existence of in-

channel transient storage. More recent studies have recognized the prevalence of in-stream

transient storage zones and have presented evidence of their importance for nutrient retention

(Gucker and Boechat 2004, Ensign and Doyle 2005). Comparisons between transport times and

transient storage metrics under low and high biomass conditions in the present study indicated

that essentially all storage zones exist in the main channel, not in the subsurface.

That transient storage zones may be abundant in ditches presents such systems in a new

light. Channelized systems are often presumed to have minimal transient storage compared to

natural streams due to the lack of structural complexity needed to create pockets of slowly-

moving water or the lack of porous substrates to allow surface-subsurface hydrologic exchanges.

However, results from this study found some of the highest rates of exchange (a) that have been

reported to date. Interestingly, another study that reported similar a values (43-58 E-4 s l) was a

channelized agricultural stream with in-stream vegetation (Ensign and Doyle 2005). As

suggested by Ensign and Doyle (2005), these high a values may represent exchanges with

turbulent eddies that would tend to be more rapid than exchanges occurring with subsurface

environments. The Koell site had the highest As/A value of all sites in the presence of channel

debris. However, this site also had the lowest a value compared to other the high transient

storage zone cases. This slower exchange rate between main channel advective flow and storage









zones helps to explain why this site had the most pronounced BTC tail. The conservative solute

spent more time circulating in dead zones formed by tires and other debris than in the turbulent

eddies created by macrophytes in the other sites.

The Koell site also provided evidence of the effect of channel morphology on the

prevalence of transient storage zones. While vegetation clearly had an impact on transport

characteristics in San Martin and Gustayson, it was much less pronounced, especially compared

to the transient storage zones created by the accumulation of trash. The effect of vegetation was

likely to be insignificant in the upper reaches ofKoell where the channel is most narrow and

deeply incised supporting only bank vegetation.

Abundant, year-round plant growth in the study region makes macrophytes good

candidates for promoting in-stream retention in wastewater ditches. The assumed zeroth order

biomass growth model used to calculate uptake rates is likely to be adequate for the 30 and 60

day periods. Linear growth of macrophytes was observed by Tanner et al. (1996) over a 60 day

period for 5 of 8 species growing in dairy farm wastewater. After 60 days, growth rates tended to

increase for some species and decrease for others depending on initial propagule vigor (Tanner et

al. 1996). Linear growth was observed over the 117 day sampling period for one species,

Cyperus involucratus. Therefore, assumption of linear growth in Gustayson over a 90 day period

may not have grossly under or overestimate growth rates but instead may have yielded an

average value over the sampling period.

Approximated N and P uptake rates are on the order of those reported for macrophytes in

constructed wetland treatment systems (DeBusk et al. 1995, Greenway 1997, DeBusk et al.

2001). However, this retention mechanism is severely limited in ditches such as Koell where

channel morphology marginalizes macrophytes to the banks of streams, restricting plant-water









contact and thus limiting direct uptake of nutrients (Reddy et al. 1999). Nutrient content of

macrophytes growing in wastewater tended to be higher than those out of wastewater particularly

for plants commonly found in the marsh in Frey. With the exception of Frey, the maj ority of the

macrophytes in the present study are likely to obtain their nutrients from sediment porewater

rather than directly from the water column. Results of plant tissue N and P content of the ditch

macrophytes indicate high P enrichment, with average belowground N:P ratios around 4 for all

sites. N:P ratios below 14 to 16 are commonly associated with P enrichment relative to N

(Koerselman and Meuleman 1996, Richardson et al. 1999). One possible explanation for

belowground enrichment of biomass P are advective gradients established by plants roots that

can drive the exchange of water and dissolved material from the water column into porewater

spaces allowing nutrients to become available for direct uptake (Reddy et al. 1999). The

estimates of N and P retention by uptake in this study do not consider the eventual release of

stored nutrients and thus it is likely that there is minimal net nutrient retention by biomass. Kim

and Geary (2001) found that biomass harvesting removed less than 5% of TP present in

microcosms planted with two species of wetland macrophytes. More than 95% was stored in

substrates. In wastewater drainage ditches, it is likely that the same situation is true: rather than

stored in plants, the maj ority of P is associated with benthic sediments.

Conclusions

This study demonstrated the influence of macrophytes on transport characteristics in

ditches by creating zones of transient storage. By slowing transport times, these storage zones

promote reductions of E coli bacteria, TSS and SRP. However, large within-site variability of

water quality limited the ability of this study to isolate more direct evidence of in-stream

retention. Future studies should take this variability into account by performing mass-balance

studies over longer time periods such as 24 hours. Net nutrient retention by plant uptake is










minimal in wastewater ditches; instead, the treatment function of plants arises from their ability

to create in-stream transient storage zones, resulting in lower-velocity, depositional

environments.











































KE = kinetic energy
Sed =sediment
Path = pathogens
OM = organic matter
N =nutrients
DO = dissolved oxygen
M =microbes


Wastewater ditch


Figure 3-1. Systems diagram illustrating the influence of macrophyte harvesting on water quality
in wastewater ditches. Plants reduce water velocities (expressed as kinetic energy)
thereby promoting retention mechanisms such as sedimentation and sorption. Plants
also take up and eventually return nutrients.





























)l I








A Frey sampling locations



Figure 3-2. Map of Oxapampa showing the locations of the sampling sites in each of the study ditches. The San Martin, Gustayson
and Koell sites discharge to the river while Frey often does not flows from the marsh (location 2). Numbers represent
sampling locations from upstream ("1") to downstream ("4").


ON.. .am pa,. Pe ru
Scale 1:5000









Table 3-1. Dominant macrophyte species common to the study sites
Species Common name San Martin Gustayson Koell Frey
Cynadon dactylon Bermuda grass X X X X
Eichornia cra~ssipes Water hyacinth X
Equisetem spp. Horsetail X
Hedychium coronarium Ginger lily X X X X
Hydrocotyle ranunculoides Water pennywort X X
M~yriphyllum aquaticum Parrot feather X
Polygonum punctatum Knotweed X X X


Figure 3-3. Ditch site San Martin looking upstream from sampling location 4 in A) November
2005 and B) June 2006
























































Figure 3-4. Gustayson site looking downstream at times of solute transport experiments in
November 2005 at A) three weeks after plants were harvested and B) lower reach,
location 4 before plant removal; C) in June 2006; and D) in July 2006, one week after
vegetation removal and herbicide application.
































Figure 3-5. Koell site looking downstream from sampling location 1 at the time of tracer
experiments A) in November 2005 and B) July 2006


Figure 3-6 Frey site at the marsh location at the time of biomass sampling in 2005










Table 3 -2. Description of tracer experiments conducted at each site
Site Date Experimental conditions of tracer experiment
Gustayson 15 November 2005 Reaches 1-3 cleared of vegetation; reach 4 not cleared
22 June 2006 Reaches 1-4 with vegetation
13 July 2006 Reach 1-4 cleared of vegetation
San Martin 24 November 2005 Low biomass
23 June 2006 High biomass
Koell 24 November 2005 Low biomass
7 August 2006 High biomass, trash removed from channel


Figure 3-7. View of the Mariotte siphon used for delivering the conservative tracer at a constant
flowrate










Table 3-3. Channel characteristics and mean water quality parameters for wastewater drainage


ditch sites

Channel slope (%)
Channel length (m)
Wetted width (m)
Discharge (L s-')

DO (mg L-1)
Temperature (oC)
pH
Conductivity (CIS cml)

TSS (mg L-')
CBOD5 (mg L-')
E. coli (CFU 100 mL ')

SRP (mg L-')
TP (mg L-')
Organic P (mg L ')d
NO3-N (mg L-')
NH4-N (mg L-')
TN (mg L-')
Organic N (mg L ')e


San Martin
1.0
356
1.09
45+20 (97)e


Gustayson
1.4
360
0.92
19+11 (64)


Koell
1.9
194
0.83
12+7 (33)


Frey
4.3"
210
0.70b
5+6 (11)


2.5+1.6 (87)
17.6+2.9 (87)
6.7+0.3 (75)
155+49 (64)

32.9+29 (88)
4.6+1.3 (5)
3684+2322 (45)

0.31+0.28 (97)
1.82+0.88 (33)
1.48+0.88 (33)
0.03+0.03 (96)
2.67+1.67 (97)
4.83+2.10 (34)
1.55+2.11 (34)


1.0+0.9 (80)
18.7+1.0 (80)
6.7+0.2 (63)
243+44 (57)

42+47 (83)
18.1+12.1 (15)
4189+4519 (36)

0.56+0.31 (90)
4.23+1.56 (33)
3.59+1.60 (33)
0.02+0.02 (89)
5.96+2.81 (89)
8.25+3.78 (38)
2.36+3.06 (38)


1.4+1.0 (28)
20.9+2.2 (32)
7.0+0.2 (26)
244+69 (32)

63+50 (40)
10.6+6.1 (8)
7485+7861 (33)

0.42+0. 14 (57)
3.06+0.98 (16)
2.62+0.91 (16)
0.02+0.01 (53)
6.00+2.33 (53)
7.11+2.06 (16)
0.57+1.87 (16)


1.9+1.8 (24)
19.3+2.4 (28)
6.8+0.3 (25)
255+46 (25)

83+79 (24)
8.5+6.3 (11)
6406+6263 (27)

0.36+0.18 (28)
3.40+2.86 (10)
3.00+2.77 (10)
0.06+0.09 (28)
5.80+2.02 (28)
6.09+2.82 (10)
1.40+2.56 (10)


a This average slope includes the escarpment. Above and below this point the channel slope is 1.6% and 0.8%,
respectively. b Averaged for the entire site and when the marsh (site 2) was ditched. Values are averages + 1 SD.
Values in parentheses are numbers of samples. d Organic P is calculated as the difference between the TP and SRP
and includes polyphosphates and acid hydrolyzable P. eOrganic N is calculated as the difference between TN and
NO3-N and NH4-N.











Koell (distance = 73 m)
5900 2400**
30000

20000

10000


Upstream Downstream

11 55.5

150

100

50


Upstream Downstream


42.5 37.8








Upstream Downstream


-1
E
o12000

0 8000

=4000
o


Upstream Downstream


Figure 3-8. Box plots of San Martin and Koell at upstream and downstream sampling locations
for A) E. coli and B) TSS. Solid circles and numbers indicate median concentration
values. Open circles denote outlier values and boxes and whiskers delimit the inter-
quartile range (IQR) and 1.5 IQR, respectively. and ** following values indicate
p < 0.05 and p < 0.01.


San Martin (distance = 255 m)

600 4400*











Koell (distance = 73 m)
3.7 2.3








Upstream Downstream

8.4 5.9*









Upstream Downstream

Koell (distance = 73 m)
0.42 0.48








Upstream Downstream

5.6 5.6








Upstream Downstream


San Martin (distance = 255 m)

5 A2.2 1.3








Upstream Downstream


12 4.8i 4.3







Upstream Downstream

San Martin (distance = 255 m)
1.0
0.8 O0.27 0.19**


0.8 Ct



0.0
Upstream Downstream


10 2.5 2.4







Upstream Downstream


Figure 3-9. Box plots of San Martin and Koell at upstream and downstream sampling locations
for A) TP, B) TN, C) SRP and D) NH4-N. Solid circles and numbers indicate median
concentration values. Open circles denote outlier values and boxes and whiskers
delimit the inter-quartile range (IQR) and 1.5 IQR, respectively. and **
following values indicate p < 0.05 and p < 0.01.











-+ San Martin (low biomass)
San Martin (high biomass)
-:a Koell (low b iom ass, with tra sh)
tKoell (high biomass, no trash)
-a- Gustayson (low biomass)
GGustayson (high biomass)

5.3 m






S0.5-
O,


6000






S2000

O


j T.5
z

E~ 5'
z

= 25:
r


0 100 200 300 400


O 100 200 300 400


Dista~nce (m) Distance (m)]


Figure 3-10. Longitudinal water quality trends of E coli, TSS, SRP and NH4-N during tracer
experiments performed in sites San Martin (circles), Koell (stars) and Gustayson
(triangles, shown for 2006 tracer experiments only) under both high (bold lines) and
low (thin lines) biomass conditions. Note: lines between symbols are used only to aid
visualization











1.5
r With plants (22 June 2006) o No plants (13 July 2006)









0.0 0.5 1.0 1.5 2.0 2.5 3.0
Time since injection (hours)

Figure 3-11. Comparison of conservative tracer breakthrough curves in Gustayson before and
after removal of ditch vegetation in 2006


o Low biomass (November 2005) High biomass (June 2006)
1.5



o

ce 0 o
.9 B 0.


0.0 0.5 1.0 1.5 2.0 2.5
Time since injection (hours)

Figure 3-12. Comparison of conservative tracer breakthrough curves in San Martin with low and
high ditch vegetation biomass


a Low biomass (November 2005)
1.5
High biomass, no trash (July 2006)
o 5



00o


0~~. "o 0 o oo oo o onoone0
0.0 1.0 2.0 3.0
Time since injection (hours)

Figure 3-13. Comparison of breakthrough curves in Koell under different biomass conditions and
after trash was removed from the stream










Table 3-4. Biomass and hydrologic conditions of each tracer experiment and resulting transient storage zone parameter values
Site Aboveground Discharge Velocity Tmed Fmed Fmed20)( A As As/A a Dal
Biomass
(g m ) (L s-') (m s ) (min) (%) (%) (E-2 m ) (E-2 m ) (E-4 s ')
Gustayson (2005)" Negligible 13.2 0.23 nab 18 24 8.07 24.0 2.97 4.53 0.36
(upstream reach)
1823 & 655 25.5 0.14 na 62 58 12.2 142 11.60 10.4 1.20
(downstream reach)

Gustayson (2006) Negligible 8.6-17.9 0.32 27.5 Oc 0" 4.47 0.00034 0.008 0.0006 0.81
2021 &413 7.4-11.8 0.11 66.7 21 21 7.92 2.13 0.27 27.9 47

San Martin 731 + 172 32.4 0.53 25.6 5.8 3.4 4.64 1.66 0.36 3.69 0.94
1962 & 659 38.0 0.25 48.4 56 56 28.8 38.4 1.33 30.5 7.6

Koell 274 & 71 2.4-3.1 0.11 27.3 52 69 3.93 59.4 15.1 7.26 0.86
(with trash)
1995 A 1030 5.0-7.9 0.17 24.0 0.89 0.92 3.84 0.27 0.07 1.26 2.2
o\ (no trash)
a See Table 3-2 for distinction between 2005 and 2006 tracer experiments in Gustarson. bCOmparisons of median travel were not made because reach lengths
were not comparable in this experiment or with the following 2006 Gustarson experiment. Fmed and Fmed")() were 4.79 E-9 and 2.62 E-9, respectively.









Table 3-5. Estimated P and N uptake rates based on measured above and belowground biomass and plant tissue nutrient storage at
each study site
San Martin Gustayson Koell Frey
Above Below Above Below Above Below Above Below
Biomass (g m-) 730 & 172 406 & 194 1823 & 655 1157 & 768 274 & 71 322 & 96 311 + 194 177 & 45
Biomass growth (g m-2d )a" 12 & 3 nab 20 & 7 na 9 &2 na 7 & 6 na

P content (mg P g- ) 11.4 & 3.8 4.6 & 1.0 10.6 & 1.5 4.5 A 1.5 7.2 & 1.1 3.6 & 0.6 9.0 & 2.5 4.3 A 1.3
N content (mg N g ) 66.8 & 9.2 19.5 & 8.0 52.4 & 7.5 15.9 & 4.6 51.7 & 13.3 13.3 +4.1 60.2 & 18.9 16.6 & 2.9


P uptake (g P d )0 53 na 70 na 10 na 19 na
(27-88) (39-108) (7-15) 2-44
N uptake (g N d ) 311 na 347 na 75 na 125 na
(201-383) (193-535) (43-115) (12-304)
"Daily biomass averaged as biomass at time of sampling divided by days since harvest, estimated as 90, 60, 30 and 30 days in Gustayson, San Martin, Koell and
Frey. bGrowth and uptake rates were not calculated for below ground biomass due to lack of remaining biomass data after plant harvesting. P and N uptake is
calculated by the product of biomass growth, ditch area and nutrient content of plant tissues, where area was assumed to equal the channel length multiplied by
average wetted channel width (See Table 3-1). For Frey the area used includes the wetland area with a width of 2 m.

Table 3-6. Average P and N loads and estimated gross nutrient retention by aboveground biomass assimilation at each study site
San Martin Gustayson Koell Fre
Mean P load (g P d ') 1195 924 432 162
Mean N load (g N d )b 3,225 12,740 6,190 2,633

Gross P Retention (%)" 4.4 7.6 2.3 11.7
Gross N Retention (%) 9.6 2.7 1.2 4.7
"Mean P load is the SRP concentration multiplied by discharge using mean values from Table 3-1. bMean N load is the sum of the NO3-N and NH4-N
concentration multiplied by discharge using mean values from Table 3-2. "Gross P and N retention are calculated as the sum of the above and belowground P and
N uptake values divided by the P and N loads.










Table 3-7. Mean (11SD) molar ratios of carbon, nitrogen and phosphorus of aboveground (AG)
and belowground (BG) plant tissues between sites
San Martin Gustayson Koell Frey
AG BG AG BG AG BG AG BG
N:P (molar) 13.5 4.0 10.1 3.6 14.1 3.6 12.5 4.2
(2.1) (0.8) (2.4) (0.3) (2.1) (0.7) (2.6) (1.4)

C:N (molar) 22.8 12.9 25.0 13.9 37.5 14.5 26.2 12.6
(1.1) (1.4) (1.3) (0.9) (2.8) (2.0) (9.8) (1.4)

C:P (molar) 148.1 51.3 120.8 49.7 250.7 52.7 154.5 53.8
(18.0) (13.4) (21.4) (5.4) (43.0) (16.4) (62.7) (21.4)

12 Leaves 5 Stems O Roots 7




Levs~ Specis Species7

Fiue31.Cmaio flaf tmadro/hzm ise n otn fmcoht
species~~~~~~ fon nwseae iths(asadlns ersn en D











12 wastewater a control 60







Speie Speie

Figure ~ ~ ~ ~ ~ ~ ~ ~ C 3-5 oprsno omnseisinOaap htaefudgoigi

* an *dnt 00 n .1









CHAPTER 4
PHOSPHORUS RETENTION BY SORPTION WITH BENTHIC SEDIMVENTS

Introduction

Surface waters in Oxapampa tend to have elevated phosphorus (P) concentrations due to

the use of fertilizers, phosphate detergents and discharges of raw domestic effluents (See Figure

1-8). The high P concentrations have led to rapid increases in filamentous algae cover along

stream banks of the Chorobamba River over the last several years indicating that the river system

is becoming increasingly eutrophic. While local water quality concerns are focused primarily on

those related to human health due to the presence of pathogens, there is also need for reducing

nutrient loads to the river to avoid adverse impacts to the local Eisheries upon which local

indigenous communities rely.

There are significant differences in phosphate concentrations observed between effluents

discharged from wastewater ditches and underground pipes (see Chapter 1). Lower phosphate

concentrations in ditches than in pipes may be caused by P sorption with benthic sediments.

Understanding the physico-chemical characteristics of benthic sediments and the nature of P

adsorption is important for determining the stability of P retention (Reddy et al. 1999), not only

in these ditches, but also in the Chorobamba River, the receiving water-body for ditch sediment

exports during storm events. The obj ective of this study was to evaluate the role of wastewater

ditch sediments for promoting phosphate retention. Specific research obj ectives were: (1)

Determine how much P is stored in benthic sediments and how stable it is; (2) Identify physico-

chemical factors responsible for P sorption; (3) Evaluate whether ditch sediments tend to act as P

sinks or sources; and (4) Examine the capacity of sediments for further retention.









Methods


Experimental Sites

Sediment P characteristics were studied in four wastewater ditches in 2005 (San Martin,

Gustayson, Koell and Frey) and in three ditches in 2006 (San Martin, Gustayson, Koell). As

described in Chapter 3, ditch vegetation is removed by the municipality 2 to 3 times per year and

ditch sediments are dredged approximately every two years to prevent the risk of flooding. The

ditches selected for this study are the most downstream drains in Oxapampa; thus transport along

these reaches provides the last opportunity for phosphorus (P) retention (or release) before

discharge to the Chorobamba River. The downstream position of these ditches (see Figure 3-1)

also suggests that they are likely to be the most impacted sites with the highest P loads. The sites

were also chosen to provide a range of physical, biological and chemical attributes to allow an

examination of relationships between ditch features and P sorption.

Three to four locations were sampled in each study site (see Figure 3-1) to inspect P

sorption properties along a longitudinal profile, from upstream to the most downstream point

before discharge to the river. San Martin (Figure 4-1) has the highest flows of the four study sites

but because approximately a quarter of the flow comes from a natural stream, it generally has the

lowest P concentrations. Gustayson (Figure 4-2) has both high flows and high P loads. The ditch

exists along four city blocks and is surrounded by homes that discharge wastes along the total

length of the ditch causing discharge to increase with distance downstream. Koell (Figure 4-3)

has lower flows than Gustayson and higher P concentrations than San Martin. Frey (Figure 4-4)

has the lowest flows of the four study ditches and on occasion dries out. The lower reaches of

Frey are often ponded creating a marsh environment with connection to the river only during

high flow events (see Chapter 3 for more detailed site descriptions).









Field Procedures

Benthic sediments (0-5 cm) for each sampling location were collected. Using a trowel,

sediments were dug out from the middle and sides of the channel and composite. A total of 15

samples (4 each in San Martin, Gustayson and Koell; 3 in Frey) were collected in 2005 and 12

samples in 2006 (Frey was not sampled in 2006). Grab samples of water were taken following

the field procedures described in Chapter 3 at the time of sediment collection for the P sorption

experiments (described below) and also during a 13-month period (March 2005 to September

2006) to determine general water quality characteristics for each site over time. Temperature,

pH, dissolved oxygen and conductivity were measured in-situ as described in Chapter 3.

Laboratory Procedures

Water chemistry and sediment characterization

Water samples were analyzed for total suspended solids (TSS), soluble reactive P (SRP),

nitrate and ammonium as described in Chapter 3. A sediment sub-sample from each sampling

site was air dried for particle size determination, soil pH and percent organic matter (%OM).

Particle size determination for sands (0.06 2 mm) and gravels (> 2mm) were determined using

nested sieves following ASTM standard methods (ASTM 1985). For determination of silt and

clay fractions the pipette method was used as described by Gee and Bauder (1986). Sediment pH

was determined using the method described by Thomas (1996). To approximate % OM,

sediments were oven dried for 4 hours at 1050C, weighed at room temperature and then

combusted at 5500C for 4 hours and reweighed to determine the percent loss by combustion.

Phosphorus sorption index

The P sorption index (PSI) provides a simple measure of the P buffering capacity of

sediments (Bache and Williams 1971). High PSI values indicate that large quantities of P can be

removed by the sediments without increasing the P concentration in the water at equilibrium









(Klotz 1985). The PSI was determined using the approach presented by Klotz (1985) and later

modified by Haggard et al. (1999). Fresh sediments were sieved to retain only particles < 2 mm.

A 100 mL solution spiked with a P concentration (50 mg P L^1) equivalent to 1 mg P g sediment

(dry) was added to approximately ten grams (dry weight) of sediments. Equilibration of

sediments at this high P concentration was assumed to saturate available sorption sites, thus

providing a measure of maximum sorption (Reddy et al. 1995). The equilibrations for each

sampling site were performed in triplicate and shaken for 1 hour after which a 15 mL aliquot was

removed, filtered and analyzed for SRP. Sediment sorption was determined from the difference

between the P in solution before and after the equilibration period. The sediments were dried

overnight at 1050C to express P sorbed per dry weight of sediment. The PSI was calculated as

X/(logloC) where Xis the P sorbed per dry weight of sediment (Cpg P g- ) and C is the final P

concentration in solution (Cpg P L^).

The relative role of biotic versus abiotic processes in P sorption was examined for the

sediment samples collected in 2005. The same procedure was followed for determining PSI as

described above except that for each site an additional sub-sample was analyzed after being

autoclaved for 20 minutes (Haggard et al. 1999). The difference in sorption capacities between

the intact and autoclaved sediments is assumed to represent the contribution of P sorption from

biotic processes alone.

Equilibrium phosphorus concentration

The sediment equilibrium P concentration (EPCo) is the sediment porewater P

concentration at which adsorption of P to sediment surfaces is in equilibrium with P release from

sediments such that adsorption equals desorption (Reddy et al. 1995). EPCo measurements

provide insight into whether sediments tend to function as P sinks or sources of P to the

overlying water. If the EPCo value of the sediment porewater is less than the P concentration in









the overlying water, the concentration gradient results in sorption to sediments. If the EPCo

value is greater than the P concentration in the water, the gradient is reversed and P is released to

the overlying water. The EPCo is determined from the y-intercept of the linear portion of a P

adsorption isotherm (Figure 4-5). To create P isotherms for each site, ditch water samples were

filtered with 0.45 Cpm membrane filters and subsequently spiked with +0.00, +0.10, +0.25, +0.50,

and +2.00 mg P L (Smith et al. 2005). A volume of 100 mL of spiked sample water was added

to approximately 25 g wet sediments (particle sizes <2 mm) and shaken for one hour. After the

equilibration period, samples were filtered and analyzed for SRP. Each equilibration was

performed in triplicate for a total of 15 samples for each sampling site (five different P

concentrations, three repetitions). The sediments for each sample were dried overnight at 1050C

and weighed to calculate P sorbed per dry weight of sediment. These values were regressed

against corresponding P concentrations to determine the EPCo value for each sampling site.

Phosphorus fractionation

The chemical forms of P adsorbed to sediments indicate the stability of P retention in

aquatic systems. The quantities of the various forms of sediment P thus indicate how much P is

stored and whether sediments function as long-term P sinks or sources. The basis of fractionation

procedures is that various forms of sorbed P exhibit differential solubilities. Using various

extracts, P can be solubilized and quantified. P fractions were determined by sequential

extractions following the chemical fractionation scheme described by Moore (2000):

* 1 M~KCl to remove the most labile inorganic P pool

* 0.1 M~NaOH to remove inorganic P sorbed by amorphous oxyhydroxides (bioavailable and
readily desorbed) and crystalline Fe/Al oxides (Fe-bound P desorbed only under anoxic
conditions). This extraction step also removes organic-bound P.

* 0.5 M HCI to remove inorganic P bound with calcium and magnesium

* persulfate acid digestion to remove residual inorganic and recalcitrant organic P.









Sediments from each sampling site were collected from the middle and sides of the

channel and composite in the Hield (approximately 20 g dry weight) in pre-weighed centrifuge

tubes and filled with ditch sample water just above the sediment surface to avoid sediment

oxidation. The tubes were capped and returned to the laboratory where they were centrifuged for

30 minutes. The supernatant was filtered using a 0.45 Cpm membrane fi1ter and analyzed

immediately for determination of porewater soluble reactive phosphorus (SRP). After

reweighing the tubes to determine the water removed, sediments were homogenized using a

spatula and two subsamples (approximately 1 g dry weight each) were removed to perform

extractions in duplicate. Sediments were added to pre-weighed centrifuge tubes along with 20

mL of 1 M~KC1. The tubes were shaken for 2 hours, centrifuged for 30 minutes and the

supernatant was filtered and analyzed for SRP for determination of the loosely-sorbed P fraction.

A similar procedure was followed for the next extraction step using 0. 1 MNaOH and an

equilibration time of 17 hours. However, after centrifuging, half of the supernatant collected was

filtered and analyzed for SRP for determination of Fe and Al-bound P while the other half was

digested and analyzed for total P. The difference between TP and SRP represented the organic-

bound P. The next extraction step used 0.5 M~KCl and a shaking time of 24 hours to remove Ca

and Mg-bound P. Supernatants were analyzed for SRP. The remaining sediments were air dried

and collected in 20 mL scintillation vials where they were later analyzed for residual P by

persulfate digestion followed by SRP analysis. Total adsorbed P was determined by summing the

various P fractions.

Statistical Analyses

Sediment characteristics and water quality data were In-transformed to meet normality

requirements. An arcsine-square root transformation was used for values expressed as

percentages. Comparisons of sediment and P sorption characteristics from years 2005 and 2006









were made using paired t-tests. An a priori significance level was set at 0.90 due to low sample

sizes resulting from grouping the data into 3 sites.

Results

The four ditches differed from one another with respect to their channel characteristics,

flowrates and water chemistry (see Table 3-3). Sampling of sediments in 2005 took place

between November 11 and November 15 which coincide with the beginning of the rainy season

in Oxapampa. In 2006, sediments were collected between June 26 and July 11 which correspond

with the latter part of the dry season (Figure 4-6).

Sediment characteristics varied between the different ditches and between years 2005 and

2006 (Table 4-1). There are notable differences in the particle size distributions between years in

San Martin and Koell. In 2005, there tended to be higher proportions of sands (94% and 87% in

San Martin and Koell, respectively) than in 2006 (50% and 26%). A similar trend was seen in

Gustayson when the second sampling site which was extremely mucky and had a very high

percentage of clays (over 40%) was omitted from the analysis. Overall, sediments collected in

2006 had significantly higher clay (paired t-test, t = -6.6, p = 0.01) and organic matter content

(paired t-test, t= -2.18, p < 0.1) than in 2005. Average sediment pH ranged between 5.5 and 6.7

for San Martin, Gustayson and Koell. Sediment pH in Frey was 7.0. Sediment pH was lower in

2006 than in 2005 (paired t-test, t = 4.06, p = 0.03). However, these values were determined

from air-dried sediments and therefore may be lower than pH values measured in situ due to a

change in redox conditions as the sediments were removed from an anaerobic environment

(Sallade and Sims 1997a, b)

Results of the single point P isotherm experiments indicate that the sediments are capable

of buffering additional P loads, particularly in the 2006 sampling period (Table 4-2). Koell

tended to have the highest overall PSI values of the three sites in 2006; however differences in









PSI values between sites were not significant (p > 0.10) due to high within-site variability

(Figures 4-7 and 4-8).

Comparison of the longitudinal trends observed for PSI and %OM at each sampling site

(Figures 4-7 and 4-8) suggests a positive association between PSI and %OM, especially for the

San Martin and Koell sites. Gustayson sediments appear to deviate from this relationship

suggesting that controls, other than OM content govern P buffer capacity at this site.

Nevertheless, strong relationships exist between particle size, OM content and K, an alternative

measure of P buffer capacity, when the data for all sites and years are pooled together (Figure 4-

9).

Despite the relationships observed between particle size and OM content, comparisons of

the contributions of abiotic and biotic sorption performed in 2005 suggest that biotic sorption

largely accounts for P buffer capacity rather than abiotic factors (Figure 4-10). Interpretation of

this result is likely to be misleading due to the use of autoclaved sediments in the PSI

experiment. Inhibition of microbial activity via autoclaving sediments has been shown to result

in an underestimation of the importance of abiotic sorption (Klotz 1985). This issue is described

in more detail in the discussion section below.

Comparison of ditch water SRP and sediment EPCo values (Figure 4-1 1) suggest that at

the time of sampling San Martin, Koell and Frey sediments tended to remove SRP from the

water column, thus acting as a P sink. In 2005, three of the 4 sampling points in Gustayson had

higher sediment EPCo values relative to ditch water SRP, indicating that sediments were

functioning as a source of P to overlying water. In 2006, one site in both Gustayson and San

Martin functioned as a P source while the remaining three sites were sinks for P. Sediments in

Koell appeared to act as a P sink in both 2005 and 2006. No strong trends were found between









EPCo and sediment particle size or %OM. However, increases in sediment EPCo values were

observed to be related to water column SRP values; this trend was strengthened by grouping the

data into sites that act as sinks and those that act as sources (Figure 4-12). Moreover, EPCo

values appear to be indirectly related to particle sizes and OM content due to the relationship

between PSI and EPCo (Figure 4-13). As the sediment P buffer capacity decreases, EPCo values

tended to increase.

Phosphorus fractionation performed on San Martin, Gustayson and Koell sediments in

2006 reveal marked differences in the distribution of sorbed P forms despite the close proximity

of the study sites (Figure 4-14). The readily labile P pool (KCl-Pi) comprised an insignificant

fraction of the total P (< 1% for all sites). Sorption with Fe/Al oxyhydroxides made up the largest

P fraction for San Martin (44.6%) and Gustayson (74.6%) and the second largest fraction for

Koell (31.6%). Most likely, the readily labile P fraction in situ was larger but oxidation of

sediments resulted in this fraction becoming adsorbed to freshly oxidized ferric surfaces (Moore

2000). Organic-bound P was a significant fraction for San Martin (37.0%) and Koell (38.3%) but

not for Gustayson (4.3%). Ca and Mg-bound P was an important P fraction in Gustayson

(17.0%) and Koell (17.3%) but not for San Martin (0.06%). The recalcitrant P fraction,

representing a long-term P sink was 18.3% and 12.8% in San Martin and Koell, respectively and

was markedly lower in Gustayson (3.3%).

The P fractions were summed to estimate the total sediment P for each site (Figure 4-15).

This total is an underestimate due to the absence of the HCl-P organic fraction. Gustayson had

approximately 10 times more adsorbed P than San Martin and Koell; however, all sites exhibited

very high P concentrations (site means between 2000 and 22,000 mg P kg )~. Longitudinal trends









of TP, from upstream to downstream, illustrate high variability with distance along each ditch

site (Figure 4-16).

Discussion

Effect of Seasonal Trends in Particle Size and Organic Matter Content on P Buffering
Capacity

The significant difference in particle-size distributions and organic matter content in

November 2005 and July 2006 in all study ditches reveals the dynamic nature of these systems

through time and space. Heavy rains between October and April affect ditch sediments

characteristics by exporting fine particles and organic matter and depositing coarser materials

resulting in higher sand fractions observed in each of the sampling sites in November. Ditch

management practices likely exacerbate sediment exports by controlling vegetation towards the

beginning of the rainy season to avoid potential flooding. In contrast, sediment erosion and

export is markedly decreased during the dry season which is characterized by infrequent and less

intense rain events. Hence, during this period, ditch vegetation is minimally managed due to the

low risk of flooding. The resulting proliferation of ditch vegetation likely promotes deposition of

fine sediments and the accumulation of organic matter as observed in the 2006 sampling.

Changes in particle-size and organic matter content between the wet and dry season appear

to have directly influenced the overall P retention capacity of the ditch sediments. Not only are

buffer capacities significantly higher in the dry season but EPCo values were much lower,

indicating that sediments will tend to remove P rather than release P from the water column.

Strong relationships between fine particle sizes and organic matter content on P buffering

capacity have been reported by many studies (e.g. Meyer 1979, Nguyen et al. 1997, Axt and

Walbridge 1999). By comparing these relationships between two distinct seasons, this study has









shown that the operating mode for P retention in wastewater ditches may be highly variable

through time in response to alterations in sediment characteristics.

Phosphorus Sorption Index

To understand how the buffering capacity of the wastewater ditches compares with that

reported for other systems, PSI values from 2005 and 2006 were tabulated with PSI values

observed for other systems including wetlands and streams (Table 4-3). This comparison

revealed that sediment PSI values in 2005 were similar to values reported for other systems and

in 2006, values were within the upper limits or exceeded reported values. The PSI values from

this study were most similar to other enriched drainage ditch systems and wetlands. The

similarity of high P buffering capacity between ditches and wetlands supports the idea that it

may be possible to manage drainage ditches as wetlands for mitigation of nutrient loads to

downstream receiving bodies (Bowmer et al. 1994, Nguyen and Sukias 2002).

Summarizing PSI across multiple studies also brought to light several issues regarding

both the reporting of PSI values and PSI methodologies that affect the interpretation and

comparison of results. The first of these issues to be addressed is the units used to report PSI

values. As illustrated by Table4-3, the choice of units used to report PSI values strongly affects

the range of values reported. Due to the formulation of the PSI equation [PSI = X (log C) '],

some studies likely report values for C using Cpg L^1 or Cpmol L^1 instead of mg L-1 to avoid taking

the log of a value less than or equal to 1. However, the use of varying units for C may preclude

direct comparison between studies unless other data is provided to allow unit conversions of the

(log CJ~' factor. As a result of the inability to directly convert PSI values from other studies to

one common set of units, PSI values from this study were instead converted to three different

sets of frequently reported units in Table 4-3









The second difficulty in comparing PSI values between systems is related to the differing

methodologies used to measure PSI. Table 4-3 provides a general comparison of methodological

approaches used in each study that may significantly influence the values reported. These

approaches include choice of sediment-water equilibration time, the initial P concentration used

to spike sediments/soils, and inhibition of microbial activity.

The choice of sediment-water equilibration time is an important factor that varied between

the studies included in Table 4-3. Most P sorption isotherm studies maintain sediment-water

contact times between 16-24 hours to ensure maximum soil-water contact. The present study

used a one-hour equilibration time ( Klotz 1985, Haggard et al. 1999) due to logistical

difficulties associated with the remoteness of the study area (i.e. acquiring a working shaker

table). Thus, sorption measurements in the present study represent an underestimation of the total

sorption potential of the ditch sediments. However, the magnitude of the difference in sorption

between a 1-hour and a 24-hour equilibration time may be strongly related to soil texture. Meyer

(1979) noted that for sorption experiments using silty sediments, 93% of the total P in solution

was removed in the first five minutes. Using sandy sediments, only 19% was removed in five

minutes; however, after 24 hours, approximately the same amount of P had been sorbed onto

both the silty and sandy sediments. These findings suggest that the underestimation of sorbed P

in the present study may be particularly significant for the sandier sediments that characterized

the study sites in 2005 and perhaps less so for the finer sediments from 2006. This finding may

have some bearing on the low PSI values reported by Haggard et al. (1999) and (2001) since the

systems in both studies were dominated by sandy sediments. It is worthy to note that despite the

one-hour equilibration time and the dominance of sandy sediments in 2005, the wastewater

ditches in the present study exhibited some of the highest PSI values reported.









Despite the likely underestimations of total P sorption capacity, the sorption tests

performed in this study provide invaluable insight into the rate of P sorption and how this factor

is influenced by changes in sediment particle-size distribution. The importance of the ability to

quickly retain P is especially highlighted in systems that are subj ect to short-term, concentrated

pulses. Pulses of high P loads associated with human activities, such as laundry washing with

phosphate detergents, are common in wastewater drainage ditches in Peru. Therefore, an

understanding of the sediment response to such flashy perturbations is likely to be as important

as knowing the total sorption capacity when examining the environmental fate of P in ditch

systems.

A second factor that is likely to influence PSI values is the concentration of P used to spike

the sediments/soil porewater. The concentrations used in the studies reported in Table 4-4 varied

between 2 mg L^1 and 1000 mg L^1 and not surprisingly, the studies using the lowest

concentrations also tended to have the lowest PSI values. To make PSI values comparable across

systems, the same initial P concentration should be used.

A final important difference between the studies reported in Table 4-3 is whether or not the

study inhibited microbial activity before performing the PSI analysis. The choice of microbial

inhibition is strictly a matter of the research goals of the particular study. Some studies are

interested in examining total P sorption capacity of the sediments while others focus solely on

abiotic sorption and to do so, microbial activity is inhibited by autoclaving sediments or adding a

toxic chemical such as toluene, chloroform, or carbonylcyanide m-cholorophenylhydrozone (CP)

before equilibration. The effect of microbial inhibition on comparing PSI values for various

systems is a factor of the importance of biotic sorption to overall sorption capacity. Results of

PSI values in this study in 2005 indicated that biotic sorption tended to dominate over abiotic










sorption. Similar findings were reported by Haggard et al. (1999). Other studies have reported

much lower contributions of biotic sorption to total sorption (Meyer 1979, Klotz 1985). It would

have been useful to compare biotic versus abiotic PSI values for 2005 and 2006. The relative

influence of abiotic sorption would likely to be greater in 2006 due to the higher percentage of

silts, clays and organic matter in ditch sediments.

However, it is extremely likely that abiotic sorption was significantly underestimated in the

ditch sediments in 2005 because autoclaving sediments releases cellular P (Klotz 1985). In two

of the sampling sites in Koell, autoclaved sediments, performed in triplicate, resulted in negative

PSI values (note the error bars for abiotic sorption in Figure 4-11). In other words, more P was in

solution after the equilibration period than was in the initial spiked P solution thus implying that

ditch sediments released P during the sorption experiment. These findings support that the use of

autoclaved sediments will likely underestimate the importance of abiotic sorption and will thus

overestimate the importance of biotic sorption (e.g. this study, Haggard et al. 1999).

The PSI is a valuable index for describing the P buffer capacity of a system. However, to

optimize its utility, especially for cross-study comparisons, standardization of units used for

reporting values, equilibration times and initial P concentrations should be adopted. The use of

the PSI to further investigate the rate of P uptake for different systems and sediment textures

would be useful for understanding how well sediments in lotic systems can buffer pulses of P

loading.

Sediment Equilibrium Phosphorus Concentrations

Sediments from the Gustayson ditch tended to have the highest EPCo values and were

furthest from aqueous P-sediment P equilibrium, represented by the 1:1 line in Figure 4-12. This

departure from equilibrium suggests that sediments exert little control over regulating solution P

concentrations in the system. This finding is contrary to studies performed in pristine systems









(Froelich 1988) but similar to studies from impacted systems (Ekka et al. 2006). The regression

between SRP and EPCo was strengthened by separating out the ditch sites that were shown to act

as sinks from those that act as sources as shown in Figure 4-12. The slopes of the EPCo versus

SRP regression differ between the two groups with the slope of the P sink sites being notably

less than that for P source sites. This difference suggests that an underlying factor distinguishing

the two groups is their P buffer capacities. A lower slope indicates that EPCo changes less given

an increase in water SRP. Conversely, the steep slope shown for the P source sites suggests that

EPCo is sensitive to changes in SRP, thereby exhibiting a more limited P buffer capacity. The

negative power law relationship between EPCo and PSI supports this hypothesis (Figure 4-13).

The two P source sites in 2006 were Gustayson site 1 and San Martin site 4 which correspond to

the lowest PSI values for each of these sites as shown in Figure 4-8.

Given the relationship between EPCo and PSI, it is not surprising that PSI and sediment

EPCo values changed significantly between the 2005 and 2006 sampling periods. This finding

further supports that sediment P sorption characteristics are extremely variable through time.

This variability suggests that the common practice of comparing water SRP over time with a

single EPCo measurement to examine seasonal or yearly sorption/desorption trends may not be

appropriate for dynamic systems such as drainage ditches.

Phosphorus Fractionation to Examine P Status of Ditch Sediments

Phosphorus fractionation results indicate that despite the close proximities of the San

Martin, Gustayson and Koell sites, they nevertheless reveal quite different distributions of labile

and non-labile pools of P. Gustayson sediments had the highest TP concentrations (19,227 mg

kg- ) and also the highest levels of bioavailable P due to non-occluded Fe/Al oxyhydroxides

(14,317 mg kg- ). Comparably high TP values (25,261 mg kg- ) were reported by Nguyen (2000)

for sewage-impacted wetland sediments in New Zealand; however the dominant P fraction was









carbonate-P (15,121 mg kg- ), although the non-occluded Fe/Al-bound P fraction was also high

(2,342 mg kg- ). Near saturation of Gustayson sediments may explain the deviation of this site

from the trend between PSI values and OM content found for the other sites. The PSI of

Gustayson ditch sediments (site 4) declined to nearly half that of the upstream sampling location

despite having higher percent OM. This decline in buffer capacity may be due to an increase in

sediment TP (Figure 4-16) that caused sorption sites to become more limited.

While the most readily bioavailable fraction (loosely sorbed KCl-P) only represented 1%

of the sediment TP in Gustayson, it nevertheless was high (185 mg kg- ) relative to the San

Martin and Koell sites (<1 mg kg- ). One possible explanation for this difference is that San

Martin and Koell sediments were oxidized either during sediment collection or during the first

extraction step such that porewater and loosely sorbed P reacted with fresh ferric oxides surfaces.

However, handling of sediments was carried out in an identical manner for all sites so it is

unlikely that Gustayson sediments remained anaerobic during the first extraction step. Most

likely, sediments from all sites were equally oxidized. Ferrous oxides have much greater surface

areas available for P sorption than those for ferric oxides (Reddy et al. 1999). Perhaps extremely

high concentrations of P associated with ferrous oxides saturated newly-available ferric oxide

surfaces once Gustayson sediments were exposed to the atmosphere. The P remaining after

saturation of ferric oxide surfaces would then have been detected as loosely exchangeable P.

While Fe and Al were not specifically measured in the study, observation of the filters used to

measure SRP after the NaOH extraction step indicate the presence of oxidized iron (Figure 4-

17).

While ferrous oxide surfaces can retain more P via sorption, the low binding energy of this

reaction results in a high potential energy for re-release to the water column (Reddy et al. 1999).









Less binding surface areas are available under aerobic conditions but sorption is much more

stable. Therefore, anaerobic sediments high in iron content will tend to have high EPCo values,

which was indeed the case for the Gustayson sites.

The sediments from Koell and San Martin also had high fractions of Fe/Al-P pools but

contained high percentages of organic P associated with humic and fulvic acids. Organic P

bound with humic substances may be readily re-released to the water column under aerobic

conditions; however, under anaerobic conditions this fraction has been shown to be more

resistant to biological breakdown (Reddy et al. 1995, Reddy and D'Angelo 1997). Therefore, this

fraction in the anaerobic ditch sediments may be considered a possible P sink.

Approximately 17% of TP pools in both Koell and Gustayson consisted of Ca/Mg-bound

P. This finding was somewhat surprising given the low sediment pH values observed. Low pH

values tend to be associated with the presence of iron and aluminum whereas higher sediment pH

indicates the presence of calcium and magnesium. Limestone outcrops exist in the surrounding

slopes and may be draining to the valley. However, precipitation of P with Ca occurs when pH

exceeds 8. Measurements of water pH at midday when values are expected to be highest due to

photosynthetic activity revealed pH values rarely exceeding 7.2. Furthermore, hardness

calculations (data not shown) indicated a higher presence of magnesium than calcium. A more

likely explanation of this P pool is fertilization of fields using commercial calcium

superphosphates.

The levels of labile P, most often defined as KCl-Pi + NaOH-Pi (Sallade and Sims 1997a),

signify the stability of sediment P retention. P fraction analysis of the wastewater drainage ditch

sediments indicates that there is extremely high potential for release and biological uptake,










especially in Gustayson where labile P represented 74% of sediment TP. The potential for

mobilizing P in sediments was also high for Koell and San Martin (34 and 33%, respectively).

Management Implications

Anaerobic conditions found in wastewater drainage ditches limit the potential for stable

Fe-bound P. However, researchers have shown that complexation of ferric oxides with humic

substances may protect against solubilization of Fe3+ to Fe2+ under reducing conditions (Reddy et

al. 1999, Nguyen and Sukias 2002). Promoting the accumulation of organic matter in drainage

ditches via proper management of vegetation may serve to increase the potential for this

mechanism to occur. For low oxygen waters, such as those in the current study, P retention may

be enhanced by promoting aeration via structural modifications such as installing riffle zones.

Conclusions

This study demonstrated the dynamic nature of wastewater drainage ditches with respect to

sediment characteristics and associated phosphorus sorption capacity. High P loads to the ditch

systems regulate the equilibrium between sediment porewaters and aqueous SRP, not the benthic

sediments. Sediment retention of P was largely regulated by sorption with Fe/Al oxyhydroxides

which for Fe-bound P is highly unstable in these anaerobic systems. Nonetheless, sediments

undoubtedly play an important role in providing a temporary storage of P that may dampen the

timing and magnitude of P exports to the Chorobamba River. This buffering mechanism appears

to be greatest during the dry season when ditches tend to trap and accumulate finer sediments

and organic matter. The timing of this increased buffering capacity is likely to be important for

mitigating P loads to the Chorobamba River at a time when flowrates are lowest and the river is

most vulnerable to eutrophication.

PSI values determined using a one-hour equilibration time demonstrated the capacity of

the ditch sediments to rapidly remove high concentrations of aqueous P, an important









consideration for these flashy systems with limited residence times. The findings from this study

suggest that the potential exists for managing drainage ditches as wetlands for mitigation of P

loads to receiving bodies. Key to this approach is likely to be the proper management of

vegetation in order to promote sedimentation and organic matter accumulation. Modifying the

ditches to increase aeration may help to take advantage of the large potential for P sorption with

ferric oxides.


Figure 4-1. Looking upstream
July 2006


am sampling site 4 in San Martin in A) November 2005 and B)





























Figure 4-2. Looking downstream from sampling site 1 in Gustayson in A) November 2005 and
B) July 2006


Figure 4-3. Looking downstream from sampling site 1 in Koell in A) November 2005 and B)
July 2006































Figure 4-4. Sampling sites in Frey (only sampled in 2005). A) looking downstream from site 1
and B) looking upstream at site 2, the marsh site after being ditched by the city. Note
that despite ditching at site 2 the water remains stagnant. (Photos taken at the time of
sampling in November 2005).


10.00

y= 12.911x- 2.4939
8.00 -I R2 = 0.9963


3 6.00
0, ~Slope= K
(P buffering
a- 4.00 -y-intercept EPCo capacity)


2.00

a.0.00
0.)00 0.20 0.40 0.60 0.80 1.00
-2.00


-4.00
Initial SRP in solution (mg P/I)

Figure 4-5. Example of the approach used to determine the sediment EPCo value and K, a
measure of P buffering capacity, from P sorption isotherms.


















, 1. Ill


II 11


.ti I. I


, 11.l lI


.





Figure 4-6. Rainfall in Oxapampa from September 2005 September 2006. Sediment sampling
in 2005 took place during the rainy season while in 2006 samples were collected
towards the end of the dry season. Rainfall data was provided by the Andean Amazon
Research Station and ProPachitea.


IIIUIY II


.


I


2005 sampling days


2006 sampling days


40


20










Table 4-1. Site averages of particle size distributions and sediment pH for the wastewater ditches in October-November 2005 and
June-August 2006
San Martin Gustayson Koell Frey
2005 2006 2005 2006 2005 2006 2005 2006
% sand 94.4 50.2 73.7 (83.7)a 76.2 87.1 26.0 55.1 n.d.b
% silt 4.1 34.4 13.3 (12.4) 9.9 10.8 49.1 29.9 n.d.
% clay 1.4 15.4 13.0 (3.9) 13.9 2.1 24.9 15.1 n.d.
% OM 1.9 3.9 3.1 (3.0) 4.7 2.9 13.2 7.4 n.d.
pH 6.0 5.6 6.7 6.0 6.5 5.5 7.0 n.d.
"Values in parentheses are averages without Gustarson sampling site 2 which had extremely high clays (40%) that bias the overall average for the ditch.
bn.d.= no data. Frey was not sampled in 2006.

Table 4-2. Comparison of P sorption measures determined in each site in years 2005 and 2006
San Martin Gustayson Koell Freya
2005 2006 2005 2006 2005 2006 2005
PSI 22.9b 98.5 17.8 181.4 22.6 233 44.9
(L g- ) (21.7-24.2) (66.2-173) (10.2-25.4) (88.1-303) (10.2-29.7) (174-342) (32.5-62.7)


Smax
(mg P kgl)

EPCo
(mg P L1)


75.8
(73.0-78.6)

0.29
(0.28-0.29)


306.0
(260.0-417.6)

0.08
(0.05-0.14)


60.8
(38.0-92.6)

1.43
(0.51-2.31)


577.3
(358.9-811.5)

0.18
(0.06-0.28)


72.1
(38.5-92.6)

0.4
(0.23-0.55)


794
(613-1124)

0.09
(0.06-0.12)


84.8
(78.2-97.0)

0.5
(0.3-0.6)


K 2.4 7.4 3.7 8.2 3.9 15.8 3.10
(kg- ) (2.3-2.6) (5.5-8.9) (2.5-4.8) (4.6-13.6) (2.8-6.0) (10.0-21.7) (2.6-4.2)
PSI = phosphorus sorption index: Smax = P sorption maximum after 1 hour; EPCo = equilibrium P concentration: K = P buffer capacity
aFrey was sampled in 2005 only. bValues are averages. Numbers in parentheses are ranges of observed values.












400


SSan Martin a Gustayson a Koell


300



200



100



0


1 2 3 4
Sampling location (upstream to downstream) for each site

Examination of longitudinal variation in 2006 PSI values for each ditch from
upstream (1) to downstream (4) sampling locations. Values are means (n = 3) and
bars represent +1 SD.


Figure 4-7.


* San Martin -m Gustayson t-Koell


c ~,
\ \
--I--~


Sampling location ( upstream to downstream) for each site


Figure 4-8. Longitudinal trends of percent organic matter in 2006 for each ditch from upstream
("1") to downstream ("4") sampling locations.











OMV
25 y-o o2x2 +061x +2.7

20 -

ct15 I /X

S10 -1 d *


Clay
y = O 2X2 -02x +4 3
F9=0 60


m sand
silt
.clay
aOM



Sand
onsX2-0O5x +24 5





'100


Silt
y = O cuseX2- 0 18x + 5 4


m
x
x m
-- -r c
r ~ m


0 20 40 60


Percent content of sand, silt, clay or organic matter

Figure 4-9. Relationships between K (a measure of P buffer capacity determined from EPCo
experiments) and percentage of sand, silt, clay and organic matter for all sites.


45




30


SBiotic n Abiotic


Koell


San Martin


Gustayson


Frey


Site


Figure 4-10. Contributions of biotic and abiotic sorption for phosphorus sorption indices (PSI)
determined for sediments in 2005. The PSI determined from autoclaved sediments
represents abiotic sorption. The difference in PSI for intact and autoclaved sediments
represents biotic sorption. The sum of the bars for each site thus represents total PSI.
Values are means (n=3) and bars represent + 1 SD.












1: line


P sorption to


G P release frorn sediments
'3 1.2-


c F
c~oG G

OK


a. G
o~ S = San Martin
K/ G =Gustayson
0.4 GK=Kol
F =Frey

i~S S
Blue, underlined = 2005
Red = 2006

0.0
0.0 0.5 1.0 1.5 2.0 2.5
Sediment EPCo (mg L 1)

Figure 4-11. Comparison of ditch water SRP concentrations with sediment equilibrium
phosphorus concentration (EPCo) experiments. Values plotted above the 1:1 line
indicate that P is removed from the water column and sorbed to sediments. Values
below the 1:1 line indicate sites where P is likely desorbed from sediments and
released to the overlying water. Letters denote sites: S = San Martin, G = Gustayson,
K= Koell, F = Frey. Blue, underlined letters indicate EPCo analyses performed in
2005 and red letters represent analyses from 2006.










2.40


*P sources

y = 1.3722x + 0.332
/ R2 = 0.6322


1.80



1.20



0.60


P sinks
y = 0.4678x 0.0027
R2 = 0.7162






2.00


*~ *
*


0.00
0.00


0.50


1.00


1.50


Ditch water SRP (mg P L )

Figure 4-12. Sediment EPCo values as a function of water column SRP concentrations grouped
by whether sampling site acted as a P sink or source.

1000


100 y=28.473x-0 6395
R2 = 0.3864
s** *


0.00


0.50


1.00


1.50


2.00


2.50


EPCo (mg L 1)
Figure 4-13. Relationship between P sorption index and sediment EPCo indicates that highest
EPCo values (and thus greater potential to act as P sources) tend to be associated with
lowest P buffering capacity.











San Martin


a18.3%


S0.0%


03 0.1%


O 37.0% n 44.6%


Gustayson


17.0% .%H1.0%


a

O 4.4% ,


9 74.5%


Koell


S12.8%


S0.0%


a 31.5%


o 38.3%


Readily bioavailable (KCl-Pi)
a Fe/Al bound P (NaOH-Pi)
O Bound w/ humic substances (NaOH-Po)
o Ca/Mg bound P (HCl-Pi)
Recalcitrant


Figure 4-14.Percentages of sediment pools of sorbed phosphorus for each site. Values presented
are averages of extractions performed on duplicate sediment samples from each
sampling location for each site.











3.E +04
19227

2.E +04
O

S1.E+04
5265
2171

(n 0.E+00
San Martin Gustayson Koell

Site

Figure 4-15. Average total sediment P for San Martin, Gustayson and Koell calculated from the
sum of the P fractions for each ditch (bars represent 1 SD about the mean).



San Martin -5 Gustayson + Koell
25000


15 000 -


S10000


5000

1 2 3 4

Sampling location (upstream to downstream) for each site

Figure 4-16. Average sediment TP for each sampling location (1 = upstream, 4 = downstream) at
San Martin, Gustayson and Koell.











Table 4-3. Phosphorus sorption index (PSI) values for wastewater ditches calculated using three different sets of units to compare with
values reported for ditches, streams and wetlands
System type PSI PSI units Spiked P Equilibration Microbial Reference
X (log C)-' concentration time (hr) activity


(mg P 1~)


inhibited?
No


Headwater stream New Hampshire, USA
comparing silty and sandy sediments
4 streams, New York, USA

3 agriculturally influenced streams,
Oklahoma, USA
5th Order stream, Arkansas, USA

Wastewater drainage ditches, Peru 2005/2006


26 agricultural drainage ditches, New Zealand
17 agricultural drainage ditches, Delaware.
USA

2 riparian wetlands, North Carolina, USA

15 isolated herbaceous wetlands, Minnesota,
USA
10 streams and 9 wetlands, Florida, USA

Wastewater drainage ditches, Peru 2005/2006


10.3 and 2.1


1.89-12.32

~ 3.5-5.5** X = pLg g '
C = pLg 1-
~1.5-2**
17.8-44.9/
98.5-233.7


(Meyer 1979)


(Klotz 1985)
(Haggard et al. 1999)


(Haggard et al. 2001)
This study


931-2328


1000


(Nguyen and Sukias 2002)
(Sallade and Sims 1997a)


(Bruland and Richardson 2004)

(Bruland and Richardson 2006)


(Dunne et al. 2006)

This study


1664 and
858*
1012*


82.5 and
173.4
88-171/
319-1818


X = mg kg'
C = mg 1-1


8 early and 8 late successional mitigated 0.13 and 9.3 48 No (D'Angelo 2005)
wetlands, Kentucky, USA 0.59 (0.3mmol 1 )

6 palustrine wetlands, Virginia, USA ~40** mg (1009)' 130 24 No (Axt and Walbridge 1999)
pLmol 1'
Wastewater drainage ditches, Peru 2005/2006 2.7-3.7/ 50 1 No This study
12.5-40.8
Note: Values are reported as means or ranges of means if multiple groups were compared. If PSI for multiple depths were considered, only PSI values from the
upper-most sediments are reported. *PSI values converted from units originally reported as X = mg (100 g)-'.**Estimated visually from diagram.






















gure 4-17. Gustayson sediments collected on filters after the NaOH extraction step indicate the
presence of iron oxides.









CHAPTER 5
EVALUATION OF CHANNEL MODIFICATIONS TO EXISTING DITCHES FOR
IMPROVED WATER QUALITY AND AESTHETICS

Introduction

Conventional, centralized wastewater treatment systems are costly to construct and manage

and consequently are often not practical realities for many communities in developing countries

(CEPIS 2002). Decentralized, small-scale systems that utilize technologies appropriate to the

region's level of resource availability are necessary for providing better management of

wastewaters (Ho 2003). The municipality of Oxapampa has plans for the construction of a

stabilization pond (Municipalidad Provincial de Oxapampa 2003) but shortages of funds prevent

the realization of the proj ect. Several stabilization ponds have been constructed in other

municipalities in the region over the last several years but have notoriously failed due to under-

sizing of the systems for combined sewer and stormwater flows and by the lack of technical and

financial resources necessary for properly managing the large, centralized treatment operations

(Arteaga, personal communication).

Currently, Oxapampa is moving towards replacing wastewater ditches with underground

pipes in order to reduce human contact with wastewater and to improve the city's image to

promote greater tourism to the area. However, even replacement of ditches with pipes is cost-

prohibitive. The total cost (materials plus labor) of installing pipe is approximately $37.50 per

meter (Arteaga, personal communication). To replace the existing 18 km of ditches with pipes

would therefore cost $675,000. In contrast, the current cost to manage ditches is $375 per 4,000

meters or about $0.10 per meter (Arteaga, personal communication).The pipe networks function

as combined sewer systems but are undersized for high flow events, occasionally resulting in

ruptures. The pipes are also damaged by heavy trucks due to the shallow depths at which they are

installed due to high water tables (Arteaga, personal communication).









The first obj ective of this study was to implement simple channel modifications to three

existing wastewater ditches (Gustayson, Koell and Frey) to test whether the modifications served

to improve water quality. The second objective was to promote community participation in the

proj ect to evaluate whether modified ditches could function as a viable, decentralized water

treatment approach. Ditches are often located in the middle of roads in order to receive

wastewater discharged from homes on either side of the street. Due to their central position,

ditches are thus a unifying feature of each neighborhood. However, ditches are regarded with

much disdain as they are perceived to mar the visual and olfactory aesthetics of the community

and symbolize a lack of economic progress. Therefore, to promote community support, it was

also important to emphasize improvements to the aesthetics of the ditch systems.

Methods

Government and Community Involvement

The support and participation of local residents was deemed to be a critical component to

the success of the channel modifications. Actions were taken at various stages of the project to

engage and inform the public and local government directly about the goals and potential

benefits of ditch modifications (Table 5-1). The first stage consisted of activities to assess

whether support for modified ditches existed or could be fostered through dialogs with

stakeholders through presentations to government officials and local NGOs (Figure 5-1),

dissemination of interpretive folios describing the purpose of studying modified ditches (Figure

H-1), and meeting with residents to determine how to implement the modifications to best meet

their needs.

The second stage involved participation of residents living along each of the study ditches

(Gustayson, Koell and Frey) during implementation of modifications. Participants were paid the

standard local wage of $5 per day, were provided with boots, gloves and equipment such as









shovels, hoes and machetes (Figure 5-2). Signs were painted by a local artist asking citizens not

to dump trash into the ditches and announcing the existence of the community-based project for

modified ditches (Figure H-2). Rugged wooden planks or bundles of sticks had been used

previously as bridges to cross ditches when meeting with neighbors and had resulted in several

children falling into ditches and being injured. Therefore, new sturdier bridges were constructed

in each of the ditches as a safety measure, to increase neighborhood aesthetics and to foster a

greater sense of community among residents.

The final stage of activities to promote community participation of the modified ditches

consisted of capacity-building for participatory management of ditches. A non-technical guide

was created and disseminated to residents and government officials that explained the purpose of

the proj ect and provided instructions and images detailing how the modifications were

performed, costs of modifications and how to manage the modified channels. Residents assisted

in the construction of compost bins for accumulated sediments and plants from the ditches

(described below). Finally, residents were provided with tools to simplify ditch management

such as cleaning screens and collecting plants and trash (also described below).

Modification Designs

Criteria for modification approaches were low cost, use of only locally available materials

and plants, and an aesthetically pleasing design. All three study ditches were first widened to a

minimal width of 1.5 m to increase width-to-depth ratios. Channel banks were reinforced with

large cobbles and boulders to prevent undercutting and bank slumping. A sedimentation basin

was dug at the most upstream location of each study ditch. The three neighborhood groups were

opposed to the use of common ditch species such as papyrus and ginger lily which were

perceived as noxious weeds. Therefore, locally-available water tolerant ornamental plants such

as canna lily (Canna flacida), calla lily (Zantedeschia aethiopica) and yellow iris (Iris










pseudacorus) were used predominantly in all three study ditches. The fast-growing rhizomes of

the canna, calla and iris plants were also expected to stabilize banks. Canna and calla lilies were

transplanted from other wastewater ditches (Figure 5-3) while irises could only be found at the

market as seeds.

The different flowrates of the study ditches prompted the use of two separate designs for

channel modifications: an open-water flow design for the high water flows characteristic of the

Gustayson and Koell ditches and an alternating subsurface flow/open-water design for the low

flow ditch of Frey.

Open-water flow design

The primary obj ectives of the open-water flow design were to extend channel residence

time and increase dissolved oxygen levels. The open-water flow design consisted of four main

design elements: a sedimentation basin, riffles and cascades, gravel bars for creating channel

sinuosity and the introduction of ornamental plants (Table 5-2).

The sedimentation basins were oval-shaped (approximately 2 m in length, 1.5 m in width

and 0.8 m deep). The width of the sedimentation basin in all sites was limited to avoid blockage

of the road to traffic on either side of the ditch and basin depth was limited by a thick clay layer.

The overall volume of the basin had to be limited to avoid water from overcoming the

impoundment structure which consisted of wood boards reinforced with stakes, cobbles and clay.

Limited resources prevented the basin from being constructed with a more robust structure made

of reinforced concrete.

Water exited the basin through a 1 cm gap between two horizontal boards and made

contact with rocks in the channel below to induce water oxygenation (Figures 5-4 and 5-5). To

avoid clogging of the exit and to improve sediment removal in the basin that would otherwise be

limited by low residence times, the basin exit was fitted with two sizes of screens: a coarse (0.5










cm) mesh followed by a fine (0.5 mm) mesh. The basins were planted with water hyacinths

(Eichornia cra~ssipes) that were transplanted from the marsh at the Frey site.

The second design element used in the open-water flow designs in Gustayson and Koell

were riffles and cascades to increase aeration of the water and to promote greater transient

storage through creation of recirculation zones and pools upstream of the riffles (Figure 5-6).

Lateral gravel bars were created to increase channel sinuosity and complexity (Figure 5-7).

The bars were intended to allow water to pass through the gravel interstices, thus promoting

physical and biochemical filtration and denitrification. The gravel bars and channel banks were

planted primarily with canna lilies, calla lilies, and irises among other attractive, water-tolerant

plants common to the area. The plants were grown hydroponically in the gravel bars to increase

available reactive surface areas of subsurface flowpaths. Along banks, small holes were dug

close to the water surface to allow plant roots to make direct contact with ditch water.

Subsurface flow design

The second modification approach, the alternating subsurface flow/open water design, was

applied to the ditch in Frey. The primary design elements used in the approach included a

sedimentation basin, planted gravel beds for subsurface flowpaths, open water sections planted

with water hyacinth (Table 5-3).

The sedimentation basin in Frey followed the same design approach as that described

above for Koell and Gustayson except that it was smaller (1.5 m length, 1 m width, and 0.6 m

depth). After passing through the screens (same mesh sizes as described above) and exiting the

sedimentation basin, water entered the first subsurface flow section. In all, there were three

subsurface flow sections and two open water sections (Figure 5-8). The lengths of the subsurface

flow and open water sections were dictated by the pre-existing locations of pipes that discharge

wastes from homes on either side of the ditch. Open water zones existed where wastes enter and









subsurface flow reaches were located in between waste discharge points (Figure 5-9). The slow

water velocities in the deep (0.5 m), open water areas trap the solids brought in by pipes. Before

entering subsurface flow sections, water from the open water flow zones first pass through a fine

mesh (0.5 mm) screen to prevent clogging of gravel interstices.

The sedimentation basin of Frey allowed flows to be regulated to the system. Once the

water levels in the basin and the inflow pipe (diameter = 25 cm) carrying wastewater from the

intersecting street to Frey were equal, the pipe would not permit more flow to enter the basin and

the remaining flow from the intersecting street continued downstream to enter subsequent

ditches. When adding the cobble and gravel layers, flow to Frey was maximized to ensure

enough layers were in place to prevent overflow during high flow events.

Subsurface flow sections were also designed to be extremely porous to extend the lifetime

of the system before clogging. Cobbles (10-20 cm) were added as the bottom layer followed by

progressively smaller gravel at the uppermost layers to prevent bed surface sediments and plant

litter from entering and obstructing pore spaces. The gravel sections were planted hydroponically

(Figure 5-10) allowing ample space for oxygen exchange from the atmosphere (Wallace et al.

2001).

Monitoring and Statistical Analyses

Tracer experiments were performed one to two weeks after completion of the channel

modifications following the field and analytical methods described in Chapter 3. Breakthrough

curves of the conservative solute were determined at two sampling locations: after the

sedimentation basin and at the most downstream location. Median transport times (Tmed) WeTO

determined from the tracer breakthrough curves and compared with previous values (before

modifications) to evaluate the effect of the modifications on channel residence times.









Grab samples of water were collected every 1 to 2 weeks for a 6 month period (September

2006 through February 2007) in 3 to 4 locations along each ditch and analyzed for fecal coliform

bacteria, phosphate, ammonium, nitrate and total suspended sediments following field and

laboratory procedures described in Chapter 3.

Comparisons between water quality parameters before and after channel modifications

were made using t-tests assuming equal variances after the data had been In-transformed to meet

normality requirements. Significant values were reported for a water quality parameter if t-tests

indicated that influent concentrations were not statistically different before and after channel

modifications but effluent concentrations after modifications were statistically less than those

observed prior to modifications.

Results

Effects of Modifications on Residence Times and Water Quality

Heavy rain events associated with El Nifio Southern Oscillation activity readily

undermined the integrity of the sedimentation basin in Gustayson and therefore only results

related to the other two sites Koell and Frey are presented here.

Tracer experiments before and after channel modifications revealed that the modifications

in Koell did not substantially lengthen the residence time or the shape of the breakthrough curve

(Figure 5-1 1, Table 5-4). Determination of Tmed at the exit of the sedimentation basin revealed

that the basin had very little effect in extending the overall channel transport times as the median

basin residence time was less than 2 minutes. In contrast, residence times in Frey increased from

approximately 6 minutes to 46 minutes after channel modifications, (Figure 5-12, Table 5-4).

However, median solute residence time in the sedimentation basin accounted for only 4.2

minutes.









The open water flow design used to modify the channel of Koell did not result in

significant improvements to water quality with respect to any of the water quality parameters

except for dissolved oxygen (Figures 5-13 to 5-18). The spread ofE. coli distributions was

reduced compared to those before channel modifications but median values were not statistically

different (Figure 5-13). Influent TSS values were slightly higher during the sampling period

following channel modifications (Figure 5-14). This finding was expected due to the prevalence

of heavy rain events carrying high sediment loads. Median TSS values did not differ between the

three sampling locations indicating that net sedimentation did not tend to occur during transport.

The creation of riffles and cascades were effective at increasing DO levels (p < 0.05)

between the first and second sampling sites (Figure 5-15) where channel slope was greatest. DO

levels tended to decrease after the second sampling site due to the lower channel gradient that

substantially reduced the aeration efficiencies of the riffles. However, despite increased DO

levels, nitrification was not promoted to a degree that resulted in reduced NH4-N concentrations

(Figure 5-16). However, nitrification was indeed occurring as evidenced from relatively high

NO3-N values compared to those observed prior to channel modification (Figure 5-17).

Moreover, denitrification was also likely to have occurred as high NO3-N concentrations

decreased between sampling locations. However, coupled nitrification-denitrification reactions

were unsuccessful at overall nitrogen retention.

Concentrations of SRP did not change significantly during transport either before or after

channel modifications (Figure 5-18). However, this should not necessarily be interpreted as the

lack of retention because additional phosphorus inputs from wastewater discharges occur

between locations 1 and 2, indicating that net retention must be occurring if concentrations do

not increase downstream. Influent concentrations of NH4-N and SRP tended to lower than those









before channel modifications and are likely to be the result of dilution from rain events. Lower

water column SRP concentrations during periods of dilution may be causing the sediments to act

as P sources (see Chapter 4 discussion of EPCo).

Comparisons of water quality parameters before and after channel modifications in Frey

indicated that the modifications were effective at removal of E. coli (t = 2. 1209, df = 14, p-value

< 0.05) and TSS (t = -4.3447, df = 22, p-value < 0.01) but not for reductions of nutrient

concentrations (Figures 5-19 through 5-24). Due to wastewater inputs along the length of Frey

(see Figure 5-8), prior to modifications, concentrations ofE coli and TSS tended to increase

with downstream distance. However, after modifications E. coli and TSS levels decreased

downstream despite additional wastewater inputs to the system (Figures 5-19 and 5-20).

Channel modifications were not effective for promoting oxygenation as evidenced by

decreased DO concentrations at the downstream sampling location (Figure 5-21). However,

concentrations of NO3-N were often higher at the downstream sampling location (Figure 5-22)

indicating that nitrification was nevertheless occurring in the system, consuming the dissolved

oxygen in the process. However, nitrification was not an effective retention process as NH4-N

concentrations tended to increase downstream (Figure 5-23). Concentrations of SRP were

effectively unchanged between the two sampling points (Figure 5-24) indicating that some P

retention occurred since concentrations would have otherwise increased from additional

wastewater mnputs.

Management Requirements

Management requirements for channel modifications included periodic removal of

sediments collected in sedimentation basins and, in the case of Frey, from the open water

sections planted with hyacinth. To facilitate sediment collection, the culvert that delivers water to

the basin from the intersecting street was blocked such that water by-passed the basin and









continued to the next intersecting ditch downstream. While the basin was being drained,

approximately 90% of the water hyacinth plants were collected and wheel-barrowed to the

compost bins. The remaining plants were returned to the basin after sediment removal.

Sediments were extracted with shovels and buckets, placed on a large tarp and mixed with

sawdust and harvested water hyacinth plants and added to the compost bins.

Sediment removal in the Koell basin was necessary three times over a four month period.

The first two cleaning were necessary after the city removed vegetation from upstream ditches

delivering high loads to Koell, one the most downstream ditches before discharge to the river.

The third cleaning was necessary after two months of heavy rainfall. The sedimentation basin

and open water sections in Frey were cleaned once in four months. Two to three cleaning per

year was the anticipated frequency of sediment removal and was considered acceptable by the

participating residents.

More frequent management requirements included monthly removal of water hyacinth and

weekly cleaning of the screens and removal of trash. A simple tool was provided to a

representative participant of each modified ditch to facilitate plant/trash removal and screen

cleaning. The tool consisted of a net on one end and a coarse brush on the other (Figure 5-25).

The brush effectively removed algae and trapped solids to prevent clogging of the screens and

exit structure. The long handle of the tool prevented the user from needing to enter the ditch.

Government and Community Response

Support by the community and government officials was high for modified ditches due to

increased visual aesthetics (Figures 5-26 and 5-27), reduced odors, perceived improvement to the

environment and low implementation costs (Table 5-5). Costs of the channel modifications

varied between $2.5 and $5.0 per meter, which were considerably lower than the cost of

installing pipe.









Six residents regularly assisted with modifications in Gustayson, 12 in Koell and 10 in

Frey. Other residents who were not directly involved in channel modifications demonstrated

support by providing drinks and snacks. Community participation provided an evident sense of

ownership as residents agreed amongst themselves that each was responsible for caring for the

section of ditch in front of their home. Interviews with residents suggested a sense of

empowerment was felt among those who participated in the proj ect because they were able to

improve the aesthetics of their neighborhood. For example, increased oxygenation in Koell

vastly improved odors emanating from the ditch. Residents commented that prior to the

modifications, the smells were so unpleasant that they were unable to sit outside in the evenings.

However, after channel modifications, several families had set up benches outside their homes

and said they could no longer smell odors. The residents of Frey and Koell hosted a celebration

to inaugurate completion of the modified ditches (Figure 5-28).

The modified ditches received much attention from the community in general. The

modified ditches were visited regularly by residents from other neighborhoods and government

officials from other cities. High school students were asked by teachers to write a report

discussing the water quality functions of the different design elements of the ditches. The

modified ditches were on the news several times and on one occasion resulted in a live television

interview. Another unexpected response from the community was that other neighborhoods

began attempting to modify their ditches (Figure 5-29) thus prompting the creation of the non-

technical guide describing methods used to modify ditches.

The municipality of Oxapampa supported the proj ect by donating rocks and equipment and

by agreeing to provide trash pick-up services to residents living at the end of Koell to prevent









future dumping of refuse into the ditch. A local NGO, Instituto de Bien Comun hopes to expand

on the idea to modify ditches in other communities.

Discussion

The modified ditches received very positive responses from the public as they were

perceived as a potential solution towards improving local wastewater management. However,

improved aesthetics resulting from modifications may foster a false sense of security that water

quality was concomitantly improved. This was certainly true for the modified ditch in Koell.

Comparisons of water quality before and after channel modifications in Koell indicated that the

open water flow approach did little to improve overall water quality. Nevertheless the approach

was very successful in providing improved aesthetics to the neighborhood. Increased dissolved

oxygen concentrations significantly reduced odors, fulfiling an important criterion for residents.

Visually, the system was very much improved as indicated by it being referred to as a garden

rather than as a ditch by passers-by.

The lack of treatment occurring in the system in Koell is the result of low retention times

that were exacerbated over the monitoring period due to the high frequency of large storm

events. Although the sedimentation basin was undersized and thus not capable of increasing

overall residence times, it was nonetheless effective at trapping sediments due to the use of the

screens. The small volume of the basin and high solids removal meant that the basin required

frequent cleaning. The basin filled with sediments on three occasions within four months. The

high management demand of the system eventually discouraged participants. The same basin

design in Gustayson filled with sediments during a large storm event and the following day, a

second storm event washed out the entire impoundment structure. For the basins to function

effectively without requiring excessive management and to be able to withstand high loads,

larger more robust structures are required.









Retention times in Koell were likely to be relatively low due to the lack of vegetation in

the ditch channel. The ornamental plants chosen in this study colonize only channel banks or are

grown hydroponically on gravel bars and thus are of limited importance for reducing water

velocities. However, the gravel bars and riffles appear to have a positive impact on increasing

solute residence times. Median transport times in Koell increased by approximatelyl2% after

channel modifications despite flows that were 44% higher. These results point to the possibility

of improved retention in Koell during the dry season. Future modifications may also want to

consider promoting submerged macrophytes for improved attenuation of velocity and direct

nutrient uptake from the water column (Mars et al. 1999, Tanaka et al. 2006)

In contrast to the open water flow design, the alternating subsurface flow/open water

design in Frey was effective at reducing concentrations of E coli and solids. This result was

expected as the design was modeled as a quasi-subsurface flow wetland which are commonly

recognized for their water treatment capabilities (Tanner et al. 1995, Billore et al. 1999).

However, it was not surprising that nutrient removal was low in the system due to limited

residence times compared to typical subsurface flow wetlands which are normally on the order of

weeks rather than minutes (e.g. Akratos and Tsihrintzis 2007).

Subsurface flow (SSF) wetland systems are commonly planted with cattails (Typha) and

reeds (Phragmites). However, more studies are reporting that ornamental, water-tolerant plants

provide comparable treatment performance. Irises have been shown to provide similar biomass

and nitrogen storage characteristics as Typha and Phragmites in constructed wetlands (Tuncsiper

et al. 2006, Zaimoglu 2006). Similar findings were reported by Ayaz and Akca (2001) who

found that Iris was more effective than Typha and Phragmites for COD, TN and TP removal and

that Canzna was highly effective at ammonium removal. DeBusk et al. (1995) reported that









Canzna demonstrated the highest foliar phosphate uptake of 10 emergent macrophytes. Belmont

and Metcalfe (2003) found that calla lilies significantly increased ammonium retention in SSF

wetlands. These studies suggest that nutrient retention in the modified ditches was not limited by

the choice of ornamental plants but instead by high loading rates coupled with low residence

times. Phosphorus removal may have been improved with the use of other filter materials (Adam

et al. 2007).

The lack of oxygen was another important factor limiting efficient retention in the

subsurface flow system in Frey. DO concentrations below 1.5 mg L-1 have been shown to limit

nitrifieation (Wolverton 1987). Values in Frey were consistently below 1.0 mg L^1. These values

suggest that oxygen translocation by plant roots was not significant enough to support nitrifying

bacteria to promote effective nitrifieation-denitrifieation reactions (Brix 1997). Oxygenated

rhizospheres have also been shown to be important sites for reducing fecal coliform and BOD

concentrations in SSF (Karathanasis et al. 2003). The treatment performance of the subsurface

flow system would be amplified if better oxygenation could be achieved (Ouellet-Plamondon et

al. 2006). However, it is expected that overall retention will improve in the subsurface flow

system as bionilms and other reactive surface areas continue to increase with time (Bigambo and

Mayo 2005, Soto et al. 2007).

The alternating subsurface flow/open water flow design proved to be successful as a

simple, low cost approach to provide primary treatment and perhaps irrigation-quality water. The

fecal coliform standard for water reuse for irrigation is 1,000 CFU mL-1 (WHO 1989). The

design did not demonstrate consistent effluent E. coli concentrations below this value but one

must take into account that reductions approaching this value occurred despite additional fecal

inputs along the length of the ditch. The reuse of treated ditch water for irrigation purposes and









for aquaculture (currently in practice in Oxapampa for raising trout) would be advantageous for

capitalizing on the bioavailable nutrients, particularly given the current dependence upon

chemical fertilizers in the region. Moreover, the subsurface flow design could be coupled with

the existing aquaculture facility to take advantage of the artificial aeration mechanisms already in

place.

Another distinct advantage of the subsurface flow design over the open water design is the

minimization of human contact with wastewater. In addition, maintaining water in the subsurface

prevents health and nuisance issues related to mosquito breeding. The maj or disadvantages of the

approach are that it is limited to ditches with low flowrates and requires large quantities of gravel

that is extracted from the river. Gravel mining is an unsustainable practice that undermines that

the self-purifying capacity of the river system by reducing the extent of beneficial hyporheic

zones (Hancock 2002).

Widespread support for testing the effects of modified ditches for improved water quality

and community aesthetics revealed an openness and readiness by the government and public to

explore alternative water treatment approaches. However, without long-term commitment from

the local government to support ditch management efforts, the modified ditches are likely to fail

as participant support wanes with time.

The approaches implemented and tested in this study are not likely to solve the

problematic wastewater management situation that currently exists in Oxapampa. However,

modified ditches nevertheless point the way for exploring smaller-scale, decentralized systems

such as neighborhood-scale treatment wetlands (Griffin and Pamplin 1998, Greenway and

Woolley 1999).









Conclusions

The experimental ditch modifications resulted in minimal improvements to concentrations

of nutrients; however fecal coliform bacteria and total suspended solids showed significant

decreases in concentrations during the first two months of sampling. Beginning in the third

month of monitoring, heavy rain events associated with ENSO activity resulted in pulses of

extremely elevated discharges that undermined the treatment function of the modified systems.

Preliminary findings suggest that the modified ditches in this study have minimal impact

on reducing downstream loads of nutrients. Nonetheless, they show potential for providing

primary treatment to otherwise raw sewage. The experimental approach used in this study was

successful at involving the community and provided residents with a viable alternative for

improving their environment, from both a public-health and an aesthetic perspective. Effective

long-term management will likely require commitment from the local government for assistance

in ditch management, particularly with respect to periodic dredging of sediments. Future

modifications should utilize designs better equipped for heavy periods of rain and sediment

loads.










Table 5-1. Activities performed to promote local support and participation in channel


modifications
Stage Activity
1. Support assessment Presented research to city
representatives and local NGOs
(3 occasions)


Purpose
Determine whether agencies
supported project and to receive
feedback

Inform community members
about modified ditches as an
alternative to pipes

Engage residents in decision-
making process; determine
whether residents were willing
to participate in proj ect
implementation and
management stages; address
specific needs and concerns

Foster sense of ownership of
proj ect

Inform interested parties about
the proj ect; discourage littering
in the ditches

Improve safety of ditch
crossings; increase aesthetics
and sense of community

Provide government and
residents with documentation of
procedures used

Provide residents with a
valuable end-product

Facilitate management of ditches


Spoke at a neighborhood meeting
and presented interpretive folios


Met with residents in each of the 3
neighborhoods


2. Implementation


Hired residents to assist in channel
modifications

Erected signs



Constructed bridges across modified
ditches


Created a non-technical guide
detailing how to modify and manage
ditches

Constructed a compost bin for ditch
plants and sediments

Provided residents with tools for
ditch management


3. Participatory
management


























Figure 5-1. Neighborhood meetings were attended to assess whether the community supported
implementation of modified ditches










































































gure 3-L. Kesiaents parucipalea In all aspects or me cnannel moanilcanlon acuvines









k' '' ,-'
ri. i


Figure 5-3. Ornamental plants such calla lily were transplanted from other sewage ditches to the
modified ditches

Table 5-2. Primary design elements used in open-water flow design and their intended functions
for improving water quality and aesthetics
Design elements Intended function
Sedimentation basin Trap solids, grease and trash


Sections of riffles and cascades s


Aerate water, increase bio-reactive surface
areas

Extend travel times, promote water
filtration, denitrification

Reduce bank and channel erosion, provide
bio-reactive surface areas, assimilate
nutrients, improve aesthetics


Lateral gravel bars


Vegetation












Basin outfall and oxygenation of water


Water hvacinth


Water oxygenation upon exit from the basin


Figure 5-4. Schematic of the sedimentation basin design used in channel modifications A) top
view and B) side view


Figure 5-5. A) Front view and B) side views of the sedimentation basin in Koell

























Figure 5-6. Sections of cascades and riffles created to improve oxygenation and increase pockets
of transient storage in A) Koell and B) Gustayson


Figure 5-7. Creation of gravel bars to increase channel sinuosity and filtration in Koell










Table 5-3. Primary design elements used in the alternating subsurface flow/open water design
and their intended functions for improving water quality and aesthetics


Design element
Sedimentation basin

Gravel beds for subsurface flow


Intended function
Trap solids, grease and trash


Increase residence times and reactive surface
areas for physical and biochemical filtration


Sedimentation of solids


Open water zones


Vegetation


Increase reactive surface areas, assimilate
nutrients, reduce bank erosion, improve
aesthetics


A portion of flow from an intersecting street enters Frey and the
remaining flow continues to the following block.








1< 83.2 m \
Note: Not to scale
Figure 5-8. Schematic of the alternating subsurface flow/open-water design implemented in Frey


Figure 5-9. Views of the alternating subsurface flow/open water design used in Frey A) water
exits the sedimentation basin and flows into the first subsurface flow section and B)
water exits the first subsurface flow section and enters the first open water section

























Water








Figure 5-10. Cross-sectional view of the subsurface flow design


SBefore modifications After modifications




0.8-.
**
0 6 "



t;S0.2 + f



0.0 0.5 1.0 1.5 2.0

Time since injection (hours)


Figure 5-11. Comparison of breakthrough curves before and after channel modifications in Koell





















































Figure 5-13. Comparison of E coli concentrations at upstream, middle and downstream
sampling locations in Koell before and after ditch modifications (dark circles
represent median values, open circles are outliers, boxes delineate inter-quartile range
(IQR), and whiskers are 1.5* IQR)


r"


rF*

r
*



jl


,,

~d c3 L~


* Before modifications After modifications


Q
Qo
L
~CCI


LL


0.25







i


Time since injection (hours)


Figure 5-12. Comparison of breakthrough curves before and after channel modifications in Frey


Table 5-4. Comparison of median travel times in Koell and Frey before and after channel
modifications


Koell
Flowrate (L sl)
5.0
7.2


Frey
Flowrate (L s')
2.8
2.4


Tmed (min)
24.0
27.0


Tmed (min)
6.8
45.6


Before
After


Koell before modifications


Koell after modifications


25000
20000
15000
10000
5000


Upstream Middle Downstream


Upstream Middle Downstream





Koell before modifications


Koell after modifications


200

150

(1 100
1-
50

0


Upstream Middle Downstream


Upstream Middle Downstream


Figure 5-14. Comparison of TSS concentrations at upstream, middle and downstream sampling
locations in Koell before and after ditch modifications (dark circles represent median
values, open circles are outliers, boxes delineate inter-quartile range (IQR), and
whiskers are 1.5* IQR)


Koell before modifications


Koell after modifications


Uostream Middle Downstream


Upstream Middle Downstream


Figure 5-15. Comparison of dissolved oxygen (DO) concentrations at upstream, middle and
downstream sampling locations in Koell before and after ditch modifications (dark
circles represent median values, open circles are outliers, boxes delineate inter-
quartile range (IQR), and whiskers are 1.5* IQR)





Koell before modifications














Upstream Middle Downstream


Koell after modifications














Uostream Middle Downstream


Figure 5-16. Comparison of NH4-N concentrations at upstream, middle and downstream
sampling locations in Koell before and after ditch modifications (dark circles
represent median values, open circles are outliers, boxes delineate inter-quartile range
(IQR), and whiskers are 1.5* IQR)


Koell before modifications


Upstream Middle Downstream


Upstream Middle Downstream


Figure 5-17. Comparison of NO3-N concentrations at upstream, middle and downstream
sampling locations in Koell before and after ditch modifications (dark circles
represent median values, open circles are outliers, boxes delineate inter-quartile range
(IQR), and whiskers are 1.5* IQR)


Koell after modifications











Koell before modifications


Koell after modifications


1.0

0.8

0.6




0.2

0.0


Upstream Middle Downstream


Upstream Middle Downstream


Figure 5-18. Comparison of SRP concentrations at upstream, middle and downstream sampling
locations in Koell before and after ditch modifications (dark circles represent median
values, open circles are outliers, boxes delineate inter-quartile range (IQR), and
whiskers are 1.5* IQR)


25000

20000

15000

10000

5000


Upstream Downstream


Upstream Downstream


Figure 5-19. Comparison of E coli concentrations at upstream, middle and downstream
sampling locations in Frey before and after ditch modifications (dark circles represent
median values, open circles are outliers, boxes delineate inter-quartile range (IQR),
and whiskers are 1.5* IQR)


Frey before modifications


Frey after modifications














,,
n


C ~


Frey before modifications


300

200

100


Upstream Downstream


Upstream Downstream


Figure 5-20. Comparison of TSS concentrations at upstream, middle and downstream sampling
locations in Frey before and after ditch modifications (dark circles represent median
values, open circles are outliers, boxes delineate inter-quartile range (IQR), and
whiskers are 1.5* IQR)


Frey before modifications


Frey after modifications


Upstream Downstream


Upstream Downstream


Figure 5-21. Comparison of dissolved oxygen (DO) concentrations at upstream, middle and
downstream sampling locations in Frey before and after ditch modifications (dark
circles represent median values, open circles are outliers, boxes delineate inter-
quartile range (IQR), and whiskers are 1.5* IQR)


Frey after modifications











Frey after modifications


I 1


Frey before modifications


0.5

0.4

0.3

0.2

0.1

0.0


Upstream Downstream


Upstream Downstream


Figure 5-22. Comparison of NO3-N concentrations at upstream, middle and downstream
sampling locations in Frey before and after ditch modifications (dark circles represent
median values, open circles are outliers, boxes delineate inter-quartile range (IQR),
and whiskers are 1.5* IQR)


Frey after modifications


Frey before modifications


Upstream Downstream


Upstream Downstream


Figure 5-23. Comparison of NH4-N concentrations at upstream, middle and downstream
sampling locations in Frey before and after ditch modifications (dark circles represent
median values, open circles are outliers, boxes delineate inter-quartile range (IQR),
and whiskers are 1.5* IQR)











Frey after modifications


Frey before modifications


2.5

2.0






0.5

0.0


Upstream Downstream


Upstream Downstream


Figure 5-24. Comparison of SRP concentrations at upstream, middle and downstream sampling
locations in Frey before and after ditch modifications (dark circles represent median
values, open circles are outliers, boxes delineate inter-quartile range (IQR), and
whiskers are 1.5* IQR)


Figure 5-25. Simple tool consisting of a combined net and brush was provided to residents to
facilitate ditch maintenance
















































Table 5-5. Approximate costs of ditch modifications for each site
Ditch Modification length Modification cost ($
(m) m )>*
Gustayson 135 2.5
Koell 195 5.0
Frey 85 3.2
Cost to install underground pipe ($ m l) 37.5
* Includes labor (typical local wage is $5 day '), construction materials (lumber, screens and nails), plants and
collection and delivery of cobbles and gravel


Figure 5-26. Images of the open water flow design in Koell ditch four months after modifications


gure 5-27. Images of the alternating subsurface
after modifications


aw/open water design in Frey four months










































Figure 5-29. Other residents began attempts at modifying their ditches


Figure 5-28. Residents from Frey held a celebration to inaugurate the ditch and new bridge









CHAPTER 6
SYNTHESIS AND FUTURE WORK

This study provided a unique perspective of drainage ditches in the landscape by

examining the treatment potential of ditches receiving domestic wastewaters in a rural Peruvian

community. Drainage ditches are prevalent landscape features throughout the world yet their role

for influencing contaminant fate and transport between terrestrial and aquatic ecosystems has

been largely ignored. Likewise, the use of vegetated ditches for wastewater disposal is poorly

documented despite their widespread use in developing countries such as Peru.

The ability of ditches to mitigate contaminant exports was shown to be predicated on the

interconnected relationships between contaminant loads, flow and channel characteristics, plant

communities and sediment properties. Plants were shown to play a crucial role in affecting

transport characteristics by creating pockets of slower-moving water that increased residence

times and allowed sedimentation and organic matter deposition to occur. These vegetation-

induced transient storage zones were likely to have been the sites where the reductions of

sediments and pathogens occurred during transport. The depositional environments created by

vegetation during the dry season in Oxapampa resulted in increased clay and organic matter

content of sediments. In turn, fine particles and organic-rich sediments were shown to be related

to much higher phosphorus retention capacity compared to the coarser benthic sediments

observed in the rainy season after plant harvesting. Given the high sorption affinity of many

contaminants to organic matter and Eine particles, depositional conditions promoted by

vegetation in ditches were also likely to retain other potential contaminants such as pathogens,

pesticides, and metals. Ditch macrophyte biomass functioned as a large storage of nitrogen and

phosphorus yet uptake processes by plants were not estimated to be a significant net retention

mechani sm.










High loads of nitrogen and phosphorus coupled with low channel residence times

overwhelmed the intrinsic processing capacity of the wastewater ditches. While evidence of

contaminant retention in ditches was observed through comparisons of water quality with pipes

and by comparisons between upstream and downstream ditch sampling sites, the ditches were

nevertheless ineffective at reducing nitrogen and phosphorus concentrations to near-acceptable

discharge levels.

The treatment performance of wastewater ditches both before and after channel

modifications was compromised by the lack of oxygen. Large quantities of retained phosphorus

due to sorption with iron oxyhydroxides were readily released to overlying water under

anaerobic conditions. Low dissolved oxygen levels prevented net nitrogen removal by limiting

nitrifieation of ammonium. Channel modifications in Koell served to increase oxygen levels to

concentrations as high as 5 mg L1 on several occasions; however, ammonium loads were too

great to result in overall nitrogen retention through coupled nitriaication-denitrifieation reactions.

These Eindings suggest that ditches, especially those receiving high nutrient loads, cannot

be relied upon as the sole treatment mechanism but instead should be used in conjunction with

other management practices that serve to attenuate contaminant pulses to ditches. For wastewater

drainage ditches these practices could include simple household septic tanks or large

neighborhood sedimentation basins for solids removal prior to discharge to the channels.

The creation of subsurface flowpaths is particularly advantageous for reducing human

contact, thus addressing a maj or health issue related to the use of ditches for waste disposal.

Moreover, the alternating subsurface flow/open water design was shown be effective at sediment

and E. coli removal and thus shows promise for treating wastewaters to meet irrigation-standards

for water reuse. The goal of ditch modifications and "wastewater" treatment systems in general









should be the removal of sediments, oxygen demand and pathogens to levels that are acceptable

for water reuse in order to recycle and utilize the valuable nutrients while preventing detrimental

effects to aquatic ecosystems caused by eutrophication.

Many questions remain regarding the use of ditches as viable treatment systems. Questions

include how ditches should be managed to optimize retention while avoiding risks of floods.

Another practical research question is what length of ditch, given different plant communities

and biomass conditions and under various hydrologic and loading conditions, is sufficient to

achieve a desired treatment response? Future ditch studies should continue to focus on how

ditches can be modified to amplify their treatment capabilities. Retrofitting existing ditches with

designs aimed at treatment, rather than solely water transport, could result in significant

improvements to water quality in regions throughout the world.










APPENDIX A
ENERGY CIRCUIT LANGUAGE

Syrmbo Name Description
Defines the sysltem beiug diagrammed. Lines that cross
the sysfte~m boi~undr- todicate i~nflos and outflows of
Boundary
O System he sy~stem.

I Eerg Cicui Aparhyal ia hich has a flow proportional to the quanti~ty
Eoergy ~min the s~torae or source -uprstreamt~.


Source Outside source of energy deliverin orce~rs according in
a program controlled from outside; a fcing limet nlon

0~~ Flow~ Limited Outside source of ener:y sith a Floui tha is externally

com3t Tgk A npartm~nt o~f enemy~ storage withilrn the syst~em
Storage Tank tom a qluantgity as thre baalance ofmflowrs and
ou~tilows.
The senso~r (tiny square box on storage) iuggests th
Senorstorage tank controls some other olow bu~t do~es not
supply the main energy for it.

P rodce Unit that collects and transforms low-qualty~ energy;
unlder Control interactions clhigh-quaity flowrs.


Cons~umer Unit that ~transfoun1S energy Iquality, stores it, and feeds
it back autocir~)iaralyical ro imp~roe inflowu.

Dispersion ofpotential. ener~gy4 into hear rhar
Het in accomanpi~e~s all real transforanation presses and
storage; loss of potential energy from further use ~byg the
systemn.

Figure A-1. Description of the symbols used in energy circuit diagrams (from Odum 1994)










































0 12 34 5 6
(m)


Section 3 (Distance = 62.3m)


APPENDIX B
DITCH CHANNEL CROSS-SECTIONAL AND LONGITUDINAL PROFILES


Section 1 (Distance =Om)


0 1 2 3 4 5 6 7
(m)


Section 2 (Distance = 33m)


10


(m)


Section 4 (Distance = 151.1m)

11
10





01234567
(m)


Figure B-1. San Martin cross-sectional profiles for sections 1 through 4













Section 5 (Distance = 200.1m)


12




E 0 1 2 3 4 5 6 7



(m)


Seccion 6 (Distance = 255.3m)


234

(m)


Section 7 (Dis tance = 315.8m )


12





9

8


-


0 12 34 56 7

(m)


Section 8 (Distance = 355.6m)


12







9.5


0 12 34 56 7

(m)


Figure B-2. San Martin cross-sectional profiles for sections 5 through 8












Longitudinal profile

12
10


E 4


0 50 100 150 200
meters


Figure B-3. San Martin longitudinal profile

Block 1 (Distance = )


250 300 350 400


10

9

E 8



6


0 1 2 3 4
m


Block 1 (Distance = 12.6m)


10

9

E 8

7

6


0 1 2
m


3 4


Block 1 (Distance =39.9m)


9
E 8
7
6


0 1 2 3 4
m


Figure B-4. Gustayson cross-sectional profiles for sections along Block 1













Block 2 (Distance = 48.2m)


0 1 2 3 4 5
m


Block 2 (Distance = 55.9m)






E 8




0 1 2 3 4 5


m


Block 2 (Distance = 80.2m)


0 1 2 3 4 5
m


Block 2 (Distance = 95.3m)


0 1 2 3 4 5


Figure B-5. Gustayson cross-sectional profiles for sections along Block 2





m


Block 3 (Distance = 187.8m)



E 8-




m


Block 3 (Distance = 212.8m)


Block 3 (Distance = 104.3m)


10

E 9

8


m


Block 3 (Distance = 137m)


0 1 23 456
m


Figure B-6. Gustayson cross-sectional profiles for sections along Block 3












Block 4 (Distance = 227.3m)
10







012345678910
m

Block 4 (Distance = 277.9m)

10



7 -


012345678910



Block 4 (Distance = 319.4m)




E 7 --




012345678910
m

Block 4 (Distance = 359.9m)



E 6-



012345678910
m

Figure B-7. Gustayson cross-sectional profiles for sections along Block 4










Gustayson longitudinal profile (Block 1)


E 7

6


L~t_


meters


Gustayson longitudinal profile (Block 2)


E7


~


meters


Gustayson longitudinal profile (Block 3)


6
100 110 120 130 140 150 160 170 180 190 200 210 220
meters


Gustayson longitudinal profile (Block 4)


6

220 230 240 250 260 270 280 290 300 310 320 330 340 350 360 370
meters


Figure B-8. Gustayson longitudinal profiles for Blocks 1-4










Section 1 (Distance = 0)


Section 2 (Distance = 39.9m)


E 8.5 J



0 2 4 6 8 10 12 14 16 18
m


Section 4 (Distance = 126.8m)


E 9 -- -



0 1 2 3 4
m


Section 3 (Distance = 113.4m)


11




8


13




10


)1 2 3 4


1 2
m


3 4


Section 5 (Distance = 210m)

13

E 12

11
0 1 2 3 4
m

Figure B-9. Frey cross-sectional profiles


Longitudinal profile


100


150


200


250


meters

Figure B-10. Frey longitudinal profile














































0 12 34 56 78 9
m

Figure B-11. Koell cross-sectional profiles


7


Section 1 (Distance = 0)


Section 2 (Distance = 28.8m)


9

E8


0123456789
m


Section 4 (Distance = 90.3m)


0123456789
m


Section 3 (Distance = 60.8m)


9



6


U


9



5


-t


0 12 34 56 78 9


0 1 23 45 67 89


Seccion 5 (Distancia = 121.5m)









Longitudinal profile

----ZI_ _


10



E4


100


150


200


250


meters


Figure B-12. Koell longitudinal profile












Table Cl Summary of QA/QC results for nutrient analyses
SRP NH4-N NO3 TN TP
(mg P L-1) (mg N L-1) (mg NO3 L ) (mg N L-1) (mg P L-1)
na 32 24 18 12 10
accuracyb 10.04 10.03 10.02 10.3 10.4

n 63 49 45 20 12
reproducibilityc & 0.04 10.02 10.02 10.2 10.3
an is the number of times accuracy and reproducibility analyses were performed. bAccuracy was
determined by the difference from the known standard concentration. cReproducibility was
calculated by the standard deviation of duplicate measurements


APPENDIX C
QA/QC RESULTS










APPENDIX D
OTIS-P SIMULATION RESULTS


500-

E 400-

300 *** observed

200~~ predicted
1 00-

S00

0.0 0.5 1 .0 1.5
Time (hours)


Figure D-1. OTIS-P modeling results for 2005 tracer experiment in San Martin


180-
S 170-
t 160-
~3150-
observed
S140-
'5 predicted
S130-
S120-
cj 110-
100
0 0.5 1 1 .5 2 2.5
Time (hours)


Figure D-2. OTIS-P modeling results for 2006 tracer experiment in San Martin


350
-

v,300


O


* observed
- predicted


0 0.5


1
Time (hours)


Figure D-3. OTIS-P modeling results for 2005 tracer experiment in Gustayson


_i~_


















-predicted
a observedI


0 0.5 1 1.5 2 2.5 3
Time (hours)


Figure D-4. OTIS-P modeling results for 2006 tracer experiment in Gustayson


6000-

5000-

v,4000-
observed
S3000-
'5- predicted
S2000-




0 1 2 3 4
Time (hours)


Figure D-5. OTIS-P modeling results for 2005 tracer experiment in Koell


600


~500


S400


S300
o


.Observed|
-Predicted


Time (hours)


Figure D-6. OTIS-P modeling results for 2006 tracer experiment in Koell










APPENDIX E
PHOSPHORUS FRACTIONATION DATA


Table E-1. Phosphorus fractionation data for Koell Site
Fractionation Procedure Rep 1


1
Rep 2

0.0883
0.0922
0.0039
0.14


Porewater
Filter wt initial (g)
Filter wt final (g)
Sediment collected (g)
SRP in pore water


0.0843
0.0875
0.0032
0.13


KCl-bioavailable
Filter wt initial (g)
Filter wt final (g)
Sediment collected (g)
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg))

NaOH-Fe and Al oxyhydroxides
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg))

HCl-Ca and Mg bound
Sediment collected (g)
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg)


0.0833
0.1
0.0167
0.03
0.0006
1.14
0.526315789

inorganic
143
2.86
1.14
2508.77193


0
78
1.56
1.14
1368.421053


0.0872
0.1031
0.0159
0.05
0.001
1.14
0.877192982

total organic
307.5 164.5
6.15 3.29
1.14 1.14
5394.736842 2885.96491



88
1.76
1.14
1543.859649


Recalcitrant
P (mg/kg))

Summary
KCl
NaOH-Pi
NaOH-Po
HCI
Recalcitrant
Total


average P (mg/kg)
0.701754386
2508.77193
2885.964912
1456.140351
125
6976.170413










Table E-2. Phosphorus fractionation data for Koell
Fractionation Procedure Rep 1
Porewater
Filter wt initial (g) 0.0845
Filter wt final (g) 0.09
Sediment collected (g) 0.0055
SRP in pore water 0.1


Site 2
Rep 2


0.0877
0.0906
0.0029
0.08


KCl-bioavailable
Filter wt initial (g)
Filter wt final (g)
Sediment collected (g)
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg))

NaOH-Fe and Al oxyhydroxides
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg))

HCl-Ca and Mg bound
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg))


0.0843
0.1003
0.016
0.03
0.0006
1.7885
0.33547666

inorganic
120
2.4
1.7885
1341.90663


98
1.96
1.7885
1095.89041


0.0874
0.1033
0.0159
0.02
0.0004
1.7885
0.223651104

total
305
6.1
1.7885
3410.67934


96
1.92
1.7885
1073.525301


organic
185
3.7
1.7885
2068.7727


Recalcitrant
P (mg/kg))

Summary
KCl
NaOH-Pi
NaOH-Po
HCI
Recalcitrant
Total


average P(mg/kg)
0.27956388
1341.90663
2068.77271
1084.70786
687
5182.95877











Table E-3. Phosphorus fractionation data for Koell Site 3
Fractionation Procedure Rep 1 Rep 2
Porewater
Filter wt initial (g) 0.0873 0.0865
Filter wt final (g) 0.0898 0.0899
Sediment collected (g) 0.0025 0.0034
SRP in pore water 0.06 0.08


KCl-bioavailable
Filter wt initial (g)
Filter wt final (g)
Sediment collected (g)
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg))

NaOH-Fe and Al oxyhydroxides
Filter wt initial (g)
Filter wt final (g)
Sediment collected (g)
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg))


0.0861
0.104
0.0179
0.04
0.0008
0.9535
0.839014158

inorganic
0.0844
0.1002
0.0158
128
2.56
0.9377
2730.084249


0.0844
0.1062
0.0218
0.03
0.0006
0.9533
0.629392636

total
0.1007
0.1062
0.0055
315
6.3
0.9478
6646.971935


organic




3.74
0.9478
3945.98


HCl-Ca and Mg bound
Filter wt initial (g)
Filter wt final (g)
Sediment collected (g)
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg))

Recalcitrant
P (mg/kg))


0.0861
0.1063
0.0202
0.1
0.002
0.8173
2.4471


0.0839
0.1142
0.0303
0.18
0.0036
0.8113
4.4373


Summary
KCl
NaOH-Pi
NaOH-Po
HCI
Recalcitrant
Total


average P (mg/kg)
0.734203397
2730.084249
3945.980165
3.442202335
436
7116.352274











Table E-4. Phosphorus fractionation data for Koell Site 4
Fractionation Procedure Rep 1 Rep 2
Porewater
Filter wt initial (g) 0.0867 0.0863
Filter wt final (g) 0.0894 0.0886
Sediment collected (g) 0.0027 0.0023
SRP in pore water 0.05 0.09


KCl-bioavailable
Filter wt initial (g)
Filter wt final (g)
Sediment collected (g)
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg))

NaOH-Fe and Al oxyhydroxides
Filter wt initial (g)
Filter wt final (g)
Sediment collected (g)
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg))

HCl-Ca and Mg bound
Filter wt initial (g)
Filter wt final (g)
Sediment collected (g)
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg))


0.0863
0.1063
0.02
0.03
0.0006
3.0712
0.195363376

inorganic
0.0847
0.1447
0.06
88
1.76
3.0112
584.4845909


0.0861
0.1063
0.0202
0.1
0.002
2.991
0.668672685


0.0853
0.1067
0.0214
0.08
0.0016
3.0617
0.522585492

total
0.0849
0.1253
0.0404
190
3.8
3.0213
1257.736736


0.0839
0.1142
0.0303
0.18
0.0036
2.991
1.203610832


organic




2.04
3.0213
675.206


Recalcitrant
P (mg/kg))

Summary
KCl
NaOH-Pi
NaOH-Po
HCI
Recalcitrant
Total


average P(mg/kg)
0.358974434
584.4845909
675.2060371
0.936141759
525
1785.851816











Table E-5. Phosphorus fractionation data for San Martin Site 1
Fractionation Procedure Rep 1 Rep 2
Porewater
Filter wt initial (g) 0.0865 0.0857
Filter wt final (g) 0.0898 0.0882
Sediment collected (g) 0.0033 0.0025
SRP in pore water 0.03 0.05


KCl-bioavailable
Filter wt initial (g)
Filter wt final (g)
Sediment collected (g)
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg))


0.0854
0.1071
0.0217
0.05
0.001
1.7894
0.558846541

inorganic
0.09
0.2301
0.1401
72
1.44
1.6493
873.0976778


0.0919
0.0962
0.0043
0.05
0.001
1.645
0.607902736


0.0887
0.1129
0.0242
0.04
0.0008
1.7256
0.463606861

total
0.0842
0.1542
0.07
140
2.8
1.6556
1691.229766


0.0885
0.0991
0.0106
0.13
0.0026
1.645
1.580547112


NaOH-Fe and Al oxyhydroxides
Filter wt initial (g)
Filter wt final (g)
Sediment collected (g)
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg))


organic




1.36
1.6556
821.4544576


HCl-Ca and Mg bound
Filter wt initial (g)
Filter wt final (g)
Sediment collected (g)
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg))


Recalcitrant
P (mg/kg))

Summary
KCl
NaOH-Pi
NaOH-Po
HCI
Recalcitrant
Total


444


average P(mg/kg)
0.511226701
873.0976778
821.4544576
1.094224924


2140.130464











Table E-6. Phosphorus fractionation data for San Martin Site 2
Fractionation Procedure Rep 1 Rep 2
Porewater
Filter wt initial (g) 0.0879 0.0862
Filter wt final (g) 0.0945 0.0933
Sediment collected (g) 0.0066 0.0071
SRP in pore water 0.06 0.05


KCl-bioavailable
Filter wt initial (g)
Filter wt final (g)
Sediment collected (g)
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg))

NaOH-Fe and Al oxyhydroxides
Filter wt initial (g)
Filter wt final (g)
Sediment collected (g)
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg))


0.0857
0.1116
0.0259
0.04
0.0008
1.4485
0.552295478

inorganic
0.088
0.1106
0.0226
162
3.24
1.4259
2272.249106


0.108
0.1114
0.0034
0.17
0.0034
1.4225
2.390158172


0.0897
0.1055
0.0158
0.02
0.0004
1.5092
0.265041081


total
0.0999
0.1866
0.0867


organic


185
3.7 0.46
1.4225 1.4225
2601.054482 323.3743409


HCl-Ca and Mg bound
Filter wt initial (g)
Filter wt final (g)
Sediment collected (g)
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg))


0
0.11
0.0022
1.4225
1.546572935


Recalcitrant
P (mg/kg))

Summary
KCl
NaOH-Pi
NaOH-Po
HCI
Recalcitrant
Total


average P(mg/kg)
0.40866828
2272.249106
323.3743409
1.968365554
678
3276.03997










Table E-7. Phosphorus fractionation data for
Fractionation Procedure Rep 1
Porewater
Filter wt initial (g) 0.0888
Filter wt final (g) 0.0971
Sediment collected (g) 0.0083
SRP in pore water 0.07


San Martin Site 3
Rep 2


0.0899
0.0961
0.0062
0.05


KCl-bioavailable
Filter wt initial (g)
Filter wt final (g)
Sediment collected (g)
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg))

NaOH-Fe and Al oxyhydroxides
Filter wt initial (g)
Filter wt final (g)
Sediment collected (g)
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg))

HCl-Ca and Mg bound
Filter wt initial (g)
Filter wt final (g)
Sediment collected (g)
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg))


0.0867
0.1207
0.034
0.1
0.002
2.6589
0.752190756

inorganic
0.0876
0.1279
0.0403
20
0.4
2.6186
152.7533797


0.087
0.0946
0.0076
0.25
0.005
2.611
1.914975105


0.0863
0.1227
0.0364
0.07
0.0014
2.6691
0.524521374

total
0.0864
0.1213
0.0349
252
5.04
2.6342
1913.294359


0.0869
0.1101
0.0232
0.15
0.003
2.611
1.148985063


organic



232
4.64
2.6342
1761.446


Recalcitrant
P (mg/kg))

Summary
KCl
NaOH-Pi
NaOH-Po
HCI
Recalcitrant
Total


average P (mg/kg)
0.638356065
152.7533797

1.531980084
276
430.860863










Table E-8. Phosphorus fractionation data for
Fractionation Procedure Rep 1
Porewater
Filter wt initial (g) 0.086
Filter wt final (g) 0. 1011
Sediment collected (g) 0.0151
SRP in pore water 0.07


San Martin Site 4
Rep 2


0.086
0.1018
0.0158
0.09


KCl-bioavailable
Filter wt initial (g)
Filter wt final (g)
Sediment collected (g)
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg))


0.0865
0.1415
0.055
0.03
0.0006
0.6374
0.941324129

inorganic
0.0876
0.1279
0.0403
34
0.68
1.1866
573.0659026


0.087
0.0946
0.0076
0.04
0.0008
1.179
0.678541137


192


0.0869
0.1081
0.0212
0.02
0.0004
0.6476
0.617665225

total
0.0864
0.1213
0.0349
156
3.12
1.2022
2595.242056


NaOH-Fe and Al oxyhydroxides
Filter wt initial (g)
Filter wt final (g)
Sediment collected (g)
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg))


organic




2.44
1.179
2069.55047


HCl-Ca and Mg bound
Filter wt initial (g)
Filter wt final (g)
Sediment collected (g)
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg))

Recalcitrant
P (mg/kg))


0.0869
0.1101
0.0232


1.179



192


Summary
KCl
NaOH-Pi
NaOH-Po
HCI
Recalcitrant
Total


average P(mg/kg)
0.779494677
573.0659026
2069.550466
0.678541137
192
2836.23734










Table E-9. Phosphorus fractionation data for Gustayson Site 1
Fractionation Procedure Rep 1 Rep 2
Porewater
Filter wt initial (g) 0.0928 0.092
Filter wt final (g) 0.1233 0.1229
Sediment collected (g) 0.0305 0.0309
SRP in pore water 1.8 0.133333333


KCl-bioavailable
Filter wt initial (g)
Filter wt final (g)
Sediment collected (g)
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg))

NaOH-Fe and Al oxyhydroxides
Filter wt initial (g)
Filter wt final (g)
Sediment collected (g)
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg))

HCl-Ca and Mg bound
Filter wt initial (g)
Filter wt final (g)
Sediment collected (g)
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg))


0.0936
0.095
0.0014
2.3
0.046
0.1906
241.343127

inorganic
0.0846
0.0895
0.0049
110
2.2
0.1857
11847.06516


0.0909
0.0971
0.0062
9
0.18
0.1795
1002.785515


0.0927
0.0928
1E-04
2
0.04
0.1796
222.7171492

total
0.0895
0.0895
0
123
2.46
0.1796
13697.10468


0.092
0.0921
0.0001
22
0.44
0.1795
2451.253482


organic




0.26
0.1796
1447.6615


Recalcitrant
P (mg/kg))

Summary
KCl
NaOH-Pi
NaOH-Po
HCI
Recalcitrant
Total


average P(mg/kg)
232.0301381
11847.06516
1447.66147
1727.019499
665
15919.13894










Table E-10. Phosphorus fractionation data for Gustayson Site 2
Fractionation Procedure Rep 1 Rep 2
Porewater
Filter wt initial (g) 0.0894 0.0909
Filter wt final (g) 0.1047 0.1142
Sediment collected (g) 0.0153 0.0233
SRP in pore water 6.2 4.933333333


KCl-bioavailable
Filter wt initial (g)
Filter wt final (g)
Sediment collected (g)
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg))

NaOH-Fe and Al oxyhydroxides
Filter wt initial (g)
Filter wt final (g)
Sediment collected (g)
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg))

HCl-Ca and Mg bound
Filter wt initial (g)
Filter wt final (g)
Sediment collected (g)
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg))


0.0935
0.0939
0.0004
0.5
0.01
0.2101
47.59638267

inorganic
0.0923
0.1136
0.0213
150
3
0.1888
15889.83051


0.0926
0.1019
0.0093
41
0.82
0.1795
4568.245125


0.0928
0.0949
0.0021
1.5
0.03
0.2066
145.2081317

total
0.0869
0.108
0.0211
170
3.4
0.1855
18328.84097


0.0917
0.0977
0.006
48
0.96
0.1795
5348.189415


organic




0.4
0.1855
2156.3342


Recalcitrant
P (mg/kg))

Summary
KCl
NaOH-Pi
NaOH-Po
HCI
Recalcitrant
Total


average P(mg/kg)
96.40225717
15889.83051

4958.21727
527
21471.37817










Table E-11. Phosphorus fractionation data for Gustayson Site 3
Fractionation Procedure Rep 1 Rep 2
Porewater
Filter wt initial (g) 0.0911 0.0874
Filter wt final (g) 0.1317 0.1291
Sediment collected (g) 0.0406 0.0417
SRP in pore water 2.5 0.8


KCl-bioavailable
Filter wt initial (g)
Filter wt final (g)
Sediment collected (g)
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg))

NaOH-Fe and Al oxyhydroxides
Filter wt initial (g)
Filter wt final (g)
Sediment collected (g)
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg))

HCl-Ca and Mg bound
Filter wt initial (g)
Filter wt final (g)
Sediment collected (g)
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg))


0.0934
0.0936
0.0002
2.7
0.054
0.2091
258.2496413

inorganic
0.0897
0.1112
0.0215
125
2.5
0.1876
13326.22601


0.0925
0.1006
0.0081
31
0.62
0.1795
3454.038997


0.0936
0.0943
0.0007
2.8
0.056
0.2096
267.1755725

total
0.0869
0.108
0.0211
137.5
2.75
0.1885
14588.85942


0.09
0.099
0.009
26
0.52
0.1795
2896.935933


organic




0.25
0.1885
1326.2599


Recalcitrant
P (mg/kg))

Summary
KCl
NaOH-Pi
NaOH-Po
HCI
Recalcitrant
Total


average P(mg/kg)
262.7126069
13326.22601
1326.259947
3175.487465
657
18747.65363










Table E-12. Phosphorus fractionation data for Gustayson Site 4
Fractionation Procedure Rep 1 Rep 2
Porewater
Filter wt initial (g) 0.0894 0.0901
Filter wt final (g) 0.1013 0.1088
Sediment collected (g) 0.0119 0.0187
SRP in pore water 1.5 0.9


KCl-bioavailable
Filter wt initial (g)
Filter wt final (g)
Sediment collected (g)
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg))

NaOH-Fe and Al oxyhydroxides
Filter wt initial (g)
Filter wt final (g)
Sediment collected (g)
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg))

HCl-Ca and Mg bound
Filter wt initial (g)
Filter wt final (g)
Sediment collected (g)
SRP (mgP/1)
P (mg)
sample wt dry (g)
P (mg/kg))


0.0929
0.0929
0
1.1
0.022
0.2028
108.4812623

inorganic
0.0861
0.1013
0.0152
152
3.04
0.1876
16204.69083


0.0925
0.1006
0.0081
31
0.62
0.1795
3454.038997


0.0936
0.0936
0
2
0.04
0.2094
191.0219675

total
0.088
0.1089
0.0209
157.5
3.15
0.1885
16710.87533


0.09
0.099
0.009
26
0.52
0.1795
2896.935933


organic




0.11
0.1885
583.55438


Recalcitrant
P (mg/kg))

Summary
KCl
NaOH-Pi
NaOH-Po
HCI
Recalcitrant
Total


average P (mg/kg)
149.7516149
16204.69083
583.5543767
3175.487465
657
20770.12855










APPENDIX F
PHOSPHORUS SORPTION INDEX (PSI) DATA

Table F-1. Year 2005 PSI data for Koell sites


TOTAL
K 1(1)
30.70

2.95

905.10
10.10

89.60
30.04

K 2(1)
30.70

18.05
412.52
10.45

39.49
10.48

K 3(1)
30.70

4.56
852.43
11.27

75.67
23.85

K 4(1)
30.70

6.08
802.87
10.59

75.78
22.98


(BIOTIC+ABIOTIC)
K 1(2) K 1(3)
30.70 30.70


DEAD (ABIOTIC)


K 1(1)
30.70

22.42

270.01
7.74

34.89
9.03

K 2(1)
30.70

44.08
-436.33
9.14

-47.73
-11.48

K 3(1)
30.70

45.98
-498.29
5.54

-89.91
-21.53

K 4(1)
30.70

25.65
164.68
10.40

15.83
4.04


K 1(2)
30.70

15.58

493.07
7.84

62.89
16.97

K 2(2)
30.70

50.92
-659.38
9.94

-66.34
-15.72

K 3(2)
30.70

36.48
-188.49
11.10

-16.99
-4.17

K 4(2)
30.70

29.83
28.37
10.28

2.76
0.69


K 1(3)
30.70

21.47

300.99
8.46

35.58
9.25

K 2(3)
30.70

49.59
-616.01
10.26

-60.02
-14.26

K 3(3)
30.70

42.37
-380.56
10.95

-34.76
-8.40

K 4(3)
30.70

28.31
77.94
10.44

7.46
1.88


Ci (mg PO4/1)
Cf Live (mg
PO4/1)
P sorbed (ug)
Dry mass (g)
ug P sorbed/g
sed
PSI


Ci (mg PO4/1)
Cf Live (mg
PO4/1)
P sorbed (ug)
Dry mass (g)
ug P sorbed/g
sed
PSI


Ci (mg PO4/1)
Cf Live (mg
PO4/1)
P sorbed (ug)
Dry mass (g)
ug P sorbed/g
sed
PSI


Ci (mg PO4/1)
Cf Live (mg
PO4/1)
P sorbed (ug)
Dry mass (g)
ug P sorbed/g
sed
PSI


4.56

852.43
7.75


5.13

833.85
10.68


109.99 78.09
34.67 24.23


K 2(2)
30.70

16.34
468.28
9.76


K 2(3)
30.70

21.47
300.99
10.70


47.97 28.13
12.87 7.32


K 3(2)
30.70

2.47
920.59
10.91


K 3(3)
30.70

4.18
864.83
11.21


84.36 77.15
29.03 24.61


K 4(2)
30.70

2.28
926.78
10.31


K 4(3)
30.70

8.93
709.93
10.28


89.86 69.04
31.29 19.93










Table F-2. Year 2005 PSI data for Frey sites
TOTAL (BIOTIC+ABIOTIC)


DEAD (ABIOTIC)


F 1(1)
30.70


F 1(2)
30.70


F 1(3)
30.70


F 1(1)
30.70

11.70
619.60
8.41

73.70
20.58

F 2 (1)
30.70

9.90
678.29
4.15

163.33
46.55

F 3 (1)
30.70

22.50
267.40
9.02

29.64
7.67


F 1(2)
30.70

9.99
675.36
8.08

83.58
23.79

F 2 (2)
30.70

9.63
687.10
4.71

145.82
41.70

F 3 (2)
30.70

21.42
302.62
8.56

35.37
9.20


F 1(3)
30.70

13.32
566.77
8.02

70.65
19.42

F 2 (3)
30.70

9.72
684.16
4.46

153.37
43.81

F 3 (3)
30.70

15.84
484.59
8.00

60.59
16.32


Ci (mg PO4/1)
Cf Live (mg
PO4/1)
P sorbed (ug)
Dry mass (g)
ug P sorbed/g
sed
PSI


Ci (mg PO4/1)
Cf Live (mg
PO4/1)
P sorbed (ug)
Dry mass (g)
ug P sorbed/g
sed
PSI


Ci (mg PO4/1)
Cf Live (mg
PO4/1)
P sorbed (ug)
Dry mass (g)
ug P sorbed/g
sed
PSI


2.97 6.48 1.71
904.28 789.82 945.37


9.06

99.80
33.42

F 2 (1)
30.70

6.48
789.82
4.46


9.07

87.05
26.18


9.09

104.04
37.88

F 2 (3)
30.70

3.06
901.35
4.47


F 2 (2)
30.70

3.24
895.48
4.39


177.09 203.98 201.78
53.26 67.46 67.28

F 3 (1) F 3 (2) F 3 (3)


30.70

1.62
948.31
8.67


30.70

1.08
965.92
8.80


30.70

2.43
921.89
8.96

102.88
35.49


109.38 109.81
40.17 43.12









Table F-3. Year 2005 PSI data for Gustayson sites
TOTAL (BIOTIC+ABIOTIC)


DEAD (ABIOTIC)


Gl (1)
30.70


G1 (2)
30.70


G1 (3)
30.70


Gl (1)
30.70

20.90
319.58
8.35

38.27
9.98

G4 (1)
30.70

27.93
90.33
10.03

9.00
2.27


Gl (2)
30.70

18.81
387.74
8.30

46.72
12.33

G4 (2)
30.70

29.83
28.37
10.11

2.81
0.70


G1 (3)
30.70

26.60
133.70
8.47

15.79
4.01

G4 (3)
30.70

30.59
3.59
11.04

0.33
0.08


Ci (mg PO4/1)
Cf Live (mg
PO4/1)
P sorbed (ug)
Dry mass (g)
ug P sorbed/g
sed
PSI


Ci (mg PO4/1)
Cf Live (mg
PO4/1)
P sorbed (ug)
Dry mass (g)
ug P sorbed/g
sed
PSI


4.94 4.75 11.02
840.04 846.24 641.77


9.26

90.70
28.28

G4 (1)
30.70

11.97
610.79
10.47

58.34
16.24


9.25

91.45
28.67

G4 (2)
30.70


9.34

68.73
19.33

G4 (3)
30.70


22.61 21.28
263.82 307.19


10.22

25.81
6.67


10.30

29.84
7.77









Table F-4. Year 2005 PSI data for San Martin sites
TOTAL ABIOTIC
SM2 (1) SM2 (1)
Ci (mg PO4/1) 30.70 30.70
Cf Live (mg
PO4/1) 5.40 18.81
P sorbed (ug) 825.04 387.74
Dry mass (g) 10.49 10.26
ug P sorbed/g
sed 78.64 37.78
PSI 24.23 9.97

SM4 (1) SM4 (1)
Ci (mg PO4/1) 30.70 30.70
Cf Live (mg
PO4/1) 7.22 16.34
P sorbed (ug) 765.69 468.28
Dry mass (g) 10.49 10.26
ug P sorbed/g
sed 72.99 45.63
PSI 21.65 12.24









Table F-5. Year 2006 PSI data for Koell sites
TOTAL (BIOTIC+ABIOTIC)
K 1(1) K 1(2) K 1(3)
Ci (mg P/1) 50.00 50.00 50.00
Cuaderno 0.33 0.24 0.17
Cf (mg P/1) 3.30 2.40 1.70
P sorbed (ug) 4670.00 4760.00 4830.00
mass i (g) 109.077 112.258 108.622
mass f (g) 116.468 119.94 116.863
samples mass (g) 7.39 7.68 8.24
ug P sorbed/g sed 631.85 619.63 586.09
PSI 179.58 183.31 181.43

K2 (1) K2 (2) K3 (3)
Ci (mg P/1) 50.00 50.00 50.00
Cuaderno 0.29 0.39 0.39
Cf (mg P/1) 2.90 3.90 3.90
P sorbed (ug) 4710.00 4610.00 4610.00
mass i (g) 108.413 110.52 104.985
mass f (g) 116.078 118.141 112.333
samples mass (g) 7.67 7.62 7.35
ug P sorbed/g sed 614.48 604.91 627.38
PSI 177.47 168.45 174.71

K3 (1) K3 (2) K3 1(3)
Ci (mg P/1) 50.00 50.00 50.00
Cuaderno 0.28 0.25 0.34
Cf (mg P/1) 2.80 2.50 3.40
P sorbed (ug) 4720.00 4750.00 4660.00
mass i (g) 108.17 110.36 106.204
mass f (g) 113.962 116.104 111.831
samples mass (g) 5.79 5.74 5.63
ug P sorbed/g sed 814.92 826.95 828.15
PSI 236.40 243.37 234.51

K4 (1) K4 (2) K4 (3)
Ci (mg P/1) 50.00 50.00 50.00
Cuaderno 0.18 0.26 0.16
Cf (mg P/1) 1.80 2.60 1.60
P sorbed (ug) 4820.00 4740.00 4840.00
mass i (g) 107.95 111.214 111.45
mass f (g) 112.165 115.392 115.869
samples mass (g) 4.22 4.18 4.42
ug P sorbed/g sed 1143.53 1134.51 1095.27
PSI 351.29 332.22 341.83









Table F-6. Year 2006 PSI data for Gustayson sites
TOTAL (BIOTIC+ABIOTIC)
Gl (1) Gl (2) G1 (3)
Ci (mg P/1) 50.00 50.00 50.00
Cf (mg P/1) 14.40 10.40 12.00
P sorbed (ug) 3560.00 3960.00 3800.00
mass i (g) 108.60 111.72 106.96
mass f (g) 121.31 121.83 116.35
samples mass (g) 12.71 10.11 9.39
ug P sorbed/g sed 280.16 391.77 404.82
PSI 67.37 97.53 99.24

G2 (1) G2 (2) G2 (3)
Ci (mg P/1) 50.00 50.00 50.00
Cf (mg P/1) 4.00 3.60 2.80
P sorbed (ug) 4600.00 4640.00 4720.00
mass i (g) 104.99 106.45 112.25
mass f (g) 111.27 112.75 118.71
samples mass (g) 6.29 6.29 6.46
ug P sorbed/g sed 731.90 737.21 731.10
PSI 203.19 207.30 212.09

G3 (1) G3 (2) G3 (3)
Ci (mg P/1) 50.00 50.00 50.00
Cf (mg P/1) 0.90 1.30 0.90
P sorbed (ug) 4910.00 4870.00 4910.00
mass i (g) 205.84 110.08 106.18
mass f (g) 211.74 115.86 112.65
samples mass (g) 5.90 5.78 6.47
ug P sorbed/g sed 832.77 842.85 758.77
PSI 281.89 320.81 307.49

G4 (1) G4 (2) G4 (3)
Ci (mg P/1) 50.00 50.00 50.00
Cf (mg P/1) 6.90 3.50 5.30
P sorbed (ug) 4310.00 4650.00 4470.00
mass i (g) 109.07 106.77 111.22
mass f (g) 120.48 117.72 122.02
samples mass (g) 11.41 10.95 10.80
ug P sorbed/g sed 377.81 424.77 413.89
PSI 112.70 138.93 127.84









Table F-7. Year 2006 PSI data for San Martin sites
TOTAL (BIOTIC+ABIOTIC)
SM1 (1) SM1 (2) SM1 (3)
Ci (mg P/1) 50.00 50.00 50.00
Cf (mg P/1) 4.00 0.40 13.20
P sorbed (ug) 4600.00 4960.00 3680.00
mass i (g) 112.34 111.74 112.26
mass f (g) 128.06 128.14 127.44
samples mass (g) 15.72 16.41 15.18
ug P sorbed/g sed 292.58 302.33 242.44
PSI 81.23 116.19 58.84

SM2 (1) SM2 (2) SM2 (3)
Ci (mg P/1) 50.00 50.00 50.00
Cf (mg P/1) 11.00 10.60 11.20
P sorbed (ug) 3900.00 3940.00 3880.00
mass i (g) 107.96 110.88 108.41
mass f (g) 122.62 125.27 123.22
samples mass (g) 14.67 14.39 14.81
ug P sorbed/g sed 265.92 273.86 262.07
PSI 65.80 68.03 64.72

SM3 (1) SM3 (2) SM3 (3)
Ci (mg P/1) 50.00 50.00 50.00
Cf (mg P/1) 0.40 0.40 1.20
P sorbed (ug) 4960.00 4960.00 4880.00
mass i (g) 104.99 110.09 109.90
mass f (g) 116.41 121.70 122.36
samples mass (g) 11.43 11.62 12.45
ug P sorbed/g sed 434.14 426.92 391.84
PSI 166.84 164.07 127.26

SM4 (1) SM4 (2) SM4 (3)
Ci (mg P/1) 50.00 50.00 50.00
Cf (mg P/1) 12.80 22.80 21.20
P sorbed (ug) 3720.00 2720.00 2880.00
mass i (g) 106.77 109.95 111.26
mass f (g) 118.64 121.97 123.27
samples mass (g) 11.86 12.02 12.01
ug P sorbed/g sed 313.61 226.35 239.78
PSI 76.36 51.94 55.42











APPENDIX G
EQUILIBRIUM PHOSPHORUS CONCENTRATION (EPCO) DATA


6.00
5.00
4.00
3.00
2.00
1.00
0.00


y = 2.5506x 0.8473
R2 = 0.9904


0.00 0.50 1.00 1.50 2.00 2.50 3.00

P in solution (mg/I)


Figure G-1. 2005 EPCO data for Koell site 1


10.00

-08.00
6.00

4.00
0 2.00
a-0.00

-2.0d)


y = 4.2103x 2.4939
R2 = 0.99631'


0.50 1.00 1.50 2.00 2.50 3.00


P in solution (mg/I)


Figure G-2. 2005 EPCO data for Koell site 2


7.00-
6.00 -
5.00 -
4.00-
3.00-
2.00 -
1.00-
0.00-
0.00


y = 2.5565x 1.6468
R2 = 0.9183





*r


0.50 1.00 1.50 2.00 2.50 3.00 3.50


P in solution


Figure G-3. 2005 EPCO data for Koell site 3












7.00
6.00
5.00
4.00
3.00
2.00
1.00
0.00


y = 2.7791x 1.3723
R2 = 0.98 /


0.00 0.50 1.00 1.50 2.00 2.50 3.00

P in solution (mg/I)


Figure G-4. 2005 EPCO data for Koell site 4


3.00
2.00
1.00
0.00
-1.00C
-2.00
-3.00
-4.00
-5.00


y = 2.9154x 6.7276
R2 = 0.8211 /


0.50


2.50 3.00 3.50


P in solution (mg/1)


Figure G-5. 2005 EPCO data for Gustayson site 1


10.00
8.00
6.00
4.00
2.00
0.00
-2.00)
-4.00


y = 4.8375x 6.2693


2.00 3.00


4.00


P in solution (mgll)


Figure G-6. 2005 EPCO data for Gustayson site 2












y = 2.4813x 1.2772
R2 = 0.9629


8.00

6.00

4.00

2.00

0.00


0.50 1.00 1.50 2.00 2.50


3.00 3.50


P in solution (mg/I)


Figure G-7. 2005 EPCO data for Gustayson site 3


ony = 4.4838x -7.0869


8.00
6.00
4.00

0.00
-2.000.
-4.00


R2 =0.853


0.50 1. 0 /.0 2.00 2.50


3.00 3.50


-6.00

P in solution (mg/I)


Figure G-8. 2005 EPCO data for Gustayson site 4


7.00
-a 6.00
m, 5.00
-" 4.00
S3.00
5~2.00
S1.00
0.00


y = 2.5786x 0.7148
R2 = 0.9808


0.00 0.50 1.00 1.50 2.00 2.50 3.00

P in solution (mg/1)


Figure G-9. 2005 EPCO data for San Martin site 2











y = 2.2826x 0.6539
R2 = 0.9963


6.00
5.00
4.00
3.00
2.00
1.00
0.00


0.00 0.50 1.00 1.50 2.00

P in solution (mg/1)


2.50 3.00


Figure G-10. 2005 EPCO data for San Martin site 4


12.00
10.00
8.00
6.00
4.00
2.00


y = 3.8476x 2.117
R2 = 0.9596


0.00
0.00 0.50 1.00 1.50 2.00 2.50 3.00 3.50


P in solution (mg/1)

Figure G-11. 2005 EPCO data for Frey site 1


20.00


y = 6.0421x 1.9738
R2 = 0.992


1 0.00

(n 5.00


0.00
0.00 0.50 1.00 1.50 2.00 2.50 3.00 3.50

P in solution (mg/1)


Figure G-12. 2005 EPCO data for Frey site 2












10.00 y = 2.8625x 0.6496
8 8.00 -R2 = 0.9906

? 6.00-

-84.00-

0. 2.00-
0.00
0.00 0.50 1.00 1.50 2.00 2.50 3.00 3.50

P in solution (mg/1)


Figure G-13. 2005 EPCO data for Frey site 3


40.00 -y =13.973x -1.2979
(n 30.00 -R2 = 0.9961

a 20.00-

58 10.00-

0.00
0.00 0.50 1.00 1.50 2.00 2.50

P in solution (mg/1)


Figure G-14. 2006 EPCO data for Koell site 1


-a25.00 -y=10.015x 1.1658
v, 20.00 -R2 = 0.99
15.00-
a,10.00-
v, 5.00-
0.00
0.00 0.50 1.00 1.50 2.00 2.50

P in solution (mg/1)


Figure G-15. 2006 EPCO data for Koell site 2












y =17.417x -1.6786
R2 = 0.9979


40.00 _

30.00-

20.00-

10.00-

0.00-


0.00 0.50 1.00 1.50

P in solution (mg/1)


Figure G-16. 2006 EPCO data for Koell site 3


2.00


2.50


y = 21.713x 1.2012
R2 = 0.9999


50.00 -
40.00

30.00-
20.00-

10.00-
0.00-


0.00 0.50 1.00 1.50

P in solution (mg/1)


Figure G-17. 2006 EPCO data for Koell site 4


5.00-
4.00 y = 8.119x 2.3081
R2 = 0.9866
m, 3.00-
2.00-
S1.00-
mj 0.00
-1.000.30 *Bfl 0.40 0.60
-2.00-

P in solution (mg/1)


Figure G-18. 2006 EPCO data for Gustayson site 1


2.00


2.50
















1.00


0.80











4.00 -
3.50 -
3.00-
2.50-
2.00-
1.50-
1.00-
0.50-
0.00-
0.00


y = 4.629x 0.2785
R2 = 0.9993~/


0.20


0.40 0.60


0.80


P in solution (mg/1)


Figure G-19. 2006 EPCO data for Gustayson site 2


y = 2.0728x 0.6432
R2 = 0.9814


6.00
5.00
4.00
3.00
2.00
1.00


-0. 50. Ocp 0.50 1.00 1.50 2.00 2.50 3.00 3.50
-2.00~

P in solution (mg/1)


Figure G-20. 2006 EPCO data for Gustayson site 3


14.00
12.00
10.00
8.00
6.00
4.00
2.00
0.00


y =13.6x -3.5117
R2 = 0.9941


0.00 0.20 0.40 0.60 0.80 1.00 1.20 1.40

P in solution (mg/1)


Figure G-21. 2006 EPCO data for Gustayson site 4











y = 5.4606x 0.3245
R2 = 0.9994


14.00-
12.00-
10.00-
8.00-
6.00-
4.00-
2.00-
0.00-
0.00


0.50


1.50


2.00


2.50


P in solution (mg/1)


Figure G-22. 2006 EPCO data for San Martin site 1


.025.00 -
S20.00 -
15.00-

10 .00-
5.00-
0.00-


y = 8.905x 0.613
R2 = 0.9993


0.50


2.00


2.50


P in solution (mg/1)


Figure G-23. 2006 EPCO data for San Martin site 2


y = 8.2791x 1.1447
R2 = 0.9965


20.00

v,15.00

S10.00

v,5.00


0.00
0.00


0.50 1.00 1.50 2.00


2.50


P in solution (mg/1)

Figure G-24. 2006 EPCO data for San Martin site 3











20.00y = 6.8756x 0.3777
R2
v,15.00-

S10.00-

52 5.00-

0.00
0.00 0.50 1.00 1.50 2.00 2.50
P in solution (mg/1)


Figure G-25. 2006 EPCO data for San Martin site 4




















































r. )ll Bar canrmn lmc 6n del mad i amben le
I aes~~coslanaconsrlrlrur ad mrnianimsons
hrqiegr brk esidimndel vearndanapra prl s reden Ie
ybrslurslrls
i.pro~vachrr h;hlsursenis cn el agua purR cll~rvrr
planlrr IIInsy~fornrmoniales
li'Brinds r ano verlan drd pr albad ucace 6hmbmalall


Figure H-1. Pamphlet distributed to citizens to inform and receive their input on the
implementation of modified ditches. A) Front and back pages and B) inside pages


o:~Ib ~ r 6~d n o a ar s maI d h deg~
So peden calapur can las Ilvuvi y (Iraocanes r mono
que haya dos r slamrr se~parades f une pr ra l des gus y
alradealcunlrrlldld ol ~s lubscn maygrandes.
STutue son calou (qro ximrdr men I+ 121als- per melm
pllrl Insulrlb ymon diklespareI reParrlosY madifrrrlos.


Ikta d+primpr proayta pilatae Jir~n Kaall

un~rdn rdual c~ardai


Ly n;;a d .5 und.rr~.U

'er Jian Lr ,Orp ..
lureafed u


0
A


Intr=L.=2.I p y~s=
).uchrr comendrdes del mande no Iunnn h: rearm

Ew la associa dr 1as To-wusa su fcldn M plre tra twlan to
canwwkmal,, as Imparstant eplarl apcicmes de
depuaracn nacanwnciansla-.
Losrealladcade I esle sdi demosR rard l pblldrd do
~usr 1ss angus no sal come crnales P.T ra uladrr los
des~hoche sino que rrn meanismo de Inralmtnlo para

-m II ~rl r Ir zal Ir -adular .).

muchrs com mal des porque aon eon.5m rr no requ ren
mdsesprce ni Ihnfrassrlclrlls Compblg hic1ra.


Cyede n-.. cma66dficd


SPazospar1 abs cimulacI~en de ldes.
Zabnrs d e yo veloadrd can pblanu fblanlaiesra
rumenlrr I~empo d+ Iransparle, p~r l llanic pr~veer
major arveld+ rala1menlo.
SVeglace n phanlrdr on lo bordes y rden Ira de 1ss an as
rerrovacher hnlannurseniesndargua era su c reampania


Z onasdeplaJdraparaaoxngerclagu ~aJrsdcekhcobrsy
Plad ra de crlulz re~hen ~ leso.
E qrgn quepas porrdmilredo bor~prcessareles
plednr s esd fllrad par erosasr tIkrr blalgics y


Ish. 0...1.;~;pl~."o... .n... ...E<'1=
Ts Appid uam


APPENDIX H
INTERPRETIVE MATERIALS USED FOR COMMUNITY PARTICIPATION


para

Depuraciajn de Agues
Residuales

en Znjas





Or, 8VOr

No Arrojor Basura

rrogecto riloto
de Part~icipacion Comunitaria para la
Deruracion de Ag~uas Residuales en Zanjas


Figure H1-2. Image of the sign painted and installed at each of the modified ditches, translated as
"Please do not throw trash: Pilot proj ect for community participation of wastewater
treatment in ditches."










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BIOGRAPHICAL SKETCH

Lynn Velisha Saunders was born May 6, 1977 in Anderson, Indiana and grew up in El

Paso, Texas. She attended UT Austin for a year before deciding that Texas just wasn't big

enough. To her parent' s dismay she left school and headed west to California to work in

Yosemite National Park. After a year, she moved north to Arcata, California where she attended

College of the Redwoods and later transferred to Humboldt State University (HSU). During her

three years in the Environmental Resources Engineering program at HSU, Lynn had several

opportunities to work and travel in Central America where she became aware of the human and

ecosystem health effects resulting from the lack of access to adequate water and sanitation.

Working as a student research assistant for two years at the Arcata Marsh and Wildlife

Sanctuary opened her eyes to the field of ecological engineering. She was inspired by the idea

that water treatment systems could be designed to provide multiple and mutual benefits such as

wetland habitat, aquaculture, recreation and aesthetics. After completing her B.S. degree at HSU,

she began her graduate program in Environmental Engineering Sciences at the University of

Florida with an emphasis on Ecological Engineering.

Her hope is that her Ph.D. research on wastewater ditches in Peru will serve as an example

of a low-tech approach of decentralized and community-based wastewater management in a

region that currently lacks the resources to implement conventional water treatment. After

completion of their degrees, Lynn and her husband Tom plan to continue working on applied

water management issues in the U.S. and abroad.





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1 TREATMENT POTENTIAL OF WASTEWATER DRAINAGE DITCHES IN A RURAL COMMUNITY OF THE ANDEAN AMAZON By LYNN VELISHA SAUNDERS A DISSERTATION PRESENTED TO THE GRADUATE SCHOOL OF THE UNIVERSITY OF FLOR IDA IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF DOCTOR OF PHILOSOPHY UNIVERSITY OF FLORIDA 2007

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2 2007 Lynn Velisha Saunders

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3 To my parents and to my namesake, Dr. Mary Velisha McIndoo

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4 ACKNOWLEDGMENTS I am grateful to the many special people in my lif e that have directly or indirectly led me to this point in time and space. Endless gratitude is deserved to my parents, Bill and Brenda McIndoo, who have always been incredible beacons of love and support, even throughout those rough adolescent years. I thank Andrew McIndoo for always being an awesome and caring brother, not to mention a great travel companion. I also thank Kori Jacobs, a wonderful friend whos kept me laughing over these last five year s during the Bartow days, field work in Peru, and the many nights just hanging out over dinner and wine. I express my great appreciation to my me ntors at Humboldt State University: Bob Gearheart, Beth Eschenbach and Brad Finney. It was with their guidance, support and inspiration that I chose to continue w ith my formal education. I thank my current advisor, Mark Brown for allowing and trusting me to spread my wings and develop my own research proj ect; for providing me with fi nancial support for conferences and equipment; and most importantly, for introd ucing me to the world of systems ecology. I thank Joe Delfino for his support by loaning me much-needed laboratory equipment. I also express my appreciation to Mich ael McClain for his inspiration a nd support as well as for taking the time out of his busy schedule to come up to Gainesville from Miami to attend my qualifying exams. I also thank Jim Jawitz for helpful sugges tions on my research and for his wry humor that made subsurface contaminant hydrology lectures entertaining. Thanks to Mark Clark for his support and insightful inquiri es into this research. Thanks to Ed Dunne for his thoughtful review of my chapter related to phosphorus sorption. My sincere gratitude goes to Joaquin Arteaga, former director of public works related to water in Oxapampa and later an assistant for this research, who provi ded me with important background data that were crucia l for this research. I thank my hard-working and wastewater

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5 ditch-tolerant field assistants, Jaime Guerovich and Gino Arteaga Koo. I also thank to Percy Summers, director of ProPachitea at the Instituto de Bien Comn for allowing me to use their research facilities in Oxapampa. Special thanks goes to all my friends in Oxapampa, Peru who supported and worked with me on this study an d made my experiences there so treasured. And finally, and most of all, to my husband Tom, the gr eatest husband, friend, mentor, collaborator and lifelong partner that I could ever imagine.

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6 TABLE OF CONTENTS page ACKNOWLEDGMENTS...............................................................................................................4 LIST OF TABLES................................................................................................................. ..........9 LIST OF FIGURES................................................................................................................ .......11 ABSTRACT....................................................................................................................... ............18 CHAPTER 1 INTRODUCTION..................................................................................................................20 Statement of the Problem....................................................................................................... .20 Study Overview and Objectives.............................................................................................21 Study Setting.................................................................................................................. .........22 Water Quality in the Chorobamba River.........................................................................23 Water Quality in Wastewater Ditches and Pipes.............................................................24 Research Approach.............................................................................................................. ...25 2 LITERATURE REVIEW.......................................................................................................34 Introduction Ditches in the Landscape................................................................................34 Conceptual Basis for Pollutant Retention in Ditches.............................................................37 Characteristics of Ditch Ecosystems...............................................................................37 Contaminant Exposure and Retention Efficiency...........................................................38 Effect of Ditch Structure on Contaminant Retention.............................................................40 Vegetation in Ditches......................................................................................................40 Effect of Vegetation on Velocity Attenuation and Transient Storage.............................41 Influence of Hydrologic Re tention on Wa ter Quality.....................................................43 Sedimentation...........................................................................................................45 Retention processes influenced by interactions with the benthos............................46 Synthesis and research needs..................................................................................................52 Conclusions.................................................................................................................... .........56 3 INFLUENCE OF MACROPHYTES ON TRANSPORT CHARACTERISTICS AND NUTRIENT STORAGES.......................................................................................................66 Introduction................................................................................................................... ..........66 Methods........................................................................................................................ ..........67 Site Descriptions..............................................................................................................67 San Martin................................................................................................................67 Gustavson.................................................................................................................68 Koell.........................................................................................................................68 Frey...........................................................................................................................68

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7 Field Procedures for Water Chemistry, Flow and Channel Characterization.................69 Laboratory Analysis of Water Quality............................................................................69 Transport Characteristics of Ditches...............................................................................70 Tracer experiment field procedures..........................................................................71 Analysis of tracer data..............................................................................................71 Vegetation Sampling.......................................................................................................73 Collection of plants at sampling locations...............................................................73 Comparison of biomass and nutrient conten t of plants grown out of wastewater...74 Statistical Analyses..........................................................................................................74 Results........................................................................................................................ .............74 Water Quality of Wastewater Ditches.............................................................................74 General characteristics.............................................................................................74 Water quality during tracer experiments..................................................................76 Transport Characteristics.................................................................................................77 Plant Biomass and Nutrient Storages..............................................................................78 Discussion..................................................................................................................... ..........80 Conclusions.................................................................................................................... .........83 4 PHOSPHORUS RETENTION BY SORP TION WITH BENTHIC SEDIMENTS.............100 Introduction................................................................................................................... ........100 Methods........................................................................................................................ ........101 Experimental Sites.........................................................................................................101 Field Procedures............................................................................................................102 Laboratory Procedures...................................................................................................102 Water chemistry and sediment characterization.....................................................102 Phosphorus sorption index.....................................................................................102 Equilibrium phosphorus concentration..................................................................103 Phosphorus fractionation........................................................................................104 Statistical Analyses........................................................................................................105 Results........................................................................................................................ ...........106 Discussion..................................................................................................................... ........109 Effect of Seasonal Trends in Particle Size and Organic Matter Content on P Buffering Capacity.....................................................................................................109 Phosphorus Sorption Index............................................................................................110 Sediment Equilibrium P hosphorus Concentrations.......................................................113 Phosphorus Fractionation to Examin e P Status of Ditch Sediments.............................114 Management Implications.............................................................................................117 5 EVALUATION OF CHANNEL MODIFICATIONS TO EXISTING DITCHES FOR IMPROVED WATER QUALIT Y AND AEST HETICS.....................................................131 Introduction................................................................................................................... ........131 Methods........................................................................................................................ ........132 Government and Community Involvement...................................................................132 Modification Designs....................................................................................................133 Open-water flow design.........................................................................................134

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8 Subsurface flow design..........................................................................................135 Monitoring and Statistical Analyses..............................................................................136 Results........................................................................................................................ ...........137 Effects of Modifications on Resi dence Times and Water Quality................................137 Management Requirements...........................................................................................139 Government and Community Response........................................................................140 Discussion..................................................................................................................... ........142 Conclusions.................................................................................................................... .......146 6 SYNTHESIS AND FUTURE WORK.................................................................................164 APPENDIX A ENERGY CIRCUIT LANGUAGE......................................................................................167 B DITCH CHANNEL CROSS-SECTIONAL AND LONGITUDINAL PROFILES............168 C QA/QC RESULTS................................................................................................................178 D OTIS-P SIMULATION RESULTS......................................................................................179 E PHOSPHORUS FRACTIONATION DATA.......................................................................181 F PHOSPHORUS SORPTION INDEX (PSI) DATA.............................................................193 G EQUILIBRIUM PHOSPHORUS CO NCENTRATION (EPCO) DATA............................200 H INTERPRETIVE MATERIALS USED FO R COMMUNITY PARTICIPATION.............209 LIST OF REFERENCES.............................................................................................................211 BIOGRAPHICAL SKETCH.......................................................................................................225

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9 LIST OF TABLES Table page 2-1 Summary of environmental conditio ns and findings from published studies examining sediment, nutrient, BOD and pest icide retention in drainage ditches..............60 2-2 Summary of the retention effectiveness of ditches base d on the studies reported in Table 2-1...................................................................................................................... ......64 2-3 Possible reactants and conditi ons that limit the retention cap acity in drainage ditches....65 2-4 Dominant macrophytes reported by the 15 studies out of 22 in Table 2-1 that provided plant species information....................................................................................65 3-1 Dominant macrophyte species common to the study sites................................................87 3-2 Description of tracer experi ments conducted at each site..................................................90 3-3 Channel characteristics and mean water qua lity parameters for wastewater drainage ditch sites.................................................................................................................... .......91 3-4 Biomass and hydrologic conditions of each tr acer experiment and resulting transient storage zone parameter values...........................................................................................96 3-5 Estimated P and N uptake rates based on measured above and belowground biomass and plant tissue nutrient stor ages at each study site...........................................................97 3-6 Average P and N loads and estimated gross nutrient retention by aboveground biomass assimilation at each study site..............................................................................97 3-7 Mean molar ratios of carbon, nitrogen and phosphorus of aboveground (AG) and belowground (BG) plant tissues between sites..................................................................98 4-1 Site averages of particle size distribu tions and sediment pH for the wastewater ditches in October-November 2005 and June-August 2006............................................122 4-2 Comparison of P sorption measures determined in each site in years 2005 and 2006....122 4-3 Phosphorus sorption index (PSI) values fo r wastewater ditches calculated using three different sets of units to compare with values reported for ditches, streams and wetlands....................................................................................................................... ....129 5-1 Activities performed to promote local support and participation in channel modifications.................................................................................................................. ..147 5-2 Primary design elements used in open-wa ter flow design and their intended functions for improving water quality and aesthetics......................................................................150

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10 5-3 Primary design elements used in the a lternating subsurface flow/open water design and their intended functions for impr oving water quality and aesthetics........................153 5-4 Comparison of median travel times in Koell and Frey before and after channel modifications.................................................................................................................. ..155 C1 Summary of QA/QC results for nutrient analyses..............................................................178 E-1 Phosphorus fractionation data for Koell Site 1................................................................181 E-2 Phosphorus fractionation data for Koell Site 2................................................................182 E-3 Phosphorus fractionation data for Koell Site 3................................................................183 E-4 Phosphorus fractionation data for Koell Site 4................................................................184 E-5 Phosphorus fractionation data for San Martin Site 1.......................................................185 E-6 Phosphorus fractionation data for San Martin Site 2.......................................................186 E-7 Phosphorus fractionation data for San Martin Site 3.......................................................187 E-8 Phosphorus fractionation data for San Martin Site 4.......................................................188 E-9 Phosphorus fractionation da ta for Gustavson Site 1........................................................189 E-10 Phosphorus fractionation da ta for Gustavson Site 2........................................................190 E-11 Phosphorus fractionation da ta for Gustavson Site 3........................................................191 E-12 Phosphorus fractionation da ta for Gustavson Site 4........................................................192 F-1 Year 2005 PSI data for Koell sites...................................................................................193 F-2 Year 2005 PSI data for Frey sites....................................................................................194 F-3 Year 2005 PSI data for Gustavson sites...........................................................................195 F-4 Year 2005 PSI data for San Martin sites..........................................................................196 F-5 Year 2006 PSI data for Koell sites...................................................................................197 F-6 Year 2006 PSI data for Gustavson sites...........................................................................198 F-7 Year 2006 PSI data for San Martin sites..........................................................................199

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11 LIST OF FIGURES Figure page 1-1 Map of Peru showing the location of Oxapampa...............................................................26 1-2 View of the Chorobamba River valley and the city of Oxapampa....................................27 1-3 The urban center of Oxapampa and the Chorobamba River..............................................27 1-4 Simplified map of the urban center of Oxapampa, indicating the locations of wastewater ditches (solid lines) and underg round pipes (dashed lines) all of which drain to the Chorobamba River..........................................................................................28 1-5 Daily discharge in the Chorobamba Ri ver between December 2001 and January 2006...29 1-6 Summary of E. coli concentrations in the Chorobamba River (2005-2006).....................29 1-7 Summary of SRP concentrations in the Chorobamba River (2005-2006).........................30 1-8 Summary of dissolved inorganic nitrogen (sum of nitrate and ammonium) concentrations in the Ch orobamba River (2005-2006)......................................................30 1-9 Dilution-corrected E. coli concentrations in A) ditch e ffluents and B) pi pe effluents......31 1-10 Dilution-corrected carbonaceous biochemical oxygen demand (CBOD5) concentrations in A) ditch effl uents and B) pipe effluents................................................31 1-11 Dilution-corrected total suspended solids (T SS) concentrations in A) ditch effluents and B) pipe effluents..........................................................................................................32 1-12 Dilution-corrected soluble reactive phos phorus (SRP) concentrations in A) ditch effluents and B) pipe effluents...........................................................................................32 1-13 Dilution-corrected ammonium (NH4-N) concentrations in A) ditch effluents and B) pipe effluents................................................................................................................. .....33 2-1 Hypothetical relationships for A) contamin ant processing efficiency versus discharge proposed by Meyer and Likens (1979) a nd B) contaminant processing efficiency versus contaminant loading, as proposed by Odum et al. (1979)......................................58 2-2 Simplified systems diagram of the inte ractions between macr ophyte growth, kinetic energy (KE, a proxy for water velocity), sediment transport and contaminant retention in a generalized ditch system..............................................................................59 3-1 Systems diagram illustrating the influence of macrophyte harvesting on water quality in wastewater ditches.........................................................................................................85

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12 3-2 Map of Oxapampa showing the locations of the sampling sites in each of the study ditches........................................................................................................................ ........86 3-3 Ditch site San Martin looking upstream from sampling location 4 in A) November 2005 and B) June 2006......................................................................................................87 3-4 Gustavson site looking downstream at tim es of solute transport experiments in November 2005 at A) three weeks after plan ts were harvested and B) lower reach, location 4 before plant removal; C) in June 2006; and D) in July 2006, one week after vegetation removal a nd herbicide application...........................................................88 3-5 Koell site looking downstream from sa mpling location 1 at the time of tracer experiments A) in November 2005 and B) July 2006.......................................................89 3-6 Frey site at the marsh location at the time of biomass sampling in 2005..........................89 3-7 View of the Mariotte sip hon used for delivering the conser vative tracer at a constant flowrate....................................................................................................................... .......90 3-8 Box plots of San Martin and Koell at upstream and downstream sampling locations for A) E. coli and B) TSS...................................................................................................92 3-9 Box plots of San Martin and Koell at upstream and downstream sampling locations for A) TP, B) TN, C) SRP and D) NH4-N........................................................................93 3-10 Longitudinal water quality trends of E. coli TSS, SRP and NH4-N during tracer experiments performed in sites San Mart in, Koell and Gustavson under both high and low biomass conditions...............................................................................................94 3-11 Comparison of conservative tracer break through curves in Gustavson before and after removal of ditch vegetation in 2006..........................................................................95 3-12 Comparison of conservative tracer breakthrough curves in San Martin with low and high ditch vegetation biomass............................................................................................95 3-13 Comparison of breakthrough curves in Ko ell under different biomass conditions and after trash was remove d from the stream...........................................................................95 3-14 Comparison of leaf, stem and root/rhi zome tissue N and P content of macrophyte species found in wastewater ditches..................................................................................98 3-15 Comparison of common species in Oxapampa that are found growing in (wastewater) and out (control) of wastewater....................................................................99 4-1 Looking upstream from sampling site 4 in San Martin in A) November 2005 and B) July 2006...................................................................................................................... ....118

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13 4-2 Looking downstream from sampling site 1 in Gustavson in A) November 2005 and B) July 2006................................................................................................................... ..119 4-3 Looking downstream from sampling site 1 in Koell in A) November 2005 and B) July 2006...................................................................................................................... ....119 4-4 Sampling sites in Frey..................................................................................................... .120 4-5 Example of the approach used to dete rmine the sediment EPCo value and K, a measure of P buffering capacity, from P sorption isotherms...........................................120 4-6 Rainfall in Oxapampa from September 2005 September 2006....................................121 4-7 Examination of longitudinal variation in 2006 PSI values for each ditch from upstream (1) to downstream (4) sampling locations........................................................123 4-8 Longitudinal trends of percent organi c matter in 2006 for each ditch from upstream () to downstream () sampling locations................................................................123 4-9 Relationships between K (a measure of P buffer capacity determined from EPCo experiments) and percentage of sand, silt, clay and organic matter for all sites..............124 4-10 Contributions of biotic and abiotic so rption for phosphorus sorption indices (PSI) determined for sediments in 2005....................................................................................124 4-11 Comparison of ditch water SRP concen trations with sediment equilibrium phosphorus concentration (EPCo) experiments...............................................................125 4-12 Sediment EPCo values as a function of water column SRP concentrations grouped by whether sampling site acted as a P sink or source......................................................126 4-13 Relationship between P sorption index a nd sediment EPCo indicates that highest EPCo values (and thus greater potential to act as P sources) tend to be associated with lowest P buffering capacity......................................................................................126 4-14 Percentages of sediment pools of sorbed phosphorus for each site.................................127 4-15 Average total sediment P for San Marti n, Gustavson and Koell calculated from the sum of the P fracti ons for each ditch...............................................................................128 4-16 Average sediment TP for each sampli ng location at San Martin, Gustavson and Koell.......................................................................................................................... .......128 4-17 Gustavson sediments collected on filters after the NaOH extraction step indicate the presence of iron oxides....................................................................................................130 5-1 Neighborhood meetings were attended to assess whether the community supported implementation of modified ditches................................................................................148

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14 5-2 Residents participated in all aspect s of the channel modification activities....................149 5-3 Ornamental plants such calla lily were tr ansplanted from other sewage ditches to the modified ditches...............................................................................................................150 5-4 Schematic of the sedimentation basi n design used in channel modifications.................151 5-5 Views of the sediment ation basin in Koell......................................................................151 5-6 Sections of cascades and riffles created to improve oxygenation and increase pockets of transient storage...........................................................................................................152 5-7 Creation of gravel bars to increase channel sinuosity and filtration in Koell..................152 5-8 Schematic of the alternating subsurface flow/open-water design implemented in Frey.153 5-9 Views of the alternat ing subsurface flow/open water design used in Frey......................153 5-10 Cross-sectional view of the subsurface flow design........................................................154 5-11 Comparison of breakthrough curves before and after channel modifications in Koell...154 5-12 Comparison of breakthrough curves before and after channel modifications in Frey.....155 5-13 Comparison of E. coli concentrations at upstream, middle and downstream sampling locations in Koell before and after ditch modifications...................................................155 5-14 Comparison of TSS concentrations at upstream, middle and downstream sampling locations in Koell before and after ditch modifications...................................................156 5-15 Comparison of dissolved oxygen (DO) concentrations at upstream, middle and downstream sampling locations in Koell before and after di tch modifications...............156 5-16 Comparison of NH4-N concentrations at upstream, middle and downstream sampling locations in Koell before and after ditch modifications...................................................157 5-17 Comparison of NO3-N concentrations at upstream, middle and downstream sampling locations in Koell before and after ditch modifications...................................................157 5-18 Comparison of SRP concentrations at upstream, middle and downstream sampling locations in Koell before and after ditch modifications...................................................158 5-19 Comparison of E. coli concentrations at upstream, middle and downstream sampling locations in Frey before a nd after ditch modifications....................................................158 5-20 Comparison of TSS concentrations at upstream, middle and downstream sampling locations in Frey before a nd after ditch modifications....................................................159

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15 5-21 Comparison of dissolved oxygen (DO) concentrations at upstream, middle and downstream sampling locations in Frey before and after ditch modifications................159 5-22 Comparison of NO3-N concentrations at upstream, middle and downstream sampling locations in Frey before a nd after ditch modifications....................................................160 5-23 Comparison of NH4-N concentrations at upstream, middle and downstream sampling locations in Frey before a nd after ditch modifications....................................................160 5-24 Comparison of SRP concentrations at upstream, middle and downstream sampling locations in Frey before a nd after ditch modifications....................................................161 5-25 Simple tool consisting of a combined ne t and brush was provided to residents to facilitate ditch maintenance.............................................................................................161 5-26 Images of the open water flow design in Koell ditch four months after modifications...162 5-27 Images of the alternating subsurface fl ow/open water design in Frey four months after modifications...........................................................................................................162 5-28 Residents from Frey held a celebration to inaugurate the ditch and new bridge.............163 5-29 Other residents began attempts at modifying their ditches..............................................163 A-1 Description of the symbols used in energy circuit diagrams (from Odum 1994)...........167 B-1 San Martin cross-sectional pr ofiles for sections 1 through 4...........................................168 B-2 San Martin cross-sectional pr ofiles for sections 5 through 8...........................................169 B-3 San Martin longitudinal profile........................................................................................170 B-4 Gustavson cross-sectional prof iles for sections along Block 1........................................170 B-5 Gustavson cross-sectional prof iles for sections along Block 2........................................171 B-6 Gustavson cross-sectional prof iles for sections along Block 3........................................172 B-7 Gustavson cross-sectional prof iles for sections along Block 4........................................173 B-8 Gustavson longitudinal profiles for Blocks 1-4...............................................................174 B-9 Frey cross-sectional profiles............................................................................................175 B-10 Frey longitudinal profile................................................................................................. .175 B-11 Koell cross-sectional profiles...........................................................................................176 B-12 Koell longitudinal profile................................................................................................ .177

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16 D-1 OTIS-P modeling results for 2005 tracer experiment in San Martin...............................179 D-2 OTIS-P modeling results for 2006 tracer experiment in San Martin...............................179 D-3 OTIS-P modeling results for 2005 tracer experiment in Gustavson................................179 D-4 OTIS-P modeling results for 2006 tracer experiment in Gustavson................................180 D-5 OTIS-P modeling results for 2005 tracer experiment in Koell........................................180 D-6 OTIS-P modeling results for 2006 tracer experiment in Koell........................................180 G-1 2005 EPCO data for Koell site 1......................................................................................200 G-2 2005 EPCO data for Koell site 2......................................................................................200 G-3 2005 EPCO data for Koell site 3......................................................................................200 G-4 2005 EPCO data for Koell site 4......................................................................................201 G-5 2005 EPCO data for Gustavson site 1..............................................................................201 G-6 2005 EPCO data for Gustavson site 2..............................................................................201 G-7 2005 EPCO data for Gustavson site 3..............................................................................202 G-8 2005 EPCO data for Gustavson site 4..............................................................................202 G-9 2005 EPCO data for San Martin site 2.............................................................................202 G-10 2005 EPCO data for San Martin site 4.............................................................................203 G-11 2005 EPCO data for Frey site 1.......................................................................................203 G-12 2005 EPCO data for Frey site 2.......................................................................................203 G-13 2005 EPCO data for Frey site 3.......................................................................................204 G-14 2006 EPCO data for Koell site 1......................................................................................204 G-15 2006 EPCO data for Koell site 2......................................................................................204 G-16 2006 EPCO data for Koell site 3......................................................................................205 G-17 2006 EPCO data for Koell site 4......................................................................................205 G-18 2006 EPCO data for Gustavson site 1..............................................................................205 G-19 2006 EPCO data for Gustavson site 2..............................................................................206

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17 G-20 2006 EPCO data for Gustavson site 3..............................................................................206 G-21 2006 EPCO data for Gustavson site 4..............................................................................206 G-22 2006 EPCO data for San Martin site 1.............................................................................207 G-23 2006 EPCO data for San Martin site 2.............................................................................207 G-24 2006 EPCO data for San Martin site 3.............................................................................207 G-25 2006 EPCO data for San Martin site 4.............................................................................208 H-1 Pamphlet distributed to citizens to inform and receive their input on the implementation of modified ditches................................................................................209 H-2 Image of the sign painted and installed at each of the modified ditches, translated as Please do not throw trash: Pilot project fo r community participation of wastewater treatment in ditches........................................................................................................210

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18 Abstract of Dissertation Pres ented to the Graduate School of the University of Florida in Partial Fulfillment of the Requirements for the Degree of Doctor of Philosophy TREATMENT POTENTIAL OF WASTEWATER DRAINAGE DITCHES IN A RURAL COMMUNITY OF THE ANDEAN AMAZON Lynn Velisha Saunders August 2007 Chair: Mark T. Brown Major: Environmental Engineering Sciences Ditches are ubiquitous features in altered la ndscapes. Although ditche s are designed to be efficient at moving water across the landscape, th ere is growing evidence that ditches provide services beyond basic water trans port. The position of ditches in the landscape lends them great importance for controlling the timing and magnitude of terrestrially-deriv ed contaminant exports to downstream water bodies. My study examined the treatment function of di tches receiving domestic sewage effluents. The use of ditches for discharging wastewaters is common in regions where treatment systems are often non-existent. My study in vestigated the occurrence of in-stream contaminant retention in wastewater ditches, identified mechanisms responsible for contaminant retention, and implemented experimental modifications to th e design of existing ditches to test whether treatment performance was improved. The setting for the study was the town of Oxap ampa, Peru where approximately two-thirds of wastewater generated is discharged to th e Chorobamba River via ear then, vegetated ditches while the remainder is routed to the rive r via underground pipes. Dilution-corrected concentrations of E. coli biochemical oxygen demand, total suspended solids and soluble reactive phosphorus exported from ditches were found to be significantly lower than the same

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19 effluents discharged via underg round pipes, indicating that transport in ditches is not conservative. Conservative solute tracer experiments reveal ed the important influence of ditch vegetation on transport characteristics such as residence time and transient storage, which have positive implications for improved contaminant retent ion and decreased expor t of sediment and E. coli Deposition of silts and clays and organic matte r accumulation were shown to be important drivers of improved phosphate retention by pr omoting sorption with benthic sediments. Channel modifications to ditches were perfor med within a participatory framework that directly involved community members in pl anning, implementation and management. Two different modification approaches were tested: an open water flow design and an alternating subsurface flow/open water design. The latter de sign proved to be effective at sediment and E. coli removal and shows promise for treatment of water to irrigation standards. Reuse of treated ditch water should be promoted to prevent th e continued eutrophicat ion of the river.

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20 CHAPTER 1 INTRODUCTION Statement of the Problem The widespread deficiency of sanitation services in developi ng countries poses increasingly grave threats to human and ecosystem health. From the 49% of rural communities in Latin America that have access to sanitation services, only 14% provide some level of wastewater treatment (WHO and UNICEF 2000). Barriers to providing adequate sanitation services to developing countries include insuffici ent resources (both stru ctural and financial), inappropriate technological approaches, and lack of public awaren ess (WHO and UNICEF 2000). Innovative strategies that couple the use of existing landscape featur es and local materials with public education will yield more appropria te and sustainable wastewater management solutions that are better tailo red to a particular region. In the Andean Amazon region of Peru, untreated wastewaters are commonly discharged to open ditches that drain to nearby streams. Th e physical, chemical and biological features of these wastewater drainage ditc hes suggest that they may provide some level of pollutant attenuation prior to discharge to the river. Like small headwater streams, these unlined and vegetated ditches often possess hi gh surface-to-volume ratios as a result of their shallow water depths thus increasing contact with reactive benthic substrates. These interactions with the benthos enhance biogeochemical reactions resu lting in greater rates of nitrificationdenitrification, sorption to se diments and nutrient assimilation by plant and algal communities (Peterson et al. 2001). Plant communities co mmonly found growing in or along the banks of wastewater drainage di tches in Peru include Eichhornia crassipes (water hyacinth), Hydrocotyle spp. (pennywort), Lemna spp. (duckweed), Polygonum spp. (smartweed), Myriophyllum aquatica (parrot feather), Cynadon dactyon (bermudagrass), Hedychium coronarium (ginger

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21 lily), and numerous others (personal observation) that have been recognized for their treatment potential in constructed wetla nds (e.g. DeBusk et al. 1995, Brix 1997, Greenway 1997, Hume et al. 2002). This combination of physical, chemic al and biological features thus provide compelling evidence that material transport in wastewater ditches is not conservative. Study Overview and Objectives Drainage ditches are common interfaces betw een terrestrial and aquatic ecosystems in altered landscapes throughout the world. Yet despite this important position of ditches in the landscape, few studies have characterized the f unction of ditches to influence the fate and transport of potential contaminants such as path ogens, nutrients and sediments. Existing studies of water quality in ditches have taken place predominantly in agricultural settings (e.g. Cooper et al. 2002, Moore et al. 2005, Dunne et al. In Press ). My study evaluated the role of ditches in a different context than previous studies by examining the capacity of ditches to attenuate levels of domestic sewage contamination in a rural Peru vian community. The primary objectives of my study were the following: Investigate the occurrence of in-stream cont aminant retention in wastewater ditches Identify the mechanisms respons ible for contam inant retention Implement design modifications to existing di tches to enhance treatment performance. The end-goal of this research was to examine the feasibility of using modified ditches to mitigate water contamination, thereby offering rural communities a simple and cost-effective approach for improving wastewater management in a region that currently lacks the resources for more conventional treatment approaches. Studying the effects on water quality of the physical, chemical and biological features in wastewat er drainage ditches both before and after modifications may yield design and management im plications that are relevant not only for

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22 dealing with wastewater discha rges but also for other point and non-point sources of pollution draining into ditch systems. Study Setting The setting for this study was Oxapampa, Pe ru (10 35S, 75 24 W) located at an elevation of 1,814 m above sea level in the Andean Amazon (Figures 1-1 and 1-2). The population of Oxapampa is approximately 13,400 w ith 58% of inhabitants residing in the urban center (Figure 1-3) and the remaining 42% in the rural periphery of town (CEPID 2002). Oxapampa experiences minimal seasonal temperature fluctuations; average temperatures range between 15 and 29 C (Municipalid ad Provincial de Oxapampa 2003) Total annual precipitation is 1,400 mm per year, 80% of which falls between November and April (EPS Selva Central S.A. 2000). The economy of Oxapampa is predominately agrarian. Important agricultural products include coffee, chili peppers, avocados, plantain s and squash. Oxapampa is a major contributor of livestock production with ove r 4,910 head of cattle (CEPID 2002). These practices influence water quality as storm events flush sediments, an imal wastes, fertilizers and pesticides into drainage ditches that are ultimately discharged to the Chorobamba River. A local slaughterhouse facility is located on the rive r bank and discharges its wastes directly without treatment. The Chorobamba River is used extensively by local inhabitants for bathing, laundry, washing vehicles, recreation, fi shing, and in some cases human consumption. In 2000, 55% of Oxapampa residents had access to running water from the San Alberto watershed. Water treatment consists of a settl ing tank and chlorination. In 2004 this number increased to 67% (9,000 people). In 2004, approximately 1,170 hom es had running water and a sewage connection, 331 houses had running wa ter only, and 200 families had neither of these services (Arteaga, personal communication) Water usage rates are high (approximately 500 L per capita

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23 per day) as the result of wasteful consumption pr actices (e.g. leaving ta ps open) and leaking pipes (Arteaga, personal communication). The city of Oxapampa currently has no wastewat er treatment facility or household septic tanks. Sewage disposal consists of discharging both grey and black waters (including solids) into open, earthen and vegetated ditches. Approximately two-thirds of the wastewater generated in Oxapampa is discharged to open ditches that dr ain to the Chorobamba River. The remainder of the wastewaters is discharged directly to the Chorobamba River via networks of underground pipes. Both pipes and ditches function as co mbined sewers during storm events. The linear distance of open wastewater d itches and underground pipes is 18 km and 8 km, respectively (Arteaga, personal communication, Fi gure 1-4). The pipes and ditc hes discharge an average of 150 L s-1 to the Chorobamba River, representing 1 to 2.5% of the total rive r discharge during the dry season (Figure 1-5). While these percent contributions of wastewater to the river are small, they nevertheless have resulted in changes to the ecological inte grity of the river, notably reductions in fish populations and increasing growths of filamentous algae in shallo w channels, presumably from excess loads of phosphorus (Municipalidad Provincial de Oxapampa 2003). The impact of untreated wastes on the health of citizens is reflected by morbidity reports. Over 60% of children under the ag e of five years suffer from eith er acute or chronic diarrhea. Gastrointestinal diseases and parasites account for 37% of all hospital visits (CEPID 2002). Of the total population, approximately 40% are under the age of 15 and are thus at greater risk of waterborne diseases. Oxapampa has the only hospital facility within 1.5 hours. Water Quality in the Chorobamba River To examine the effect of wastewater discha rges on water qualit y, the Chorobamba River was sampled on multiple occasions in years 2005 and 2006 at one site upstream of the city and at

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24 three consecutive sites downstream of Oxapampa. Grab samples of water were collected at each site at several locations along a channel cros s-section and composited. Water samples were analyzed for E. coli concentrations, soluble reactive phosphorus (SRP), ammonium (NH4-N) and nitrate (NO3-N) following the procedures detailed in Chapter 3. Results of these analyses indicated that wastewater discharges from the city negatively impact water quality in the Chorobamba River by elevating pathogen concentr ations (Figure 1-6) and nutrient levels (Figures 1-7 and 1-8) downstream of the city. Th e health threat posed by fecal contamination is illustrated by E. coli concentrations that consistently exceeded limits established by the USEPA for recreational waters. Water Quality in Wastewater Ditches and Pipes To examine water quality characte ristics of wastewater efflue nts entering the river, water samples were collected from discharge point s of three underground pipes (Grau, Nueva Berna and Santa Domingo) and three ditches (Gustavs on, Koell and San Martin). Samples were analyzed for E. coli total suspended solids (TSS), 5day carbonaceous biochemical oxygen demand (CBOD5), SRP, NH4-N and NO3-N following analytical procedures described in Chapter 3. Parameter concentrations in ditches we re corrected for dilution by accounting for the percentage of total flow that originated from natural streams (approximately 45%, 37%, 20% and 5% in San Martin, Gustavson, Koell and Frey, re spectively). This correction was performed to distinguish possible in-stream tr eatment processes from dilution to provide a fair comparison against pipe water quality. Differences in wa ter quality between pipes and ditches were determined using t-tests on ln-transformed data. Concentrations of dilution-corrected E. coli TSS, CBOD5 and SRP were found to be significantly lower (p < 0.01) in ditch effluents compared to pipe effluents (Figures 1-9 through 1-13).

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25 There is little reason to believe that waste-st reams or water usage di ffers substantially in the town leading to the differences in water qua lity observed between ditches and pipes (Arteaga, personal communication). These findings thus imply that in-stream retention mechanisms influence concentrations of E. coli sediments, oxygen demand and nutrients during transport in ditches. This central hypothesis forms the basi s for the remainder of this study as described below. Research Approach A general assessment was made of the potential for treatment to occur in drainage ditches by synthesizing findings from the literature. The conceptual framework for understanding potential treatment mechanisms occurring in ditc hes was then applied to several wastewater drainage ditches in Oxapampa, Peru. The potential for treatment in wastewater ditches was first examined with respect to the function of vegeta tion for influencing transport characteristics, sedimentation processes and nutrient uptake. The significant differences in SRP concentrations observed between ditches and pipes prompted an ev aluation of the role of benthic sediments for mitigating phosphorus exports via sorption proce sses. Finally, utilizing insights gained from studying treatment influences of plants and sedi ments, channel modifications were implemented in an effort to amplify treatment processes.

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26 Figure 1-1. Map of Peru showing the location of Oxapampa, situated in the Andean Amazon at an elevation of 1,814 masl. Map created using ArcGIS 9.1 using ESRIs publicly available Shaded World Relief metadata (ESRI Data & Maps, 2004) Study site location: Oxapampa, Peru

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27 Figure 1-2. View of the Chorobamba Rive r valley and the city of Oxapampa Figure 1-3. The urban center of Ox apampa and the Chorobamba River

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28 Figure 1-4. Simplified map of the urban center of Oxapampa, indicating the locations of wastewater ditches (solid lines) and underg round pipes (dashed lines) all of which drain to the Chorobamba River

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29 0 50 100 150 200 18-Dec-0118-Dec-0218-Dec-0318-Dec-0418-Dec-05 DateDischarge (m3 s-1) Figure 1-5. Daily discharge in the Chorobamb a River between December 2001 and January 2006 (Note: missing data between September and December 2004; data source: AARS/ProPachitea) 0 400 800 1200 1234 Sampling siteE. coli concentration (CFU 100 mL-1) Figure 1-6. Summary of E. coli concentrations in the Chorob amba River (2005-2006), indicating that sampling sites downstream of Oxapampa (sampling sites 2-4) exceed maximum allowable E. coli concentrations for recreational wate rs. Points are average values (n = 6) and bars are SD. 235 CFU 100 mL-1 Water quality standard applied to recreational waters (EPA 2003)

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30 0 0.05 0.1 0.15 0.2 1234 Sampling siteSRP (mg P L-1) Figure 1-7. Summary of SRP c oncentrations in the Choroba mba River (2005-2006), showing that sites downstream of Oxapampa (s ampling sites 2-4) have elevated SRP concentrations relative to ups tream of the town (Site 1). Points are average values (n=6) and bars are SD. 0 0.1 0.2 0.3 0.4 0.5 1234 Sam p lin g siteDIN (NO3-N + NH4-N) (mg N L-1) Figure 1-8. Summary of dissolv ed inorganic nitrogen (sum of nitrate and ammonium) concentrations in the Chorobamba River (2005-2006), showing th at sites downstream of Oxapampa (sampling sites 2-4) have el evated concentrations relative to upstream of the town (Site 1). Points are aver age values (n=6) and bars are SD.

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31 G u s t a v s o n K o e ll S a n M a r t inDitch sites 0 10000 20000 30000 40000 50000E. coli (CFU 100 mL-1) G r a u N u e v a B e r n a S t a D o m in g oPipe sites Figure 1-9. Dilution-corrected E. coli concentrations in ditch effluents (A) compared to E. coli in pipe effluents (B) where filled circles represent median values(n=12), open circles are outliers, boxes delineate inter-quartile range (IQR), and whiskers are 1.5* IQR G u s t a v s o n K o e l l S a n M a r t i nDitch sites 0 50 100 150 200 250CBOD5 (mg L-1) G r a u N u e v a B e r n a S t a D o m i n g oPipe sites Figure 1-10. Dilution-corre cted carbonaceous biochemical oxygen demand (CBOD5) concentrations in ditch effluents (A) compared to CBOD5 in pipe effluents (B) where filled circles represent median values (n=5 ), open circles are outliers, boxes delineate inter-quartile range (IQR), and whiskers are 1.5* IQR A B A B

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32 G u s t a v s o n K o e l l S a n M a r t i nDitch sites 0 50 100 150 200 250TSS (mg L-1) G r a u N u e v a B e r n a S t a D o m i n g oPi p e sites Figure 1-11. Dilution-corrected tota l suspended solids (TSS) concentr ations in ditch effluents (A) compared to TSS in pipe effluents (B) wher e filled circles represent median values (n=12), open circles are outli ers, boxes delineate inte r-quartile range (IQR), and whiskers are 1.5* IQR G u s t a v s o n K o e l l S a n M a r t i nDitch sites 0 2 4 6SRP (mg L-1) G r a u N u e v a B e r n a S t a D o m i n g oPipe sites Figure 1-12. Dilution-corrected soluble reactive phosphor us (SRP) concentrations in ditch effluents (A) compared to SRP in pipe e ffluents (B) where filled circles represent median values (n=12), open circles are out liers, boxes delineate inter-quartile range (IQR), and whiskers are 1.5* IQR A B A B

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33 G u s t a v s o n K o e l l S a n M a r t i nDitch sites 0 10 20 30NH4-N (mg L-1) G r a u N u e v a B e r n a S t a D o m i n g oPipe sites Figure 1-13. Dilution-corrected ammonium (NH4-N) concentrations in ditch effluents (A) compared to NH4-N in pipe effluents (B) where fill ed circles represent median values (n=12), open circles are outli ers, boxes delineate inte r-quartile range (IQR), and whiskers are 1.5* IQR B A

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34 CHAPTER 2 LITERATURE REVIEW Introduction Ditches in the Landscape Ditches are ubiquitous features in altered landscapes. Their presence grows in proportion to the expansion of anthropogenic land use and as a consequence, ditches have become an inevitable component of ecosystems around the world. In developed nations, their primary purpose is to route water for irri gation or to improve drainage from agricultural fields and roadways. Drainage ditches are equally pervas ive in many parts of the developing world; however, in the absence of wastew ater treatment facilities, draina ge ditches often serve as the primary transport conduits for untreated human sewage. Ditches are designed and managed to be effici ent at moving water acro ss the landscape and as a consequence, are commonly perceived as li ttle more than conduits for routing water and materials. Until recently, little interest had been given to understanding and promoting potential ancillary functions of drainage ditches. Howeve r, there is growing evidence that ditches provide services beyond basic water trans port. Studies have shown that so me ditches serve as important habitat for maintaining species richness (Armitage et al. 2003), function as wildlife corridors (Mauritzen et al. 1999, Mazerolle 2005) and provide recharge of surficial aquifers (Fernald and Guldan 2006). The position of ditches in the landscape suggests their unique importance as interface ecosystems (Odum and Odum 2003), connecting te rrestrial and aquatic environments and controlling the timing and magnitude of terrestria lly-derived contaminant exports to downstream water bodies. Surprisingly little at tention has been given to inve stigating the treatment function of drainage ditches despite th eir downstream connection with r eceiving water bodies. While there are studies that point to di scharges from drainage ditches as major sources of water quality

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35 contamination (e.g. Hunt et al. 1999, Fletcher et al. 2004, Cahoon et al. 2006) there is also growing evidence that some ditches mitigate ne gative impacts to downstream water quality. Ditch networks may be viewed as low-order st ream systems that are often responsible for draining a disproportionate area of many watershe ds, especially in ag ricultural landscapes (Alexander et al. 2000). The a ssociation between land-use prac tices, diffuse pollution and the continued deterioration of water quality is well recognized and has been discussed in great detail in the literature (see Howarth et al. 1996). Despite the large numb er of studies addressing runoff issues, mitigating the effects of diffuse polluti on remains a somewhat elusive goal in practice (Sharpley and Tunney 2000, McDowell et al. 2004). A number of best management practices (BMPs) such as buffer strips have been recomme nded to reduce runoff loads to ditches. It makes good sense to couple these load-reducing prac tices with ditch design and management approaches that amplify the abilities of draina ge systems to retain pollutants (Cooper et al. 2004, Dabney et al. 2006). The objective of this review is to prov ide a general, conceptual framework for understanding potential retention mechanisms in ditches and to synthesize findings from the literature that have reported on the treatment capab ilities of drainage ditc h systems. The purpose of this exercise is to gain an understanding of the degree to wh ich ditches have been shown to mitigate contaminant exports to receiving bodies and to examine the conditions under which retention has been shown to occur. The term re tention is used here to mean the difference between imports and exports in a system resu lting from transformation and storage processes which ultimately affect the timing and magnit ude of downstream exports. This review also highlights where future research efforts s hould be focused to expand upon our current understanding of the treatment f unction of drainage ditches.

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36 The studies covered in this re view are strongly biased toward s agricultural drainage ditches because this is the setting in which the vast majority of the studies focusing on contaminant processing in ditches have been conducted. Regard less of this bias towards agricultural ditches, the lessons learned from examining the physical and biogeochemical processes taking place in agricultural systems may be transfer red and applied to other settings such as roadside ditches and ditches associated with aquacultu re, forestry, industrial practic es and even with routing of wastewater effluents. The exis tence of open channels for rou ting untreated sewage is poorly documented. However, in regions such as th e Andean Amazon of Peru, these ditches are currently the norm for handling wastewaters. Stud ies of the treatment function of ditches thus yield important implications for improved wate r management across diverse land-use types and across the globe. For purposes of clarification, ditches will be referred to as long, linear water conveyance features that are earthen, unlin ed and vegetated (unless managed otherwise). Of the 22 studies identified in this review that examined c ontaminant retention in ditches, six employed approaches developed originally to understand material cycling in stream ecosystems and referred to their study systems as streams rather than ditches (Macrae et al. 2003, Royer et al. 2004, Schaller et al. 2004, Bernot et al. 2006, Ensign et al. 2006, Gucker and Pusch 2006). Ditch characteristics are as diverse as the settings in which they ar e found (Moore et al. 2005), thus complicating their strict definition. However, if the systems are channelized and convey drainage waters from altered landscapes, such as agricultural lands, they are referre d to here as ditches. The take-home message from this review is that the consideration of drainage ditches in the landscape should not be limited strictly to thei r conveyance function of water and materials but

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37 should more broadly recognize their position in th e landscape as important interface ecosystems between the terrestrial environm ent and downstream water bodies. Conceptual Basis for Pollutant Retention in Ditches Characteristics of Ditch Ecosystems It is not surprising that some ditches possess the capacity for significantly reducing concentrations of contaminants such as pest icides, nutrients and pathogens. The treatment function of ditches arises from their physical, biological and chemical characteristics, many of which are shared with other sy stems recognized for their impor tance in pollutant retention. Many earthen and vegetated ditches share important physical ch aracteristics with headwater streams. For example, many ditches ha ve relatively shallow water depths and thus tend to have high surface area to volume ratios, which may enha nce biogeochemical reactions (Alexander et al. 2000) resulting in increased rates of denitrif ication, sorption to reactive substrates and assimilation by plant and algal co mmunities. High nutrient loads and colonization by emergent and submerged aquatic plants often yield a system whose bi ological features are reminiscent of nutrient-rich treat ment wetlands. Accumulation of plant litter and deposition of fine organic-rich materials creates a benthic environment with redox conditions that are more similar to a wetland than to a stream system. Ho wever, ditches are lotic systems and experience pulses of water and materials that may frequently disturb ecosystem struct ure thus re-setting or maintaining the system in early stages of su ccession. This blend of physical, biological, and chemical attributes creates a unique ecosystem consisting of hydrogeomorphic features common to many headwater streams and chemical and biol ogical characteristics reminiscent of treatment wetlands.

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38 Contaminant Exposure and Retention Efficiency The physical, chemical and biological charact eristics expressed by a ditch reflect the conditions to which it is exposed. Ditch exposure to contaminants is sometimes direct such as during the application of herbicides to control vegetation. Other time s, it is incidental, as in the case of drift during aerial applic ations of fertilizers or pesticid es. Most often, however, exposure in ditches occurs when there is water movement, particularly dur ing storm events that generate runoff draining to ditches or from subsurface tile drains that remove excess water to improve agricultural field conditi ons. For wastewater drainage ditche s, exposure is often greatest during hours of peak water use. The patte rns of water flow within a ditch thus dictate the regularity and the intensity of the contaminant exposure. The exposure regime of potential contaminants in a ditch governs its treatment efficiency. The exposure regime in a ditch consists of two important factors that function as controls on pollutant retention: discharge and contaminan t loading. Discharge gove rns the residence time within the ditch. Meyer and Li kens (1979) proposed that discha rge controls whether a stream system is operating in a processing mode or a throughput mode. In the processing mode, the system is efficient at storing, utilizing and transforming potential environmental pollu tants. This mode tends to occur at lower flows when cont act time is greatest, thereby promoting more physical and biogeochemical interactions during tr ansport. For higher rates of flow, the system switches to a throughput mode, whereby potenti al contaminant retention mechanisms are bypassed due to low residence times in the channel. Contaminant loading is the second important control on the efficiency of pollutant retention. The effect of loading on treatment effi ciency can be conceptualized using the subsidystress gradient theory of Odum et al. (1979). A ditch ecosystem may initially become stimulated by the availability of macronutrients or organic ma terials until a certain th reshold is reached at

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39 which the loading exceeds the processing capacity of the system (see Figure 2-1). At this point, loadings no longer serve to subsidize system meta bolism but instead become a stressor to system and reduce its processing efficiency. In this sense, contaminant loadi ng vis--vis the subsidystress gradient theory may embody a switch between a processing and throughput mode similar to that hypothesized for discharge by Meyer and Likens (1979). The subsidy-stress gradient of Odum et al (1979) also consider s inputs of toxic substances (not depicted in Figure 2-1), which in ditch ecosystems could represent toxicity limits of pesticides and metals in plan ts and other organisms. In the cas e of either a usable or a toxic input to the system, once the loading has exceeded the processing capability of the system, the system is no longer efficient at pollutant rete ntion and instead switches to a throughput mode, thus exporting environmental pollu tants to downstream water-bodies (Haggard et al. 2001, Marti et al. 2004, Gucker and Pusch 2006). The obvious implication behind these two hypothese s is that ditches will be of limited use for reducing pollutant loads to downstream wa ter-bodies if they are overwhelmed by high discharges and contaminant loadings. It woul d be insightful to develop a general understanding of the discharge and loading le vels at which this switch between operating in a processing versus a throughput mode tends to occur in order to pred ict the treatment reliability of a system under different conditions. However, such threshol ds may be very site-specific and thus less appropriate for extrapolating to other drainage ditch systems. Perhaps a more widely-reaching contribution by scientists would be to investig ate approaches to bett er modify and manage ditches with the objective of maintaining or extending the effective operating range of the processing mode under elevated discharge or load ing scenarios. Modifications and management

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40 affect and re-organize system structure. Therefor e, for such approaches to be effective, it is imperative to understand how the structure of ditches influences its treatment function. Effect of Ditch Structure on Contaminant Retention In a very broad sense, ditch ecosystems can be viewed as consisting of two dominant physical/biological components: ve getation and benthic substrates. These two components affect and are affected by flows of water, material s and energy thereby creat ing a dynamic three-way relationship between water flow, sediment trans port and plant growth (Figure 2-3, Clarke 2002). It is within this dynamic relationship that th e potential for contaminant retention exists via transformations and storages governed by plants a nd sediments that control contaminant fate and transport in ditch ecosystems. In many ditch ecosystems, vegetation and ofte n to a lesser degree, sediments are removed to improve drainage and prevent the risk of flooding. This review emphasizes the role of vegetation in ditches, not only because this is the component most easily controlled by managers, but as depicted in Fig. 2-2, the removal of vegeta tion typically results in the subsequent export of sediments. Vegetation in Ditches Ditches are often unshaded, tend to be relativ ely shallow and often experience low water velocities except perhaps during pulse events such as storms. These characteristics, often in conjunction with high concentratio ns of bioavailable nutrients, support abunda nt growth of instream vegetation. In-stream vegetation is a di stinctive feature of ma ny drainage ditches; therefore, understanding how vege tation influences the fate and transport of contaminants is particularly germane in ditch ecosystems. While much attention has been paid to the tr eatment function of vegetation in constructed wetland ecosystems (Brix 1997, Richardson and Qian 1999, Dierberg et al. 2002, Karathanasis et

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41 al. 2003, Collins et al. 2005), littoral and riparian zones (Tabacchi et al. 1998, Spruill 2000, Jia et al. 2006), and other wetland ecosystems (Odum et al. 2000), relatively few studies have specifically examined this function within lotic ecosystems (Clarke 2002). Perhaps this is derived from a perception of limited plant-water c ontact time in streams and ditches compared to very low-velocity systems. Another explanation that vegetation is often overlooked for uptake of organic contaminants such as pesticides in ditc hes is that partitioning is often assumed to take place only between the water and sediments (Coope r et al. 2004). However, increasing numbers of studies are examining the eff ects of vegetation in flowing-wa ter ecosystems and reporting that the presence of vegetation yields significant implications for c ontaminant fate and transport (e.g. Wilcock et al. 1999, Champion and Tanner 2000, Clar ke 2002, Salehin et al. 2003, Cooper et al. 2004, Scholz 2005, Dabney et al. 2006). The potential treatment processes promoted by the presence of ditch vegetation are varied and interrelated. Many studies have shown the importance of direct uptake by plants for removal of potential environmental contaminants such as nutrients (e.g. Brix and Schierup 1989, DeBusk et al. 1995, DeBusk et al. 2001). However, rele ase of assimilated nutrients during plant decomposition often makes this re tention process ineffective unl ess plants are periodically harvested (Reddy et al. 1995, DeBusk et al. 2001). Often, the role of vegetation for contaminant retention in aquatic systems is less by their direct uptake and more by the conditions they create which in turn promote retention (Brix 1997, Sobolewski 1999, Kim and Geary 2001, Schulz et al. 2003). Arguably the most importa nt contribution of in-stream ve getation for retention is the attenuation of water velocity. Effect of Vegetation on Velocity Attenuation and Transient Storage Stands of emergent and submerged vegetati on increase channel roughness (Wilcock et al. 1999, Champion and Tanner 2000, Madsen et al. 2001) and provide structural complexity

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42 (Clarke 2002, Salehin et al. 2003, Grimm et al. 2005) that serve to attenuate water velocity and increase solute residence time. Flow resist ance by transport thr ough and around stands of vegetation results in flow becomi ng generally deeper and slower. Patches of vegetation increase channel heterogeneity, creating e ddies and pockets of slower moving water that impede solute transport. This delay of downstream advective transport (or hydrologic retention) due to time spent in stream channel quiescent zones is term ed in-channel transient storage (Bencala and Walters 1983). The impedance of flow due to stands of in-stream vegetation may also promote the subsurface component of transient storage, ca lled hyporheic exchange. Hyporheic exchange is the exchange of surface water with shallow groundwater (Harvey and Wagner 2000) due to increased hydraulic pressure which drives water into and through porous channel beds, a process termed advective pumping by Wrman et al. (2 002). However, the importance of hyporheic exchange in many drainage ditches may be severe ly limited by the presence of fine sediments that clog interstices and impede the transfer of water across the hyporheic zone (Grimaldi and Chaplot 2000, Hancock 2002). Hyporheic zones may also have reduced importance in ditches as a result of channel straightening which disconnect s the stream channel from parafluvial zones, reduces channel complexity and increas es water velocities (Hancock 2002). The capacity of in-stream vegetation to attenua te flow velocities and promote transient storage is predicated on macrophyte morphology a nd growth characteris tics (Clarke 2002, Dodds and Biggs 2002, Schulz et al. 2003). For example, submerged macrophytes with finely divided leaves that tend to grow in dense stands will ex ert greater influence over flow and sedimentation patterns than those in more open stands with broader or ribbon-like l eaves (Sand-Jensen and Mebus 1996, Clarke 2002). While there is no single variable that can be used to determine

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43 velocity attenuation of different community types under varying hydrologic conditions, Dodds and Biggs (2002) found that the mass density of macrophytes and periphyton served as an adequate proxy, explaining 81% of the variance in velocity attenuation in their study. However, it is necessary to note that the presence of m acrophytes in lotic systems does not always ensure greater velocity a ttenuation. The velocity attenuation cap acity of macrophytes is linked to the spatial distribution of their stands as illustrated by Wilcock et al. (2004) who found that emergent plants colonizing stream margins reduced the cha nnel cross-sectional area, resulting in increased water velocities. In contrast, submerged plants growing within the channel served to impede flow, thereby reducing velocitie s and increasing water depths. Influence of Hydrologic Retention on Water Quality Hydrologic retention, the averag e time spent by solutes in tr ansient storage zones per unit downstream distance (Harvey et al. 2003), increase s solute contact (in time and space) with biogeochemically reactive substrates and has be en linked to significant changes of downstream water chemistry (Mulholland et al. 1997, Harv ey and Fuller 1998, Hill et al. 1998). The importance of in-stream obstructions such as vegetation and woody-debris dams for increasing channel residence time and promoting contamin ant retention has been documented in recent years (Valett et al. 2002, Wilcoc k et al. 2004, Ensign et al. 2006). Wilcock et al. (2004) showed that ammonium retention in f our agricultural stream s with varying cove rage of macrophytes could be approximated by the stream length ( X ) and velocity ( v ) such that % retention X / v The link between transient storage and nutrien t retention in streams is commonly made using nutrient spiraling methods (Newbold et al. 1981, Stream Solute Workshop 1990). In this approach, solute injection expe riments are performed to fit a 1-D transport model to observed solute breakthrough curves (BTCs). Several transport models have been used in this approach and all incorporate transient storage parameters that relate exchange ra tes and size of storage

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44 zones (see Stream Solute Workshop 1990, Runkel 1998). Conservative solute BTCs allow the estimation of transient storage and other hydrau lic parameters while changes in the dilutioncorrected concentrations of r eactive solutes with downstream distance provide evidence of retention processes occurring during transport. Spiraling metrics, such as uptake length, uptake rate and uptake velocity, have been used as prox ies for the self-purifying capacity of stream ecosystems (Marti et al. 2004) and are used to compare nutrient retention over time, space and between different systems. Nutrient spiraling theory has its roots in understanding and comparing nutrient dynamics in pristine stream systems. However, the methodol ogy has been applied in recent years to eutrophic systems as well (Marti et al. 2004, Royer et al. 2004, Haggard et al. 2005, Bernot et al. 2006, Ensign et al. 2006, Gucker and Pusch 2006). The nutrient spiraling con cept by itself does not reveal the mechanisms responsible for retention but instead provides a w hole-reach measure of retention with respect to the so lute in question. This approach has allowed nutrient retention in drainage ditch systems to be compared w ith that in less-altered stream systems. Ensign et al. (2006) used this approach to compare spiraling metrics on various dates to create an empirical model to predict uptake in four drainage systems under various discharge regimes. The authors reported spiraling metrics that fell within the same ranges reported for headwater streams (see Ensign and Doyle 2006) a nd predicted relatively high retention (65 and 37% for ammonium and phosphate, respectively) dur ing storm flows when re tention tends to be lowest. These encouraging results contrast with results from other ditch spiraling studies which concluded that despite high nutri ent uptake rates and uptake veloci ties, overall nut rient retention in ditches tended to be low relative to high c ontaminant loads (Table 2-1, Macrae et al. 2003, Royer et al. 2004, Gucker and Pusch 2006).

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45 Low nutrient retention in ditche s is often hypothesized to be the result of the lack of channel structural complexity (e.g. Grimm et al 2005, Bernot et al. 2006) that limits retention times in transient storage zones and availabi lity of reactive surface areas. To explicitly investigate the relationship between channel comp lexity, transient storage and nutrient retention, Ensign and Doyle (2005) performed channel manipula tions in an agricultura l drainage canal and a channelized blackwater stream in North Carolin a, USA. Transient storage was measured under 3 different channel conditions: unaltered, remova l of all vegetation and coarse woody debris (CWD) and the addition of PVC baffles. Remova l of vegetation and CWD reduced transient storage zones by 61% and 43%, while baffles increased storage zones by 227% and 119%. Phosphate and ammonium retention were quantified using the uptake velocity ( Vf) spiraling metric, which represents the re lative demand for the solute by the benthos (Bernot et al. 2006). After vegetation removal, Vf values for phosphate and ammonium decreased by 38% and 88% and with baffles in place they increased by approximately 3000% (from -1.5 to 53 mm min-1) and 143%. The Ensign and Doyle ( 2005) study supports the importance of maintaining structural complexity for improved drainage ditch treatm ent function. The following sections examine other study results to provide an overview of th e effects of sedimenta tion and contact with reactive substrates in ditches for contaminant processing and retention. Sedimentation Suspended sediments and BOD removal. Sedimentation is an important retention mechanism in ditches as many potential contaminants are in particulate form or are associated with sediment surfaces. Lecce et al. (2006) report ed that vegetated ditches in North Carolina functioned as efficient sediment traps, reduci ng sediment exports from the basin by 2,978 Mg yr1 or 3.83 Mg ha-1yr-1. Hargreaves (2005) reported that approximately 97%, 67% and 69% of high loads of TSS, TN and TP were removed within a 150-200 m distance along a vegetated ditch in

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46 Mississippi, USA. Scholtz and Trepel (2004) re ported decreases in BOD from approximately 15 mg L-1 to less than 1 mg L-1 within 90 m. However, between 90 and 150 m BOD values tended to fluctuate with final co ncentrations around 6.0 mg L-1 on average. Unless particulate matter is buried and incorporated into the sediment (Saunders and Kalff 2001), the process of sedimentation unfortunately does not ensure permanent contaminan t removal. Erosion can resuspend and transport particulate and sediment -bound contaminants to downstream water-bodies. The presence of in-stream vege tation reduces the potential for re-suspension (Madsen et al. 2001) but during senescence this function de creases markedly (Schulz et al. 2003). Besides removing suspended materials from th e water column, sedimentation can promote other retention processes by allowing depositi on and accumulation of orga nic matter within and downstream of plant stands (Sc hulz et al. 2003). The accumulation of organic matter in addition to the presence of root systems increases the porosity and permeability of benthic sediments and therefore can affect sediment pore water fluxe s (Salehin et al. 2003). Sedimentation in the presence of vegetation thus results in significan t changes to the chemistry of benthic sediments leading to important implicati ons for contaminant retention. Retention processes influenced by interactions with the benthos Sedimentation and deposition of seston and plant litter occurring in the presence of macrophytes or woody debris dams create benthic environments that set in motion important processes such as deni trification and sorption. Nitrate removal. Denitrification is often the primary mechanism for nitrogen retention in aquatic systems (Howarth et al 1996, Alexander et al. 2000) and is of particular importance because it represents a permanent loss from th e system. Saunders and Kalff (2001) in their review of nitrogen retention of wetlands, lakes and rivers found that from the sites studied, rivers provided the least amount of N rete ntion and wetlands the greatest. Hence, rates of retention due

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47 to denitrification in vegetated di tches are likely to fall somewhere in the middle. Denitrification rates in streams are often limited by contact time, zones of anoxia and the availability of organic carbon or nitrate (Hedin et al. 1998). Macrophyte presence in ditc hes provides extensive surface areas for microbial growth, increases residence time and promotes the accumulation of organic matter thus providing the necessary carbon source. The importance of promoting this retention mechanism is highlighted in agricultural ditc hes because TN loading to ditches is often dominated by nitrate loads that ha ve bypassed riparian cont rols due to the use of subsurface tiles. Reports concluding on the effectiv eness of denitrification in d itches are varied and appear in some cases to be largely a function of di scharge. Scholtz and Tr epel (2004) reported high losses of nitrate (~79%), presumably due to denitrification, along a 150 m reach in Germany. Royer et al. (2004) and Schaller et al. (2004) examined denitrif ication under several discharge regimes in Illinois and conclude d that although denitrif ication rates tended to be high, retention was insignificant since the highest nitrate load s coincided with highest flowrates Nitrate concentrations for all 3 studies were within the same ranges (13-18 mg L-1); however flowrates reported by Scholtz and Trepel (2004) were ~0.5 L s-1 whereas those for Royer et al. (2004) and Schaller et al. (2004) were as high as 1,840 and 1,370 L s-1. High nitrate loads in conjunction with high flows switches the sy stem to a throughput mode (a s noted by Royer et al. 2004) rendering in-stream processes incapable of mitig ating downstream exports. Bernot et al. (2006) also reported that nitrate demand in six ditches in Indiana and Michigan appeared to be saturated; however no significant correlations were found betw een retention and nitrate concentration or discharge. The results from these studies sugge st that while the poten tial for high nitrate retention exists in ditches, it is likely to be severely compro mised under high flowrates that limit residence time, especially in systems with extremely limited channel complexity.

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48 Hyporheic exchange is often considered to be of minor signifi cance for improving water quality in ditches due to channel straighteni ng and clogging of interstices by fine sediments. However, it has been shown that nitrate can be depleted by denitrification within the first few centimeters (Hill et al. 1998, Revsbech et al. 2005) making even small vertical fluxes potentially significant. No studies were iden tified that examined the role of hyporheic exchange specifically in drainage ditches; however, Le Febvre et al. (2004) reported hi gh rates of denitrification in a straightened river reach in Fran ce. Denitrification rates were co rrelated to the presence of high total organic matter and fine sediments resulti ng from direct sediment exports from surrounding fields. Ammonium retention. High concentrations of amm onium are common in ditches receiving waste inputs of manure or with high ra tes of decaying plant material (e.g. Dukes and Evans 2006). Retention of ammonium is an impo rtant goal in many ditc h systems because of ammonia toxicity to aquatic organisms. Macr ophytes translocate oxygen to their root systems providing aerobic microsites that promote mineralization of or ganic N and nitr ification of ammonium thus setting the stage for coupled nitr ification-denitrification interactions (Reddy et al. 1989). Ammonium may also be utilized by al gae and heterotrophic microbes that make up reactive biofilms on sediments and other benthic s ubstrates. Physical adsorption of ammonium to sediment surfaces is most important in pristine systems with low background ammonium concentrations (Triska et al. 1994). For ditches, the more likely pathway for ammoni um retention is nitrification-denitrification and temporary storage by macrophytes and biofil ms. Findings from Gucker and Pusch (2006) support this hypothesis by reporting that nitrif ication represented th e dominant ammonium retention mechanism in channeli zed, eutrophic systems in Germa ny. Scholtz and Trepel (2004)

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49 reported no net change in ammonium concentra tion along ditches in northern Germany despite increasing dissolved oxygen levels with downstrea m distance. One possible explanation is that decay of plant material contri buted ammonium to the water co lumn; this hypothesis is supported by fluctuations in BOD. Similar to the results re ported for nitrate enriched systems by Royer et al. (2004), Schaller et al. (2004), Inwood et al. (2005) and Grimm et al. (2005), Gucker and Pusch (2006) found that while abso lute uptake rates of ammonium and nitrate were higher in the eutrophic systems than in pristine streams, the percent retention was minimal due to comparatively much greater N loads. Phosphorus retention. Retention of phosphorus (P) presen ts another formidable challenge in drainage ditch systems. While biological assimilation is an important retention mechanism, it is short-term as P stored in tissues eventually becomes re-mineralized and re-released to the water column (DeBusk et al. 2001). Dierberg et al. (2002) reported that high water column pH occurring during photosynthesis by submerged aquatic vegetation resulted in co-precipitation of calcium and phosphorus that rem oved 50-79% of P in constructed wetlands for agricultural runoff. However, in most stream and wetland systems, sediment sorption is the dominant retention process (Reddy et al. 1995). The capacity of benthic sediments to temporarily retain dissolved P from the water column by sorption is governed by physiochemical proper ties of the sediment such as soil composition (mineralogy and texture), pH, organic matter content and redox condi tions (Reddy et al. 1995, Sallade and Sims 1997a, Reddy et al. 1999, N guyen and Sukias 2002). Biotic processes occurring on sediment surfaces may enhance sorption capacity (Haggard et al. 1999). Nguyen and Sukias (2002) determined fractions of sorbed P in 26 agricultural drainage ditches in New Zealand to determine the extent of and mech anisms responsible for sediment P sorption.

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50 Although 64-68% of sediment surfaces were found to be saturated, the ditches still demonstrated high P retention capacity by removing 44-84% of the 5,000 mg P kg-1 added during P buffer capacity experiments. The authors speculated that large presence of fine particles (63-67%) and high organic carbon concentrations (3.8-14.1%) likely result ed in sorption of Al and Fe which could have formed humic-Fe/Al complexes that are effective at promoting even greater P sorption. From the P fractionation analysis Ngyugen and Sukias (2002) found that approximately half of P sorbed represented a te mporary storage as those fractions were loosely bound and hence could be re-released to the water column. However, some of the ditches were also long-term sinks as 6% to 39% of to tal sorbed P was stored as residual P. The concentration gradient between water co lumn soluble reactive P (SRP) and sorbed sediment P controls whether sediments release or remove phosphorus from the water column. The aqueous concentration at which P is neither released nor sorbed from sediments is termed the equilibrium phosphorus concen tration (EPCo). This value is used as a benchmark to determine whether sediments tend to function as a sink or a source of SRP to the water column. If SRP is greater than EPCo, sediments act as a temporary P sink and vi ce-versa. Smith et al. (2005) reported that the addition of alum to drainage ditches reduced EPCo values (but not the P buffering capacity) due to phosphorus precipitation of with aluminum. Th is precipitate is generally more robust than that of iron because under reducing condition ferric iron is reduced and releases P to the water column. In a sepa rate study Smith et al. (2006) found that alum treatments were not effective at reducing ex changeable P and EPCo and hypothesized that a storm event mobilized alum-treated sediments a nd deposited new sediments. While the addition of alum is common to water treatment facilities and in controlling P concentrations in lakes, the practice may not prove economically viab le to many farmers or land managers.

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51 Pesticides and heavy metals retention While the organic matter content of sediments has been shown to promote greater sorption of P, it is also responsible for promoting retention of heavy metals (Gambrell 1994, Debusk et al. 1996, Hammer and Keller 2002) and pesticides (Margoum et al. 2006). Retention of metals and pest icides have also been linked to direct uptake by macrophytes and sorption to macrophyte-associat ed substrates (e.g. Moore et al. 2001, Collins et al. 2005). While no studies on heavy metal retention in vegeta ted ditches were found, grassed swales are nevertheless common BMPs for retain ing stormwater and have been shown to be moderately effective at reduc ing sediment-bound heavy metals (Backstrom 2003, Zanders 2005). Few studies have explicitly st udied pesticide retention in dr ainage ditches; however, the studies that do exist illustrate that vegetated ditches show great promise as viable BMPs for mitigation of pesticide exports. Bennett et al. (2005) found that ditch lengths of 120 and 280 m were sufficient to reduce two pyrethroid pes ticides (bifenthrin and lamda-cyhalothrin) concentrations to 1.0% and 0.1% of the initia l concentrations, respec tively, with macrophytes, rather than sediments, func tioning as the dominant pesticid e sink. In similar experiments performed by Cooper et al. (2002) 96-97% and 99% of added lamd a-cyhalothrin and bifentrhrin was associated with macrophytes within only 3 h ours of the exposure event and concentrations for both pesticides were below ecotoxicological limits within a downstream distance of 200 m. Cooper et al. (2004) reported th at a ditch length of 510 meters would be sufficient to remove 99.9% of initial exposure concen tration of the pesticide esfe nvalerate. Using microcosm experiments, Bouldin et al. (2005) found that Juncus effuses was more resilien t to exposure of atrazine and lamda-cyhalothrin compared to Ludwigia peploides However, L. peploides tended to accumulate more pesticides in its tissues (up to 2,461 g kg-1 for atrazine).

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52 While no studies of metals retention were identified specifically in drainage ditch ecosystems, studies from wetlands show that the ability of macrophytes to remove heavy metals also differs significantly by sp ecies (e.g. Debusk et al. 1996, Mire tzky et al. 2004, Maine et al. 2006), with some species exhibiting internal de toxification mechanisms that allow them to accumulate metals beyond accepted plant toxicity lim its (Deng et al. 2004). It is evident that some plant species may be more effective for pe sticides and metals retention based on their accumulation capacity and toxicity tolerance. Synthesis and research needs The studies reviewed here paint varying pictures of the ability of ditch systems to mitigate exports of potential environmental contaminants such as sediments, nutrients and pesticides (Tables 2-1 and 2-2). In genera l, current research supports th at vegetated ditches show high potential for retention of sediments and pest icides. Dissolved nutrients present a greater challenge due to inadequate residence times and high nutrient loads relative to available reactive surface areas. Unfortunately, the timing of high con centration events is often in conjunction with the greatest levels of discharge, thereby exacer bating the ineffectiveness of in-stream processing. Increasing residence times in ditches pose a ri sk for flooding during hi gh flow events thus creating an incompatibility between promoting both in-stream treatment and efficient water transport. To reconcile these conflicting goals, several researchers have repor ted on diversified and integrated buffer measures that serve as multiple checks for dampening runoff rates and concentrations (Bouldin et al. 2004, Wang et al 2005, Dabney et al. 2006). Front-end mitigation measures in agricultural settings include BMPs such as buffer strips, conservation tillage and winter cover crops (Cooper et al. 2004). Dukes et al. (2003) re ported that controlled drainage was shown to dampen the flashiness of storm ev ents, which could have important implications

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53 for improved processing within ditches. Magnitude of loading is also a f unction of the timing of fertilizer and pesticide applicati ons. Jia et al. (2006) found that scheduling of irri gation played a more important role in controlli ng losses of nutrients from fields than the BMPs in place. For urban systems, front-end mitigation measures include the reduction of impervious surfaces contributing to drains (Walsh et al. 2005). Continued research on load-reduction s hould then be coupled with innovative experimentation to seek approaches that amplif y the efficiency of in-stream processing in ditches. An appropriate theoretical template is that of hot spots and hot moments presented by McClain et al. (2003). Mechanisms for reten tion, particularly thos e governed by rates of biogeochemical reactions, are controlled by one or more limiting reactants in space or time (Table 2-3). Identifying limiting factors as a starting point, ecol ogical engineers can then begin the creative design process of implementing thos e missing reactants into ditch systems to enhance overall retention. Hedin et al. (1998) provided a conceptual ex ample of this approach by offering the advice that managers interested in promoting denitrific ation should consider prom oting natural inputs of labile carbon to the near-stream region. Groffman et al. (2005) f ound that local inputs of labile carbon could be provided by promoting the presence of organic debris jams which functioned as denitrification hot spots in ur ban steams. Taking the idea of providing local carbon sources one step further, Cooke et al. (2001) presented the idea of a denitrif ication bioreactor consisting of gravel, sawdust and corncobs for the outfalls of subsurface tiles. Where anoxic conditions are limited, water levels or hydraulic gradients in some cases may be adjusted as proposed by Grimaldi and Chaplot (2000). Sorption sites fo r P retention may be enhanced by using soil amendments such as calcium carbonate, hematite, vermiculate among others (Dunne et al. In

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54 Press ). Ann et al. (2000) suggested aluminum or calcium carbonate amendments that are less influenced by redox conditions. Recycled concrete gravel or other sources of limestone could be added to the benthic substrat e to create riffle sections for improved oxygenation, mixing and possible co-precipitation of calcium and phosphorus. Other thought-provoking desi gn ideas can be gained fr om innovative stormwater infiltration trenches such as the Ecology Ditch described by Barber et al. (2003). This design consists of vegetation that is underlain by a layer of compost, followed by sand and finally by a perforated drainage pipe su rrounded by gravel. Vegetation trap s pollutants and organic matter provides adsorption and filtration sites. Infiltration to the sand and gravel layers provides further filtration and dampens peak discharges. The pipe provides drainage when the storage capacity of the sands and gravels is exceeded. In most ditches, efficient rete ntion is primarily constrained by lack of channel complexity and low residence times. Creating or restoring channel meanders may be possible in some cases to increase channel heterogeneity and lengthen transport time. While this approach would require additional land, the total area requi red may still be less than that needed for a constructed wetland. For ditches that have sufficient hydrau lic capacity, baffles or other structural modifications might be installed in place of cr eating meanders. Materials such as straw bales (except where grazing animals have access) might be useful as permeable baffle structures that also provide a source of labile carbon and reactive surface areas. Although some findings contradict the contaminant retention benefits of vegetation in ditches (e.g. Barlow et al. 2003, D ukes and Evans 2006), there appear s to be a general consensus that plants are key components due to their direct uptake and sorp tion of potential contaminants and especially by the hydraulic and microbial cond itions they promote. Despite their importance,

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55 relatively few studies have speci fically addressed macr ophyte management strategies in ditches and other macrophyte-rich lotic systems (van Stri en et al. 1991, Milsom et al. 2004, Wilcock et al. 2004, Vereecken et al. 2006). Fifteen of the 22 studies presented in Table 2-1 identified dominant macrophyte species present at the study sites. Ten species were prevalent in more than one study and are summarized in Table 2-4. Future ditch studies should determine which species are best suited for mitigating target contaminan ts given the geomorphic and climatic conditions present. While some macrophyte species have hi gher uptake rates and tolerances to toxic substances, studies have shown th at polycultures nevertheless te nd to be more effective than monocultures for removal of nutrients (Kadlec and Knight 1996, Picard et al. 2005) and pesticides (Bouldin et al. 2005). Future studies should also exam ine at what biomass densities macrophytes are most effective at promoting the treatment function of ditches without compromising the risk of flooding during large st orm events. Improved understanding of which species to promote and how often to harvest wi ll enable managers to develop more informed strategies for management of ditches as potential BMPs. Some ditches may not support in-stream m acrophyte growth due to their geomorphology. These ditches could be widened to reduce their wa ter depth and velocity or where land is limited, one option for introducing macrophytes to the system is the use of floating platforms that support hydroponic growth of macrophytes. Constructed floa ting islands are frequently used in lagoons to support hydroponic growth of fast-growing speci es with extensive root systems such as Vetiveria zizanioides for improved water quality (Lavania et al. 2004). For ditches with perennial water flow, floating platforms could be made to freely move vertically with changing water levels yet be fixed in horizontal space.

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56 Examination of Table 2-2 also indicates wh ere research in ditches is most lacking. Despite concerns about water qua lity impairment due to pathogen contamination (Rodgers et al. 2003, Kay et al. 2007), few studies have addresse d BMPs for pathogen reduction in agriculture (Oliver et al. 2005) and surprisingl y, no studies were found that spec ifically addressed the ability of vegetated ditches to mitigate pathogens. Sim ilarly, no studies were found that examined heavy metal retention during transport in vegetated ditc hes. Another important research area for ditch management is the impact of sediment dredging on retention mechanisms. Macrae et al. (2003) found no difference between phosphorus retention be fore and after dredging; however others have reported decreases in retention due to th e loss of accumulated organic matter (Sallade and Sims 1997a, Nguyen and Sukias 2002). Furthermore, there is the question of what should be done with the extracted sediments. Van Strien et al. (1991) recommended utilizing organic and nutrient-rich sludge by re-applyi ng it to surrounding fields. Experimental approaches for management a nd modifications to existing channels may not always be feasible due to possible interferen ces with channel operations. To provide insights into the effects of differing vege tation and sediment management strategies on ditch performance under controlled yet realistic field conditions, expe rimental drainage ditch research facilities have been constructed (Drent and Kersting 1993 Strock et al. 2005). Whether research is conducted in experimental or actua l field sites, there is clearly the need to achieve a greater understanding of how ditches can be better design ed and managed to reduce impacts to receiving water bodies. Conclusions This review presented an overview of the potential treatment mechanisms occurring in ditches and provided a synopsis of the current lite rature that has reporte d on the use of ditches for reducing loads to downstream water bodies. Fi ndings from 22 individual studies suggest that

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57 as stand-alone systems, vegetated ditches are likely to be unreliable as effective mitigation measures, especially for reducing loads of di ssolved nutrients during high discharge events. However, if coupled with a load-reduction sc heme, vegetated ditches may serve as a costeffective and efficient BMP for mitigating cont aminant exports to receiving water bodies. Recognition of their potential role in mitigating contaminant loads may help to realign design objectives for ditches to achieve goals beyond that of water transport. E xperimental modification of drainage ditch structure for amplifying the treatment function of ditches represents a virtually untapped and much-needed realm of research that could have signi ficant implications for improved downstream water quality in urban a nd rural environments around the world.

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58 0 1 01 dischargeretention efficiency 0 1 01 loadretention efficiency Figure 2-1. Hypothetical relationshi ps for A) contaminant processing efficiency versus discharge proposed by Meyer and Likens (1979) a nd B) contaminant processing efficiency versus contaminant loading, as proposed by Odum et al. (1979) Processing mode Throughput mode Processing mode Throughput mode A B

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59 Figure 2-2. Simplified systems diagram of the in teractions between macrophyte growth, kinetic energy (KE, a proxy for water velocity), sediment transport and contaminant retention in a generalized ditch system : Rain events or water usage deliver environmental contaminants to the system (N=nitrogen, P=phosphorus, Pest=pesticides, metals and sediments). Gr eater plant biomass leads to lower water KE and more sediment accumulation. Lowe r KE promotes contaminant retention, represented here by storage in benthic se diments and uptake and temporary storage by macrophytes. When the uptake capacity of nutrients is reache d, they no longer contribute to plant biomass. For simplicity, to xicity effects of metals and pesticides are not represented here. When plant biom ass reaches the maximum density threshold (as determined by the land manager) the pl ants are harvested, providing a permanent sink of accumulated contaminants but temporarily diminishing their KE attenuation function. Higher KE results in reduced reten tion, represented here as re-suspension from and the lack of deposition to benthic sediment storages.

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60Table 2-1. Summary of environmental condi tions and findings from published studies examining sediment, nutrient, BOD and pesticide retention in drainage ditc hes (Note: NP = data not provided) Contaminant Ditch type (source of water) Discharge (L s-1) Concentration(s) (mg L-1) Substrate Vegetation Treatment effect(s) Represent significant effect? Citation Sediment 4 agricu ltural ditches, Coastal Plain, North Carolina, USA 804-8940 TSS = 107-1345 Sandy Variable cover depending on season Ditches were sinks for loads from fields Sediment storage = 8.6-130.1 kg m-1yr-1 Yes (Lecce et al. 2006) Sediment TN, TP Receiving pond water from catfish farm, Mississippi, USA 95-127 TSS=740-4549 TN=13.5-32.5 TP=0.42-3.89 NP 50-75% cover Polygonum, Ludwigia, Sagittaria, Typha, Eleocharis Avg.% removal (n=3) TSS = 96.5 TP = 68.8 TN = 67.1 Yes (Hargreaves et al. 2005) PO4 Agricultural and pastoral ditches, Florida, USA NP NP sandy,siliceous OM=11-22% NP Degree of P saturation > 25% indicating potential of P release Not clear (Dunne et al. In Press ) PO4 7 agricultural ditches, Indiana, USA NP SRP=0.02-0.09 silt+clay = 593% TC= 0.9-11.4% NP Some sites acted as P sinks and others as P sources Not clear (Smith et al. 2005) TP, PO4 2 dairy farm irrigation ditches, Australia 0-20 TP=0.1-5.7 SRP=0.07-5.23 silt+clay=7389% 5-85% cover Lolium, Trifolium Paspalum, Cyperus Estimated 14% decrease in P exports using a bare drain Yes (Barlow et al. 2003) PO4 Agricultural ditch, Canada 1.5-7.3 SRP=0.7-2.1 sands a nd silts Tall grasses and periphyton P areal uptake rates in same range pristine systems but represented only 510% retention No (Macrae et al. 2003) TP, PO4 12 pastoral farms, New Zealand <2 TP = 0.05-0.40 SRP= 0.015-0.03 silt+clay >60% TC=4.7-14.1% Potamogeton, Polygonum, Glyceria, Nitella, Lemna, Ranunculus, Bromus Sediments acted as a long-term P sink (residual P 6-39%) Yes (Nguyen and Sukias 2002)

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61Table 2-1. Continued Contaminant Ditch type (source of water) Wetted width (m) Discharge (L s-1) Concentration(s) (mg L-1) Substrate Vegetation Treatment effect(s) Represent significant effect? Citation PO4 17 agricultural ditches, Delaware, USA NP NP TP = 0.21-6.14 SRP = 0.04-0.74 silt+clay =24% OM = 8% NP Sediments were enriched with bioavailable P but still had high buffer capacity Not clear (Sallade and Sims 1997a) P (forms not specified) Wet meadow ditches, The Netherlands NP NP NP NP Heavy cover Fontinalis 90-95% removal capacity of P Yes (Meuleman and Beltman 1993) TN, TP NO3, PO4 5 irrigation ditches, Australia ~2-4 2-90 TP=0.11-1.02 SRP = 0.036-0.53 TN=1.56-3.47 NO3-N=0.02-2.14 NP Variable cover Paspalum, Typha, Potamogeton, Schoenoplectus, Ludwigia, Vallisneria, Elodea, Sagitaria Reductions in all parameters in 4 of 5 ditches Yes (Bowmer et al. 1994) NH4 NO3 BOD 4 groundwater-fed agricultural ditches, Germany 1.5 > 0.5 NH4-N= 0.13.05 NO3-N= 13.1.0 BOD= 14.8 NP >50% flow obstruction Phalaris, Carex ~79% and 60% reduction in NO3-N and BOD in 150 m, no net change in NH4-N Yes (Scholz and Trepel 2004) NO3 NH4 PO4 6 agricultural ditches, Michigan and Indiana, USA NP 1.5 575 NH4-N=0.0120.144 NO3-N=0.2-5.1 SRP=0.0062-0.262 NP Presence of algae and macrophytes Uptake velocities within same range as those reported in relatively pristine systems Not clear (Bernot et al. 2006) PO4 NH4 2 agricultural row crops ditches and 2 canals, North Carolina, USA 0.9 and 2.6 12 and 18 NH4 0.086-0.648 SRP 0.039-0.604 Coarse to mediumgrained sand, accumulations of silt. No CWD Ditches Ludwigia Canals Potamogeton Uptake velocities within ranges reported for pristine systems Estimated 46-75% and 13-66% NH4 and PO4 removed during high flow periods Yes (Ensign et al. 2006)

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62Table 2-1. Continued Contaminant Ditch type (source of water) Wetted width (m) Discharge (L s-1) Concentration(s) (mg L-1) Substrate Vegetation Treatment effect(s) Represent significant effect? Citation PO4 NH4 NO3 2 agricultural ditches, receiving additional point source inputs of septic and WWTP effluents, Germany NP 22-23 164-511 NO3-N = 0.816.4 NH4-N = 0.030.3 SRP = 0.01 0.27 fine sands OM = 5.9% Potamogeton, Sparganium, Phalaris, Glyceria Uptake ra tes were high compared to pristine systems but retention was low due to high loads No (Gucker and Pusch 2006) NO3 Agricultural tiledrained row crops, Illinois, USA 7.4 20 -1370 NO3-N= 3-17.7 Gravel, sand, some organic sediments. 27-86% cover Cladophora Potamogeton Denitrification rates ranged from ~0 to 16 mgN m-2 hr-1 No (Schaller et al. 2004) NO3 5 agricultural tiledrained ditches, Illinois, USA NP 10-1840 NO3-N = 0.615.8 Gravel and sand >95% Open canopy in 4 of 5 sites Denitrification rates ranged from <0.1 to 15 mgN/hr/m2 No (Royer et al. 2004) Pesticides Row crops, Mississippi Delta, USA 2.8 29.4* pyrethroid = 3.74 and 6.66 NP ~88% cover (up to 6000 g m-2), Ludwigia, Lemna Polygonum Pyrethroid concentrations decreased to 0.1% of initial concentration within 280 m Yes (Bennett et al. 2005) Pesticides Row crops, Mississippi Delta, USA 2.8 16.2 pyrethroid = 0.15 NP Ludwigia, Polygonum Leersia Pyrethroid concentrations decreased to 0.1% of initial concentration within 510 m Yes (Cooper et al. 2004) Pesticides 3 experimental ditches, The Netherlands 1.6 NP lambacyhalothrin = 0.0025 Silty clay loam Variable densities Myriphyllum, Elodea, Sagittaria 94-98% insecticide added was removed from water column by day 3 Yes (Leistra et al. 2004)

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63Table 2-1. Continued Contaminant Ditch type (source of water) Wetted width (m) Discharge (L s-1) Concentration(s) (mg L-1) Substrate Vegetation Treatment effect(s) Represent significant effect? Citation Pesticides 2 ditches, row crops, Mississippi Delta, USA 2.8 1.0 lamba-cyhalothrin = 0.06 bifentrhin = 0.11 NP Polygonum,Leersia Lemna lamba-cyhalothrin and bifenthrin reduced to below ecotoxicological levels within 200 m Yes (Cooper et al. 2002) Pesticides 2 agricultural ditches draining irrigation runoff, Canada NP 0 to 2000 No initial/upstream concentrations provided Loams to fine sandy loams NP 2 herbicides occasionally exceeded aquatic life guidelines 1.9 and 2.2% N and P applied to fields exited ditches Not clear (Cessna et al. 2001) Pesticides Row crops, Mississippi Delta, USA NP 61.3 pyrethroid = 0.46 atrazine = 28.9 NP 16-881 g m2 Polygonum, Leersia Sporobolus Atrazine and pyrethroid concentrations were decreased to noeffects level within 50 m Yes (Moore et al. 2001) Pesticides 8 experimental ditches, The Netherlands 3.4 NP linuron = 0.00050.5 sandy loam OM = 26% 20-110g m2 Myriophyllum, 40-60% of applied linuron was discharged from the 40-m ditches Not clear (Crum et al. 1998) *estimated from channel and velocity information

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64Table 2-2. Summary of the retention effectiveness of ditches based on the studies reported in Table 2-1 Contaminant Number of cases* Number concluding ditches to be effective Number concluding ditches to be ineffective Inconclusive Sediments 2 2 0 0 Pesticides 7 5 0 2 BOD 1 1 0 0 Total N 2 2 0 0 Total P 5 4 0 1 Phosphate 10 4 2 4 Nitrate 6 2 3 1 Ammonium 5 1 3 1 Metals 0 0 0 0 Pathogens 0 0 0 0 38 55% 21% 24% Individual studies from Table 2-1 were counted more than once if they considered multiple contaminants

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65 Table 2-3. Possible reactants and co nditions that limit the retention capacity in drainage ditches Contaminant Retention limited by Nitrate Labile carbon Anoxic zones Assimilation capacity Ammonium Oxygen for nitrification Assimilation capacity Phosphate Sorption sites Appropriate redox conditions Assimilation capacity Sediments Quiescent zone s and channel friction Pathogens Sorption sites Quiescent zones Pesticides Sorption sites Assimilation capacity Table 2-4. Dominant macrophytes re ported by the 15 studies out of 22 in Table 2-1 that provided plant species information Scientific name Common name O ccurrence frequency Plant type Polygonum spp. Knotweed 6 Emergent Potamogeton spp. Pondweed 5 Submerged or floating leaves Ludwigia spp. Primrose 4 Emergent Leersia spp. Cutgrass 3 Emergent Lemna spp. Duckweed 3 Floating Sagittaria spp. Arrowhead 3 Emergent Elodea spp. Pondweed 2 Submerged Glyceria spp. Manna grass 2 Emergent Myriophyllum spp. Water milfoil 2 Submerged Typha spp. Cattail 2 Emergent

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66 CHAPTER 3 INFLUENCE OF MACROPHYTES ON TRANSPORT CHARACTERISTICS AND NUTRIENT STORAGES Introduction Macrophytes are important features affec ting the transport and fate of potential environmental contaminants in running water ecosy stems. Plants provide temporary retention of potential contaminants such as nutrients dir ectly through uptake and assimilation into tissues. They also indirectly influence in-stream retention mechanisms by providing structural complexity that creates zones of in-stream tr ansient storage. These zones of slowly-moving water increase channel residence time, promot e sedimentation of suspended materials, and amplify biogeochemical reactions in both time and space due to greater contact with reactive substrates (Figure 3-1). Vegetation is likely to influence the trans port and fate of pathogens, sediments and nutrients in wastewater draina ge ditches in Oxapampa. The mild climate coupled with high nutrient conditions characteristic of these ditches results in year-round abundant growth of vegetation. Due to the risk of flooding, potential mosquito and other pest problems and the local negative perception of weedy plan ts species, vegetation in ditc hes is periodically removed. The current management practice by the city of Ox apampa and other municipalities in the region consists of harvesting aboveground ditch vegetati on with machetes followed by application of a non-selective glyphosate herbicide (Roundup ) For wastewater ditches with medium to high daily water flows (>0.5 liters/s), these procedures are carried ou t two to four times per year (Arteaga, personal communication). Few studies have explicitly li nked the presence of plants with transient storage in lotic ecosystems (Harvey et al. 2003). This study exam ines how plant management in wastewater ditches influences properties of water transpor t by comparing transport times and transient

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67 storage parameters under different biomass conditi ons. It also quantifies nutrient storages in macrophyte tissues to examine whether plant uptake serves as an important retention mechanism. The specific hypotheses tested in my study are: (1 ) transport in vegetate d ditches results in decreased concentrations of suspended sediments, E. coli bacteria and nutrients; (2) the presence of ditch vegetation significantly increases transient storage and channel residence time; and (3) aboveground biomass harvesting provi des an important nutrient sink. Methods Site Descriptions Four wastewater ditches were chosen as st udy sites for examination of water quality and plant biomass characteristics: San Martin, Gust avson, Koell and Frey. At the times of sampling, the lower reach of Frey was ponded with no fl ow exiting the site for discharge to the Chorobamba River. As a result, tr ansport studies to examine transient storage were limited to the San Martin, Gustavson and Koell sites. The site names are derived from the names of the streets along which the ditches are located. Wastewater ditches within the city run parallel to road s (Figure 3-2) and consequently are straight and typically incised. Most sites were dominated primarily by Cynadon dactylon and Hedychium coronarium with the exception of Frey, which co nsisted of a larger variety of both floating and emergent macrophytes due to the pr esence of a marsh environment in the lower reach (Table 3-1). San Martin Four locations were sampled in San Martin, the first of which was located 30 m below the junction between a natural stream and the inflow of wastewaters. The second sampling location receives occasional inputs from a small ditch fed by wastewater from a single house. There are no wastewater inputs between locations 2 and 4, the most downstream location, which is

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68 approximately 6 m upstream from the discharge poi nt to the river (Figur e 3-3). All of the San Martin sampling sites are surrounde d by heavily grazed pasture. Gustavson The Gustavson ditch exists along four city blocks and is surrounded by homes that discharge wastewaters along the to tal length of the d itch (Figure 3-4). Samp les were collected along each block for a total of four sampling site s before discharge to the river. Additional wastewaters enter Gustavson from two perpendicu lar streets. Hence, discharge in Gustavson tends to increase at each consecutive sampling site. Koell Koell is located along a densely-inhabite d city block and th en passes through an abandoned pasture before discharging to the river (Figure 3-5). The first two sampling locations are at the top and bottom of the city block where homes contribute wastewater that results in increased flowrates between the two locations. Th e channel along this section is deeply incised and subject to severe slumping. The third and f ourth sampling locations are within a pasture where the channel gradient is lower and the channel bed is wider and shallower. The Koell ditch site has a history of trash dumping into the mid -reach of the ditch due to the lack of a trash collection service for residents li ving at the end of the block. Th erefore between sampling sites 2 and 3 there existed an accumulation of predominan tly plastic items such as bags and bottles as well as miscellaneous items such as tires, pieces of scrap wood and metal. Frey Frey has the lowest flows of the four study ditches and on occasion dries out. The first sampling location in Frey is in the upper porti on of ditch where the channel is surrounded by homes that discharge wastes. At the end of the city block, the channel abruptly drops 2 m and drains into a marsh covered where the second sa mpling site is located (Figure 3-6). The city

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69 occasionally attempts to ditch the marsh to prom ote flow towards the river to avoid mosquito problems. However, the sites low topography re sults in ponding despite the citys efforts. During high flows, a portion of the water in Frey flows out of the marsh and joins with the Koell site. Just before this confluence is the final sampling location in Frey. Field Procedures for Water Chemistry, Flow and Channel Characterization Grab samples of ditch water for each site were collected periodically between May and November 2005 and March through July 2006 to examin e water quality over distance traveled in each ditch (hereafter longitudinal samplings) and through time during the dry season. Water samples were collected in acid-washed, amber 500 mL Nalgene bottles. Samples were placed on ice in a cooler and taken to the research stati on in Oxapampa for analysis. Water temperature and dissolved oxygen (DO) were measured in situ using a handheld DO meter (Model 55, YSI Environmental, Yellow Springs, OH). Water pH and specific conductivity were also determined in situ using a combined pH and conductiv ity meter (ExStik EC500, Extech Instruments, Waltham, MA). Water velocity was measured at each sampling location using an impeller flowmeter (Model 2030, General Oceanics, Miami, FL) and used with channel width and depth measurements for calculating discharge. Sites were surveyed to evaluate channel characteristics such as length, slope and cross-sectional areas. Laboratory Analysis of Water Quality After collection, water samples we re immediately filtered through 0.45 m membrane filters and analyzed for soluble r eactive phosphorus (SRP), nitrate (NO3-N) and ammonium (NH4-N) with a portable Hach DR/890 colorimeter and powder pillow reagents (Hach Company, Loveland, CO). SRP was analyzed with the PhosVer3 ascorbic acid method (equivalent to USEPA Method 365.2 and Standard Method 4500-PE; detecti on limit 0.05 mg L-1 as PO4, precision 0.03 mg PO4 L-1). NO3-N was analyzed using the cadmium reduction method

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70 (detection limit of 0.01 mg L-1 as NO3, precision .03 mg NO3 L-1). Liquid wastes from the NO3 analysis were collected and di sposed of through Environmenta l Health and Safety at the University of Florida. NH4-N was analyzed with the salicyl ate method (detection limit 0.02 mg L-1 as N, precision 0.02 mg N L-1). Unfiltered samples were analyzed for total phosphorus (TP) using the Hach acid persulfate digestion method followed by analysis of SRP as described above. Total nitrogen (TN) was measured using Hach Te st N Tube digestion me thod (detection limit of 2 mg N L-1, precision 0.5 mg N L-1). Proper operator and equipment performance was tested for each set of samples through the use of known standards (purchased as ampules from Hach Company), blanks and replicates. Sediments collected on 0.45 m membrane filters were used for gravimetric determination of total suspended sediments. Concentrations of E. coli bacteria were measured using Coliscan Easy Gel kits (Micrology Laboratories, Gosh en, IN). Carbonaceous biochemical oxygen demand (CBOD5) was analyzed following Standard Meth od 5210 B (APHA 1992) and measured with a YSI 58 DO meter fitted with a YSI 5010 BOD probe (YSI Environmental, Yellow Springs, OH). Transport Characteristics of Ditches Transport characteristics of each ditch were determined through the analysis of breakthrough curves of conserva tive solute tracer experiments (described below, Table 3-2). Tracer experiments were performed on three occa sions in Gustavson. The first experiment took place approximately 3 weeks following the removal of plants from the upper 3 reaches of the ditch and before plant removal in the lower reach Therefore, the analys is of the first tracer experiment in Gustavson compares transport properties between the cleared and vegetated reaches. Tracer experiments were performed twi ce in San Martin and Koell; once in November 2005 and again in June 2006 to compare transp ort characteristics under low and high biomass

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71 conditions, respectively. In Koell, the second experiment also co incided with the removal of approximately 0.5 metric tons of trash from the stream channel. Tracer experiment field procedures Constant rate injections of NaCl were c onducted following experime ntal design protocols detailed by Wagner and Harvey (1997). Five and ten gallon Mariotte siphons (Figure 3-7) were constructed and used to deliver the conservative tracer at a c onstant flowrate such that the resulting breakthrough curve (BTC) reached a plateau concentration at the downstream monitoring location. For most experiments a constant rate injection of 0.9 L min-1 (accuracy .02 L min-1) was used. However, flows could be adju sted if desired by the changing the head differential between the vent t ube and the outlet from the Mari otte siphon. Quantities of the conservative tracer (between 20-40 g NaCl L-1) were pre-weighed at th e laboratory and dissolved at the site using ditch water th at had been sieved to remove any solids. The solution was then added to the Mariotte siphon. Tracer breakthrough curves were determined by measurements of specific conductivity (YSI 556 MPS, YSI Enviro nmental, Yellow Springs, OH). Grab samples of water were collected at 3 to 4 points duri ng tracer experiments to analyze water quality following procedures described above. Analysis of tracer data Transport and transient storage parameters were estimated from the resulting tracer breakthrough curve data using th e One-dimensional Transport with Inflow and Storage with Parameter Estimation (OTIS-P) model (Runkle 1998). The OTIS-P model conceptualizes flowing water systems as having two distinct hydrologic regimes (Wagner and Harvey 1997): surface water flow in the main channel and im mobile or slowly moving water in surface quiescent zones and in subsurface flowpaths. The mathematical model is a 1-D advectivedispersive transport model that accounts for a dvection, dispersion, lateral inflow, and transient

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72 storage. Transient storage is represented as a firs t-order mass transfer that yields two parameters. The first parameter represents the cross-sec tional area of the tran sient storage zone ( As) and the second is a first-order mass transfer coefficient ( ) that describes the rate of hydrologic exchange between storage zones and the main ch annel. The governing equations are: C C C C A q x C AD x A x C A Q t Cs L L 1 (3-1) C C A A t Cs s s (3-2) where t and x are time and distance along the stream; C CS and CL are the solute concentrations in the stream, storage zones and groundwater (mg L-1), Q is the stream volumetric flow rate (m3 s-1), A and As are the cross-sectional areas of th e stream channel and storage zone (m2), D is the dispersion coefficient (m2 s-1), qL is the lateral inflow rate (m3 s-1 m-1), and is the storage exchange coefficient (s-1). The OTIS-P model uses a least squares al gorithm to adjust parameter values of D, A, and As to fit the observed tracer BTCs. Confident in terpretation of model results is predicated on the reliability parameter estimations. Wagner and Harvey (1997) suggest reporting of the Damkohler number (DaI) as an i ndicator of parameter uncertainty where values on the order of 0.1 to 1.0 indicate high parameter relia bility. DaI is calculated as: v L A A I Das s 1 (3-3) where L is the reach length (m) and v is mean water velocity (m s-1). Parameter values and metrics that are co mmonly reported to facilitate cross-study interpretations of transient storage studies include As, As/A, Tmed, Fmed and Fmed 200. The ratio of As/A shows the relative importance of storag e zone and main channe l cross-sectional areas (Harvey et al. 2003). Tmed represents the median transport time, the time at which 50% of the

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73 plateau curve is realized (Runkel 2002). The fraction of transport ti me spent in transient storage zones is denoted as Fmed and may be approximated by (Runkel 2002): s s u LA A A e Fmed ) / (1 (3-4) Fmed is also calculated using a stan dardized distance of L = 200 m (Fmed 200) to allow comparison with other systems (Runkel 2002). Vegetation Sampling Collection of plants at sampling locations Above and below-ground biomass in ditches were sampled in November 2005 and June 2006 in San Martin, Gustavson and Koell in conj unction with tracer experiments as described above. Biomass sampling in Frey took place only in November 2005. Representative biomass samples were collected using a 0.25m x 0.25m qu adrat at three random locations within a m distance of each water quality sampling location in each study ditch for a total of 12 biomass samples per ditch. Only plants growing within the channel or with plant parts exposed to the water were collected. If the plant had no contact with the water or channel it was not included in the sample. This sampling approach was chosen to only include plants that might have an effect on transport characteristics or contaminant retention. At the research station, the plant tissues were carefully rinsed to remove trapped sediments, air dried for an hour and then se parated into leaves, stems and r oots/rhizomes and weighed. The plant tissues were then dried at 105C for 48 hours and reweighed. Moisture content was determined from the difference between the wet and dry weights. Biomass was calculated as the tissue dry weight multiplied by the total sampled area as determined by use of the quadrat. Dried plant tissue was ground to a 40 mesh si ze using a Wiley mill for analysis of total carbon, nitrogen and phosphorus. TN and TC cont ent were determined using approximately 15

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74 mg of dry sediments. Samples were analyzed by dry combustion using a CNS analyzer (Carlo Erba Model NA-1500). Total P content was determined on 0.5 g of dry sample that was combusted at 550C in a muffle furnace for four hours. The remaining ash was then dissolved in 6 M HCl (Andersen 1976) and the digestate was an alyzed for P on a Technicon autoanalyzer III using an automated ascorbic ac id method (Method 365.1, USEPA 1993). Plant tissue nutrient content was used to estimate above and belo wground nutrient storages at each study site. Comparison of biomass and nutrient content of plants grown out of wastewater Biomass samples of dominant macrophyte sp ecies found in wastewater ditches in Oxapampa were collected and compared with th e same species found growing out of wastewater following the protocols given a bove. The purpose of this comparison was to examine whether nutrient content and allocation of nutrients diff ered in plants found growing in wastewater. Statistical Analyses Comparisons of upstream and downstream concen trations of water quality parameters were made using paired one-tail t-tests of data that were ln-transformed to meet normality requirements. Plant tissue N and P of macrophyte species growing in wastewater ditches were compared with the same species growing out of wastewater using the same statistical approach. Results Water Quality of Wastewater Ditches General characteristics The ditch sites differed from one another with respect to their channel and flow characteristics as well as their water chemistry (Table 3-3). San Martin experiences the highest water flows and has the lowest nutrient and su spended sediment con centrations. The lower temperature and specific conductivity (17.6C, 155 S cm-1) in San Martin compared to that of the other sites (18.7-20.9C, 243-255 S cm-1) further supports that San Martin experiences more

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75 dilution than the other sampling sites. DO values for all sites were low (average DO values for the sites varied between 1 and 2.5 mg L-1) and CBOD5 values ranged from 4.6 mg L-1 in San Martin to 18 mg L-1 in Gustavson. These low oxygen environments limit nitrification, as evidenced from the high ammonium and orga nic N concentrations and low nitrate concentrations. Longitudinal water quality samplings indicated that concentrations of nutrients, TSS, and E. coli tended to increase with downstream distance in Gustavson due to wastewater inputs along the length of the ditch. Similar results were seen in the upper reach of Frey In the marsh of Frey, water was ponded on all sampling occasions, precluding water quality parameter measurements after passage thr ough the marsh. The lower reaches of the San Martin and Koell sites (reach lengths 255 m and 73 m, respectively) have no additional wastewater inputs and therefore allow observation of potential retention of TSS, E. coli and nutrient exports to the Choroba mba River. However, there were only 8 and 10 sampling occasions for San Martin and Koell, respectively, for which discharge measurements in the lower reaches did not vary si gnificantly (i.e. the change in flowrate between sampling sites was less than the flowmeter accuracy of 10 cm s-1). Seven of the 8 sampling dates in San Martin were characterized by high bioma ss conditions. Eight of the 10 sampling dates in Koell had high biomass as well. These small samp le sizes of low biomass conditions prevents a statistical comparison to made between upstream and downstream trends in water quality based on biomass conditions. Attempts were made to sample the same parcel of water at each sampling location by measuring mean water velocity to estimate arriva l times to downstream sites. It is recognized that this approach does not ensure that th e same water was sampled; therefore, these

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76 measurements were grouped to examine overa ll trends in upstream and downstream water quality to seek evidence of in-stream E. coli TSS and nutrient retention in San Martin and Koell during transport (Figures 3-8 and 3-9). Paired t-tests indicated a significant decrease in E. coli concentrations between upstream and downstream sampling sites for both Koell (t = 3.38, df = 6, p<0.01) and San Martin (t = 2.29, df = 6, p-value < 0.05). Downward trends in TSS were observed for both sites but were not significant at p < 0.05. TN and TP concentrations tended to decrease at both sites as well but only TN in Koell was significantly lower at the dow nstream location (t = 3.80, df = 3, p < 0.05). In San Martin, downstream SRP concentrations tended to be significantly lower than the upstream sampling location (t = 4.24, df = 7, p< 0.01) while downstream SRP values at Koell tended increase slightly compared to the ups tream sampling location. Ammonium and nitrate concentrations did not change significan tly between upstream and downstream sampling locations (NO3-N values not shown). Water quality during tracer experiments Water samples collected during tracer experi ments have the advantage of allowing the same parcel of water to be sampled through space. However, samples collected during the experiments revealed no consistent patterns that clearly distinguished th e effects of high biomss from low biomass conditions on contaminant rete ntion (Figure 3-10). TSS was variable in all sites but tended to be lower under high bi omass conditions in Koell and San Martin. E. coli concentrations consistently decreased with di stance downstream in all cases except during the low biomass conditon experiment in Gustavson. Nitrate values for all sites on all tracer experiment dates were less than 0.05 mg N L-1 and did not vary more than 0.03 mg N L-1 and therefore were not included in Figure 3-10. Phosphate and am monium concentrations were

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77 variable or increased with dow nstream distance for all sites. The only exception was phosphate which decreased from 0.9 to 0.55 mg P L-1 in San Martin under high biomass conditions. Transport Characteristics Comparison of breakthrough curves for sites und er different biomass conditions revealed the importance of macrophytes for providing structural complexity that served to extend travel times and increase sizes and exchanges with transi ent storage zones (Figures 3-11, 3-12 and 3-13 and Table 3-4). Median transport times (Tmed) in San Martin and Gustavson increased by 89 and 143% under high biomass conditions. The difference in Tmed between tracer experiments in Gustavson was likely exaggeratd by the higher flowrates during the low biomass experiment (17.9 versus 11.8 L s-1). However, the opposite was true for the experiments in San Martin (32 versus 38 L s-1 for the low and high biomass experiment s, respectively) indicating that the calculated percent increase in Fmed was likely underestimated for this site. Estimated parameter values revealed notable differences in transient storage between the low and high biomass tracer experiments (Table 3-4). The ratio of the transient storage and channel cross-sectional areas (As/A) in the 2005 Gustavson experiment increased from 3.0 to 11.6 between the low biomass and high biomass reaches, respectively. The fraction of median transport time spent in transient storage zones (Fmed) also increased from 18 to 62%. Similar results were observed in Gustavs on the following year in the sec ond set of tracer experiments. The As/A ratio increased from 0.008 to 0.27 in Gustavson and 0.36 to 1.33 in San Martin for the low and high biomass tracer experiments. Fmed increased from zero to 21% in Gustavson and 5.5 to 56% in San Martin. Exchange ra tes between the stream and storage zones ( ) also increased under high biomass conditions from 6.0 E-8 to 27.9 E-4 s-1 in Gustavson and from 3.7 E-4 to 30.5 E-4 s-1 in San Martin. Damkohler numbers were on the order of 0.1 to 1.0 for all

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78 tracer experiments except for the second high bi omass experiment in Gustavson (DaI = 47) suggesting high reliability of the parameter estimations (Wagner and Harvey 1997). The tracer studies performed in Koell reve aled an additional source of structural complexity common to wastewater drainage ditches in the study area: garbage. Before removal of approximately 0.5 metric tons of refuse from the channel, the conservative tracer BTC displayed a long tail implying greater contact with storage zones compared to the second experiment which took place shortly after tras h removal but during high biomass conditions (Figure 3-13). Values of As/A and Fmed after garbage removal decreased from 15.1 to 0.07 and from 52% to 0.9%. As seen in Table 3-4, medi an transport times are approximately the same between the two experiments in Koell despite hi gher transient storage during the low biomass conditions. The majority of the refuse had accumu lated in the channel be tween reach lengths of 93 and 121 m. The channel above and below this section was relatively clear of other flow obstructions. Therefore, transport times were likel y to be distributed between rapid transport in the upper and lower reaches with slow transport due to mixing with storage zones in the section with trash accumulation. Plant Biomass and Nutrient Storages At the times of plant sampling in 2005, mean total biomass (sum of above and belowground biomasses, Table 3-5) was highest in Gustavson (2,980 g m-2), followed by San Martin (1,137 g m-2), Koell (596 g m-2) and Frey (487 g m-2). In 2006, the sites showed a more even distribution in biomass with m ean total biomass values of 2,574 g m-2 in Gustavson, 2,725 g m-2 in San Martin and 2,826 g m-2 in Koell. Vegetation in 2005 had been harvested by the city from Gustavson and San Martin approximately 3 months and 2 months prior to biomass collection and from Koell and Frey approximately one month prior (Arteaga, pers onal communication). Using these approximate

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79 time periods of 90, 60 and 30 days together with biomass areal estimates and measured values of biomass and plant tissue nutrient content, daily N and P uptake by plants was estimated (Table 3-5). These calculations assumed a zeroth order growth model due to a lack of intermediary biomass data. Belowground biomass growth rate s could not be calculated due to an unknown quantity of roots and rhizomes remaining even after herbicide applications. Therefore the contribution of belowground biomass was not consid ered in daily uptake ca lculations. Estimated nutrient uptake rates for aboveground biomass were used with average dissolved inorganic N and P loads to each of the ditches to arrive at a rough approximation of th e potential gross N and P retention in ditches by plant upt ake (Table 3-6). The approximati ons made in Table 3-6 suggests that aboveground biomass of ditc h vegetation assimilates the equi valent of up to 9.6 and 11.7% of daily N and P loads in ditches. While these estimates do not include uptake contributions by belowground biomass, nutrient content of roots and rhizomes tended be minimal compared that stored in aboveground tissues. Nevertheless, the ove rall plant assimilations rates are therefore underestimates. However, N and P retention valu es do not account for nutrient releases from decaying plant material. Mean molar ratios of above and belowground C, N and P indicate high P enrichment at all sites (Table 3-7), particularly in Gustavs on (molar N:P of 10.1 and 3.6 in above and belowground plant tissues, respectively). Nutrie nt content per species (Figure 3-14) indicated that Cynadon had the highest foliar P content (7.0 mg g-1) while highest P in roots was Myriphyllum (9.4 mg g-1). Hydrocotyle had the highest foliar N content (47.4 mg g-1) followed by Cynadon (37.1 mg g-1). Eichornia and Myriophyllum contained the highest N in roots (42.3 and 40.1 mg g-1). Polygonum and Hydrocotyle had the highest P in stems (8.2 and 8.0 mg g-1) while Equisetum contained the hi ghest N in stems (32.2 mg g-1).

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80 Comparisons of whole plant nutr ient content of species growi ng in and out of wastewater show that plants in wastewater ditch to have elevated N and P tissue content (Figure 3-15). Both N and P in Hydrocotyle tissues were higher (p < 0.01) in wa stewater. P content was significantly higher (p < 0.01) in Myriophyllum while N content was significantly higher for Hedychium However, caution must be noted that the nutrient status of the soils of the control species was not determined. Discussion Waste streams discharged to wastewater drai nage ditches are highly variable through time and space reflecting punctuated and diverse uses of water by residents. The variability of inputs to a system leads to high uncertainty during longitudinal sampling campaigns that the same parcel of water is being captured over space in order to fairly assess in-stream retention. A decrease in downstream concentra tion may easily be misinterpreted as retention when instead a non-representative slug of water had been sample d. The problem of teasing apart within-stream variability from evidence of retention is a ddressed to some degree by lumping the data to evaluate changes in upstream and downstream concentrations. The resulting set of box plots illustrate that though there are tre nds in the data suggesting concen tration decreases with distance downstream, these differences may often be due to chance alone given the wide spread of the data. Notwithstanding, results do strongly imply potential for retention in the study ditches by sedimentation of E. coli TSS, TN and TP. The San Martin site also presents compelling evidence of SRP retention. Similar sampling cam paigns were not carri ed out during the wet season when flows tend to be highest. Therefor e, the results of the dry season conditions presented here are likely to overes timate overall system performance. While a growing number of studi es are reporting on the effect s of in-stream vegetation for reducing velocities and prom oting sedimentation (Wilcock et al. 1999, Champion and Tanner

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81 2000, Clarke 2002, Schulz et al. 2003, Hargreaves et al. 2005, Lecce et al. 2006), only a few to date have explicitly quantified the importance of macrophytes fo r creating zones of transient storage (Harvey et al. 2003, Salehin et al. 2003, Ensign and Doyle 2005). Analysis of BTCs using OTIS-P does not distinguish between surfac e and subsurface storage zones. Historically, the majority of OTIS-P interp retations assumed that exchanges with storage zones took place between the main channel and subsurface flowpa ths and often negated the existence of inchannel transient storage. More recent studies have recognized the prevalence of in-stream transient storage zones and have presented evid ence of their importance for nutrient retention (Gucker and Boechat 2004, Ensign and Doyle 2005) Comparisons between transport times and transient storage metrics under low and high biomass conditions in the present study indicated that essentially all storages zones exist in the main channel, not in the subsurface. That transient storage zones may be abundant in ditches presents such systems in a new light. Channelized systems are often presumed to have minimal transient storage compared to natural streams due to the lack of structural complexity needed to cr eate pockets of slowlymoving water or the lack of porous substrates to allow surface-subsurface hydrologic exchanges. However, results from this study found so me of the highest rates of exchange ( ) that have been reported to date. Interestingly, another study that reported similar values (43-58 E-4 s-1) was a channelized agricultural stream with in-stream vegetation (Ensign and Doyle 2005). As suggested by Ensign and Doyle (2005), these high values may represent exchanges with turbulent eddies that would tend to be more rapid than exchanges occurring with subsurface environments. The Koell site had the highest As/A value of all sites in the presence of channel debris. However, this site also had the lowest value compared to other the high transient storage zone cases. This slower exchange rate between main cha nnel advective flow and storage

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82 zones helps to explain why this site had the mo st pronounced BTC tail. The conservative solute spent more time circulating in dead zones formed by tires and other debris than in the turbulent eddies created by macrophytes in the other sites. The Koell site also provided evidence of the effect of channel morphology on the prevalence of transient storage zones. While vegetation clearly had an impact on transport characteristics in San Martin and Gustavson, it was much less pronounced, especially compared to the transient storage zones created by the accumu lation of trash. The effect of vegetation was likely to be insignificant in the upper reaches of Koell where the channel is most narrow and deeply incised supporting only bank vegetation. Abundant, year-round plant growth in the study region makes macrophytes good candidates for promoting in-stream retention in wastewater ditches. Th e assumed zeroth order biomass growth model used to ca lculate uptake rates is likely to be adequate for the 30 and 60 day periods. Linear growth of macrophytes was observed by Ta nner et al. (1996) over a 60 day period for 5 of 8 species growing in dairy farm wastewater. After 60 days, growth rates tended to increase for some species and decrease for othe rs depending on initial pr opagule vigor (Tanner et al. 1996). Linear growth was observed over th e 117 day sampling period for one species, Cyperus involucratus Therefore, assumption of linear grow th in Gustavson over a 90 day period may not have grossly under or overestimate gr owth rates but instead may have yielded an average value over the sampling period. Approximated N and P uptake rates are on the order of those reported for macrophytes in constructed wetland treatment systems (DeBus k et al. 1995, Greenway 1997, DeBusk et al. 2001). However, this retention mechanism is seve rely limited in ditches such as Koell where channel morphology marginalizes macrophytes to th e banks of streams, restricting plant-water

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83 contact and thus limiting direct uptake of nutrients (Reddy et al. 1999). Nutrient content of macrophytes growing in wastewater tended to be higher than those out of wastewater particularly for plants commonly found in the marsh in Frey. With the exception of Frey, the majority of the macrophytes in the present study are likely to obt ain their nutrients from sediment porewater rather than directly from the water column. Results of plant tissue N and P content of the ditch macrophytes indicate high P enrichment, with average belowground N:P ratios around 4 for all sites. N:P ratios below 14 to 16 are commonly associated with P enrichment relative to N (Koerselman and Meuleman 1996, Richardson et al. 1999). One possi ble explanation for belowground enrichment of biomass P are advec tive gradients established by plants roots that can drive the exchange of water and dissolved ma terial from the water column into porewater spaces allowing nutrients to become availabl e for direct uptake (Reddy et al. 1999). The estimates of N and P retention by uptake in this study do not consider th e eventual release of stored nutrients and thus it is likely that there is minimal net nutrient retention by biomass. Kim and Geary (2001) found that biomass harvesting removed less than 5% of TP present in microcosms planted with two species of wetla nd macrophytes. More than 95% was stored in substrates. In wastewater drainage ditches, it is likely that the same situation is true: rather than stored in plants, the majority of P is associated with benthic sediments. Conclusions This study demonstrated the influence of macrophytes on transport characteristics in ditches by creating zones of transi ent storage. By slowing transport times, these storage zones promote reductions of E. coli bacteria, TSS and SRP. However, large within-site variability of water quality limited the ability of this study to isolate more direct evidence of in-stream retention. Future studies shoul d take this variability into account by performing mass-balance studies over longer time periods such as 24 hours. Net nutrient re tention by plant uptake is

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84 minimal in wastewater ditches; instead, the treatm ent function of plants arises from their ability to create in-stream transient storage zone s, resulting in lower-velocity, depositional environments.

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85 SEDIMENT WATERTO RIVER WATER HUMANS KE = kinetic energy Sed = sediment Path = pathogens OM = organic matter N = nutrients DO = dissolved oxygen M = microbes OM PATH N N OM PATH RAIN SED OM PATH N AG/ PASTURE SED SUN FLOATING EMERGENT M SEDIMENT WATER WIND DO low KE KE SED PATH OM N high KE DO M and worms Harvest Wastewater ditch Figure 3-1. Systems diagram illustrating the infl uence of macrophyte harvesting on water quality in wastewater ditches. Plants reduce wate r velocities (expressed as kinetic energy) thereby promoting retention mechanisms such as sedimentation and sorption. Plants also take up and eventual ly return nutrients.

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86 Figure 3-2. Map of Oxapampa showing the lo cations of the sampling sites in each of th e study ditches. The San Martin, Gustavson and Koell sites discharge to the river while Frey often does not flows from th e marsh (location 2). Numbers represent sampling locations from upstream () to downstream ().

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87 Table 3-1. Dominant macrophyte species common to the study sites Species Common name San Martin Gustavson Koell Frey Cynadon dactylon Bermuda grass X X X X Eichornia crassipes Water hyacinth X Equisetem spp Horsetail X Hedychium coronarium Ginger lily X X X X Hydrocotyle ranunculoides Water pennywort X X Myriphyllum aquaticum Parrot feather X Polygonum punctatum Knotweed X X X Figure 3-3. Ditch site San Martin looking upstr eam from sampling location 4 in A) November 2005 and B) June 2006 A B

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88 Figure 3-4. Gustavson site looki ng downstream at times of solute transport experiments in November 2005 at A) three weeks after plan ts were harvested and B) lower reach, location 4 before plant removal; C) in June 2006; and D) in July 2006, one week after vegetation removal and he rbicide application. A B C D

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89 Figure 3-5. Koell site looking downstream from sampling location 1 at the time of tracer experiments A) in November 2005 and B) July 2006 Figure 3-6 Frey site at the marsh locati on at the time of biomass sampling in 2005 A B

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90 Table 3-2. Description of tracer experiments conducted at each site Site Date Experimental conditions of tracer experiment Gustavson 15 November 2005 Reaches 1-3 clea red of vegetation; reach 4 not cleared 22 June 2006 Reaches 1-4 with vegetation 13 July 2006 Reach 1-4 cleared of vegetation San Martin 24 November 2005 Low biomass 23 June 2006 High biomass Koell 24 November 2005 Low biomass 7 August 2006 High biomass, trash removed from channel Figure 3-7. View of the Mariotte siphon used for delivering the c onservative tracer at a constant flowrate

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91 Table 3-3. Channel characteristics and mean wate r quality parameters for wastewater drainage ditch sites San Martin Gustavson Koell Frey Channel slope (%) 1.0 1.4 1.9 4.3a Channel length (m) 356 360 194 210 Wetted width (m) 1.09 0.92 0.83 0.70b Discharge (L s-1) 4520 (97)c 1911 (64) 127 (33) 56 (11) DO (mg L-1) 2.51.6 (87) 1.00.9 (80) 1.41.0 (28) 1.91.8 (24) Temperature (C) 17.62.9 (87) 18.71.0 (80) 20.92.2 (32) 19.32.4 (28) pH 6.70.3 (75) 6.70.2 (63) 7.00.2 (26) 6.80.3 (25) Conductivity ( S cm-1) 15549 (64) 24344 (57) 24469 (32) 25546 (25) TSS (mg L-1) 32.929 (88) 4247 (83) 6350 (40) 8379 (24) CBOD5 (mg L-1) 4.61.3 (5) 18.112.1 (15) 10.66.1 (8) 8.56.3 (11) E. coli (CFU 100 mL-1) 36842322 (45) 41894519 (36)74857861 (33)64066263 (27) SRP (mg L-1) 0.310.28 (97) 0.560.31 (90) 0.420.14 (57) 0.360.18 (28) TP (mg L-1) 1.820.88 (33) 4.231.56 (33) 3.060.98 (16) 3.402.86 (10) Organic P (mg L-1)d 1.480.88 (33) 3.591.60 (33) 2.620.91 (16) 3.002.77 (10) NO3-N (mg L-1) 0.030.03 (96) 0.02+0.02 (89) 0.020.01 (53) 0.060.09 (28) NH4-N (mg L-1) 2.671.67 (97) 5.962.81 (89) 6.002.33 (53) 5.802.02 (28) TN (mg L-1) 4.832.10 (34) 8.253.78 (38) 7.112.06 (16) 6.092.82 (10) Organic N (mg L-1)e 1.552.11 (34) 2.363.06 (38) 0.571.87 (16) 1.402.56 (10) a This average slope includes the escarpment. Above and below this point the channel slope is 1.6% and 0.8%, respectively. b Averaged for the entire site and when the marsh (site 2) was ditched. c Values are averages 1 SD. Values in parentheses are numbers of samples. d Organic P is calculated as the difference between the TP and SRP and includes polyphosphates and acid hydrolyzable P. eOrganic N is calculated as the difference between TN and NO3-N and NH4-N.

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92 UpstreamDownstream 0 4000 8000 12000E. coli (CFU 100 mL-1) San Martin (distance = 255 m) UpstreamDownstream 0 10000 20000 30000 Koell (distance = 73 m)5900 UpstreamDownstream 0 50 100 150TSS (mg L-1) UpstreamDownstream 0 50 100 150 60004400* 2400** 42.537.8 55.5 101 Figure 3-8. Box plots of San Ma rtin and Koell at upstream a nd downstream sampling locations for A) E. coli and B) TSS. Solid circles and numbe rs indicate median concentration values. Open circles denote outlier values and boxes and whiskers delimit the interquartile range (IQR) and 1.5 IQR, respectively. and ** following values indicate p < 0.05 and p < 0.01. A B

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93 UpstreamDownstream 0 1 2 3 4 5TP (mg L-1) San Martin (distance = 255 m) UpstreamDownstream 0 1 2 3 4 5 Koell (distance = 73 m)3.7 UpstreamDownstream 0 4 8 12TN (mg L-1) UpstreamDownstream 0 4 8 12 2.21.3 2.3 4.8 4.3 5.9* 8.4 UpstreamDownstream 0.0 0.2 0.4 0.6 0.8 1.0SRP (mg L-1) San Martin (distance = 255 m) UpstreamDownstream 0.0 0.2 0.4 0.6 0.8 1.0 Koell (distance = 73 m)0.42 UpstreamDownstream 0 2 4 6 8 10NH4-N (mg L-1) UpstreamDownstream 0 2 4 6 8 10 0.270.19** 0.48 2.52.4 5.6 5.6 Figure 3-9. Box plots of San Ma rtin and Koell at upstream a nd downstream sampling locations for A) TP, B) TN, C) SRP and D) NH4-N. Solid circles and numbers indicate median concentration values. Open circles denot e outlier values and boxes and whiskers delimit the inter-quartile range (IQR) a nd 1.5 IQR, respectively. and ** following values indicate p < 0.05 and p < 0.01. A D C B

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94 Figure 3-10. Longitudinal water quality trends of E. coli TSS, SRP and NH4-N during tracer experiments performed in sites San Martin (circles), Koell (stars) and Gustavson (triangles, shown for 2006 tracer experime nts only) under both high (bold lines) and low (thin lines) biomass conditions. Note: li nes between symbols are used only to aid visualization 5.3 mg P L-1

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95 0 0.5 1 1.5 0.00.51.01.52.02.53.0 Time since injection (hours)Fraction of plateau concentration With plants (22 June 2006) No plants (13 July 2006) Figure 3-11. Comparison of conser vative tracer breakthrough curves in Gustavson before and after removal of ditch vegetation in 2006 0 0.5 1 1.5 0.00.51.01.52.02.5 Time since injection (hours)Fraction of plateau concentration Low biomass (November 2005) High biomass (June 2006) Figure 3-12. Comparison of conser vative tracer breakthrough curves in San Martin with low and high ditch vegetation biomass 0 0.5 1 1.5 0.01.02.03.0 Time since injection (hours)Fraction of plateau concentration Low biomass (November 2005) High biomass, no trash (July 2006) Figure 3-13. Comparison of breakthrough curves in Koell under different biomass conditions and after trash was removed from the stream

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96Table 3-4. Biomass and hydrologic conditions of each tracer experiment a nd resulting transient storage zone parameter values Site Aboveground Biomass (g m-2) Discharge (L s-1) Velocity (m s-1) Tmed (min) Fmed (%) Fmed 200 (%) A (E-2 m2) As (E-2 m2) As/A (E-4 s-1) DaI Gustavson (2005)a Negligible (upstream reach) 13.2 0.23 nab 18 24 8.07 24.0 2.97 4.53 0.36 1823 655 (downstream reach) 25.5 0.14 na 62 58 12.2 142 11.60 10.4 1.20 Gustavson (2006) Negligible 8.6-17.9 0.32 27.5 0c 0c 4.47 0.00034 0.008 0.0006 0.81 2021 413 7.4-11.8 0.11 66.7 21 21 7.92 2.13 0.27 27.9 47 San Martin 731 172 32.4 0.53 25.6 5.8 3.4 4.64 1.66 0.36 3.69 0.94 1962 659 38.0 0.25 48.4 56 56 28.8 38.4 1.33 30.5 7.6 Koell 274 71 (with trash) 2.4-3.1 0.11 27.3 52 69 3.93 59.4 15.1 7.26 0.86 1995 1030 (no trash) 5.0-7.9 0.17 24.0 0.89 0.92 3.84 0.27 0.07 1.26 2.2 a See Table 3-2 for distinction between 2005 and 2006 tracer experiments in Gustavson. bComparisons of median travel were not made because reach lengths were not comparable in this experiment or with the following 2006 Gustavson experiment. c Fmed and Fmed 200 were 4.79 E-9 and 2.62 E-9, respectively.

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97Table 3-5. Estimated P and N uptake rates based on measured a bove and belowground biomass and pl ant tissue nutrient storages at each study site San Martin Gustavson Koell Frey AboveBelowAboveBelow AboveBelowAboveBelow Biomass (g m-2) 730 172406 1941823 6551157 768 274 71322 96311 194177 45 Biomass growth (g m-2d-1)a 12 3nab20 7na 9 2na7 6na P content (mg P g-1) 11.4 3.84.6 1.010.6 1.54.5 1.5 7.2 1.13.6 0.69.0 2.54.3 1.3 N content (mg N g-1) 66.8 9.219.5 8.052.4 7.515.9 4.6 51.7 13.313.3 .160.2 18.916.6 2.9 P uptake (g P d-1)c 53 (27-88) na70 (39-108) na 10 (7-15) na19 2-44 na N uptake (g N d-1) 311 (201-383) na347 (193-535) na 75 (43-115) na125 (12-304) na aDaily biomass averaged as biomass at time of sampling divided by days since harvest, estimated as 90, 60, 30 and 30 days in Gus tavson, San Martin, Koell and Frey. bGrowth and uptake rates were not calculated for belowground biomass due to lack of remaining biomass data after plant harvestin g. c P and N uptake is calculated by the product of biomass growth, ditch area and nutrient content of plant tissues, where area was assumed to equal the channel length multiplied by average wetted channel width (See Table 3-1). For Frey the ar ea used includes the wetland area with a width of 2 m. Table 3-6. Average P and N loads and estimated gross nutrient retention by aboveground biomass assimilation at each study site San Martin Gustavson Koell Frey Mean P load (g P d-1)a 1195 924 432 162 Mean N load (g N d-1)b 3,225 12,740 6,190 2,633 Gross P Retention (%)c 4.4 7.6 2.3 11.7 Gross N Retention (%) 9.6 2.7 1.2 4.7 aMean P load is the SRP concentration multiplied by discharge using mean values from Table 3-1. bMean N load is the sum of the NO3-N and NH4-N concentration multiplied by discharge using mean values from Table 3-2. cGross P and N retention are calculated as the sum of the above and belowground P and N uptake values divided by the P and N loads.

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98 Table 3-7. Mean (SD) molar ratios of car bon, nitrogen and phosphorus of aboveground (AG) and belowground (BG) plan t tissues between sites San Martin GustavsonKoell Frey AGBG AGBGAGBGAG BG N:P (molar) 13.5 (2.1) 4.0 (0.8) 10.1 (2.4) 3.6 (0.3) 14.1 (2.1) 3.6 (0.7) 12.5 (2.6) 4.2 (1.4) C:N (molar) 22.8 (1.1) 12.9 (1.4) 25.0 (1.3) 13.9 (0.9) 37.5 (2.8) 14.5 (2.0) 26.2 (9.8) 12.6 (1.4) C:P (molar) 148.1 (18.0) 51.3 (13.4) 120.8 (21.4) 49.7 (5.4) 250.7 (43.0) 52.7 (16.4) 154.5 (62.7) 53.8 (21.4) 0 4 8 12Cynadon Ei c hornia Equ ise tu m He d yc h i u m Hy d ro c oty l e Myr i ophyllum P o l y gonumSpeciesTissue P content (mg P g-1) Leaves Stems Roots 0 25 50 75Cynadon Ei chorn i a Equ i s e tum He d y c hium Hyd ro cot y le My ri o p h yll u m PolygonumSpeciesTissue N content (mg N g-1) Figure 3-14. Comparison of leaf, stem and roo t/rhizome tissue N and P content of macrophyte species found in wastewater ditches (b ars and lines represent means 1 SD)

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99 0 4 8 12Ei ch o rn ia Hed ychiu m Hydr o cot yle M yr i o p hyllum P o lyg onumSpeciesP content (mg g-1) wastewater control 0 20 40 60Eicho rn ia Hed ych iu m Hydrocotyle M yr iop h yl lu m Polyg o nu mSpeciesN content (mg g-1) Figure 3-15. Comparison of common species in Oxapampa that are found growing in (wastewater) and out (control) of wastewat er (bars and lines represent means 1 SD, and ** denote p < 0.05 and p < 0.01) ** ** **

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100 CHAPTER 4 PHOSPHORUS RETENTION BY SORPTION WITH BENTHIC SEDIMENTS Introduction Surface waters in Oxapampa tend to have el evated phosphorus (P) concentrations due to the use of fertilizers, phosphate detergents and discharges of raw domestic effluents (See Figure 1-8). The high P concentrations have led to ra pid increases in filamentous algae cover along stream banks of the Chorobamba River over the last several years indicating that the river system is becoming increasingly eutrophic. While local water quality concerns are focused primarily on those related to human health du e to the presence of pathogens, there is also need for reducing nutrient loads to the river to avoid adverse impacts to the local fisheries upon which local indigenous communities rely. There are significant differences in phosphate concentrations observed between effluents discharged from wastewater ditches and underground pipes (s ee Chapter 1). Lower phosphate concentrations in ditches than in pipes may be caused by P sorption with benthic sediments. Understanding the physico-chemical characteristic s of benthic sediments and the nature of P adsorption is important for determining the stabil ity of P retention (Reddy et al. 1999), not only in these ditches, but also in the Chorobamba Ri ver, the receiving water-body for ditch sediment exports during storm events. The objective of this study was to evaluate the role of wastewater ditch sediments for promoting phosphate reten tion. Specific research objectives were: (1) Determine how much P is stored in benthic sedi ments and how stable it is; (2) Identify physicochemical factors responsible for P sorption; (3) Ev aluate whether ditch sediments tend to act as P sinks or sources; and (4) Examine the capaci ty of sediments for further retention.

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101 Methods Experimental Sites Sediment P characteristics were studied in four wastewater ditches in 2005 (San Martin, Gustavson, Koell and Frey) and in three ditche s in 2006 (San Martin, Gustavson, Koell). As described in Chapter 3, ditch vegetation is remove d by the municipality 2 to 3 times per year and ditch sediments are dredged approximately every two years to prevent th e risk of flooding. The ditches selected for this study are the most downs tream drains in Oxapampa; thus transport along these reaches provides the last opportunity for phosphorus (P) retention (or release) before discharge to the Chorobamba Ri ver. The downstream position of these ditches (see Figure 3-1) also suggests that they are likely to be the most impacted sites w ith the highest P loads. The sites were also chosen to provide a range of physical, biological and ch emical attributes to allow an examination of relationships betwee n ditch features and P sorption. Three to four locations were sampled in each study site (see Figure 3-1) to inspect P sorption properties along a longitudinal profile, from upstream to the most downstream point before discharge to the river. San Martin (Figure 4-1) has the highest flow s of the four study sites but because approximately a quarter of the flow comes from a natural stre am, it generally has the lowest P concentrations. Gustavson (Figure 4-2) has both high flows and high P loads. The ditch exists along four city blocks and is surrounded by homes that discharge wastes along the total length of the ditch causing discharge to increase with distance downstream. Koell (Figure 4-3) has lower flows than Gustavson and higher P concen trations than San Martin. Frey (Figure 4-4) has the lowest flows of the four study ditches and on occasion dries out. The lower reaches of Frey are often ponded creating a marsh environm ent with connection to the river only during high flow events (see Chapter 3 for more detailed site descriptions).

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102 Field Procedures Benthic sediments (0-5 cm) for each sampli ng location were collected. Using a trowel, sediments were dug out from the middle and side s of the channel and composited. A total of 15 samples (4 each in San Martin, Gustavson and Ko ell; 3 in Frey) were collected in 2005 and 12 samples in 2006 (Frey was not sampled in 2006). Gr ab samples of water were taken following the field procedures described in Chapter 3 at the time of sediment collection for the P sorption experiments (described below) and also dur ing a 13-month period (March 2005 to September 2006) to determine general water quality character istics for each site over time. Temperature, pH, dissolved oxygen and conductivity were measur ed in-situ as described in Chapter 3. Laboratory Procedures Water chemistry and sedime nt characterization Water samples were analyzed for total susp ended solids (TSS), soluble reactive P (SRP), nitrate and ammonium as described in Chapter 3. A sediment sub-sample from each sampling site was air dried for particle size determina tion, soil pH and percen t organic matter (%OM). Particle size determination for sands (0.06 2 mm) and gravels (> 2mm) were determined using nested sieves following ASTM standard methods (ASTM 1985). For determination of silt and clay fractions the pipette method was used as described by Gee and Bauder (1986). Sediment pH was determined using the method described by Thomas (1996). To approximate % OM, sediments were oven dried for 4 hours at 105C, weighed at room temperature and then combusted at 550C for 4 hours and reweighed to determine the percent loss by combustion. Phosphorus sorption index The P sorption index (PSI) provides a simple measure of the P buffering capacity of sediments (Bache and Williams 1971). High PSI values indicate that large quantities of P can be removed by the sediments without increasing the P concentration in the water at equilibrium

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103 (Klotz 1985). The PSI was determined using th e approach presented by Klotz (1985) and later modified by Haggard et al. (1999). Fresh sediment s were sieved to retain only particles < 2 mm. A 100 mL solution spiked with a P concentration (50 mg P L-1) equivalent to 1 mg P g-1 sediment (dry) was added to approximately ten grams (dry weight) of sediments. Equilibration of sediments at this high P concentration was assume d to saturate available sorption sites, thus providing a measure of maximum sorption (Red dy et al. 1995). The equilibrations for each sampling site were performed in triplicate and sh aken for 1 hour after which a 15 mL aliquot was removed, filtered and analyzed for SRP. Sedime nt sorption was determined from the difference between the P in solution before and after th e equilibration period. The sediments were dried overnight at 105C to express P sorbed per dry we ight of sediment. The PSI was calculated as X/(log10C) where X is the P sorbed per dry weight of sediment ( g P g-1) and C is the final P concentration in solution ( g P L-1). The relative role of biotic versus abiotic processes in P sorption was examined for the sediment samples collected in 2005. The same procedure was followed for determining PSI as described above except that for each site an a dditional sub-sample was analyzed after being autoclaved for 20 minutes (Haggard et al. 1999 ). The difference in so rption capacitie s between the intact and autoclaved sediments is assumed to represent the contribut ion of P sorption from biotic processes alone. Equilibrium phosphorus concentration The sediment equilibrium P concentrati on (EPCo) is the sediment porewater P concentration at which adsorption of P to sediment surfaces is in equilibrium with P release from sediments such that adsorption equals deso rption (Reddy et al. 1995). EPCo measurements provide insight into whether se diments tend to function as P si nks or sources of P to the overlying water. If the EPCo value of the sediment porewater is less than the P concentration in

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104 the overlying water, the concentration gradient results in sorption to sediments. If the EPCo value is greater than the P concentration in the wate r, the gradient is reversed and P is released to the overlying water. The EPCo is determined from the y-intercept of the linear portion of a P adsorption isotherm (Figure 4-5). To create P is otherms for each site, ditch water samples were filtered with 0.45 m membrane filters and subsequently spiked with +0.00, +0.10, +0.25, +0.50, and +2.00 mg P L-1 (Smith et al. 2005). A volume of 100 mL of spiked sample water was added to approximately 25 g wet sediments (particle si zes <2 mm) and shaken for one hour. After the equilibration period, samples were filtered and analyzed for SRP. Each equilibration was performed in triplicate for a total of 15 samp les for each sampling site (five different P concentrations, three repetitions). The sediment s for each sample were dried overnight at 105C and weighed to calculate P sorbed per dry wei ght of sediment. These values were regressed against corresponding P concentrations to determine the EPCo value for each sampling site. Phosphorus fractionation The chemical forms of P adsorbed to sediment s indicate the stability of P retention in aquatic systems. The quantities of the various fo rms of sediment P thus indicate how much P is stored and whether sediments function as long-term P sinks or sources. Th e basis of fractionation procedures is that various forms of sorbed P exhibit differential solubilities. Using various extracts, P can be solubilized and quantified. P fractions were determined by sequential extractions following the chemical fractio nation scheme described by Moore (2000): 1 M KCl to remove the most labile inorganic P pool 0.1 M NaOH to remove inorganic P sorbed by amorphous oxyhydroxides (bioavailable and readily desorbed) and crysta lline Fe/Al oxides (Fe-bound P desorbed only under anoxic conditions). This extraction step also removes organic-bound P. 0.5 M HCl to remove inorganic P bound with calcium and magnesium persulfate acid digestion to remove residua l inorganic and recalcit rant organic P.

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105 Sediments from each sampling site were co llected from the middle and sides of the channel and composited in the field (approximate ly 20 g dry weight) in pre-weighed centrifuge tubes and filled with ditch sample water just above the sediment surface to avoid sediment oxidation. The tubes were capped and returned to the laboratory where they were centrifuged for 30 minutes. The supernatant was filtered using a 0.45 m membrane filter and analyzed immediately for determination of porewate r soluble reactive phos phorus (SRP). After reweighing the tubes to determine the water re moved, sediments were homogenized using a spatula and two subsamples (approximately 1 g dry weight each) were removed to perform extractions in duplicate. Sediments were adde d to pre-weighed centrifuge tubes along with 20 mL of 1 M KCl. The tubes were shaken for 2 hours, centrifuged for 30 minutes and the supernatant was filtered and analyzed for SRP for determination of the loosely-sorbed P fraction. A similar procedure was followed for the next extraction step using 0.1 M NaOH and an equilibration time of 17 hours. However, after cen trifuging, half of the supernatant collected was filtered and analyzed for SRP for determination of Fe and Al-bound P while the other half was digested and analyzed for total P. The differ ence between TP and SRP represented the organicbound P. The next extraction step used 0.5 M KCl and a shaking time of 24 hours to remove Ca and Mg-bound P. Supernatants were analyzed fo r SRP. The remaining sediments were air dried and collected in 20 mL scintillation vials wher e they were later analyzed for residual P by persulfate digestion followed by SRP analysis. Total adsorbed P was determined by summing the various P fractions. Statistical Analyses Sediment characteristics and water quality da ta were ln-transformed to meet normality requirements. An arcsine-square root transf ormation was used for values expressed as percentages. Comparisons of sediment and P sorption characteristics from years 2005 and 2006

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106 were made using paired t-tests. An a priori significance level was set at 0.90 due to low sample sizes resulting from grouping the data into 3 sites. Results The four ditches differed from one another with respect to their channel characteristics, flowrates and water chemistry (see Table 33). Sampling of sediments in 2005 took place between November 11 and November 15 which coin cide with the beginning of the rainy season in Oxapampa. In 2006, sediments were collected between June 26 and July 11 which correspond with the latter part of th e dry season (Figure 4-6). Sediment characteristics varied between th e different ditches a nd between years 2005 and 2006 (Table 4-1). There are notable differences in the particle size distributions between years in San Martin and Koell. In 2005, there tended to be higher proportions of sands (94% and 87% in San Martin and Koell, respectively) than in 2006 (50% and 26%). A similar trend was seen in Gustavson when the second sampling site whic h was extremely mucky and had a very high percentage of clays (over 40%) was omitted from the analysis. Overall, sediments collected in 2006 had significantly higher clay (paired t-test, t = -6.6, p = 0. 01) and organic matter content (paired t-test, t= -2.18, p < 0.1) than in 2005. Average sediment pH ranged between 5.5 and 6.7 for San Martin, Gustavson and Koell. Sediment pH in Frey was 7.0. Sediment pH was lower in 2006 than in 2005 (paired t-test, t = 4.06, p = 0.03) However, these values were determined from air-dried sediments and therefore may be lowe r than pH values measured in situ due to a change in redox conditions as the sediments we re removed from an anaerobic environment (Sallade and Sims 1997a, b) Results of the single point P is otherm experiments indicate th at the sediments are capable of buffering additional P loads, particularly in the 2006 sampling period (Table 4-2). Koell tended to have the highest overall PSI values of the three sites in 2006; however differences in

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107 PSI values between sites were not significant (p > 0.10) due to high within-site variability (Figures 4-7 and 4-8). Comparison of the longitudinal trends observe d for PSI and %OM at each sampling site (Figures 4-7 and 4-8) suggests a positive associ ation between PSI and %OM, especially for the San Martin and Koell sites. Gustavson sediment s appear to deviate from this relationship suggesting that controls, other than OM c ontent govern P buffer cap acity at this site. Nevertheless, strong relationships exist between particle size, OM content a nd K, an alternative measure of P buffer capacity, when the data for a ll sites and years are pool ed together (Figure 49). Despite the relationships obser ved between particle size and OM content, comparisons of the contributions of abiotic a nd biotic sorption performed in 2005 suggest that biotic sorption largely accounts for P buffer capacity rather than abiotic factors (Figure 4-10). Interpretation of this result is likely to be misleading due to the use of autoclaved sediments in the PSI experiment. Inhibition of microbi al activity via autoclaving sedi ments has been shown to result in an underestimation of the importance of abioti c sorption (Klotz 1985). This issue is described in more detail in the discussion section below. Comparison of ditch water SRP and sediment EP Co values (Figure 4-11) suggest that at the time of sampling San Martin, Koell and Fr ey sediments tended to remove SRP from the water column, thus acting as a P sink. In 2005, three of the 4 sampling points in Gustavson had higher sediment EPCo values relative to ditc h water SRP, indicating that sediments were functioning as a source of P to overlying water. In 2006, one s ite in both Gustavson and San Martin functioned as a P source while the remaini ng three sites were sinks for P. Sediments in Koell appeared to act as a P sink in both 2005 and 2006. No strong trends were found between

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108 EPCo and sediment particle size or %OM. Howeve r, increases in sediment EPCo values were observed to be related to wate r column SRP values; this tren d was strengthened by grouping the data into sites that act as sinks and those that act as sources (Figure 4-12). Moreover, EPCo values appear to be indirectly related to par ticle sizes and OM content due to the relationship between PSI and EPCo (Figure 4-13). As the sedi ment P buffer capacity decreases, EPCo values tended to increase. Phosphorus fractionation performed on San Ma rtin, Gustavson and Koell sediments in 2006 reveal marked differences in the distribution of sorbed P fo rms despite the close proximity of the study sites (Figure 4-14). The readily labile P pool (KCl-P i) comprised an insignificant fraction of the total P ( 1% for all sites). Sorption with Fe/Al oxyhydroxides made up the largest P fraction for San Martin (44.6%) and Gustavso n (74.6%) and the second largest fraction for Koell (31.6%). Most likely, the readily labile P fraction in situ was larger but oxidation of sediments resulted in this fraction becoming adso rbed to freshly oxidized ferric surfaces (Moore 2000). Organic-bound P was a significan t fraction for San Martin (37.0%) and Koell (38.3%) but not for Gustavson (4.3%). Ca and Mg-bound P was an important P fraction in Gustavson (17.0%) and Koell (17.3%) but not for San Mart in (0.06%). The recalcitrant P fraction, representing a long-term P sink was 18.3% and 12 .8% in San Martin and Koell, respectively and was markedly lower in Gustavson (3.3%). The P fractions were summed to estimate the total sediment P for each site (Figure 4-15). This total is an underestimate due to the absen ce of the HCl-P organic fraction. Gustavson had approximately 10 times more adsorbed P than San Martin and Koell; however, all sites exhibited very high P concentrations (site means between 2000 and 22,000 mg P kg-1). Longitudinal trends

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109 of TP, from upstream to downs tream, illustrate high variability with distance along each ditch site (Figure 4-16). Discussion Effect of Seasonal Trends in Particle Si ze and Organic Matter Content on P Buffering Capacity The significant difference in particle-size distributions and organic matter content in November 2005 and July 2006 in all study ditches reveals the dynamic nature of these systems through time and space. Heavy rains between October and April affect ditch sediments characteristics by exporting fine particles and organic matter and depositing coarser materials resulting in higher sand fractions observed in each of the sampling si tes in November. Ditch management practices likely exacerbate sediment exports by controlling vegetation towards the beginning of the rainy season to avoid potential flooding. In contrast, sediment erosion and export is markedly decreased during the dry seas on which is characterized by infrequent and less intense rain events. Hence, during this period, ditch vegetation is minimally managed due to the low risk of flooding. The resulting proliferation of ditch vegetation likely promotes deposition of fine sediments and the accumulation of orga nic matter as observed in the 2006 sampling. Changes in particle-size and organic matter co ntent between the wet and dry season appear to have directly influenced th e overall P retention capacity of th e ditch sediments. Not only are buffer capacities significantly higher in the dry season but EPCo values were much lower, indicating that sediments will tend to remove P rather than release P from the water column. Strong relationships between fine particle sizes and organic matte r content on P buffering capacity have been reported by many studies (e.g. Meyer 1979, Nguyen et al. 1997, Axt and Walbridge 1999). By comparing these relationships between two distinct seasons, this study has

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110 shown that the operating mode fo r P retention in wastewater d itches may be highly variable through time in response to altera tions in sediment characteristics. Phosphorus Sorption Index To understand how the buffering capacity of the wastewater ditches compares with that reported for other systems, PSI values from 2005 and 2006 were tabulated with PSI values observed for other systems including wetlands and streams (Table 4-3). This comparison revealed that sediment PSI valu es in 2005 were similar to values reported for other systems and in 2006, values were within the upper limits or exceeded reported values. The PSI values from this study were most similar to other enrich ed drainage ditch systems and wetlands. The similarity of high P buffering capacity between ditches and wetlands supports the idea that it may be possible to manage drainage ditches as wetlands for mitigation of nutrient loads to downstream receiving bodies (Bowmer et al. 1994, Nguyen and Sukias 2002). Summarizing PSI across multiple studies also brought to light several issues regarding both the reporting of PSI valu es and PSI methodologies that affect the interpretation and comparison of results. The first of these issues to be addressed is the units used to report PSI values. As illustrated by Table4-3 the choice of units us ed to report PSI values strongly affects the range of values reported. Due to the formulation of the PSI equation [ PSI = X (log C)-1], some studies likely report values for C using g L-1 or mol L-1 instead of mg L-1 to avoid taking the log of a value less than or equal to 1. However, the use of varying units for C may preclude direct comparison between studies unless other data is provided to allow unit conversions of the (log C)-1 factor. As a result of the inability to directly convert PSI values from other studies to one common set of units, PSI values from this study were instead conver ted to three different sets of frequently reported units in Table 4-3

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111 The second difficulty in comparing PSI values between systems is related to the differing methodologies used to measure PSI. Table 4-3 provides a general comparison of methodological approaches used in each study that may signi ficantly influence the values reported. These approaches include choice of sediment-water equ ilibration time, the initial P concentration used to spike sediments/soils, and inhibition of microbial activity. The choice of sediment-water equilibration time is an important factor that varied between the studies included in Table 4-3. Most P sorpti on isotherm studies maintain sediment-water contact times between 16-24 hours to ensure maximum soil-water contact. The present study used a one-hour equilibration time ( Klotz 1985, Haggard et al 1999) due to logistical difficulties associated with the remoteness of the study area (i.e. acquiring a working shaker table). Thus, sorption measurements in the presen t study represent an underestimation of the total sorption potential of the ditch sediments. Howeve r, the magnitude of the difference in sorption between a 1-hour and a 24-hour equilibration time may be strongly related to soil texture. Meyer (1979) noted that for sorption experiments using s ilty sediments, 93% of the total P in solution was removed in the first five minutes. Using sa ndy sediments, only 19% was removed in five minutes; however, after 24 hours, approximately the same amount of P had been sorbed onto both the silty and sandy sediments. These findings suggest that th e underestimation of sorbed P in the present study may be particularly significan t for the sandier sediments that characterized the study sites in 2005 and perhaps less so for th e finer sediments from 2006. This finding may have some bearing on the low PSI values reported by Haggard et al. (199 9) and (2001) since the systems in both studies were dominated by sandy se diments. It is worthy to note that despite the one-hour equilibration time and the dominance of sandy sediments in 2005, the wastewater ditches in the present stu dy exhibited some of the highest PSI values reported.

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112 Despite the likely underestimations of to tal P sorption capacity, the sorption tests performed in this study provide invaluable insight into the rate of P sorption and how this factor is influenced by changes in sediment particle-size distribution. The importance of the ability to quickly retain P is especially highlighted in syst ems that are subject to short-term, concentrated pulses. Pulses of high P loads associated with human activities, such as laundry washing with phosphate detergents, are common in wastewater drainage ditches in Peru. Therefore, an understanding of the sediment response to such flas hy perturbations is likely to be as important as knowing the total sorption capacity when exam ining the environmental fate of P in ditch systems. A second factor that is likely to influence PSI values is the co ncentration of P used to spike the sediments/soil porewater. The concentrations us ed in the studies report ed in Table 4-4 varied between 2 mg L-1 and 1000 mg L-1 and not surprisingly, the studies using the lowest concentrations also tended to have the lowest PSI values. To make PSI values comparable across systems, the same initial P concentration should be used. A final important difference between the studies reported in Table 4-3 is whether or not the study inhibited microbial activity before perfor ming the PSI analysis. The choice of microbial inhibition is strictly a matter of the research goals of the particular study. Some studies are interested in examining total P sorption capacity of the sediments while others focus solely on abiotic sorption and to do so, microbial activity is inhibited by autoclaving sediments or adding a toxic chemical such as toluene, chloroform, or carbonylcyanide m -cholorophenylhydrozone (CP) before equilibration. The effect of microbial inhibition on comp aring PSI values for various systems is a factor of the importance of biotic sorption to overall sorption capacity. Results of PSI values in this study in 2005 i ndicated that biotic sorption tended to dominate over abiotic

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113 sorption. Similar findings were reported by Hagg ard et al. (1999). Other studies have reported much lower contributions of biotic sorption to total sorption (Meyer 1979, Klotz 1985). It would have been useful to compare biotic versus abiotic PSI values for 2005 and 2006. The relative influence of abiotic sorption would likely to be greater in 2006 due to th e higher percentage of silts, clays and organic matter in ditch sediments. However, it is extremely likely that abiotic sorption was significantly underestimated in the ditch sediments in 2005 because autoclaving sedime nts releases cellular P (Klotz 1985). In two of the sampling sites in Koell, autoclaved sediment s, performed in triplicate, resulted in negative PSI values (note the error bars for abiotic sorptio n in Figure 4-11). In other words, more P was in solution after the equilibration period than was in the initial spiked P solution thus implying that ditch sediments released P duri ng the sorption experiment. These findings support that the use of autoclaved sediments will likely underestimate the importance of abiotic sorption and will thus overestimate the importance of biotic sorption (e.g. this study, Haggard et al. 1999). The PSI is a valuable index for describing the P buffer capacity of a system. However, to optimize its utility, especially for cross-study co mparisons, standardization of units used for reporting values, equilibration times and initial P concentrations should be adopted. The use of the PSI to further investigate the rate of P uptake for different systems and sediment textures would be useful for understanding how well sedi ments in lotic systems can buffer pulses of P loading. Sediment Equilibrium Phosphorus Concentrations Sediments from the Gustavson ditch tended to have the highest EPCo values and were furthest from aqueous P-sediment P equilibrium, represented by th e 1:1 line in Figure 4-12. This departure from equilibrium suggests that sediments exert little control over regulating solution P concentrations in the system. This finding is co ntrary to studies perfor med in pristine systems

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114 (Froelich 1988) but similar to studies from imp acted systems (Ekka et al. 2006). The regression between SRP and EPCo was strengthened by separati ng out the ditch sites th at were shown to act as sinks from those that act as sources as shown in Figure 4-12. The slopes of the EPCo versus SRP regression differ between the two groups with the slope of the P sink sites being notably less than that for P source sites. This differen ce suggests that an underlyi ng factor distinguishing the two groups is their P buffer capacities. A lowe r slope indicates that EPCo changes less given an increase in water SRP. Conversely, the steep slope shown for the P sour ce sites suggests that EPCo is sensitive to changes in SRP, thereby exhibiting a more limited P buffer capacity. The negative power law relationship between EPCo and PSI supports this hypothesis (Figure 4-13). The two P source sites in 2006 were Gustavson site 1 and San Martin site 4 which correspond to the lowest PSI values for each of these sites as shown in Figure 4-8. Given the relationship between EPCo and PSI, it is not surprising that PSI and sediment EPCo values changed significantly between the 2005 and 2006 sampling periods. This finding further supports that sediment P sorption charact eristics are extremely variable through time. This variability suggests that the common prac tice of comparing water SRP over time with a single EPCo measurement to examine seasonal or yearly sorption/desorption trends may not be appropriate for dynamic systems such as drainage ditches. Phosphorus Fractionation to Examin e P Status of Ditch Sediments Phosphorus fractionation results indicate that despite the close proximities of the San Martin, Gustavson and Koell sites, they neverthele ss reveal quite different distributions of labile and non-labile pools of P. Gustavson sediments had the highest TP concentrations (19,227 mg kg-1) and also the highest levels of bioava ilable P due to non-occluded Fe/Al oxyhydroxides (14,317 mg kg-1). Comparably high TP values (25,261 mg kg-1) were reported by Nguyen (2000) for sewage-impacted wetland sediments in New Zealand; however the dominant P fraction was

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115 carbonate-P (15,121 mg kg-1), although the non-occluded Fe/A l-bound P fraction was also high (2,342 mg kg-1). Near saturation of Gustavson sediments may explain the devia tion of this site from the trend between PSI values and OM c ontent found for the other sites. The PSI of Gustavson ditch sediments (site 4) declined to ne arly half that of the upstream sampling location despite having higher percent OM. This decline in buffer capacity may be due to an increase in sediment TP (Figure 4-16) that caused sorption sites to become more limited. While the most readily bioavailable fraction (loosely sorbed KCl-P) only represented 1% of the sediment TP in Gustavson, it nevertheless was high (185 mg kg-1) relative to the San Martin and Koell sites (<1 mg kg-1). One possible explanation for this difference is that San Martin and Koell sediments were oxidized either during sediment collection or during the first extraction step such that porewat er and loosely sorbed P reacted with fresh ferric oxides surfaces. However, handling of sediments wa s carried out in an identical manner for all sites so it is unlikely that Gustavson sediments remained an aerobic during the first extraction step. Most likely, sediments from all sites were equally ox idized. Ferrous oxides have much greater surface areas available for P sorption th an those for ferric oxides (Reddy et al. 1999). Perhaps extremely high concentrations of P associat ed with ferrous oxides saturate d newly-available ferric oxide surfaces once Gustavson sediments were exposed to the atmosphere. The P remaining after saturation of ferric oxide surfaces would then have been detected as loosely exchangeable P. While Fe and Al were not specifically measured in the study, observation of the filters used to measure SRP after the NaOH extraction step i ndicate the presence of oxidized iron (Figure 417). While ferrous oxide surfaces can retain more P via sorption, the low binding energy of this reaction results in a high potenti al energy for re-release to the water column (Reddy et al. 1999).

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116 Less binding surface areas are av ailable under aerobic conditions but sorption is much more stable. Therefore, anaerobic sediments high in ir on content will tend to have high EPCo values, which was indeed the case for the Gustavson sites. The sediments from Koell and San Martin al so had high fractions of Fe/Al-P pools but contained high percentages of organic P associ ated with humic and fulvic acids. Organic P bound with humic substances may be readily re -released to the water column under aerobic conditions; however, under anaerobi c conditions this fraction ha s been shown to be more resistant to biological breakdown (Reddy et al. 1995, Reddy and D'Angelo 1997). Therefore, this fraction in the anaerobic ditch sediments may be considered a possible P sink. Approximately 17% of TP pools in both Ko ell and Gustavson consisted of Ca/Mg-bound P. This finding was somewhat surprising given th e low sediment pH values observed. Low pH values tend to be associated with the presence of iron and aluminum whereas higher sediment pH indicates the presence of cal cium and magnesium. Limestone outcrops exist in the surrounding slopes and may be draining to the valley. However, precipitation of P with Ca occurs when pH exceeds 8. Measurements of water pH at midday wh en values are expected to be highest due to photosynthetic activity revealed pH values ra rely exceeding 7.2. Furthermore, hardness calculations (data not shown) i ndicated a higher presence of ma gnesium than calcium. A more likely explanation of this P pool is fertili zation of fields using commercial calcium superphosphates. The levels of labile P, most often defined as KCl-Pi + NaOH-Pi (Sallade and Sims 1997a), signify the stability of sediment P retention. P fr action analysis of the wastewater drainage ditch sediments indicates that there is extremely hi gh potential for release and biological uptake,

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117 especially in Gustavson where labile P repr esented 74% of sediment TP. The potential for mobilizing P in sediments was also high for Koe ll and San Martin (34 and 33%, respectively). Management Implications Anaerobic conditions found in wastewater drai nage ditches limit the potential for stable Fe-bound P. However, researchers have shown th at complexation of ferric oxides with humic substances may protect ag ainst solubilization of Fe3+ to Fe2+ under reducing conditions (Reddy et al. 1999, Nguyen and Sukias 2002). Promoting the accumulation of organic matter in drainage ditches via proper management of vegetation ma y serve to increase the potential for this mechanism to occur. For low oxygen waters, such as those in the current study, P retention may be enhanced by promoting aeration via structural modifications such as installing riffle zones. Conclusions This study demonstrated the dynamic nature of wa stewater drainage ditc hes with respect to sediment characteristics and a ssociated phosphorus sorption capacity. High P loads to the ditch systems regulate the equilibrium between sediment porewaters and aqueous SRP, not the benthic sediments. Sediment retention of P was larg ely regulated by sorpti on with Fe/Al oxyhydroxides which for Fe-bound P is highly unstable in these anaerobic systems. N onetheless, sediments undoubtedly play an important role in providing a temporary storage of P that may dampen the timing and magnitude of P exports to the Chorobamba River. This buffering mechanism appears to be greatest during the dry season when ditc hes tend to trap and accumulate finer sediments and organic matter. The timing of this increased buffering capacity is likely to be important for mitigating P loads to the Chorobamba River at a ti me when flowrates are lowest and the river is most vulnerable to eutrophication. PSI values determined using a one-hour equi libration time demonstrated the capacity of the ditch sediments to rapidly remove high c oncentrations of aqueous P, an important

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118 consideration for these flashy sy stems with limited residence times. The findings from this study suggest that the potential exists for managing dr ainage ditches as wetlands for mitigation of P loads to receiving bodies. Key to this approach is likely to be the proper management of vegetation in order to promote sedimentation and organic matter accumulation. Modifying the ditches to increase aeration may he lp to take advantage of the large potential for P sorption with ferric oxides. Figure 4-1. Looking upstream from sampling site 4 in San Martin in A) November 2005 and B) July 2006 A B

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119 Figure 4-2. Looking downstream from sampling s ite 1 in Gustavson in A) November 2005 and B) July 2006 Figure 4-3. Looking downstream from sampling site 1 in Koell in A) November 2005 and B) July 2006 B A A B

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120 Figure 4-4. Sampling sites in Fr ey (only sampled in 2005). A) looking downstream from site 1 and B) looking upstream at site 2, the marsh site after being ditched by the city. Note that despite ditching at site 2 the water re mains stagnant. (Photos taken at the time of sampling in November 2005). y = 12.911x 2.4939 R2 = 0.9963 -4.00 -2.00 0.00 2.00 4.00 6.00 8.00 10.00 0.000.200.400.600.801.00 Initial SRP in solution (mg P/l)P sorbed (mg P/kg DW) Figure 4-5. Example of the approach used to determine the sediment EPCo value and K, a measure of P buffering capacity, from P sorption isotherms. y-intercept = EPCo Slope = K (P buffering capacity) A B

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121 0 20 40 609/ 1/ 2 0 0 5 1 0/ 1 / 2 00 5 1 1/ 1 / 2 00 5 1 2 / 1 / 2 0 0 5 1/ 1/ 2 0 0 6 2 / 1 / 2 0 0 6 3 / 1 / 2 0 0 6 4/ 1/ 2 0 0 6 5 / 1 / 2 0 0 6 6 / 1 / 2 0 0 6 7/ 1/ 2 0 0 6 8/ 1/ 2 0 0 6 9 / 1/ 2 0 0 6Rainfall (mm) Figure 4-6. Rainfall in Oxapampa from Sept ember 2005 September 2006. Sediment sampling in 2005 took place during the rainy seas on while in 2006 samples were collected towards the end of the dry season. Rainfa ll data was provided by the Andean Amazon Research Station and ProPachitea. 2005 sampling days 2006 sampling days

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122Table 4-1. Site averages of particle size distributions and sediment pH for the wast ewater ditches in October-November 2005 and June-August 2006 San Martin Gustavson Koell Frey 2005 2006 2005 2006 2005 2006 2005 2006 % sand 94.4 50.2 73.7 (83.7)a76.287.126.055.1 n.d.b % silt 4.1 34.4 13.3 (12.4)9.910.849.129.9n.d. % clay 1.4 15.4 13.0 (3.9)13.92.124.915.1n.d. % OM 1.9 3.9 3.1 (3.0)4.72.913.27.4n.d. pH 6.0 5.6 6.76.06.55.57.0n.d. aValues in parentheses are averages without Gustavson sampling site 2 which had extremel y high clays (40%) that bias the overall average for the ditch. bn.d.= no data. Frey was not sampled in 2006. Table 4-2. Comparison of P sorption measures determined in each site in years 2005 and 2006 San Martin Gustavson Koell Freya 2005 2006 2005 2006 2005 2006 2005 PSI (L g-1) 22.9b (21.7-24.2) 98.5 (66.2-173) 17.8 (10.2-25.4) 181.4 (88.1-303) 22.6 (10.2-29.7) 233 (174-342) 44.9 (32.5-62.7) Smax (mg P kg-1) 75.8 (73.0-78.6) 306.0 (260.0-417.6) 60.8 (38.0-92.6) 577.3 (358.9-811.5) 72.1 (38.5-92.6) 794 (613-1124) 84.8 (78.2-97.0) EPCo (mg P L-1) 0.29 (0.28-0.29) 0.08 (0.05-0.14) 1.43 (0.51-2.31) 0.18 (0.06-0.28) 0.4 (0.23-0.55) 0.09 (0.06-0.12) 0.5 (0.3-0.6) K (kg-1) 2.4 (2.3-2.6) 7.4 (5.5-8.9) 3.7 (2.5-4.8) 8.2 (4.6-13.6) 3.9 (2.8-6.0) 15.8 (10.0-21.7) 3.10 (2.6-4.2) PSI = phosphorus sorption index; Smax = P sorption maximum after 1 hour; EPCo = equilibrium P concentration; K = P buffer capac ity aFrey was sampled in 2005 only. bValues are averages. Numbers in parentheses are ranges of observed values.

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123 0 100 200 300 400 1234 Sampling location (upstream to downstream) for each sitePSI (L kg-1) San Martin Gustavson Koell Figure 4-7. Examination of longitudinal variat ion in 2006 PSI values for each ditch from upstream (1) to downstream (4) sampling locations. Values are means (n = 3) and bars represent SD. 0 5 10 15 20 1234 Sampling location ( upstream to downstream) for each sitePercent OM (%) San Martin Gustavson Koell Figure 4-8. Longitudinal trends of percent organic matter in 2006 for each ditch from upstream () to downstream () sampling locations.

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124 Figure 4-9. Relationships between K (a measur e of P buffer capacity determined from EPCo experiments) and percentage of sand, silt, clay and organic matter for all sites. 0 15 30 45 San MartinGustavsonKoellFrey SitePSI (L kg-1) Biotic Abiotic Figure 4-10. Contributions of bi otic and abiotic sorption for p hosphorus sorption indices (PSI) determined for sediments in 2005. The PSI determined from autoclaved sediments represents abiotic sorption. The difference in PSI for intact and autoclaved sediments represents biotic sorption. The sum of the ba rs for each site thus represents total PSI. Values are means (n=3) a nd bars represent 1 SD.

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125 Figure 4-11. Comparison of ditch water SRP co ncentrations with sediment equilibrium phosphorus concentration (EPCo) experiment s. Values plotted above the 1:1 line indicate that P is removed from the water column and sorbed to sediments. Values below the 1:1 line indicate sites where P is likely desorbed from sediments and released to the overlying water. Letters de note sites: S = San Martin, G = Gustavson, K= Koell, F = Frey. Blue, underlined letter s indicate EPCo analyses performed in 2005 and red letters represen t analyses from 2006. 0.00.51.01.52.02.5 0.0 0.4 0.8 1.2 1.6G G G G K K K K F F F S S G G G G K K K K S S S S Sediment EPCo (mg/l)P release from sedimentsSediment EPCo (mg/l) Sediment EPCo (mg/l)Blue, underlined = 2005 Red = 2006P release from sediments P release from sedimentsSRP concentration (mg/l)S = San Martin G = Gustavson K = Koell F = Frey1:1 line P sorption to sediments Sediment EPCo (mg L-1) SRP concentration ( m g L-1 )

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126 y = 0.4678x 0.0027 R2 = 0.7162 y = 1.3722x + 0.332 R2 = 0.6322 0.00 0.60 1.20 1.80 2.40 0.000.501.001.502.00 Ditch water SRP (mg P L-1)Sediment EPCo (mg P L-1) Figure 4-12. Sediment EPCo values as a functio n of water column SRP concentrations grouped by whether sampling site acted as a P sink or source. y = 28.473x-0.6395R2 = 0.38641 10 100 1000 0.000.501.001.502.002.50 EPCo (mg L-1)PSI Figure 4-13. Relationship between P sorption inde x and sediment EPCo indicates that highest EPCo values (and thus greater potential to act as P sources) tend to be associated with lowest P buffering capacity. P sources P sinks

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127 San Martin 0.1% 44.6% 18.3% 37.0% 0.0% Gustavson 17.0% 74.5% 3.3% 1.0% 4.4% Koell 17.3% 31.5% 12.8% 0.0% 38.3% Readily bioavailable (KCl-Pi) Fe/Al bound P (NaOH-Pi) Bound w/ humic substances (NaOH-Po) Ca/Mg bound P (HCl-Pi) Recalcitrant Figure 4-14.Percentages of sediment pools of so rbed phosphorus for each site. Values presented are averages of extractions performed on duplicate sediment samples from each sampling location for each site.

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128 19227 2171 5265 0.E+00 1.E+04 2.E+04 3.E+04 San MartinGustavsonKoell SiteSediment total P (mg kg-1) Figure 4-15. Average total sediment P for San Ma rtin, Gustavson and Koell calculated from the sum of the P fractions for each ditch (b ars represent SD about the mean). 0 5000 10000 15000 20000 25000 1234 Sampling location (upstream to downstream) for each siteSediment TP (mg kg-1) San Martin Gustavson Koell Figure 4-16. Average sediment TP for each sampli ng location (1 = upstream, 4 = downstream) at San Martin, Gustavson and Koell.

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129Table 4-3. Phosphorus sorption index (PSI) va lues for wastewater ditches calculated using three differ ent sets of units to comp are with values reported for ditches, streams and wetlands System type PSI X (log C)-1 PSI units Spiked P concentration (mg P l-1) Equilibration time (hr) Microbial activity inhibited? Reference Headwater stream New Hampshire, USA comparing silty and sandy sediments 10.3 and 2.1 2 1 No (Meyer 1979) 4 streams, New York, USA 1.89-12.32 2 1 No (Klotz 1985) 3 agriculturally influenced streams, Oklahoma, USA ~ 3.5-5.5** 2 1 No (Haggard et al. 1999) 5th order stream, Arkansas, USA ~1.5-2** 2 1 No (Haggard et al. 2001) Wastewater drainage ditches, Peru 2005/2006 17.8-44.9/ 98.5-233.7 X = g g-1 C = g l-1 50 1 No This study 26 agricultural drainage ditches, New Zealand 931-2328 1000 16 No (Nguyen and Sukias 2002) 17 agricultural drainage ditches, Delaware, USA 900 15 18 Yes (Sallade and Sims 1997a) 2 riparian wetlands, North Carolina, USA 1664 and 858* 130 24 Yes (Bruland and Richardson 2004) 15 isolated herbaceous wetlands, Minnesota, USA 1012* 130 24 Yes (Bruland and Richardson 2006) 10 streams and 9 wetlands, Florida, USA 82.5 and 173.4 100 24 No (Dunne et al. 2006) Wastewater drainage ditches, Peru 2005/2006 88-171/ 319-1818 X = mg kg-1 C = mg l-1 50 1 No This study 8 early and 8 late successional mitigated wetlands, Kentucky, USA 0.13 and 0.59 9.3 (0.3mmol l-1) 48 No (D'Angelo 2005) 6 palustrine wetlands, Virginia, USA ~40** 130 24 No (Axt and Walbridge 1999) Wastewater drainage ditches, Peru 2005/2006 2.7-3.7/ 12.5-40.8 mg (100 g)-1 mol l-1 50 1 No This study Note: Values are reported as means or ranges of means if multiple groups were compared. If PSI for multiple depths were conside red, only PSI values from the upper-most sediments are reported. *PSI values converted from units originally reported as X = mg (100 g)-1.**Estimated visually from diagram.

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130 Figure 4-17. Gustavson sediments collected on filt ers after the NaOH extraction step indicate the presence of iron oxides.

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131 CHAPTER 5 EVALUATION OF CHANNEL MODIFICATIONS TO EXISTING DITCHES FOR IMPROVED WATER QUALITY AND AESTHETICS Introduction Conventional, centralized wastewater treatment systems are costly to construct and manage and consequently are often not practical realit ies for many communities in developing countries (CEPIS 2002). Decentralized, small-scale systems that utilize technologies appropriate to the regions level of resource availability are necessary for providing better management of wastewaters (Ho 2003). The municipality of Ox apampa has plans for the construction of a stabilization pond (Municipalidad Provincial de Ox apampa 2003) but shortages of funds prevent the realization of the project. Several stabil ization ponds have been constructed in other municipalities in the region over the last several years but have notoriously failed due to undersizing of the systems for combined sewer and stor mwater flows and by the lack of technical and financial resources necessary for properly mana ging the large, centralized treatment operations (Arteaga, personal communication). Currently, Oxapampa is moving towards repl acing wastewater di tches with underground pipes in order to reduce human contact with wa stewater and to improve the citys image to promote greater tourism to the ar ea. However, even replacement of ditches with pipes is costprohibitive. The total cost (mater ials plus labor) of installing pipe is approximately $37.50 per meter (Arteaga, personal communi cation). To replace the existing 18 km of ditches with pipes would therefore cost $675,000. In contrast, the cu rrent cost to manage ditches is $375 per 4,000 meters or about $0.10 per meter (Arteaga, person al communication).The pipe networks function as combined sewer systems but are undersized for high flow events, oc casionally resulting in ruptures. The pipes are also damaged by heavy trucks due to the shallow depths at which they are installed due to high water tables (Arteaga, personal communication).

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132 The first objective of this study was to implem ent simple channel modifications to three existing wastewater ditches (Gustavson, Koell and Frey) to test whether the modifications served to improve water quality. The second objective wa s to promote community participation in the project to evaluate whether modified ditches could function as a viable, decentralized water treatment approach. Ditches are often located in the middle of roads in order to receive wastewater discharged fr om homes on either side of the street. Due to their central position, ditches are thus a unifying feat ure of each neighborhood. However, ditches are regarded with much disdain as they are perceived to mar the vi sual and olfactory aesthetics of the community and symbolize a lack of economic progress. Th erefore, to promote community support, it was also important to emphasize improvements to the aesthetics of the ditch systems. Methods Government and Community Involvement The support and participation of local residents was deemed to be a critical component to the success of the channel modifications. Actions we re taken at various stages of the project to engage and inform the public and local governme nt directly about th e goals and potential benefits of ditch modifications (Table 5-1). Th e first stage consisted of activities to assess whether support for modified ditches existed or could be fostered through dialogs with stakeholders through presentations to governme nt officials and local NGOs (Figure 5-1), dissemination of interpretive folios describing th e purpose of studying modified ditches (Figure H-1), and meeting with residents to determine how to implement the modifications to best meet their needs. The second stage involved participation of re sidents living along each of the study ditches (Gustavson, Koell and Frey) during implementation of modifications. Participants were paid the standard local wage of $5 per day, were prov ided with boots, gloves and equipment such as

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133 shovels, hoes and machetes (Figure 5-2). Signs we re painted by a local ar tist asking citizens not to dump trash into the ditches and announcing the existence of the commun ity-based project for modified ditches (Figure H-2). Rugged wooden pl anks or bundles of sticks had been used previously as bridges to cross ditches when meeting with neighbor s and had resulted in several children falling into ditches and being injured. Th erefore, new sturdier br idges were constructed in each of the ditches as a safety measure, to increase neighborhood aesthetics and to foster a greater sense of community among residents. The final stage of activities to promote comm unity participation of the modified ditches consisted of capacity-building for participatory management of ditches. A non-technical guide was created and disseminated to re sidents and government official s that explained the purpose of the project and provided instru ctions and images detailing how the modifications were performed, costs of modifications and how to ma nage the modified channels. Residents assisted in the construction of compost bins for accumu lated sediments and plants from the ditches (described below). Finally, residents were pr ovided with tools to simplify ditch management such as cleaning screens and collecting plan ts and trash (also de scribed below). Modification Designs Criteria for modification approaches were low cost, use of only locally available materials and plants, and an aesthetically pleasing design. All three study d itches were first widened to a minimal width of 1.5 m to increase width-to-dep th ratios. Channel banks were reinforced with large cobbles and boulders to prevent undercut ting and bank slumping. A sedimentation basin was dug at the most upstream location of each study ditch. The three neighborhood groups were opposed to the use of common ditch species su ch as papyrus and ginger lily which were perceived as noxious weeds. Therefore, locally-ava ilable water tolerant ornamental plants such as canna lily ( Canna flacida ), calla lily ( Zantedeschia aethiopica ) and yellow iris ( Iris

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134 pseudacorus ) were used predominantly in all three study ditches. The fast-growing rhizomes of the canna, calla and iris plants we re also expected to stabilize ba nks. Canna and calla lilies were transplanted from other wastewat er ditches (Figure 5-3) while ir ises could only be found at the market as seeds. The different flowrates of the study ditches pr ompted the use of two separate designs for channel modifications: an open-water flow design for the high water flows characteristic of the Gustavson and Koell ditches and an alternati ng subsurface flow/open-water design for the low flow ditch of Frey. Open-water flow design The primary objectives of the open-water flow design were to extend channel residence time and increase dissolved oxygen levels. The openwater flow design consisted of four main design elements: a sedimentation basin, riffles and cascades, gravel bars for creating channel sinuosity and the introduction of or namental plants (Table 5-2). The sedimentation basins were oval-shaped (a pproximately 2 m in length, 1.5 m in width and 0.8 m deep). The width of the sedimentation basin in all sites was limited to avoid blockage of the road to traffic on either side of the ditc h and basin depth was limited by a thick clay layer. The overall volume of the basi n had to be limited to avoid water from overcoming the impoundment structure which consis ted of wood boards reinforced with stakes, cobbles and clay. Limited resources prevented the basin from being constructed with a more robust structure made of reinforced concrete. Water exited the basin through a 1 cm ga p between two horizontal boards and made contact with rocks in the cha nnel below to induce water oxygena tion (Figures 5-4 and 5-5). To avoid clogging of the exit and to improve sediment removal in the basin that would otherwise be limited by low residence times, the basin exit was f itted with two sizes of screens: a coarse (0.5

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135 cm) mesh followed by a fine (0.5 mm) mesh. Th e basins were planted with water hyacinths ( Eichornia crassipes ) that were transplanted from the marsh at the Frey site. The second design element used in the open-wa ter flow designs in Gustavson and Koell were riffles and cascades to increase aeration of the water and to promote greater transient storage through creation of recirc ulation zones and pools upstream of the riffles (Figure 5-6). Lateral gravel bars were crea ted to increase channe l sinuosity and complexity (Figure 5-7). The bars were intended to allow water to pass through the gravel inters tices, thus promoting physical and biochemical filtration and denitrifica tion. The gravel bars and channel banks were planted primarily with canna lilie s, calla lilies, and ir ises among other attractive, water-tolerant plants common to the area. The plants were gr own hydroponically in the gravel bars to increase available reactive surface areas of subsurface flowpaths. Along banks, small holes were dug close to the water surface to al low plant roots to make direct contact with ditch water. Subsurface flow design The second modification approach, the alterna ting subsurface flow/open water design, was applied to the ditch in Frey. The primary desi gn elements used in the approach included a sedimentation basin, planted gravel beds for subsurface flowpaths, open water sections planted with water hyacinth (Table 5-3). The sedimentation basin in Frey followed th e same design approach as that described above for Koell and Gustavson except that it was smaller (1.5 m length, 1 m width, and 0.6 m depth). After passing through the screens (same mesh sizes as described above) and exiting the sedimentation basin, water entered the first subs urface flow section. In all, there were three subsurface flow sections and two open water secti ons (Figure 5-8). The lengths of the subsurface flow and open water sections were dictated by th e pre-existing locations of pipes that discharge wastes from homes on either side of the ditch. Open water zones existed where wastes enter and

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136 subsurface flow reaches were located in between waste discharge points (Figure 5-9). The slow water velocities in the deep (0.5 m), open water ar eas trap the solids brought in by pipes. Before entering subsurface flow sections, water from th e open water flow zones first pass through a fine mesh (0.5 mm) screen to prevent clogging of gravel interstices. The sedimentation basin of Frey allowed flow s to be regulated to the system. Once the water levels in the basin and the inflow pipe (diameter = 25 cm) carrying wastewater from the intersecting street to Frey were equal, the pipe would not permit more flow to enter the basin and the remaining flow from the intersecting str eet continued downstream to enter subsequent ditches. When adding the cobble and gravel laye rs, flow to Frey was maximized to ensure enough layers were in place to prevent overflow during high flow events. Subsurface flow sections were also designed to be extremely porous to extend the lifetime of the system before clogging. Cobbles (10-20 cm ) were added as the bottom layer followed by progressively smaller gravel at the uppermost layers to prevent bed surface sediments and plant litter from entering and obstructi ng pore spaces. The gravel secti ons were planted hydroponically (Figure 5-10) allowing ample space for oxygen exch ange from the atmosphere (Wallace et al. 2001). Monitoring and Statistical Analyses Tracer experiments were performed one to two weeks after comple tion of the channel modifications following the field and analytical methods descri bed in Chapter 3. Breakthrough curves of the conservative solute were de termined at two sampling locations: after the sedimentation basin and at the most downstr eam location. Median transport times (Tmed) were determined from the tracer breakthrough curves and compared with pr evious values (before modifications) to evaluate th e effect of the modifications on channel residence times.

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137 Grab samples of water were collected every 1 to 2 weeks for a 6 month period (September 2006 through February 2007) in 3 to 4 locations along each ditch and analyzed for fecal coliform bacteria, phosphate, ammonium, nitrate and to tal suspended sediments following field and laboratory procedures described in Chapter 3. Comparisons between water quality parameters before and after channel modifications were made using t-tests assuming equal variances af ter the data had been ln-transformed to meet normality requirements. Significant values were re ported for a water quality parameter if t-tests indicated that influent concentrations were not statistically different before and after channel modifications but effluent concentrations after modifications were statistically less than those observed prior to modifications. Results Effects of Modifications on Re sidence Times and Water Quality Heavy rain events associated with El Nio Southern Oscillation activity readily undermined the integrity of the sedimentation basin in Gustavson and therefore only results related to the other two sites Koe ll and Frey are presented here. Tracer experiments before and after channel m odifications revealed that the modifications in Koell did not substantially le ngthen the residence time or the shape of the breakthrough curve (Figure 5-11, Table 5-4). Determination of Tmed at the exit of the sedimentation basin revealed that the basin had very little effect in extending the overall channel transport times as the median basin residence time was less than 2 minutes. In c ontrast, residence times in Frey increased from approximately 6 minutes to 46 minutes after ch annel modifications, (Figure 5-12, Table 5-4). However, median solute residence time in th e sedimentation basin accounted for only 4.2 minutes.

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138 The open water flow design used to modify the channel of Koell did not result in significant improvements to water quality with re spect to any of the water quality parameters except for dissolved oxygen (Figures 5-13 to 5-18). The spread of E. coli distributions was reduced compared to those before channel modifica tions but median values were not statistically different (Figure 5-13). Influent TSS values were slightly hi gher during the sampling period following channel modifications (Figure 5-14). Th is finding was expected due to the prevalence of heavy rain events carrying high sediment load s. Median TSS values did not differ between the three sampling locations indicating that net sedimentation did not te nd to occur during transport. The creation of riffles and cascades were e ffective at increasing DO levels (p < 0.05) between the first and second sampling sites (Fig ure 5-15) where channel slope was greatest. DO levels tended to decrease after th e second sampling site due to th e lower channel gradient that substantially reduced the aeration efficiencies of the riffles. However, despite increased DO levels, nitrification was not promoted to a degree that resulted in reduced NH4-N concentrations (Figure 5-16). However, nitrif ication was indeed occurring as evidenced from relatively high NO3-N values compared to those observed prio r to channel modification (Figure 5-17). Moreover, denitrification was also likely to have occurred as high NO3-N concentrations decreased between sampling locations. However, c oupled nitrification-den itrification reactions were unsuccessful at overa ll nitrogen retention. Concentrations of SRP did not change significantly during trans port either before or after channel modifications (Figure 5-18). However, th is should not necessarily be interpreted as the lack of retention because additional phosphorus inputs from wastewat er discharges occur between locations 1 and 2, indica ting that net retention must be occurring if concentrations do not increase downstream. Influe nt concentrations of NH4-N and SRP tended to lower than those

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139 before channel modifications and are likely to be the result of dilution from rain events. Lower water column SRP concentrations during periods of dilution may be causing the sediments to act as P sources (see Chapter 4 discussion of EPCo). Comparisons of water quality parameters befo re and after channel modifications in Frey indicated that the modifications were effective at removal of E. coli (t = 2.1209, df = 14, p-value < 0.05) and TSS (t = -4.3447, df = 22, p-valu e < 0.01) but not for reductions of nutrient concentrations (Figures 5-19 thr ough 5-24). Due to wastewater inpu ts along the length of Frey (see Figure 5-8), prior to modi fications, concentrations of E. coli and TSS tended to increase with downstream distance. However, after modifications E. coli and TSS levels decreased downstream despite additional wa stewater inputs to the syst em (Figures 5-19 and 5-20). Channel modifications were not effective for promoting oxygenation as evidenced by decreased DO concentrations at the downstream sampling location (Figure 5-21). However, concentrations of NO3-N were often higher at the downstr eam sampling location (Figure 5-22) indicating that nitrification wa s nevertheless occurring in the system, consuming the dissolved oxygen in the process. However, nitrification was not an effective retention process as NH4-N concentrations tended to increase downstream (Figure 5-23). Concentr ations of SRP were effectively unchanged between the two sampling points (Figure 5-24) i ndicating that some P retention occurred since concentrations would have otherwise increased from additional wastewater inputs. Management Requirements Management requirements for channel modi fications included pe riodic removal of sediments collected in sedimentation basins an d, in the case of Frey, from the open water sections planted with hyacinth. To facilitate sediment collection, the culvert that delivers water to the basin from the intersecting street was blocked such that water by-passed the basin and

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140 continued to the next inters ecting ditch downstream. While the basin was being drained, approximately 90% of the water hyacinth plants were collected and wheel-barrowed to the compost bins. The remaining plants were retu rned to the basin after sediment removal. Sediments were extracted with shovels and buc kets, placed on a large tarp and mixed with sawdust and harvested water hyacinth plan ts and added to the compost bins. Sediment removal in the Koell basin was nece ssary three times over a four month period. The first two cleanings were necessary after th e city removed vegetation from upstream ditches delivering high loads to Koell, one the most downs tream ditches before discharge to the river. The third cleaning was necessary after two months of heavy rainfall. The sedimentation basin and open water sections in Frey were cleaned on ce in four months. Two to three cleanings per year was the anticipated frequency of sediment removal and was considered acceptable by the participating residents. More frequent management requirements in cluded monthly removal of water hyacinth and weekly cleaning of the screens and removal of trash. A simple tool was provided to a representative participant of each modified ditc h to facilitate plant/tr ash removal and screen cleaning. The tool consisted of a net on one end and a coarse brush on the other (Figure 5-25). The brush effectively removed algae and trappe d solids to prevent clogging of the screens and exit structure. The long handle of the tool prevente d the user from needing to enter the ditch. Government and Community Response Support by the community and government official s was high for modified ditches due to increased visual aesthetics (Figur es 5-26 and 5-27), reduced odors, perceived improvement to the environment and low implementation costs (Table 5-5). Costs of the channel modifications varied between $2.5 and $5.0 per meter, which were considerably lower than the cost of installing pipe.

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141 Six residents regularly assisted with modifications in Gust avson, 12 in Koell and 10 in Frey. Other residents who were not directly involved in channe l modifications demonstrated support by providing drinks and sn acks. Community participation provided an evident sense of ownership as residents agreed amongst themselves that each was responsible for caring for the section of ditch in front of their home. Inte rviews with residents suggested a sense of empowerment was felt among those who participated in the project because they were able to improve the aesthetics of thei r neighborhood. For example, in creased oxygenation in Koell vastly improved odors emanating from the ditc h. Residents commente d that prior to the modifications, the smells were so unpleasant that th ey were unable to sit outside in the evenings. However, after channel modifications, several families had set up benches outside their homes and said they could no longer smell odors. The re sidents of Frey and Koell hosted a celebration to inaugurate completion of the modified ditches (Figure 5-28). The modified ditches received much attention from the community in general. The modified ditches were visited regularly by re sidents from other neighborhoods and government officials from other cities. High school stude nts were asked by teachers to write a report discussing the water quality functions of the different design elements of the ditches. The modified ditches were on the news several times and on one occasion resulted in a live television interview. Another unexpected response from the community was that other neighborhoods began attempting to modify their ditches (Figur e 5-29) thus prompting the creation of the nontechnical guide describing methods used to modify ditches. The municipality of Oxapampa supported th e project by donating rocks and equipment and by agreeing to provide trash pickup services to residents living at the end of Koell to prevent

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142 future dumping of refuse into the ditch. A local NGO, Instituto de Bien Comn hopes to expand on the idea to modify ditches in other communities. Discussion The modified ditches received very positiv e responses from the public as they were perceived as a potential solution towards impr oving local wastewater management. However, improved aesthetics resulting from m odifications may foster a false sense of security that water quality was concomitantly improved. This was cert ainly true for the modified ditch in Koell. Comparisons of water quality before and after channel modifications in Koell indicated that the open water flow approach did little to improve overall water quality. Nevertheless the approach was very successful in providi ng improved aesthetics to the neighborhood. Increased dissolved oxygen concentrations significantly reduced odors, fulfilling an importa nt criterion for residents. Visually, the system was very much improved as indicated by it being re ferred to as a garden rather than as a ditch by passers-by. The lack of treatment occurring in the system in Koell is the result of low retention times that were exacerbated over the monitoring peri od due to the high frequency of large storm events. Although the sedimentation basin was unde rsized and thus not capable of increasing overall residence times, it was nonetheless effectiv e at trapping sediments due to the use of the screens. The small volume of the basin and high solids removal meant that the basin required frequent cleanings. The basin fill ed with sediments on three occas ions within four months. The high management demand of the system eventua lly discouraged participants. The same basin design in Gustavson filled with sediments duri ng a large storm event and the following day, a second storm event washed out the entire impound ment structure. For th e basins to function effectively without requiring ex cessive management and to be able to withstand high loads, larger more robust structures are required.

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143 Retention times in Koell were likely to be rela tively low due to the lack of vegetation in the ditch channel. The ornamental plants chosen in this study colonize on ly channel banks or are grown hydroponically on gravel bars and thus are of limited importance for reducing water velocities. However, the gravel bars and riffle s appear to have a positive impact on increasing solute residence times. Median transport times in Koell increased by approximately12% after channel modifications despite flows that were 44 % higher. These results poi nt to the possibility of improved retention in Koell during the dry se ason. Future modifications may also want to consider promoting submerged macrophytes for improved attenuation of velocity and direct nutrient uptake from the water column (Mar s et al. 1999, Tanaka et al. 2006) In contrast to the open water flow design, the alternating subsurface flow/open water design in Frey was effective at reducing concentrations of E. coli and solids. This result was expected as the design was modeled as a qua si-subsurface flow wetland which are commonly recognized for their water treatment capabilitie s (Tanner et al. 1995, Billore et al. 1999). However, it was not surprising th at nutrient removal was low in the system due to limited residence times compared to typical subsurface flow wetlands which are normally on the order of weeks rather than minutes (e.g. Akratos and Tsihrintzis 2007). Subsurface flow (SSF) wetland systems are commonly planted with cattails ( Typha ) and reeds ( Phragmites ). However, more studies are reporting th at ornamental, water-tolerant plants provide comparable treatment performance. Irise s have been shown to provide similar biomass and nitrogen storage characteristics as Typha and Phragmites in constructed wetlands (Tuncsiper et al. 2006, Zaimoglu 2006). Si milar findings were reported by Ayaz and Akca (2001) who found that Iris was more effective than Typha and Phragmites for COD, TN and TP removal and that Canna was highly effective at ammonium remova l. DeBusk et al. (1995) reported that

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144 Canna demonstrated the highest foliar phosphate uptake of 10 emergent macrophytes. Belmont and Metcalfe (2003) found that calla lilies signi ficantly increased ammonium retention in SSF wetlands. These studies suggest th at nutrient retention in the m odified ditches was not limited by the choice of ornamental plants but instead by high loading rates coupled with low residence times. Phosphorus removal may have been improved with the use of other filter materials (Adam et al. 2007). The lack of oxygen was another important f actor limiting efficient retention in the subsurface flow system in Frey. DO concentrations below 1.5 mg L-1 have been shown to limit nitrification (Wolverton 1987). Values in Fr ey were consistently below 1.0 mg L-1. These values suggest that oxygen translocation by plant roots was not signifi cant enough to support nitrifying bacteria to promote effective nitrification-denitrif ication reactions (B rix 1997). Oxygenated rhizospheres have also been shown to be impor tant sites for reducing fecal coliform and BOD concentrations in SSF (Karathana sis et al. 2003). The treatment performance of the subsurface flow system would be amplif ied if better oxygenation could be achieved (Ouellet-Plamondon et al. 2006). However, it is expected that overall retention will impr ove in the subsurface flow system as biofilms and other reactive surface area s continue to increase with time (Bigambo and Mayo 2005, Soto et al. 2007). The alternating subsurface flow/open water flow design proved to be successful as a simple, low cost approach to pr ovide primary treatment and perhap s irrigation-quality water. The fecal coliform standard for water reuse for irrigation is 1,000 CFU mL-1 (WHO 1989). The design did not demonstrate consistent effluent E. coli concentrations belo w this value but one must take into account that reductions approachi ng this value occurred de spite additional fecal inputs along the length of the ditch. The reuse of treated ditch water for irrigation purposes and

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145 for aquaculture (currently in practice in Oxapam pa for raising trout) would be advantageous for capitalizing on the bioavailable nutrients, part icularly given the current dependence upon chemical fertilizers in the region. Moreover, th e subsurface flow design could be coupled with the existing aquaculture facility to take advantage of the artificial aeration mechanisms already in place. Another distinct advantage of the subsurface flow design over the open water design is the minimization of human contact with wastewater. In addition, maintaining water in the subsurface prevents health and nuisance issues related to mo squito breeding. The major disadvantages of the approach are that it is limited to ditches with lo w flowrates and requires large quantities of gravel that is extracted from the river. Gravel mining is an unsustainable practice that undermines that the self-purifying capacity of the river system by reducing the ex tent of beneficial hyporheic zones (Hancock 2002). Widespread support for testing the effects of modified ditches for improved water quality and community aesthetics revealed an openness and readiness by the gov ernment and public to explore alternative water treatment approaches. However, without long-te rm commitment from the local government to support ditch management e fforts, the modified ditc hes are likely to fail as participant support wanes with time. The approaches implemented and tested in this study are not likely to solve the problematic wastewater management situation th at currently exists in Oxapampa. However, modified ditches nevertheless point the way for exploring smaller-scale, decentralized systems such as neighborhood-scale treatment wetland s (Griffin and Pamplin 1998, Greenway and Woolley 1999).

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146 Conclusions The experimental ditch modifications resulted in minimal improvements to concentrations of nutrients; however fecal coliform bacteria and total suspended so lids showed significant decreases in concentrations during the first tw o months of sampling. Beginning in the third month of monitoring, heavy rain events associ ated with ENSO activity resulted in pulses of extremely elevated discharges that undermined the treatment function of the modified systems. Preliminary findings suggest that the modified ditches in th is study have minimal impact on reducing downstream loads of nutrients. Nonetheless, they show potential for providing primary treatment to otherwise raw sewage. The experimental approach used in this study was successful at involving the community and provide d residents with a viable alternative for improving their environment, from both a public-hea lth and an aesthetic perspective. Effective long-term management will likely require comm itment from the local government for assistance in ditch management, particularly with respect to periodic dredging of sediments. Future modifications should utilize desi gns better equipped for heavy pe riods of rain and sediment loads.

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147 Table 5-1. Activities performed to promote local support and participation in channel modifications Stage Activity Purpose 1. Support assessment Presented research to city representatives and local NGOs (3 occasions) Determine whether agencies supported project and to receive feedback Spoke at a neighborhood meeting and presented interpretive folios Inform community members about modified ditches as an alternative to pipes Met with residents in each of the 3 neighborhoods Engage residents in decisionmaking process; determine whether residents were willing to participate in project implementation and management stages; address specific needs and concerns 2. Implementation Hired reside nts to assist in channel modifications Foster sense of ownership of project Erected signs Inform interested parties about the project; discourage littering in the ditches Constructed bridges across modified ditches Improve safety of ditch crossings; increase aesthetics and sense of community 3. Participatory management Created a non-technical guide detailing how to modify and manage ditches Provide government and residents with documentation of procedures used Constructed a compost bin for ditch plants and sediments Provide residents with a valuable end-product Provided residents with tools for ditch management Facilitate management of ditches

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148 Figure 5-1. Neighborhood meetings were attended to assess wh ether the community supported implementation of modified ditches

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149 Figure 5-2. Residents particip ated in all aspects of the channel modification activities

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150 Figure 5-3. Ornamental plants such calla lily were transplanted fr om other sewage ditches to the modified ditches Table 5-2. Primary design elements used in ope n-water flow design and their intended functions for improving water quality and aesthetics Design elements Intended function Sedimentation basin Trap solids, grease and trash Sections of riffles and cascades s Ae rate water, increas e bio-reactive surface areas Lateral gravel bars Extend travel times, promote water filtration, denitrification Vegetation Reduce bank and channel erosion, provide bio-reactive surface ar eas, assimilate nutrients, improve aesthetics

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151 Figure 5-4. Schematic of the sedimentation basi n design used in channe l modifications A) top view and B) side view Figure 5-5. A) Front view a nd B) side views of the sedimentation basin in Koell Direction of flow Basin outfall and oxygenation of water Screens Water hyacinth Accumulated solids Water oxygenation upon exit from the basin A B A B

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152 Figure 5-6. Sections of cascades and riffles created to improv e oxygenation and increase pockets of transient storage in A) Koell and B) Gustavson Figure 5-7. Creation of gravel bars to increase channel sinuosity and filtration in Koell A B

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153 Table 5-3. Primary design elements used in the alternating subsurface flow/open water design and their intended functions for im proving water quality and aesthetics Design element Intended function Sedimentation basin Trap solids, grease and trash Gravel beds for subsurface flow Incr ease residence times and reactive surface areas for physical and biochemical filtration Open water zones Sedimentation of solids Vegetation Increase reactive surface areas, assimilate nutrients, reduce bank erosion, improve aesthetics Figure 5-8. Schematic of the alternating subsurfa ce flow/open-water design implemented in Frey Figure 5-9. Views of the altern ating subsurface flow/open water design used in Frey A) water exits the sedimentation basin and flows into the first subsurface flow section and B) water exits the first subsurface flow secti on and enters the first open water section 83.2 m 10 m 22.6 11 m 24.3 m 1.5 m 10 m Sedimentation basin Subsurface flow Screens Note: Not to scale A portion of flow from an inters ecting street enters Frey and the remaining flow continues to the following block. Open water with water hyacinth Wastewater p i p es

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154 Figure 5-10. Cross-sectional view of the subsurface flow design 0 0.2 0.4 0.6 0.8 1 0.00.51.01.52.0 Time since injection (hours)Fraction of plateau concentration Before modifications After modifications Figure 5-11. Comparison of breakthrough curves befo re and after channel modifications in Koell Water level

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155 0 0.25 0.5 0.75 1 00.511.52 Time since injection (hours)Fraction of plateau concentration Before modifications After modifications Figure 5-12. Comparison of breakthrough curves befo re and after channel modifications in Frey Table 5-4. Comparison of median travel times in Koell and Frey before and after channel modifications Koell Frey Flowrate (L s-1) Tmed (min) Flowrate (L s-1) Tmed (min) Before 5.0 24.0 2.8 6.8 After 7.2 27.0 2.4 45.6 UpstreamMiddleDownstreamKoell before modifications 0 5000 10000 15000 20000 25000E. coli (CFU 100 mL-1) UpstreamMiddleDownstreamKoell after modifications Figure 5-13. Comparison of E. coli concentrations at upstr eam, middle and downstream sampling locations in Koell before and after ditch modifications (dark circles represent median values, open circles are out liers, boxes delineate inter-quartile range (IQR), and whiskers are 1.5* IQR)

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156 UpstreamMiddleDownstreamKoell before modifications 0 50 100 150 200TSS (mg L-1) UpstreamMiddleDownstreamKoell after modifications Figure 5-14. Comparison of TSS concentrations at upstream, middle and downstream sampling locations in Koell before and after ditch m odifications (dark circles represent median values, open circles are outliers, boxes de lineate inter-quartile range (IQR), and whiskers are 1.5* IQR) U p streamMiddleDownstreamKoell before modifications 0 1 2 3 4DO (mg L-1) U p streamMiddleDownstreamKoell after modifications Figure 5-15. Comparison of dissolved oxygen ( DO) concentrations at upstream, middle and downstream sampling locations in Koell befo re and after ditch modifications (dark circles represent median values, open ci rcles are outliers, boxes delineate interquartile range (IQR), and whiskers are 1.5* IQR)

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157 U p streamMiddleDownstreamKoell before modifications 0 2 4 6 8 10NH4-N (mg L-1) U p streamMiddleDownstreamKoell after modifications Figure 5-16. Comparison of NH4-N concentrations at upstr eam, middle and downstream sampling locations in Koell before and after ditch modifications (dark circles represent median values, open circles are out liers, boxes delineate inter-quartile range (IQR), and whiskers are 1.5* IQR) UpstreamMiddleDownstreamKoell before modifications 0.0 0.1 0.2 0.3 0.4 0.5NO3-N (mg L-1) UpstreamMiddleDownstreamKoell after modifications Figure 5-17. Comparison of NO3-N concentrations at upstr eam, middle and downstream sampling locations in Koell before and after ditch modifications (dark circles represent median values, open circles are out liers, boxes delineate inter-quartile range (IQR), and whiskers are 1.5* IQR)

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158 UpstreamMiddleDownstreamKoell before modifications 0.0 0.2 0.4 0.6 0.8 1.0SRP (mg L-1) UpstreamMiddleDownstreamKoell after modifications Figure 5-18. Comparison of SRP concentrations at upstream, middle and downstream sampling locations in Koell before and after ditch m odifications (dark circles represent median values, open circles are outliers, boxes de lineate inter-quartile range (IQR), and whiskers are 1.5* IQR) UpstreamDownstreamFrey before modifications 0 5000 10000 15000 20000 25000E. coli (CFU 100 mL-1) UpstreamDownstreamFrey after modifications Figure 5-19. Comparison of E. coli concentrations at upstr eam, middle and downstream sampling locations in Frey before and after ditch modifications (dark circles represent median values, open circles are outliers, boxes delineate interquartile range (IQR), and whiskers are 1.5* IQR)

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159 UpstreamDownstreamFrey before modifications 0 100 200 300TSS (mg L1 ) UpstreamDownstreamFrey after modifications Figure 5-20. Comparison of TSS concentrations at upstream, middle and downstream sampling locations in Frey before and after ditch m odifications (dark circles represent median values, open circles are outliers, boxes de lineate inter-quartile range (IQR), and whiskers are 1.5* IQR) UpstreamDownstreamFrey before modifications 0 1 2 3DO (mg/l) UpstreamDownstreamFrey after modifications Figure 5-21. Comparison of dissolved oxygen ( DO) concentrations at upstream, middle and downstream sampling locations in Frey befo re and after ditch modifications (dark circles represent median values, open ci rcles are outliers, boxes delineate interquartile range (IQR), and whiskers are 1.5* IQR)

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160 UpstreamDownstreamFrey before modifications 0.0 0.1 0.2 0.3 0.4 0.5NO3-N (mg/l) UpstreamDownstreamFrey after modifications Figure 5-22. Comparison of NO3-N concentrations at upstr eam, middle and downstream sampling locations in Frey before and after ditch modifications (dark circles represent median values, open circles are outliers, boxes delineate interquartile range (IQR), and whiskers are 1.5* IQR) UpstreamDownstreamFrey before modifications 0 5 10 15NH4-N (mg L-1) UpstreamDownstreamFrey after modifications Figure 5-23. Comparison of NH4-N concentrations at upstr eam, middle and downstream sampling locations in Frey before and after ditch modifications (dark circles represent median values, open circles are outliers, boxes delineate interquartile range (IQR), and whiskers are 1.5* IQR)

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161 UpstreamDownstreamFrey before modifications 0.0 0.5 1.0 1.5 2.0 2.5SRP (mg/l) UpstreamDownstreamFrey after modifications Figure 5-24. Comparison of SRP concentrations at upstream, middle and downstream sampling locations in Frey before and after ditch m odifications (dark circles represent median values, open circles are outliers, boxes de lineate inter-quartile range (IQR), and whiskers are 1.5* IQR) Figure 5-25. Simple tool consisting of a combined net and brush was provided to residents to facilitate ditch maintenance

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162 Figure 5-26. Images of the open water flow design in Koell ditch four months after modifications Figure 5-27. Images of the altern ating subsurface flow/open water design in Frey four months after modifications Table 5-5. Approximate costs of ditch modifications for each site Ditch Modification length (m) Modification cost ($ m-1)* Gustavson 135 2.5 Koell 195 5.0 Frey 85 3.2 Cost to install underground pipe ($ m-1) 37.5 Includes labor (typical local wage is $5 day-1), construction materials (lumber, screens and nails), plants and collection and delivery of cobbles and gravel

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163 Figure 5-28. Residents from Frey held a celebration to inaugura te the ditch and new bridge Figure 5-29. Other reside nts began attempts at modifying their ditches

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164 CHAPTER 6 SYNTHESIS AND FUTURE WORK This study provided a unique perspective of drainage ditches in the landscape by examining the treatment potential of ditches rece iving domestic wastewaters in a rural Peruvian community. Drainage ditches are prevalent landscap e features throughout the world yet their role for influencing contaminant fate and transport between terrestrial and aquatic ecosystems has been largely ignored. Likewise, the use of vegeta ted ditches for wastewat er disposal is poorly documented despite their widespread use in developing countries such as Peru. The ability of ditches to mitigate contaminan t exports was shown to be predicated on the interconnected relationships between contaminant loads, flow and channel characteristics, plant communities and sediment properties. Plants were shown to play a crucial role in affecting transport characteristics by crea ting pockets of slower-moving wa ter that increased residence times and allowed sedimentation and organic matter deposition to occur. These vegetationinduced transient storage zones were likely to have been the sites wh ere the reductions of sediments and pathogens occurred during transport. The depositional environments created by vegetation during the dry season in Oxapampa re sulted in increased clay and organic matter content of sediments. In turn, fine particles and organic-rich sediments we re shown to be related to much higher phosphorus retention capacity co mpared to the coarser benthic sediments observed in the rainy season afte r plant harvesting. Given the high sorption affinity of many contaminants to organic matter and fine pa rticles, depositional conditions promoted by vegetation in ditches were also likely to retain othe r potential contaminants such as pathogens, pesticides, and metals. Ditch macrophyte biomass functioned as a large storage of nitrogen and phosphorus yet uptake processes by pl ants were not estimated to be a significant net retention mechanism.

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165 High loads of nitrogen and phosphorus coupl ed with low channel residence times overwhelmed the intrinsic processing capacity of the wastewater ditches. While evidence of contaminant retention in ditches was observed th rough comparisons of water quality with pipes and by comparisons between upstream and downstr eam ditch sampling sites, the ditches were nevertheless ineffective at redu cing nitrogen and phosphorus concen trations to near-acceptable discharge levels. The treatment performance of wastewater ditches both before and after channel modifications was compromised by the lack of o xygen. Large quantities of retained phosphorus due to sorption with iron oxyhydroxides were readily released to overlying water under anaerobic conditions. Low dissolved oxygen leve ls prevented net nitrogen removal by limiting nitrification of ammonium. Channe l modifications in Koell served to increase oxygen levels to concentrations as high as 5 mg L-1 on several occasions; however, ammonium loads were too great to result in ov erall nitrogen re tention through coupled nitrification-denitr ification reactions. These findings suggest that ditc hes, especially those receiv ing high nutrient loads, cannot be relied upon as the sole treatment mechanism but instead should be used in conjunction with other management practices that serve to attenuate contaminant pulses to ditches. For wastewater drainage ditches these practi ces could include simple househ old septic tanks or large neighborhood sedimentation basins for solids rem oval prior to discharg e to the channels. The creation of subsurface flowpaths is par ticularly advantageous for reducing human contact, thus addressing a major health issue related to the use of ditches for waste disposal. Moreover, the alternating subsurf ace flow/open water design was s hown be effective at sediment and E. coli removal and thus shows promise for treati ng wastewaters to meet irrigation-standards for water reuse. The goal of ditch modifications and wastewater treatment systems in general

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166 should be the removal of sediments, oxygen demand and pathogens to levels that are acceptable for water reuse in order to recy cle and utilize the valuable nutri ents while preventing detrimental effects to aquatic ecosystems caused by eutrophication. Many questions remain regarding the use of ditc hes as viable treatment systems. Questions include how ditches should be managed to opt imize retention while avoiding risks of floods. Another practical research question is what length of ditch, given different plant communities and biomass conditions and under various hydrolog ic and loading conditio ns, is sufficient to achieve a desired treatment response? Future ditch studies should cont inue to focus on how ditches can be modified to amplify their treatm ent capabilities. Retrofitti ng existing ditches with designs aimed at treatment, rather than solely water transport, could result in significant improvements to water quality in regions throughout the world.

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167 APPENDIX A ENERGY CIRCUIT LANGUAGE Figure A-1. Description of the symbols used in energy circuit diagrams (from Odum 1994)

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168 APPENDIX B DITCH CHANNEL CROSS-SECTIONAL AND LONGITUDINAL PROFILES Section 1 (Distance = 0m) 5 6 7 8 9 01234567 (m)(m) Section 2 (Distance = 33m) 5 6 7 8 9 10 01234567 (m)(m) Section 3 (Distance = 62.3m) 6 7 8 9 10 11 -11357 (m)(m) Section 4 (Distance = 151.1m) 6 7 8 9 10 11 01234567 (m)(m) Figure B-1. San Martin cross-sectiona l profiles for sections 1 through 4

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169 Section 5 (Distance = 200.1m) 8 9 10 11 12 01234567 (m)(m) Seccion 6 (Distance = 255.3m) 8 9 10 11 12 01234567 (m)(m) Section 7 (Distance = 315.8m) 8 9 10 11 12 01234567 (m)(m) Section 8 (Distance = 355.6m) 9.5 10 10.5 11 11.5 12 01234567 (m)(m) Figure B-2. San Martin cross-sectiona l profiles for sections 5 through 8

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170 Longitudinal profile 0 2 4 6 8 10 12 050100150200250300350400 metersmeters Figure B-3. San Martin longitudinal profile Block 1 (Distance = 0) 6 7 8 9 10 01234 mm Block 1 (Distance = 12.6m) 6 7 8 9 10 01234 mm Block 1 (Distance = 39.9m) 6 7 8 9 10 01234 mm Figure B-4. Gustavson cross-sectional profiles for sections along Block 1

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171 Block 2 (Distance = 48.2m) 7 8 9 10 012345 mm Block 2 (Distance = 55.9m) 7 8 9 012345 mm Block 2 (Distance = 80.2m) 7 8 9 012345 mm Block 2 (Distance = 95.3m) 7 8 9 012345 mm Figure B-5. Gustavson cross-sectional profiles for sections along Block 2

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172 Block 3 (Distance = 104.3m) 8 9 10 0123456 mm Block 3 (Distance = 137m) 7 8 9 10 0123456 mm Block 3 (Distance = 187.8m) 7 8 9 0123456 mm Block 3 (Distance = 212.8m) 6 7 8 9 0123456 mm Figure B-6. Gustavson cross-sectional profiles for sections along Block 3

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173 Block 4 (Distance = 227.3m) 7 8 9 10 012345678910 mm Block 4 (Distance = 277.9m) 6 7 8 9 10 012345678910 mm Block 4 (Distance = 319.4m) 6 7 8 012345678910 mm Block 4 (Distance = 359.9m) 5 6 7 012345678910 mm Figure B-7. Gustavson cross-sectional profiles for sections along Block 4

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174 Gustavson longitudinal profile (Block 1) 6 7 8 9 01020304050 metersmeters Gustavson longitudinal profile (Block 2) 6 7 8 9 405060708090100 metersmeters Gustavson longitudinal profile (Block 3) 6 7 8 9 100110120130140150160170180190200210220 metersmeters Gustavson longitudinal profile (Block 4) 6 7 8 9 220230240250260270280290300310320330340350360370 metersmeters Figure B-8. Gustavson longitudi nal profiles for Blocks 1-4

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175 Section 2 (Distance = 39.9m) 8 8.5 9 024681012141618 mm Section 1 (Distance = 0) 8 9 10 01234 mm Section 3 (Distance = 113.4m) 8 9 10 11 01234 mm Section 5 (Distance = 210m) 11 12 13 01234 mm Figure B-9. Frey cross-sectional profiles Longitudinal profile 6 8 10 12 14 050100150200250 metersmeters Figure B-10. Frey longitudinal profile Section 4 (Distance = 126.8m) 10 11 12 13 01234 mm

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176 Section 1 (Distance = 0) 7 8 9 10 0123456789 mm Section 3 (Distance = 60.8m) 5 6 7 8 9 0123456789 mm Seccion 5 (Distancia = 121.5m) 4 5 6 0123456789 mm Figure B-11. Koell cros s-sectional profiles Section 2 (Distance = 28.8m) 6 7 8 9 0123456789 mm Section 4 (Distance = 90.3m) 4 5 6 7 8 9 0123456789 mm

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177 Longitudinal profile 0 2 4 6 8 10 050100150200250 metersmeters Figure B-12. Koell longitudinal profile

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178 APPENDIX C QA/QC RESULTS Table C1 Summary of QA/QC results for nutrient analyses SRP (mg P L-1) NH4-N (mg N L-1) NO3 (mg NO3 L-1) TN (mg N L-1) TP (mg P L-1) na 32 24 18 12 10 accuracyb .04 .03 .02 .3 .4 n 63 49 45 20 12 reproducibilityc 0.04 .02 .02 .2 .3 an is the number of times accuracy and re producibility analyses were performed. bAccuracy was determined by the difference from the known standard concentration. cReproducibility was calculated by the standard devi ation of duplicate measurements

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179 APPENDIX D OTIS-P SIMULATION RESULTS 0 100 200 300 400 500 0.00.51.01.5 Time (hours)Conductivity (uS cm-1) observed predicted Figure D-1. OTIS-P modeling results fo r 2005 tracer experiment in San Martin 100 110 120 130 140 150 160 170 180 00.511.522.5 Time (hours)Conductivity (uS cm-1) observed predicted Figure D-2. OTIS-P modeling results fo r 2006 tracer experiment in San Martin 200 250 300 350 00.511.52 Time (hours)Conductivity (uS cm-1) observed predicted Figure D-3. OTIS-P modeling results fo r 2005 tracer experiment in Gustavson

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180 250 270 290 310 330 350 370 390 00.511.522.53 Time (hours)Conductivity (uS cm-1) predicted observed Figure D-4. OTIS-P modeling results fo r 2006 tracer experiment in Gustavson 0 1000 2000 3000 4000 5000 6000 01234 Time (hours)Conductivity (uS cm-1) observed predicted Figure D-5. OTIS-P modeling results for 2005 tracer experiment in Koell 200 300 400 500 600 0.00.51.01.5 Time (hours)Conductivity (uS cm-1) Observed Predicted Figure D-6. OTIS-P modeling results for 2006 tracer experiment in Koell

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181 APPENDIX E PHOSPHORUS FRACTIONATION DATA Table E-1. Phosphorus fractionation data for Koell Site 1 Fractionation Procedure Rep 1 Rep 2 Porewater Filter wt initial (g) 0.0843 0.0883 Filter wt final (g) 0.0875 0.0922 Sediment collected (g) 0.0032 0.0039 SRP in pore water 0.13 0.14 KCl-bioavailable Filter wt initial (g) 0.0833 0.0872 Filter wt final (g) 0.1 0.1031 Sediment collected (g) 0.0167 0.0159 SRP (mgP/l) 0.03 0.05 P (mg) 0.0006 0.001 sample wt dry (g) 1.14 1.14 P (mg/kg)) 0.526315789 0.877192982 NaOH-Fe and Al oxyhydroxides inorganic total organic SRP (mgP/l) 143 307.5 164.5 P (mg) 2.86 6.15 3.29 sample wt dry (g) 1.14 1.14 1.14 P (mg/kg)) 2508.77193 5394.736842 2885.96491 HCl-Ca and Mg bound Sediment collected (g) 0 0 SRP (mgP/l) 78 88 P (mg) 1.56 1.76 sample wt dry (g) 1.14 1.14 P (mg/kg) 1368.421053 1543.859649 Recalcitrant P (mg/kg)) 125 125 Summary average P (mg/kg) KCl 0.701754386 NaOH-Pi 2508.77193 NaOH-Po 2885.964912 HCl 1456.140351 Recalcitrant 125 Total 6976.170413

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182 Table E-2. Phosphorus fractionation data for Koell Site 2 Fractionation Procedure Rep 1 Rep 2 Porewater Filter wt initial (g) 0.0845 0.0877 Filter wt final (g) 0.09 0.0906 Sediment collected (g) 0.0055 0.0029 SRP in pore water 0.1 0.08 KCl-bioavailable Filter wt initial (g) 0.0843 0.0874 Filter wt final (g) 0.1003 0.1033 Sediment collected (g) 0.016 0.0159 SRP (mgP/l) 0.03 0.02 P (mg) 0.0006 0.0004 sample wt dry (g) 1.7885 1.7885 P (mg/kg)) 0.33547666 0.223651104 NaOH-Fe and Al oxyhydroxides inorganic total organic SRP (mgP/l) 120 305 185 P (mg) 2.4 6.1 3.7 sample wt dry (g) 1.7885 1.7885 1.7885 P (mg/kg)) 1341.90663 3410.67934 2068.7727 HCl-Ca and Mg bound SRP (mgP/l) 98 96 P (mg) 1.96 1.92 sample wt dry (g) 1.7885 1.7885 P (mg/kg)) 1095.89041 1073.525301 Recalcitrant P (mg/kg)) 687 687 Summary average P (mg/kg) KCl 0.27956388 NaOH-Pi 1341.90663 NaOH-Po 2068.77271 HCl 1084.70786 Recalcitrant 687 Total 5182.95877

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183 Table E-3. Phosphorus fractionation data for Koell Site 3 Fractionation Procedure Rep 1 Rep 2 Porewater Filter wt initial (g) 0.0873 0.0865 Filter wt final (g) 0.0898 0.0899 Sediment collected (g) 0.0025 0.0034 SRP in pore water 0.06 0.08 KCl-bioavailable Filter wt initial (g) 0.0861 0.0844 Filter wt final (g) 0.104 0.1062 Sediment collected (g) 0.0179 0.0218 SRP (mgP/l) 0.04 0.03 P (mg) 0.0008 0.0006 sample wt dry (g) 0.9535 0.9533 P (mg/kg)) 0.839014158 0.629392636 NaOH-Fe and Al oxyhydroxides inorganic total organic Filter wt initial (g) 0.0844 0.1007 Filter wt final (g) 0.1002 0.1062 Sediment collected (g) 0.0158 0.0055 SRP (mgP/l) 128 315 P (mg) 2.56 6.3 3.74 sample wt dry (g) 0.9377 0.9478 0.9478 P (mg/kg)) 2730.084249 6646.971935 3945.98 HCl-Ca and Mg bound Filter wt initial (g) 0.0861 0.0839 Filter wt final (g) 0.1063 0.1142 Sediment collected (g) 0.0202 0.0303 SRP (mgP/l) 0.1 0.18 P (mg) 0.002 0.0036 sample wt dry (g) 0.8173 0.8113 P (mg/kg)) 2.4471 4.4373 Recalcitrant P (mg/kg)) 436 436 Summary average P (mg/kg) KCl 0.734203397 NaOH-Pi 2730.084249 NaOH-Po 3945.980165 HCl 3.442202335 Recalcitrant 436 Total 7116.352274

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184 Table E-4. Phosphorus fractionation data for Koell Site 4 Fractionation Procedure Rep 1 Rep 2 Porewater Filter wt initial (g) 0.0867 0.0863 Filter wt final (g) 0.0894 0.0886 Sediment collected (g) 0.0027 0.0023 SRP in pore water 0.05 0.09 KCl-bioavailable Filter wt initial (g) 0.0863 0.0853 Filter wt final (g) 0.1063 0.1067 Sediment collected (g) 0.02 0.0214 SRP (mgP/l) 0.03 0.08 P (mg) 0.0006 0.0016 sample wt dry (g) 3.0712 3.0617 P (mg/kg)) 0.195363376 0.522585492 NaOH-Fe and Al oxyhydroxides inorganic total organic Filter wt initial (g) 0.0847 0.0849 Filter wt final (g) 0.1447 0.1253 Sediment collected (g) 0.06 0.0404 SRP (mgP/l) 88 190 P (mg) 1.76 3.8 2.04 sample wt dry (g) 3.0112 3.0213 3.0213 P (mg/kg)) 584.4845909 1257.736736 675.206 HCl-Ca and Mg bound Filter wt initial (g) 0.0861 0.0839 Filter wt final (g) 0.1063 0.1142 Sediment collected (g) 0.0202 0.0303 SRP (mgP/l) 0.1 0.18 P (mg) 0.002 0.0036 sample wt dry (g) 2.991 2.991 P (mg/kg)) 0.668672685 1.203610832 Recalcitrant P (mg/kg)) 525 525 Summary average P (mg/kg) KCl 0.358974434 NaOH-Pi 584.4845909 NaOH-Po 675.2060371 HCl 0.936141759 Recalcitrant 525 Total 1785.851816

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185 Table E-5. Phosphorus fractionati on data for San Martin Site 1 Fractionation Procedure Rep 1 Rep 2 Porewater Filter wt initial (g) 0.0865 0.0857 Filter wt final (g) 0.0898 0.0882 Sediment collected (g) 0.0033 0.0025 SRP in pore water 0.03 0.05 KCl-bioavailable Filter wt initial (g) 0.0854 0.0887 Filter wt final (g) 0.1071 0.1129 Sediment collected (g) 0.0217 0.0242 SRP (mgP/l) 0.05 0.04 P (mg) 0.001 0.0008 sample wt dry (g) 1.7894 1.7256 P (mg/kg)) 0.558846541 0.463606861 NaOH-Fe and Al oxyhydroxides inorganic total organic Filter wt initial (g) 0.09 0.0842 Filter wt final (g) 0.2301 0.1542 Sediment collected (g) 0.1401 0.07 SRP (mgP/l) 72 140 P (mg) 1.44 2.8 1.36 sample wt dry (g) 1.6493 1.6556 1.6556 P (mg/kg)) 873.0976778 1691.229766821.4544576 HCl-Ca and Mg bound Filter wt initial (g) 0.0919 0.0885 Filter wt final (g) 0.0962 0.0991 Sediment collected (g) 0.0043 0.0106 SRP (mgP/l) 0.05 0.13 P (mg) 0.001 0.0026 sample wt dry (g) 1.645 1.645 P (mg/kg)) 0.607902736 1.580547112 Recalcitrant P (mg/kg)) 444 444 Summary average P (mg/kg) KCl 0.511226701 NaOH-Pi 873.0976778 NaOH-Po 821.4544576 HCl 1.094224924 Recalcitrant 444 Total 2140.130464

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186 Table E-6. Phosphorus fractionati on data for San Martin Site 2 Fractionation Procedure Rep 1 Rep 2 Porewater Filter wt initial (g) 0.0879 0.0862 Filter wt final (g) 0.0945 0.0933 Sediment collected (g) 0.0066 0.0071 SRP in pore water 0.06 0.05 KCl-bioavailable Filter wt initial (g) 0.0857 0.0897 Filter wt final (g) 0.1116 0.1055 Sediment collected (g) 0.0259 0.0158 SRP (mgP/l) 0.04 0.02 P (mg) 0.0008 0.0004 sample wt dry (g) 1.4485 1.5092 P (mg/kg)) 0.552295478 0.265041081 NaOH-Fe and Al oxyhydroxides inorganic total organic Filter wt initial (g) 0.088 0.0999 Filter wt final (g) 0.1106 0.1866 Sediment collected (g) 0.0226 0.0867 SRP (mgP/l) 162 185 P (mg) 3.24 3.7 0.46 sample wt dry (g) 1.4259 1.4225 1.4225 P (mg/kg)) 2272.249106 2601.054482323.3743409 HCl-Ca and Mg bound Filter wt initial (g) 0.108 Filter wt final (g) 0.1114 Sediment collected (g) 0.0034 0 SRP (mgP/l) 0.17 0.11 P (mg) 0.0034 0.0022 sample wt dry (g) 1.4225 1.4225 P (mg/kg)) 2.390158172 1.546572935 Recalcitrant P (mg/kg)) 678 678 Summary average P (mg/kg) KCl 0.40866828 NaOH-Pi 2272.249106 NaOH-Po 323.3743409 HCl 1.968365554 Recalcitrant 678 Total 3276.03997

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187 Table E-7. Phosphorus fractionati on data for San Martin Site 3 Fractionation Procedure Rep 1 Rep 2 Porewater Filter wt initial (g) 0.0888 0.0899 Filter wt final (g) 0.0971 0.0961 Sediment collected (g) 0.0083 0.0062 SRP in pore water 0.07 0.05 KCl-bioavailable Filter wt initial (g) 0.0867 0.0863 Filter wt final (g) 0.1207 0.1227 Sediment collected (g) 0.034 0.0364 SRP (mgP/l) 0.1 0.07 P (mg) 0.002 0.0014 sample wt dry (g) 2.6589 2.6691 P (mg/kg)) 0.752190756 0.524521374 NaOH-Fe and Al oxyhydroxides inorganic total organic Filter wt initial (g) 0.0876 0.0864 Filter wt final (g) 0.1279 0.1213 Sediment collected (g) 0.0403 0.0349 SRP (mgP/l) 20 252 232 P (mg) 0.4 5.04 4.64 sample wt dry (g) 2.6186 2.6342 2.6342 P (mg/kg)) 152.7533797 1913.294359 1761.446 HCl-Ca and Mg bound Filter wt initial (g) 0.087 0.0869 Filter wt final (g) 0.0946 0.1101 Sediment collected (g) 0.0076 0.0232 SRP (mgP/l) 0.25 0.15 P (mg) 0.005 0.003 sample wt dry (g) 2.611 2.611 P (mg/kg)) 1.914975105 1.148985063 Recalcitrant P (mg/kg)) 276 276 Summary average P (mg/kg) KCl 0.638356065 NaOH-Pi 152.7533797 NaOH-Po 0 HCl 1.531980084 Recalcitrant 276 Total 430.860863

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188 Table E-8. Phosphorus fractionati on data for San Martin Site 4 Fractionation Procedure Rep 1 Rep 2 Porewater Filter wt initial (g) 0.086 0.086 Filter wt final (g) 0.1011 0.1018 Sediment collected (g) 0.0151 0.0158 SRP in pore water 0.07 0.09 KCl-bioavailable Filter wt initial (g) 0.0865 0.0869 Filter wt final (g) 0.1415 0.1081 Sediment collected (g) 0.055 0.0212 SRP (mgP/l) 0.03 0.02 P (mg) 0.0006 0.0004 sample wt dry (g) 0.6374 0.6476 P (mg/kg)) 0.941324129 0.617665225 NaOH-Fe and Al oxyhydroxides inorganic total organic Filter wt initial (g) 0.0876 0.0864 Filter wt final (g) 0.1279 0.1213 Sediment collected (g) 0.0403 0.0349 SRP (mgP/l) 34 156 P (mg) 0.68 3.12 2.44 sample wt dry (g) 1.1866 1.2022 1.179 P (mg/kg)) 573.0659026 2595.2420562069.55047 HCl-Ca and Mg bound Filter wt initial (g) 0.087 0.0869 Filter wt final (g) 0.0946 0.1101 Sediment collected (g) 0.0076 0.0232 SRP (mgP/l) 0.04 P (mg) 0.0008 sample wt dry (g) 1.179 1.179 P (mg/kg)) 0.678541137 Recalcitrant P (mg/kg)) 192 192 Summary average P (mg/kg) KCl 0.779494677 NaOH-Pi 573.0659026 NaOH-Po 2069.550466 HCl 0.678541137 Recalcitrant 192 Total 2836.23734

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189 Table E-9. Phosphorus fractionati on data for Gustavson Site 1 Fractionation Procedure Rep 1 Rep 2 Porewater Filter wt initial (g) 0.0928 0.092 Filter wt final (g) 0.1233 0.1229 Sediment collected (g) 0.0305 0.0309 SRP in pore water 1.8 0.133333333 KCl-bioavailable Filter wt initial (g) 0.0936 0.0927 Filter wt final (g) 0.095 0.0928 Sediment collected (g) 0.0014 1E-04 SRP (mgP/l) 2.3 2 P (mg) 0.046 0.04 sample wt dry (g) 0.1906 0.1796 P (mg/kg)) 241.343127 222.7171492 NaOH-Fe and Al oxyhydroxides inorganic total organic Filter wt initial (g) 0.0846 0.0895 Filter wt final (g) 0.0895 0.0895 Sediment collected (g) 0.0049 0 SRP (mgP/l) 110 123 P (mg) 2.2 2.46 0.26 sample wt dry (g) 0.1857 0.1796 0.1796 P (mg/kg)) 11847.06516 13697.10468 1447.6615 HCl-Ca and Mg bound Filter wt initial (g) 0.0909 0.092 Filter wt final (g) 0.0971 0.0921 Sediment collected (g) 0.0062 0.0001 SRP (mgP/l) 9 22 P (mg) 0.18 0.44 sample wt dry (g) 0.1795 0.1795 P (mg/kg)) 1002.785515 2451.253482 Recalcitrant P (mg/kg)) 665 665 Summary average P (mg/kg) KCl 232.0301381 NaOH-Pi 11847.06516 NaOH-Po 1447.66147 HCl 1727.019499 Recalcitrant 665 Total 15919.13894

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190 Table E-10. Phosphorus fractionatio n data for Gustavson Site 2 Fractionation Procedure Rep 1 Rep 2 Porewater Filter wt initial (g) 0.0894 0.0909 Filter wt final (g) 0.1047 0.1142 Sediment collected (g) 0.0153 0.0233 SRP in pore water 6.2 4.933333333 KCl-bioavailable Filter wt initial (g) 0.0935 0.0928 Filter wt final (g) 0.0939 0.0949 Sediment collected (g) 0.0004 0.0021 SRP (mgP/l) 0.5 1.5 P (mg) 0.01 0.03 sample wt dry (g) 0.2101 0.2066 P (mg/kg)) 47.59638267 145.2081317 NaOH-Fe and Al oxyhydroxides inorganic total organic Filter wt initial (g) 0.0923 0.0869 Filter wt final (g) 0.1136 0.108 Sediment collected (g) 0.0213 0.0211 SRP (mgP/l) 150 170 P (mg) 3 3.4 0.4 sample wt dry (g) 0.1888 0.1855 0.1855 P (mg/kg)) 15889.83051 18328.84097 2156.3342 HCl-Ca and Mg bound Filter wt initial (g) 0.0926 0.0917 Filter wt final (g) 0.1019 0.0977 Sediment collected (g) 0.0093 0.006 SRP (mgP/l) 41 48 P (mg) 0.82 0.96 sample wt dry (g) 0.1795 0.1795 P (mg/kg)) 4568.245125 5348.189415 Recalcitrant P (mg/kg)) 527 527 Summary average P (mg/kg) KCl 96.40225717 NaOH-Pi 15889.83051 NaOH-Po 0 HCl 4958.21727 Recalcitrant 527 Total 21471.37817

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191 Table E-11. Phosphorus fractionatio n data for Gustavson Site 3 Fractionation Procedure Rep 1 Rep 2 Porewater Filter wt initial (g) 0.0911 0.0874 Filter wt final (g) 0.1317 0.1291 Sediment collected (g) 0.0406 0.0417 SRP in pore water 2.5 0.8 KCl-bioavailable Filter wt initial (g) 0.0934 0.0936 Filter wt final (g) 0.0936 0.0943 Sediment collected (g) 0.0002 0.0007 SRP (mgP/l) 2.7 2.8 P (mg) 0.054 0.056 sample wt dry (g) 0.2091 0.2096 P (mg/kg)) 258.2496413 267.1755725 NaOH-Fe and Al oxyhydroxides inorganic total organic Filter wt initial (g) 0.0897 0.0869 Filter wt final (g) 0.1112 0.108 Sediment collected (g) 0.0215 0.0211 SRP (mgP/l) 125 137.5 P (mg) 2.5 2.75 0.25 sample wt dry (g) 0.1876 0.1885 0.1885 P (mg/kg)) 13326.22601 14588.85942 1326.2599 HCl-Ca and Mg bound Filter wt initial (g) 0.0925 0.09 Filter wt final (g) 0.1006 0.099 Sediment collected (g) 0.0081 0.009 SRP (mgP/l) 31 26 P (mg) 0.62 0.52 sample wt dry (g) 0.1795 0.1795 P (mg/kg)) 3454.038997 2896.935933 Recalcitrant P (mg/kg)) 657 657 Summary average P (mg/kg) KCl 262.7126069 NaOH-Pi 13326.22601 NaOH-Po 1326.259947 HCl 3175.487465 Recalcitrant 657 Total 18747.65363

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192 Table E-12. Phosphorus fractionatio n data for Gustavson Site 4 Fractionation Procedure Rep 1 Rep 2 Porewater Filter wt initial (g) 0.0894 0.0901 Filter wt final (g) 0.1013 0.1088 Sediment collected (g) 0.0119 0.0187 SRP in pore water 1.5 0.9 KCl-bioavailable Filter wt initial (g) 0.0929 0.0936 Filter wt final (g) 0.0929 0.0936 Sediment collected (g) 0 0 SRP (mgP/l) 1.1 2 P (mg) 0.022 0.04 sample wt dry (g) 0.2028 0.2094 P (mg/kg)) 108.4812623 191.0219675 NaOH-Fe and Al oxyhydroxides inorganic total organic Filter wt initial (g) 0.0861 0.088 Filter wt final (g) 0.1013 0.1089 Sediment collected (g) 0.0152 0.0209 SRP (mgP/l) 152 157.5 P (mg) 3.04 3.15 0.11 sample wt dry (g) 0.1876 0.1885 0.1885 P (mg/kg)) 16204.69083 16710.87533 583.55438 HCl-Ca and Mg bound Filter wt initial (g) 0.0925 0.09 Filter wt final (g) 0.1006 0.099 Sediment collected (g) 0.0081 0.009 SRP (mgP/l) 31 26 P (mg) 0.62 0.52 sample wt dry (g) 0.1795 0.1795 P (mg/kg)) 3454.038997 2896.935933 Recalcitrant P (mg/kg)) 657 657 Summary average P (mg/kg) KCl 149.7516149 NaOH-Pi 16204.69083 NaOH-Po 583.5543767 HCl 3175.487465 Recalcitrant 657 Total 20770.12855

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193 APPENDIX F PHOSPHORUS SORPTION INDEX (PSI) DATA Table F-1. Year 2005 PSI data for Koell sites TOTAL (BIOTIC+ABIOTIC) DEAD (ABIOTIC) K 1 (1) K 1 (2) K 1 (3) K 1 (1) K 1 (2) K 1 (3) Ci (mg PO4/l) 30.70 30.70 30.70 30.70 30.70 30.70 Cf Live (mg PO4/l) 2.95 4.56 5.13 22.42 15.58 21.47 P sorbed (ug) 905.10 852.43 833.85 270.01 493.07 300.99 Dry mass (g) 10.10 7.75 10.68 7.74 7.84 8.46 ug P sorbed/g sed 89.60 109.99 78.09 34.89 62.89 35.58 PSI 30.04 34.67 24.23 9.03 16.97 9.25 K 2 (1) K 2 (2) K 2 (3) K 2 (1) K 2 (2) K 2 (3) Ci (mg PO4/l) 30.70 30.70 30.70 30.70 30.70 30.70 Cf Live (mg PO4/l) 18.05 16.34 21.47 44.08 50.92 49.59 P sorbed (ug) 412.52 468.28 300.99 -436.33 -659.38 -616.01 Dry mass (g) 10.45 9.76 10.70 9.14 9.94 10.26 ug P sorbed/g sed 39.49 47.97 28.13 -47.73 -66.34 -60.02 PSI 10.48 12.87 7.32 -11.48 -15.72 -14.26 K 3 (1) K 3 (2) K 3 (3) K 3 (1) K 3 (2) K 3 (3) Ci (mg PO4/l) 30.70 30.70 30.70 30.70 30.70 30.70 Cf Live (mg PO4/l) 4.56 2.47 4.18 45.98 36.48 42.37 P sorbed (ug) 852.43 920.59 864.83 -498.29 -188.49 -380.56 Dry mass (g) 11.27 10.91 11.21 5.54 11.10 10.95 ug P sorbed/g sed 75.67 84.36 77.15 -89.91 -16.99 -34.76 PSI 23.85 29.03 24.61 -21.53 -4.17 -8.40 K 4 (1) K 4 (2) K 4 (3) K 4 (1) K 4 (2) K 4 (3) Ci (mg PO4/l) 30.70 30.70 30.70 30.70 30.70 30.70 Cf Live (mg PO4/l) 6.08 2.28 8.93 25.65 29.83 28.31 P sorbed (ug) 802.87 926.78 709.93 164.68 28.37 77.94 Dry mass (g) 10.59 10.31 10.28 10.40 10.28 10.44 ug P sorbed/g sed 75.78 89.86 69.04 15.83 2.76 7.46 PSI 22.98 31.29 19.93 4.04 0.69 1.88

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194 Table F-2. Year 2005 PSI data for Frey sites TOTAL (BIOTIC+ABIOTIC) DEAD (ABIOTIC) F1 (1) F1 (2) F1 (3) F1 (1) F1 (2) F1 (3) Ci (mg PO4/l) 30.70 30.70 30.70 30.70 30.70 30.70 Cf Live (mg PO4/l) 2.97 6.48 1.71 11.70 9.99 13.32 P sorbed (ug) 904.28 789.82 945.37 619.60 675.36 566.77 Dry mass (g) 9.06 9.07 9.09 8.41 8.08 8.02 ug P sorbed/g sed 99.80 87.05 104.04 73.70 83.58 70.65 PSI 33.42 26.18 37.88 20.58 23.79 19.42 F2 (1) F2 (2) F2 (3) F2 (1) F2 (2) F2 (3) Ci (mg PO4/l) 30.70 30.70 30.70 30.70 30.70 30.70 Cf Live (mg PO4/l) 6.48 3.24 3.06 9.90 9.63 9.72 P sorbed (ug) 789.82 895.48 901.35 678.29 687.10 684.16 Dry mass (g) 4.46 4.39 4.47 4.15 4.71 4.46 ug P sorbed/g sed 177.09 203.98 201.78 163.33 145.82 153.37 PSI 53.26 67.46 67.28 46.55 41.70 43.81 F3 (1) F3 (2) F3 (3) F3 (1) F3 (2) F3 (3) Ci (mg PO4/l) 30.70 30.70 30.70 30.70 30.70 30.70 Cf Live (mg PO4/l) 1.62 1.08 2.43 22.50 21.42 15.84 P sorbed (ug) 948.31 965.92 921.89 267.40 302.62 484.59 Dry mass (g) 8.67 8.80 8.96 9.02 8.56 8.00 ug P sorbed/g sed 109.38 109.81 102.88 29.64 35.37 60.59 PSI 40.17 43.12 35.49 7.67 9.20 16.32

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195 Table F-3. Year 2005 PSI da ta for Gustavson sites TOTAL (BIOTIC+ABIOTIC) DEAD (ABIOTIC) G1 (1) G1 (2) G1 (3) G1 (1) G1 (2) G1 (3) Ci (mg PO4/l) 30.70 30.70 30.70 30.70 30.70 30.70 Cf Live (mg PO4/l) 4.94 4.75 11.02 20.90 18.81 26.60 P sorbed (ug) 840.04 846.24 641.77 319.58 387.74 133.70 Dry mass (g) 9.26 9.25 9.34 8.35 8.30 8.47 ug P sorbed/g sed 90.70 91.45 68.73 38.27 46.72 15.79 PSI 28.28 28.67 19.33 9.98 12.33 4.01 G4 (1) G4 (2) G4 (3) G4 (1) G4 (2) G4 (3) Ci (mg PO4/l) 30.70 30.70 30.70 30.70 30.70 30.70 Cf Live (mg PO4/l) 11.97 22.61 21.28 27.93 29.83 30.59 P sorbed (ug) 610.79 263.82 307.19 90.33 28.37 3.59 Dry mass (g) 10.47 10.22 10.30 10.03 10.11 11.04 ug P sorbed/g sed 58.34 25.81 29.84 9.00 2.81 0.33 PSI 16.24 6.67 7.77 2.27 0.70 0.08

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196 Table F-4. Year 2005 PSI data for San Martin sites TOTAL ABIOTIC SM2 (1) SM2 (1) Ci (mg PO4/l) 30.70 30.70 Cf Live (mg PO4/l) 5.40 18.81 P sorbed (ug) 825.04 387.74 Dry mass (g) 10.49 10.26 ug P sorbed/g sed 78.64 37.78 PSI 24.23 9.97 SM4 (1) SM4 (1) Ci (mg PO4/l) 30.70 30.70 Cf Live (mg PO4/l) 7.22 16.34 P sorbed (ug) 765.69 468.28 Dry mass (g) 10.49 10.26 ug P sorbed/g sed 72.99 45.63 PSI 21.65 12.24

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197 Table F-5. Year 2006 PSI data for Koell sites TOTAL (BIOTIC+ABIOTIC) K1 (1) K1 (2) K1 (3) Ci (mg P/l) 50.00 50.00 50.00 Cuaderno 0.33 0.24 0.17 Cf (mg P/l) 3.30 2.40 1.70 P sorbed (ug) 4670.00 4760.00 4830.00 mass i (g) 109.077 112.258 108.622 mass f (g) 116.468 119.94 116.863 samples mass (g) 7.39 7.68 8.24 ug P sorbed/g sed 631.85 619.63 586.09 PSI 179.58 183.31 181.43 K2 (1) K2 (2) K3 (3) Ci (mg P/l) 50.00 50.00 50.00 Cuaderno 0.29 0.39 0.39 Cf (mg P/l) 2.90 3.90 3.90 P sorbed (ug) 4710.00 4610.00 4610.00 mass i (g) 108.413 110.52 104.985 mass f (g) 116.078 118.141 112.333 samples mass (g) 7.67 7.62 7.35 ug P sorbed/g sed 614.48 604.91 627.38 PSI 177.47 168.45 174.71 K3 (1) K3 (2) K31 (3) Ci (mg P/l) 50.00 50.00 50.00 Cuaderno 0.28 0.25 0.34 Cf (mg P/l) 2.80 2.50 3.40 P sorbed (ug) 4720.00 4750.00 4660.00 mass i (g) 108.17 110.36 106.204 mass f (g) 113.962 116.104 111.831 samples mass (g) 5.79 5.74 5.63 ug P sorbed/g sed 814.92 826.95 828.15 PSI 236.40 243.37 234.51 K4 (1) K4 (2) K4 (3) Ci (mg P/l) 50.00 50.00 50.00 Cuaderno 0.18 0.26 0.16 Cf (mg P/l) 1.80 2.60 1.60 P sorbed (ug) 4820.00 4740.00 4840.00 mass i (g) 107.95 111.214 111.45 mass f (g) 112.165 115.392 115.869 samples mass (g) 4.22 4.18 4.42 ug P sorbed/g sed 1143.53 1134.51 1095.27 PSI 351.29 332.22 341.83

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198 Table F-6. Year 2006 PSI da ta for Gustavson sites TOTAL (BIOTIC+ABIOTIC) G1 (1) G1 (2) G1 (3) Ci (mg P/l) 50.00 50.00 50.00 Cf (mg P/l) 14.40 10.40 12.00 P sorbed (ug) 3560.00 3960.00 3800.00 mass i (g) 108.60 111.72 106.96 mass f (g) 121.31 121.83 116.35 samples mass (g) 12.71 10.11 9.39 ug P sorbed/g sed 280.16 391.77 404.82 PSI 67.37 97.53 99.24 G2 (1) G2 (2) G2 (3) Ci (mg P/l) 50.00 50.00 50.00 Cf (mg P/l) 4.00 3.60 2.80 P sorbed (ug) 4600.00 4640.00 4720.00 mass i (g) 104.99 106.45 112.25 mass f (g) 111.27 112.75 118.71 samples mass (g) 6.29 6.29 6.46 ug P sorbed/g sed 731.90 737.21 731.10 PSI 203.19 207.30 212.09 G3 (1) G3 (2) G3 (3) Ci (mg P/l) 50.00 50.00 50.00 Cf (mg P/l) 0.90 1.30 0.90 P sorbed (ug) 4910.00 4870.00 4910.00 mass i (g) 205.84 110.08 106.18 mass f (g) 211.74 115.86 112.65 samples mass (g) 5.90 5.78 6.47 ug P sorbed/g sed 832.77 842.85 758.77 PSI 281.89 320.81 307.49 G4 (1) G4 (2) G4 (3) Ci (mg P/l) 50.00 50.00 50.00 Cf (mg P/l) 6.90 3.50 5.30 P sorbed (ug) 4310.00 4650.00 4470.00 mass i (g) 109.07 106.77 111.22 mass f (g) 120.48 117.72 122.02 samples mass (g) 11.41 10.95 10.80 ug P sorbed/g sed 377.81 424.77 413.89 PSI 112.70 138.93 127.84

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199 Table F-7. Year 2006 PSI data for San Martin sites TOTAL (BIOTIC+ABIOTIC) SM1 (1) SM1 (2) SM1 (3) Ci (mg P/l) 50.00 50.00 50.00 Cf (mg P/l) 4.00 0.40 13.20 P sorbed (ug) 4600.00 4960.00 3680.00 mass i (g) 112.34 111.74 112.26 mass f (g) 128.06 128.14 127.44 samples mass (g) 15.72 16.41 15.18 ug P sorbed/g sed 292.58 302.33 242.44 PSI 81.23 116.19 58.84 SM2 (1) SM2 (2) SM2 (3) Ci (mg P/l) 50.00 50.00 50.00 Cf (mg P/l) 11.00 10.60 11.20 P sorbed (ug) 3900.00 3940.00 3880.00 mass i (g) 107.96 110.88 108.41 mass f (g) 122.62 125.27 123.22 samples mass (g) 14.67 14.39 14.81 ug P sorbed/g sed 265.92 273.86 262.07 PSI 65.80 68.03 64.72 SM3 (1) SM3 (2) SM3 (3) Ci (mg P/l) 50.00 50.00 50.00 Cf (mg P/l) 0.40 0.40 1.20 P sorbed (ug) 4960.00 4960.00 4880.00 mass i (g) 104.99 110.09 109.90 mass f (g) 116.41 121.70 122.36 samples mass (g) 11.43 11.62 12.45 ug P sorbed/g sed 434.14 426.92 391.84 PSI 166.84 164.07 127.26 SM4 (1) SM4 (2) SM4 (3) Ci (mg P/l) 50.00 50.00 50.00 Cf (mg P/l) 12.80 22.80 21.20 P sorbed (ug) 3720.00 2720.00 2880.00 mass i (g) 106.77 109.95 111.26 mass f (g) 118.64 121.97 123.27 samples mass (g) 11.86 12.02 12.01 ug P sorbed/g sed 313.61 226.35 239.78 PSI 76.36 51.94 55.42

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200 APPENDIX G EQUILIBRIUM PHOSPHORUS CONCENTRATION (EPCO) DATA y = 2.5506x 0.8473 R2 = 0.9904 0.00 1.00 2.00 3.00 4.00 5.00 6.00 0.000.501.001.502.002.503.00P in solution (mg/l)P sorbed/ g sed Figure G-1. 2005 EPCO data for Koell site 1 y = 4.2103x 2.4939 R2 = 0.9963 -2.00 0.00 2.00 4.00 6.00 8.00 10.00 0.000.501.001.502.002.503.00 P in solution (mg/l)P sorbed/ g sed Figure G-2. 2005 EPCO data for Koell site 2 y = 2.5565x 1.6468 R2 = 0.9183 0.00 1.00 2.00 3.00 4.00 5.00 6.00 7.00 0.000.501.001.502.002.503.003.50 P in solutionP sorbed/ g sed Figure G-3. 2005 EPCO data for Koell site 3

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201 y = 2.7791x 1.3723 R2 = 0.98 0.00 1.00 2.00 3.00 4.00 5.00 6.00 7.00 0.000.501.001.502.002.503.00 P in solution (mg/l)P sorbed/g sed Figure G-4. 2005 EPCO data for Koell site 4 y = 2.9154x 6.7276 R2 = 0.8211 -5.00 -4.00 -3.00 -2.00 -1.00 0.00 1.00 2.00 3.00 0.000.501.001.502.002.503.003.50 P in solution (mg/l)P sorbed/g sed Figure G-5. 2005 EPCO data for Gustavson site 1 y = 4.8375x 6.2693 R2 = 0.9543 -4.00 -2.00 0.00 2.00 4.00 6.00 8.00 10.00 0.001.002.003.004.00 P in solution (mg/l)P sorbed/ g sed Figure G-6. 2005 EPCO data for Gustavson site 2

PAGE 202

202 y = 2.4813x 1.2772 R2 = 0.9629 -2.00 0.00 2.00 4.00 6.00 8.00 0.000.501.001.502.002.503.003.50 P in solution (mg/l)P sorbed/g sed Figure G-7. 2005 EPCO data for Gustavson site 3 y = 4.4838x 7.0869 R2 = 0.853 -6.00 -4.00 -2.00 0.00 2.00 4.00 6.00 8.00 0.000.501.001.502.002.503.003.50 P in solution (mg/l)P sorbed/g sed Figure G-8. 2005 EPCO data for Gustavson site 4 y = 2.5786x 0.7148 R2 = 0.9808 0.00 1.00 2.00 3.00 4.00 5.00 6.00 7.00 0.000.501.001.502.002.503.00 P in solution (mg/l)P sorbed/g sed Figure G-9. 2005 EPCO data for San Martin site 2

PAGE 203

203 y = 2.2826x 0.6539 R2 = 0.9963 0.00 1.00 2.00 3.00 4.00 5.00 6.00 0.000.501.001.502.002.503.00 P in solution (mg/l)P sorbed/g sed Figure G-10. 2005 EPCO data for San Martin site 4 y = 3.8476x 2.117 R2 = 0.9596 0.00 2.00 4.00 6.00 8.00 10.00 12.00 0.000.501.001.502.002.503.003.50 P in solution (mg/l)P sorbed/g sed Figure G-11. 2005 EPCO data for Frey site 1 y = 6.0421x 1.9738 R2 = 0.992 0.00 5.00 10.00 15.00 20.00 0.000.501.001.502.002.503.003.50 P in solution (mg/l) P sorbed/g sed Figure G-12. 2005 EPCO data for Frey site 2

PAGE 204

204 y = 2.8625x 0.6496 R2 = 0.9906 0.00 2.00 4.00 6.00 8.00 10.00 0.000.501.001.502.002.503.003.50 P in solution (mg/l)P sorbed/g sed Figure G-13. 2005 EPCO data for Frey site 3 y = 13.973x 1.2979 R2 = 0.9961 0.00 10.00 20.00 30.00 40.00 0.000.501.001.502.002.50 P in solution (mg/l)P sorbed/ g sed Figure G-14. 2006 EPCO data for Koell site 1 y = 10.015x 1.1658 R2 = 0.9994 0.00 5.00 10.00 15.00 20.00 25.00 0.000.501.001.502.002.50 P in solution (mg/l)P sorbed/ g sed Figure G-15. 2006 EPCO data for Koell site 2

PAGE 205

205 y = 17.417x 1.6786 R2 = 0.9979 0.00 10.00 20.00 30.00 40.00 0.000.501.001.502.002.50 P in solution (mg/l)P sorbed/ g sed Figure G-16. 2006 EPCO data for Koell site 3 y = 21.713x 1.2012 R2 = 0.9999 0.00 10.00 20.00 30.00 40.00 50.00 0.000.501.001.502.002.50 P in solution (mg/l)P sorbed/ g sed Figure G-17. 2006 EPCO data for Koell site 4 y = 8.119x 2.3081 R2 = 0.9866 -2.00 -1.00 0.00 1.00 2.00 3.00 4.00 5.00 0.000.200.400.600.801.00 P in solution (mg/l)P sorbed/ g sed Figure G-18. 2006 EPCO data for Gustavson site 1

PAGE 206

206 y = 4.629x 0.2785 R2 = 0.9993 0.00 0.50 1.00 1.50 2.00 2.50 3.00 3.50 4.00 0.000.200.400.600.801.00 P in solution (mg/l)P sorbed/ g sed Figure G-19. 2006 EPCO data for Gustavson site 2 y = 2.0728x 0.6432 R2 = 0.9814 -2.00 -1.00 0.00 1.00 2.00 3.00 4.00 5.00 6.00 -0.500.000.501.001.502.002.503.003.50 P in solution (mg/l)P sorbed/g sed Figure G-20. 2006 EPCO data for Gustavson site 3 y = 13.6x 3.5117 R2 = 0.9941 0.00 2.00 4.00 6.00 8.00 10.00 12.00 14.00 0.000.200.400.600.801.001.201.40 P in solution (mg/l)P sorbed/ g sed Figure G-21. 2006 EPCO data for Gustavson site 4

PAGE 207

207 y = 5.4606x 0.3245 R2 = 0.9994 0.00 2.00 4.00 6.00 8.00 10.00 12.00 14.00 0.000.501.001.502.002.50 P in solution (mg/l)P sorbed/ g sed Figure G-22. 2006 EPCO data for San Martin site 1 y = 8.905x 0.613 R2 = 0.9993 0.00 5.00 10.00 15.00 20.00 25.00 0.000.501.001.502.002.50 P in solution (mg/l)P sorbed/ g sed Figure G-23. 2006 EPCO data for San Martin site 2 y = 8.2791x 1.1447 R2 = 0.9965 0.00 5.00 10.00 15.00 20.00 0.000.501.001.502.002.50 P in solution (mg/l)P sorbed/ g sed Figure G-24. 2006 EPCO data for San Martin site 3

PAGE 208

208 y = 6.8756x 0.3777 R2 = 1 0.00 5.00 10.00 15.00 20.00 0.000.501.001.502.002.50 P in solution (mg/l)P sorbed/ g sed Figure G-25. 2006 EPCO data for San Martin site 4

PAGE 209

209 APPENDIX H INTERPRETIVE MATERIALS USED FOR COMMUNITY PARTICIPATION Figure H-1. Pamphlet distribu ted to citizens to inform a nd receive their input on the implementation of modified ditches. A) Fr ont and back pages and B) inside pages A B

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210 PorfavorNo ArrojarBasuraProyectoPiloto de ParticipaconComunitariaparala Depuraconde AguasResidualesen Zanjas Figure H-2. Image of the sign painted and installed at each of the modified ditches, translated as Please do not throw trash: Pilot project fo r community participation of wastewater treatment in ditches.

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211 LIST OF REFERENCES Adam, K., T. Krogstad, L. Vrale, A. K. Sovik, and P. D. Jenssen. 2007. Phosphorus retention in the filter materials shell sand and Filtralit e P Batch and column experiment with synthetic P solution and secondary wastewater. Ecological Engineering 29:200-208. Akratos, C. S., and V. A. Tsihrintzis. 2007. E ffect of temperature, HRT, vegetation and porous media on removal efficiency of pilot-scal e horizontal subsurface flow constructed wetlands. Ecological Engineering 29:173-191. Alexander, R. B., R. A. Smith, and G. E. Schw arz. 2000. Effect of stream channel size on the delivery of nitrogen to the Gulf of Mexico. Nature 403:758-761. Andersen, J. M. 1976. An ignition method for de termination of phosphorus in lake sediments. Water Resoures 10:329-331. Ann, Y., K. R. Reddy, and J. J. Delfino. 2000. Influence of redox potential on phosphorus solubility in chemically amended wetla nd organic soils. Ecol ogical Engineering 14:169180. APHA. 1992. Standard methods for the exam ination of water and wastewater, 5210 B. Washington D.C. Armitage, P. D., K. Szoszkiewi cz, J. H. Blackburn, and I. Nesbitt. 2003. Ditch communities: a major contributor to floodplain biodivers ity. Aquatic Conservation-Marine and Freshwater Ecosystems 13:165-185. ASTM. 1985. Standard test method for partic le-size of soil, D 422-63. Pages 117-127 Annual book of ASTM standards. American Society for Testing and Materi als, Philadelphia. Axt, J. R., and M. R. Walbridge. 1999. Phospha te removal capacity of palustrine forested wetlands and adjacent uplands in Virginia. Soil Science Society of America Journal 63:1019-1031. Bache, B. W., and E. G. Williams. 1971. Phosphate sorption index for soils. Journal of Soil Science 22:289-301. Backstrom, M. 2003. Grassed swales for stormwat er pollution control dur ing rain and snowmelt. Water Science and Technology 48:123-132. Barber, M. E., S. G. King, D. R. Yonge, a nd W. E. Hathhorn. 2003. Ecology ditch: A best management practice for storm water r unoff mitigation. Journal of Hydrologic Engineering 8:111-122. Barlow, K., D. Nash, H. Turral, and R. Grays on. 2003. Phosphorus uptake and release in surface drains. Agricultural Water Management 63:109-123.

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225 BIOGRAPHICAL SKETCH Lynn Velisha Saunders was born May 6, 1977 in Anderson, Indiana and grew up in El Paso, Texas. She attended UT Austin for a year before deciding that Texas just wasnt big enough. To her parents dismay she left school and headed west to California to work in Yosemite National Park. After a year, she moved nor th to Arcata, California where she attended College of the Redwoods and late r transferred to Humboldt Stat e University (HSU). During her three years in the Environmen tal Resources Engineering prog ram at HSU, Lynn had several opportunities to work and travel in Central Am erica where she became aware of the human and ecosystem health effects resulting from the lack of access to adequate water and sanitation. Working as a student research assistant fo r two years at the Arcata Marsh and Wildlife Sanctuary opened her eyes to the field of ecol ogical engineering. She was inspired by the idea that water treatment systems could be designed to provide multiple and mutual benefits such as wetland habitat, aquaculture, recreation and aesthe tics. After completing her B.S. degree at HSU, she began her graduate program in Environmenta l Engineering Sciences at the University of Florida with an emphasis on Ecological Engineering. Her hope is that her Ph.D. research on wastewat er ditches in Peru will serve as an example of a low-tech approach of decentralized and community-based wastewater management in a region that currently lacks the resources to implement conven tional water treatment. After completion of their degrees, Lynn and her husband Tom plan to continue working on applied water management issues in the U.S. and abroad.