EFFECTS OF OXYGEN AND NITRATE ON MERCURY CYLCING IN THE PROFUNDAL
ZONE OF LAKES AND RESERVOIRS
By
RICARDI DUVIL
A dissertation submitted in partial fulfillment of
the requirements for the degree of
DOCTOR OF PHILOSOPHY
WASHINGTON STATE UNIVERSITY
Department of Civil & Environmental Engineering
MAY 2015
To the Faculty of Washington State University:
The members of the Committee appointed to examine the dissertation of RICARDI
DUVIL find it satisfactory and recommended that it be accepted.
_________________________________
Marc M. Beutel. Ph.D., Chair
_________________________________
David Yonge, Ph.D
_________________________________
John Harrison, Ph.D
_________________________________
Tom Grizzard, Ph.D
ii ACKNOWLEDGEMENT
I would like to thank my family for their endless love and support throughout this
journey. I would especially like to thank my father for instilling in me a passion for education
and the courage and grit to do whatever it takes to accomplish my dreams. I will never forget
how many times my father told me “education is my passport to a better future.” I will always be
grateful to you, father, for coaching me to lead by example and to make it my passion to become
an advocate for the underprivileged.
I would also like to thank my first academic advisor and my friend, Dr. Walter Johnson,
who exposed me to water research projects. I would like to express my deepest gratitude to Dr.
Marc Beutel, my advisor and mentor, without you, I wouldn’t be here. You have shown me
incredible patients and afforded me with innumerable opportunities to grow as a professional
through countless hours of work on policy and technical writing. I would like to thank Dr. Tom
Grizzard and his team at Occoquan Monitoring Laboratory, whose resources made my research
possible. Many thanks to Dr. David Yonge and Dr. John Harrison for taking the time to be on my
committee and evaluating my dissertation. I would also like to thank David Drury and the Santa
Clara Valley Water District for allowing my research group to conduct its investigations. Many
thanks to Dr. Stephen Dent, whose guidance and advice helped me accomplish my laboratory
work.
Finally, I would like to give a hearty thank you to Dr. Steven Stehr, Dr. Nick Lovrich,
Dr. Greg Browder, Dr. John Reganold, Dr. Linda Klein, Andrea Hokanson, Taylor Altomare,
Abraham Ambaw and Alex Julien and his family for all the support.
iii EFFECTS OF OXYGEN AND NITRATE ON MERCURY CYLCING IN THE PROFUNDAL
ZONE OF LAKES AND RESERVOIRS
Abstract
by Ricardi Duvil, Ph.D.
Washington State University
May 2015
Chair: Marc W. Beutel
Methylmercury, a neurotoxin, has been identified as a serious public health concern. To date,
effective remediation techniques for preventing and remediating methylmercury contamination
have remained elusive, mainly due to the lack of knowledge in regard to how methylmercury is
generated and degraded in the aquatic environment. This dissertation evaluated methylmercury
accumulation in both replicate experimental chambers containing undisturbed sediment-water
interface samples from mercury-contaminated lakes in northern California, and a two-year field
study evaluating spatial and temporal patterns of nitrate, oxygen and methylmercury during
summer stratification in Occoquan Reservoir, Virginia. The primary objective of this research
was to access how oxidants, including nitrate, affect the cycling of mercury in the profundal zone
of lakes and reservoirs. Results demonstrated that methylmercury accumulation in both
experimental chamber water and field sampling was lower under aerobic and anoxic (no oxygen
but nitrate present) conditions versus anaerobic conditions. In Guadalupe Reservoir, California,
MeHg efflux measured in replicate bench-scale chambers averaged 5.5 ng/m2/d under aerobic
conditions and 22 ng/m2/d under anaerobic conditions. In Almaden Lake, California, MeHg
iv efflux was -2.3 ng/m2/d (uptake) under aerobic conditions and 11 ng/m2·d under anaerobic
conditions. Like oxygen, nitrate repressed MeHg efflux from sediments from Almaden Lake to <
0.2 ng/m2/d. In Occoquan Reservoir, Virginia, MeHg was < 0.1 ng/L in a tributary enriched with
nitrate from treated wastewater effluent compared to 0.04-1.1 ng/L in a low nitrate reference
tributary. MeHg accumulation correlated with elevated iron and manganese concentrations.
Nitrate addition to the reservoir likely kept downstream bottom waters from going highly
reduced, thereby reducing MeHg accumulation in bottom waters. This dissertation also describes
a fellowship performed with the World Bank, as part of the WSU NPSIRE IGERT program,
evaluating environmental and economic policy in the water/wastewater sector in Cartagena,
Colombia. The investigation found that Cartagena's water supply coverage has increased from
73% to 99.9% and sewer coverage from 61% to 90%. However, a lack of transparent
environmental policy regarding public participation has limited the effectiveness of achieving
meaningful benefits without negatively impacting the public.
v TABLE OF CONTENTS
ACKNOWLEDGEMENTS……………………………………………………………….…….. iii
ABSTRACT.……………………………………………………………………………….…… iv
LIST OF TABLES……………………………………………………………………………….xii
LIST OF FIGURES……………………………………………………………………………..xiii
DEDICATION…………………………………………………………………………………...xv
CHAPTER
1. INTRODUCTION…………………………………………………………………….1
1.1 Overview of Dissertation.…………..………………………………………………...1
1.2. OUTLINE AND RATIONALE FOR DISSERATION CHAPTERS……………….1
1.2.1. Chapter 2: Overview of Nitrogen Cycling in Aquatic Ecosystems…...........2
1.2.2. Chapter 3: Policy Study in Cartagena, Colombia……………………..……2
1.2.3. Chapter 4: Chamber Study in northern California…………………….……4
1.2.4. Chapter 5: Field Study at Occoquan Reservoir, Virginia…………………..5
1.3. FUNDING SOURCES………………………………………………………………7
1.4. REFERENCES………………………………………………………………………8
2. OVERVIEW OF NITROGEN CYCLING IN AQUATIC ECOSYSTEMS
2.1 INTRODUCTION…………………………………………………………….9
2.2. NITRIFICATION…………………………………………………………..10
2.2.1. Ammonia Oxidation…………………………………………........11
2.2.2. Nitrite Oxidation…………………………………………………..14
2.3. DENITRIFICATION……………………………………………………….14
2.3.1. Conventional Denitrification……………………………………...14
vi 2.3.2. Dissimilatory Reduction of Nitrate to Ammonia………………….17
2.3.3. Nitrous Oxide Production during Denitrification…………………18
2.4. ANAEROBIC AMMONIUM OXIDATION………………………………21
2.5. IMPACTS OF NITRATE ON MERCURY CYCLING……………………23
2.6. CASE STUDIES OF MANAGED NITRATE ADDITION IN LAKES
AND RESERVOIRS FOR WATER QUALITY IMPROVEMENT………25
2.6.1. Nitrate as a Pollutant of Concern………………………………….25
2.6.2. Nitrate as a Resource to Improve Water Quality………………….28
2.6.3. Sequential Treatment of Sediment with Nitrate…………………..30
2.6.4. Quarry Lakes Experiment, Habo, Sweden…………………….......31
2.6.5. Lake Lyng, Denmark……………………………………………..32
2.6.6. Schlei Fjord, Germany……………………………………………32
2.6.7. Occoquan Reservoir, USA………………………………………...33
2.6.8. Lake Tegel, Germany……………………………………………..34
2.6.9. Lake Onondaga, USA…………………………………………….35
2.7 REFERENCES..……………………………………………………………...37
3. ENVIRONMENTAL AND ECONOMIC POLICY STUDY IN CARTAGENA,
COLOMBIA: WATER SUPPLY AND ENVIRONMENTAL MANAGEMENT
PROJECT
3.1. BACKGROUND…………………………………………………………...48
3.2. PROJECT OBJECTVES…………………………………………………...51
3.3. CARTAGENA: A CULTURAL AND INDUSTRIAL CITY……………...54
3.4. CONFRONTING CARTAGENA’S ENVIRONMENTAL CRISIS………55
vii 3.5. RESETTING THE INSTITUTIONAL FRAMEWORK FOR WATER
SUPPLY AND MANAGEMENT……..…………………………………..56
3.6. DEVELOPING NEW ENVIRONMENTAL POLICY AND
MANAGEMEN……………………………………………………………58
3.7. FINANCING AND IMPLEMENTING A PHASED
INFRASTRUCTURE DEVELOPMENT PROGRAM…………………..58
3.8. SOCIAL IMPACT CONCERNS RAISED BY THE LOCAL
NON-GOVERMENTAL ORGANIZATION……………………………….63
3.9. ADDRESSING CONCERNS RAISED BY THE LNGO………………….64
3.10. IMPLEMENTATIONS METRICS AND SUMMARY
OF ACUACAR’S ACHIEVEMENTS……………………………………65
3.11. ACUACAR’S METHODS FOR OVERCOMING
IMPEDIMENTS……………………………………………………..........68
3.12. COMMISSIONING OF CARTAGENA’S WASTEWATER
SYSTEM…………………………………………………………………..70
3.13. COASTAL ENVIRONMENTAL IMPROVEMENTS……………...........70
3.14. FUTURE CHALLENGES IN COASTAL ENVIRONMENTAL
RESTORATION…………………………………………………………..78
3.15. ANALYSIS OF THE CARTAGENA WATER SUPPLY AND
ENVIRONMENTAL MANAGEMENT PROJECT……………………...79
3.16. ENVIRONMENTAL POLICY ISSUES REVEALED
IN THIS PROJECT……………………………………………………….80
3.17. SUMMARY OF KEY LESSONS IN THE CARTAGENA
viii PROJECT…………………………………………………………………82
3.18. REFERENCES…………………………………………………………...85
4. EFFECTS OF OXYGEN, NITRATE AND ALUMINUM HYDROXIDE ON
METHYLMERCURY EFFLUX FROM MINE-CONTAMINATED
PROFUNDAL LAKE SEDIMENTS
4.1. INTRODUCTION……………………………………………………..…...88
4.2. MATERIALS AND METHODS……………………………………..…….93
4.2.1. Study sites………………………………………………………..93
4.2.2. Collection of Sediment Cores……………………………............95
4.2.3 Sediment Incubations……………………………………………..95
4.2.4. Water Quality Analyses………………………………………….96
4.2.5. Sediment Quality Analyses………………………………………97
4.3. RESULTS……………………………………………………………..........98
4.3.1. Sediment Quality………………………………………………...98
4.3.2 Study 1: Total Mercury, Sulfate and Methylmercury
Effluxes under Aerobic and Anaerobic Conditions………………98
4.3.3 Study 1: Iron and Manganese in Experimental Chambers…..........99
4.3.4 Study 1: Cumulative Mass of Methylmercury, Manganese
and Iron in Anaerobic Phase…………………………………….100
4.3.5 Study 2: Efflux Rates for Methylmercury, Manganese,
Iron and Ammonia.……………………………………………..100
4.4. DISCUSSION……………………………………………………………..101
4.4.1. Comparison of Release Rates of Methylmercury Efflux………..101
ix 4.4.2. Efflux Rates in Almaden Lake versus Guadalupe Reservoir……103
4.4.3. Effects of Oxygen and Nitrate in Experimental Chambers…......108
4.4.4. Sediment Ammonia Release…………………………………….110
4.4.5. Effect of Sodium Aluminate in Chamber Water………………...112
4.4.6. Implications for Lake Management……………………………..115
4.5 REFERENCES………………………………………………………….....118
5. SPATIAL AND TEMPORAL PATTERNS OF NITRATE, OXYGEN
AND METHYLMERCURY DURING SUMMER STRATIFICATION
IN OCCOQUAN RESERVOIR, VIRGINIA
5.1. INTRODUCTION…………………………………………………………138
5.2. MATERIALS AND METHODS…………………………………………..142
5.2.1. Study Site………………………………………………………..142
5.2.2. Collection of Field Sampling…………………………………....144
5.2.3. Water Quality Analysis………………………………………….145
5.3. RESULTS………………………………………………………………….145
5.3.1. Water Quality…………………………………………………………....145
5.3.2. Correlated Variation between Methylmercury, Oxygen, Nitrate,
Manganese and Iron……………….…………………………………....149
5.4. DISCUSSION……………………………………………………………...150
5.4.1. Comparison between Stations RE-30 and RE-35,
2012 and 2013…………………………………………………………..150
5.4.2. Variation of Methylmercury, Manganese and Iron
in Downstream Region of Reservoir…...……………………………155
x 5.5. CONCLUSION……………………………………………………………157
5.6. REFERENCES……………………………………………………………159
xi LIST OF TABLES
3.1. Implementation Metrics Water Supply and Wastewater Treatment………………………...73
4.1. Sediment Quality in Lake Almaden and Guadalupe Reservoir……………………………135
4.2. Fluxes measured in experimental chambers under aerobic and anaerobic conditions
in Lake Almaden and Guadalupe Reservoir……………………………………………….136
4.3. Comparison of net flux of methylmercury rates to similar studies in freshwater
and marine systems………………………………………………………………………...137
xii LIST OF FIGURES
2.1. The nitrogen cycle in a surface flow wetland. Figure from Kadlec and Knight (1996).
Treatment Wetlands. First Edition, CRC Press, Boca Raton, Florida……………………….10
3.1. Map of project area………………………………………………………………………......58
3.2. Preliminary Water Quality Monitoring in Cartagena………………………………………..81
4.1. Map of the Guadalupe River watershed in San Jose, California. Figure is from the
Draft Final Conceptual Model Report, June 2004 by Tetra Tech, Inc……………………..130
4.2. Profiles of temperature, dissolved oxygen (DO), and methylmercury (MeHg).
A. Lake Almaden sampled September 13, 2011. B. Guadalupe Reservoir sampled
B. August 16, 2011. ……………………………………………………………………...131
4.3. 2008 dissolved oxygen (DO), nitrate and methylmercury (MeHg) concentrations
in bottom water near the sediment-water interface of Lake Almaden in 2008………….....132
4.4. Total mercury and sulfate efflux from Lake Almaden (left) and Guadalupe
Reservoir (right) to aerobic conditions (clear bars) and anaerobic conditions
(black bars). Error bars are one standard deviation (n = 4)………………………………...133
4.5. Cumulative mass of methylmercury, manganese and iron in chamber water
from anaerobic incubation of Lake Almaden sediment…………………………………….134
4.6. Efflux rates from Lake Almaden sediment for methylmercury (ng/m2·d), manganese
(mg/m2·d), iron (mg/m2·d), and ammonia (mg/m2·d) in anaerobic control and three
experimental treatments: aerobic, anoxic with nitrate addition, and anaerobic with
sodium aluminate addition.....................................................................................................135
xiii 5.1. Map of Occoquan Reservoir complements of Francisco Cubas and Tom Grizzard
of Virginia Tech. Stations monitored in 2012 include Stations 1 (RE30), 2 (RE35),
3 (RE20), 4 (RE15), 5 (RE10), 6 (RE5) and 7 (RE02). Stations monitored in 2013
include 1 (RE30), 2 (RE35), 3 (RE20)……………………………………………………..164
5.2. 2009 dissolved oxygen (DO), nitrate and phosphate in the bottom waters of the
Occoquan Reservoir, Virginia……………………………………………………………...165
5.3. Bottom water quality upstream in Occoquan Reservoir in 2012……………….................166
5.4. Bottom water quality downstream in Occoquan Reservoir in 2012……………………….167
5-5. Bottom water quality upstream in Occoquan Reservoir in 2013…………………………..168
5.6. Correlations between DO/nitrate and methylmercury (MeHg) in bottom waters…………169
5.7. Correlations between iron/manganese and methylmercury (MeHg) in bottom waters……170
xiv Dedication
This dissertation is dedicated to the memory of my beloved father, Succes Joseph, a man full of
love and compassion. You will always be profoundly appreciated for your unflinching
commitment to my career pursuit!
“To accomplish great things, you must not only act, but also dream; not only plan, but also
believe.”
xv CHAPTER 1
INTRODUCTION
1.1 Overview of Dissertation
This dissertation examines the use of nitrate as a resource to control mercury cycling in
lakes and reservoirs. The dissertation is divided into five chapters. The Introduction herein
provides an outline and rationale for chapters 2, 3, 4 and 5 and an introduction to the Nitrogen
Systems: Policy-oriented Integrated Research and Education (NSPIRE) and the Bullitt
Foundation Environmental Fellowship programs that helped to fund my research.
Chapter 2 is a discussion of the nitrogen cycle and the linkages between the nitrogen
cycle and the mercury cycle. Chapter 2 also includes a discussion of case studies where nitrate
has been used to enhance surface water quality. Chapter 3 is a policy paper based on my NSPIRE
internship with the World Bank and describes the implementation of a water and sanitation
project to enhance access to services and improve coastal water quality in Cartagena, Colombia.
Chapters 4 presents a chamber study that evaluated oxygen and nitrate control on mercury fluxes
from mercury-contaminated reservoir sediments from San Jose, CA. Chapter 5 presents results of
field work at Occoquan Reservoir, VA that focused on understanding the spatial and temporal
patterns of nitrate and mercury in the water column of a reservoir that is loaded high levels of
nitrate.
1 1.2. Outline and Rationale for Dissertation Chapters
1.2.1. Chapter 2: Overview of Nitrogen Cycling in Aquatic Ecosystems
Chapter 2 is a review of major processes and factors controlling nitrogen transport and
transformations in aquatic ecosystems. This chapter first describes the main processes controlling
nitrogen cycling in aquatic environment. The chapter emphasizes heterotrophic denitrification,
with organic carbon as an electron donor; autotrophic denitrification using reduced sulfur;
dissimilatory nitrate reduction to ammonium; and anoxic oxidation of ammonium. This chapter
also reviews environmental factors that affect denitrification rates such as nitrate concentration,
nutrient availability, microbial acclimation, pH, temperature, and presence of toxic compounds.
This chapter also discusses denitrification process that influence air quality. For example,
incomplete denitrification in aquatic ecosystems may enhance the biogenic production of the
potent greenhouse gas nitrous oxide (N2O). Furthermore, this chapter summarizes the linkages
between the nitrogen cycle and the mercury cycle, illustrating that nitrogen, in form of nitrate,
can affect mercury biogeochemistry in low-oxygen environments. Finally, this chapter
acknowledges that significant accumulation of nitrate in aquatic ecosystems can be a major
pollution source for humans and wildlife. However, a number of case studies are presented that
illustrate how managed nitrate addition can be used to enhance surface water quality.
1.2.2. Chapter 3: Policy Study in Cartagena, Colombia
Chapter 3 describes the history and development of environmental policies in
Cartagena, Colombia, where I worked for the World Bank as part of my NSPIRE fellowship.
The chapter focuses on the Cartagena Water Supply, Sewerage and Environmental Management
2 Project, which was part of the World Bank Water and Sanitation program. The objectives of the
project were to: 1) improve the water and sewage services of the city of Cartagena and the
sanitary conditions of Cartagena’s poorest population; 2) facilitate the environmental cleanup of
surrounding water bodies, including Cartagena Bay, Caribbean beaches, and Cienaga de la
Virgen Lagoon; and 3) improve the sustainability of water quality. The project had a number of
interrelated components including construction of water supply and sewerage system, the
construction of a submarine outfall for discharging treated effluent to the Caribbean Sea near
Punta Canoa, Cartagena, and the implementation of environmental and social policies to promote
project success.
Unexpectedly, the proposed submarine outfall off the coast of Punta Canoa turned out to
be especially contentious. Local citizens argued that the outfall would negatively affect the
coastal zone's fisheries, which was the primary source of food and income for the local people.
They believed that the project's untreated wastewater would be discharged into the sea,
contaminate marine life, and have serious and permanent impacts on people’s health and
livelihoods. This belief was of particular concern for the indigenous people of Punta Canoa,
whose lives are inextricably linked to the health of the Caribbean Sea.
This chapter focuses on the main efforts to improve water quality in Cartagena,
mechanisms for overcoming impediments, and key points for success. This chapter also
evaluates two fundamental policy aspects of the Cartagena effort, the lack of a water
management program and efforts to empower local citizens’ participation to enhance project
outcomes and best environmental management practices. Finally, this chapter discusses the
difficulties that any environmental policy faces when technical drivers take precedence over
3 social and cultural concerns of local stakeholders. Consequently, policy lessons from this project
are also revealed.
1.2.3. Chapter 4: Chamber Study at Almaden Lake and Guadalupe Reservoir, California
Chapter 4 of this dissertation is a summary of a sediment-water-interface chamber study
that was conducted in Almaden Lake and Guadalupe Reservoir, two mercury-impacted lakes
managed by the Santa Clara Valley Water District near San Jose, California. The study evaluated
the sediment release of methylmercury (HgCH3+) under aerobic conditions, anaerobic conditions,
anoxic conditions with nitrate addition, and anaerobic conditions with sodium aluminate
addition.
The hypothesis of this study was that anaerobic chambers would accumulate more
methylmercury than aerobic chambers or anoxic chambers with nitrate. This is because
anaerobic conditions, in which both oxygen and nitrate are depleted, tend to promote
methylmercury production in sediments through a number of potential mechanisms. For
example, anaerobic microorganisms such as sulfate-reducing bacteria (SRB) have been
implicated in methylmercury production (Benoit et al., 2003). Anoxic conditions also promote
the dissolution of iron and manganese oxides in surficial sediments, which may release sorbed
methylmercury upon reduction (Mathews et al., 2013). Only a handful of studies have
documented nitrate suppression of methylmercury accumulation in lake bottom waters, including
one by the consulting firm CH2MHill, which added nitrate to a number of small lakes in
Minnesota (Austin, 2011), and a large-scale nitrate addition pilot project in Onondaga Lake, New
York (Matthews et al., 2013).
4 An additional hypothesis was that addition of aluminum, which fosters the formation of
an aluminum hydroxide floc at the sediment-water interface, would prevent the release of
methylmercury from anaerobic sediment to overlaying water by sorbing effluxing
methylmercury. Aluminum addition as sodium aluminate and/or aluminum sulfate is a common
lake management additive to enhance the sorption capacity of surfacial sediments. Aluminum
additions are generally used to control sediment phosphorus release (DeGasperi et al., 1993).
Little research has been conducted on the effects of this common lake additive on mercury
cycling.
1.2.4. Chapter 5: Field Study at Occoquan Reservoir, Virginia
Chapter 5 presents the results of a two-year field study that evaluated the spatial and
temporal patterns of nitrate and mercury during the summer stratification period in Occoquan
Reservoir, VA. The reservoir receives large inputs of nitrate from tertiary treated wastewater that
is discharged into the upper part of the reservoir (Randall and Grizzard, 1995). This nitrate tends
to flow with cool water into the profundal zone of the lake and impact biogeochemical processes
in the bottom of the lake. The hypothesis was that methylmercury would only be observed in
substantial levels in bottom waters if oxygen and nitrate were not present at the sediment-water
interface, that is, when or where water conditions shifted from anoxic (no oxygen but some
nitrate) to anaerobic (no oxygen or nitrate). The rationale for this hypothesis was that the
production of methylmercury, as noted above in sub-section 1.2.3, is usually associated with
anaerobic conditions that enhance the activity of SRB and the dissolution of mercury-rich metal
oxides in surfacial sediments.
5 Only a limited number of field studies have documented interactions between nitrogen
and mercury cycling in large, human-impacted reservoir ecosystems. A comparable evaluation of
methylmercury accumulation was performed in Lake Onondaga in Syracuse, NY, a mercurypolluted lake that receives 20% of its annual inflow from the Syracuse Metropolitan wastewater
treatment plant (WWTP). Todorova et al. (2009) reported that the presence of nitrate from
wastewater effluent alleviated anaerobic conditions in bottom waters and repressed
methylmercury accumulation. Nitrate is currently being added to the profundal zone of the
Onondaga Lake in an effort to further inhibit methylmercury release from sediments (Matthews
et al., 2013).
1.3. Funding Sources
This dissertation was funded in part by a two-year fellowship with the Nitrogen Systems:
Policy-oriented Integrated Research and Education (NSPIRE) program at Washington State
University and a two-year Bullitt Foundation Environmental Fellowship. NSPIRE was funded by
the Integrative Graduate Education and Research Traineeship (IGERT) program. IGERT is part
of the National Science Foundation's flagship interdisciplinary training program, educating U.S.
Ph.D. scientists and engineers by building on the foundations of their disciplinary knowledge
with interdisciplinary training. Nitrogen pollution is one of world's most widespread, costly, and
challenging environmental problems. NSPIRE focuses on creating a new generation of scientists
with broad and rigorous training in nitrogen cycling and who can integrate nitrogen cycle science
for effective communication with public policy makers. This program prepares graduate students
for leadership roles in science and engineering to help solve nitrogen cycle problems. In addition,
6 diversity among the students contributes to their preparation to solve large and complex research
problems of significant scientific and societal importance at national and international levels.
In 2012, I received a Bullitt Foundation Environmental Fellowship. This program focuses
on educating environmentally knowledgeable graduate students from a community underrepresented in the environmental movement. Students awarded the fellowship must also have
demonstrated an exceptional capacity for leadership and scholarship. This fellowship embraces
the idea of stewardship, supporting projects that touch the future and promote a healthy,
sustainable environment. The Foundation supports a variety of tactics to advance its mission.
Grantees tend to be creative in devising projects and campaigns to build understanding, advance
policy innovations, and garner the political power needed to affect change. These tactics
generally include the following: developing a diverse pipeline of future environmental and
sustainability leaders; ensuring compliance with environmental laws and policies; and convening
environmental, scientific, business, and political leaders to clarify differences and seek common
ground.
7 1.4 References
Austin, D., Scharf, R., Carroll, J., Enochs, M. 2011. Suppression of mercury methylation by
hypolimnetic liquid calcium nitrate amendment in Round Lake, Eden Prairie, Minnesota.
North American Lake Management Society. Spokane, USA. November 2011.
Benoit, J. M., et al. 2003. Chapter 19: Geochemical and Biological Controls over Methylmercury
Production and Degradation in Aquatic Ecosystems. American Chemical Society. p. 262–
296.
DeGasperi, C.L., D.E. Spyridakis, E.B. Welch. 1993. Alum and nitrate as controls of short-term
anaerobic sediment P release: An in vitro comparison. Lake Reserv. Manage. 8:49–59.
Matthews, D., D. Babcock, J. Nolan, A. Prestigiacomo, S. Effler, C. Driscoll, S. Todorova, K. Kuhr.
2013. Whole-lake nitrate addition for control of methylmercury in mercury-contaminated
Onondaga Lake, NY. Environ. Res. 125, 52–60.
Randall, C.W., Grizzard, T. J. 1995. Management of the Occoquan River Basin a 20-year case
history. Water Science Technology. 32:235–243.
Todorova, S., Driscoll, C., Matthews, D., Effler, S., Hines, E. 2009. Evidence of regulation of
monomethyl mercury by nitrate in a seasonally stratified, eutrophic lake. Environmental
Science and Technology. 43:6572–6578.
8 CHAPTER 2
OVERVIEW OF NITROGEN CYCLING IN AQUATIC ECOSYSTEMS
2.1. Introduction
Nitrogen (N) cycling in aquatic ecosystems is complex. Figure 1 of N cycling in a
surface-flow wetland illustrates this complexity. The figure includes the following key
components of N cycling in aquatic ecosystems: 1) ecological compartments (sediment, water,
air, biota); 2) ecological niches (aerobic, anoxic, anaerobic); 3) N species [organic N, ammonium
(NH4+), ammonia (NH3), nitrite ions (NO2-), nitrate ions (NO3-), nitrogen gas (N2), nitrous oxide
(N2O)]; 4) N transformations (assimilation, ammonification, nitrification, denitrification,
anammox, fixation); and 5) N transfer processes (leaching, volatilization, diffusion). This
overview will focus on exploring three key N transformations of significance in aquatic
ecosystems: nitrification, denitrification, and anaerobic ammonium oxidation (anammox).
Microorganisms, particularly bacteria, play a major role in most N transformations. Bacteria
obtain metabolic energy by redox processes, for example, using nitrate, nitrite, and soluble N
oxides as terminal respiratory oxidants under oxygen-limiting conditions during denitrification
(Burging et al., 2011). Thus, this discussion will also focus on bacteria and their capacity to
transform N compounds in aquatic ecosystems. Finally, this discussion will examine the linkages
between the N cycles and mercury (Hg) cycles and present case studies illustrating how nitrate
can be used as a resource to improve water quality.
9 Figure 2-1. The nitrogen cycle in a surface flow wetland. Figure from Kadlec and Knight (1996).
Treatment Wetlands. First Edition, CRC Press, Boca Raton, Florida.
2.2. Nitrification
Nitrification, the microbial oxidation of ammonia to nitrate, is a key transformation in the
N cycle. In aquatic environments, decomposing microorganisms, such as bacteria and fungi,
facilitate ammonification, the breakdown of organic N to ammonia. Next, during the process of
nitrification, ammonium is oxidized to nitrite and then to nitrate. Thus, nitrifying
microorganisms such as nitrosomonas and nitrobacter within aquatic ecosystems reduce total
10 ammonia concentrations through nitrification according to the following overall reactions (Chin,
2006):
4 NH4+ + 6 O2 → 4 NO2- + 8 H+ + 4 H2O
4 NO2- + 2 O2 → 4 NO3
NH4+ + 2 O2 → NO3- + 2 H+ + H2O
Nitrification requires the presence of oxygen; therefore, nitrification can only happen in
aerobic environments such as surface layers of soils and sediments in micro-oxic zones (Barnard
et al., 2005). Because nitrification is central to the accumulation and ultimate loss of nitrate (via
denitrification), factors regulating nitrification are critical to eutrophication and elevated nitrate
concentrations in fresh waters. Nitrification converts excessive ammonia to nitrate, which can
then be lost from the aquatic environment via denitrification (see Section 2.3); therefore, this
process can play a potentially important role in protecting surface waters from eutrophication
(Hagopian and Riley, 1998). However, some important human health consequences must be
considered. If nitrate assimilation does not occur at a rate at least equal to nitrate production, the
aquatic environment may accumulate nitrate. Excessive nitrate in drinking water is linked to
several illnesses in humans including methemoglobinemia and the internal production of
carcinogenic nitrosamines (Vitousek et al., 1997; Carpenter et al., 1998).
2.2.1. Ammonia Oxidation
Ammonia oxidation, a reaction mediated by both ammonia-oxidizing bacteria (AOB) and
ammonia-oxidizing archaea (AOA), is the first and rate-limiting step of nitrification (Falkowski
11 et al., 2008). AOB are thought to be largely responsible for the oxidation of ammonia to nitrite in
natural environments. However, Erguder et al. (2009) summarized the current knowledge on
environmental conditions influencing the amount and presence of AOA and identified several
site-related properties and environmental factors. These properties and factors included
ammonium, organic carbon, temperature, salinity, oxygen, pH, sulfide, phosphate, and some
metal contents. The two key variables that appear to control AOA versus AOB dominance of
ammonia oxidation include temperature and pH.
Temperature exhibited different effects on the abundance and diversity of AOA and
AOB. Higher temperature tends to select for AOA because elevated temperature enhances the
decomposition of extracellular DNA or dead cells in sediment (Wu et al., 2013), which may
promote higher AOA abundance. Temperature variations could lead to the fluctuations of other
environmental factors such as organic carbon, oxygen availability, and ammonia. These
additional factors might also contribute to community shifts of ammonia oxidizers and lead to
significant variations in AOA and AOB. For example, elevated temperature could increase the
decomposition rate of organic matter and release more organic molecules, which could also
affect AOB (Park and Noguera, 2004). Racz et al. (2010) reported that organic carbons such as
peptone and glucose could affect the abundance and diversity of AOB in a mixed culture with
the heterotroph community. Elevated temperature could also increase the decomposition rates of
organic matter and decrease the dissolved oxygen availability due to heterotrophic bacteria,
which have higher affinities for dissolved oxygen than the nitrifiers (Starry et al., 2005).
Elevated decomposition rates of organic matter would also increase the ammonia availability,
which is the substrate for nitrification (Jantti et al., 2011). Consequently, variations due to
12 elevated temperature could change the abundances and community compositions of AOA and
AOB in the lake sediments and, in turn, affect the N cycling process.
Besides temperature, low pH environments may favor AOA because lower pH could
drive the conversion from ammonia to ammonium, decreasing substrate availability and
increasing energy stress that may select for AOA over AOB as the dominating ammonia oxidizer
(Valentine, 2007). As a result, AOA communities more easily adapt to low pH environments
compared to their AOB counterparts. This speculation was recently confirmed by cultivation and
characterization of an obligate, acidophilic, thaumarchaeal ammonia oxidizer from a nitrifying
acidic soil (Lehtovirta-Morley et al., 2011).
Both AOA and AOB have ammonia mono-oxygenase, one of the key enzymes
responsible for the transformation of ammonia to hydroxylamine (NH2OH) (Jin et al., 2010).
Hydroxylamine then undergoes a redox reaction with oxygen to produce nitrite and energy. This
reaction generates a small amount of energy relative to many other types of metabolism; as a
result, nitrosofiers (nitrifying bacteria) are notoriously slow growers. Additionally, aerobic
ammonia oxidizers are autotrophs, fixing carbon dioxide to produce organic carbon much like
photosynthetic organisms, but using ammonia as the energy source instead of light.
Conditions for ammonia oxidizers are best at the interface between oxic environments
with adequate oxygen concentrations and anoxic environments with high ammonium
concentrations (Voytek and Ward, 1995). This part of the N cycle is especially sensitive to pH.
Acidification forces a proton onto ammonia, producing ammonium and lowering concentrations
of ammonia, the actual substrate used by ammonia oxidizers. Lower ammonia concentrations
may result in slower rates of ammonia oxidation (Kirchman, 2012). Despite limited research, it is
well-known that AOB and AOA play a vital role in linking the input of fixed N, through N-
13 fixation and remineralization, to potential loss processes, including denitrification and anammox
(Bouskill et al., 2011).
2.2.2. Nitrite Oxidation
Nitrite oxidation is the second step of the nitrification process, during which nitrite is
quickly oxidized to nitrate by a separate group of prokaryotes, known as nitrite-oxidizing
bacteria. Some of the genera involved in nitrite oxidation include Nitrospira, Nitrobacter,
Nitrococcus, and Nitrospina (Koenneke et al., 2005). Although nitrite rarely builds up in the
environment, the nitrite oxidation reaction is an essential component of nitrification. Similar to
ammonia oxidizers, the energy generated from the oxidation of nitrite to nitrate is small, and thus
growth rates are low (Zehr et al., 2003). In fact, ammonia- and nitrite-oxidizers must oxidize
many molecules of ammonia or nitrite in order to fix a single molecule of carbon dioxide
(Bernhard, 2012). For complete nitrification, both ammonia oxidation and nitrite oxidation must
occur.
2.3. Denitrification
2.3.1. Conventional Denitrification
Denitrification is the process that converts nitrate to N gas, thus removing bioavailable N
and returning it to the atmosphere. Dinitrogen gas is the ultimate end product of denitrification,
but other intermediate gaseous form, including nitric oxide and nitrous oxide, a potent
greenhouse gas (see subsection 2.3.3). Unlike nitrification, denitrification is an anaerobic
process, occurring mostly in soils and sediments and anoxic zones in lakes and oceans. Similar to
14 N fixation, denitrification is carried out by a diverse group of prokaryotes, but recent evidence
suggests that some eukaryotes are also capable of denitrification (Risgaard-Petersen et al., 2006).
Some denitrifying bacteria include species in the genera Bacillus, Paracoccus, and
Pseudomonas. Denitrifiers are chemoorganotrophs and thus must also be supplied with some
form of organic carbon. The overall energy reactions are (Metcalf and Eddy, 2003):
6 NO3- + 2 CH3OH → 6 NO2- + 2 CO2 + 4 H2O
6 NO2- + 3 CH3OH → 3 N2 + 3CO2 + 3 H2O + 6OH6 NO3- + 5 CH3OH → 5 CO2 + 3 N2 + 7 H2O + 6 OH-
Denitrification is an important microbial process with beneficial consequences for water
quality (Burgin et al., 2012). The importance of denitrification in the N cycle in marine and
freshwater ecosystems has been the focus of an increasing number of studies in recent decades
(Seitzinger et al., 1993). Every year, nearly 50 million tonnes of reactive N enter marine
ecosystem via rivers, resulting in eutrophication of coastal ecosystems (Galloway et al., 2004).
Denitrification is the major pathway of fixed N loss from aquatic systems. Thus, the process is a
critical component of the global N budget and a balancing mechanism for removing
anthropogenic N along the terrestrial-freshwater-marine continuum (Galloway et al., 2004;
Seitzinger et al., 2006). Additionally, denitrification in wastewater treatment plays a beneficial
role by removing nitrate from wastewater effluent, thereby reducing the chances that the treated
discharge will cause undesirable environmental consequences (e.g., algal blooms). However, the
denitrification process is a major concern in agricultural areas. Significant losses from some
surface-applied N sources can occur, leading to limited plant growth (Bernhard, 2012).
15 In lakes, denitrification occurs in low-oxygen bottom waters, where rates generally range
from 0.2 to 1.9 µmol N/L/d (Seitzinger, 1988). In lakes, denitrification rates are greater in the
sediments compared to the water column. In the absence of oxygen, denitrifiers respire nitrate
sequentially to nitrite, nitric oxide, and nitrous oxides, and finally to dinitrogen gas, with organic
carbon used as the electron donor (Devol, 2008). In consuming organic matter, nitrate acts as the
terminal electron acceptor for the oxidation of organic matter under anaerobic conditions. In
aquatic sediments, nitrate is usually converted to dinitrogen gas, with a variable but small
fraction escaping as nitrous oxide (Burgin et al., 2007). During this process, microorganisms
regenerate inorganic nutrients such as carbon dioxide, ammonium, and phosphate, which are
necessary to sustain continued primary production. However, this process is controlled by a
variety of environmental parameters and their interactions, including substrate availability,
temperature, and/or oxygen concentration. Why one or the other parameter dominates is under
debate (Deutsch et al., 2010; Seitzinger, 1988). The specific issue of nitrous oxide production
during denitrification is discussed in greater detail in subsection 2.3.3.
An alternative to heterotrophic, biological denitrification is autotrophic, reduced sulfurdriven denitrification. Common autotrophic denitrifiers, Thiobacillus denitrificans and
Thiomicrospira denitrificans, can use various reduced sulfur compounds as electron donors
(Campos et al., 2008; Sengupta et al., 2007; Krishnakumar and Manilal, 1999). Sulfide,
particularly, can fuel the activity of some autotrophic denitrifiers. Also, sulfide inhibition can
reduce denitrification rates and stimulate nitrous oxide production because nitrous oxide
reductase, the enzyme catalyzing the reduction of nitrous oxide to dinitrogen, is sensitive to
sulfide (Brunet and Garcia-Gil, 1996; Joye, 2002; Porubsky et al., 2009). The overall reaction for
sulfur-driven denitrification is (Soares, 2002):
16 5S0 + 6NO3− + 2H2O → 5SO42- + 3N2− + 4H−
Hydrogen sulfide may play an important role in regulating N-cycling processes and has
the ability to repress N transformations, such as nitrification and heterotrophic denitrification.
However, in lakes, where nitrate and reduced sulfur compounds occur, chemolithoautotrophs can
use reduced sulfur as an energy source and nitrate as the electron acceptor to produce N gas, thus
enhancing denitrification rates (Burgin et al., 2012).
2.3.2. Dissimilatory Reduction of Nitrate to Ammonia
Fermentative dissimilatory nitrate reduction to ammonia (DNRA) is a process by which
nitrate reduction can produce ammonium via a bacterial-mediated heterotrophic process in
anaerobic environments. This process is different from the assimilatory process in that ammonia
is not directly incorporated into biomass. Fermentative DNRA couples organic matter oxidation
to nitrate reduction by fermentation reactions (Tiedje, 1988). The reaction has been found in
anoxic sediments with bacteria of the Thioploca and Thiomargarita species. Both types of
bacteria are able to concentrate nitrates within their own cells for the subsequent oxidation of
sulphur-containing compounds in reduced form. In this way, the bacteria are able to reduce
nitrate to ammonium passing through nitrite as an intermediary compound. This reaction,
although still needing clarification, would potentially supply nitrite and ammonium to the
anammox reaction in anoxic sediments (Burgin et al., 2007; Francis et al., 2007). This process
has been identified as an important sink for nitrate in organic-rich marine and freshwater
sediments (Gardner et al., 2006). The overall reaction is (Schumacher et al, 1992):
17 NO3- → NO2- → NH4+
DNRA represents a potential positive feedback to ecosystems because it converts nitrate
to ammonia, a biologically available form of N that can be taken up by plants (Scott et al., 2008).
However, ammonia generated by DNRA could also fuel nitrification. Thus, DNRA may be an
important pathway for nitrate removal in aquatic sediments, particularly in sediments containing
labile carbon, reduced sulfur, and reduced iron (Burgin et al., 2007). Additionally, in highly
reducing environments, sulfides rather than organic substrates can serve as electron donors for
sulfur-driven lithoautotrophic DNRA (Gardner et al., 2006). Sulfide concentrations play a critical
role in determining the dominant pathway of nitrate reduction. At low sulfide concentrations,
nitrate is reduced mainly via denitrification; however, at higher sulfide concentrations, DNRA
becomes the dominant pathway (Senga et al., 2006; Brunet and Garcia-Gil, 1996). As noted by
Brunet et al. (1996), DNRA accounted for up to 30% of the total nitrate reduction in freshwater
sediments amended with calcium nitrate.
2.3.3. Nitrous Oxide Production during Denitrification
Nitrous oxide is a denitrification byproduct of concern. In all systems, the formation of
nitrous oxide during denitrification results in tropospheric warming and stratospheric ozone
depletion (Davidson and Seitzinger, 2006). Nitrous oxide is produced during several
microbiological processes, including nitrification, denitrification, and dissimilatory nitrate
reduction to ammonium (De Wilde and De Bie, 2000). However, denitrification and
chemolithotrophic nitrification appear to be the main biological sources of nitrous oxide
18 emission in natural systems (Bonin et al., 2002). Recently, attention has been directed to redox
processes in the N cycle that operate in tandem, stressing the importance of considering total
nitrous oxide production rates from nitrification and denitrification (Meyer et al., 2008; Davidson
and Seitzinger, 2006). In addition, nitrous oxide production is reliant on a number of
environmental factors such as inorganic N concentrations, sediment redox potential, and organic
carbon availability (Oliveira Fernandes et al., 2010). Therefore, assessment of environmental
parameters and their interrelationships with net nitrous oxide production in sediments is crucial
for determining the key parameters governing nitrous oxide formation.
In many aquatic systems, most of the nitrous oxide produced during denitrification is
subsequently reduced to dinitrogen gas. However, the ratio of nitrous oxide to dinitrogen gas is
variable and remains uncertain because a comprehensive understanding of denitrification rates
and controlling factors across ecosystems is lacking (Davidson and Seitzinger, 2006; McCutchan
et al., 2003). Some key factors include dissolved oxygen concentration, pH, temperature, nitrate
concentrations, and carbon availability (de Klein et al., 2001, 2010). These factors influence
denitrifier populations and their enzyme activities, and hence, affect the production of nitrous
oxide relative to dinitrogen gas.
Because denitrification is an anaerobic reaction, oxygen availability is one of the most
important factors inhibiting this process in aquatic environments. The amount of nitrous oxide
produced via denitrification is a function of the overall denitrification rate, which is often
controlled by oxygen concentration (Mulholland et al., 2002). While denitrification is inhibited
in the presence of oxygen, a potential source of nitrous oxide under low-oxygen conditions is
nitrification by ammonium-oxidizing bacteria. On a global basis, well-oxygenated flowing
waters are a much smaller source of nitrous oxide than aquatic ecosystems such as estuaries or
19 freshwater lakes that exhibit periods of hypoxia or fluctuate between oxic and anoxic conditions
(Beaulieu et al., 2014).
In addition to oxygen concentrations, pH and temperature can affect nitrous oxide
production. For example, at lower pH levels, enhancement of abiotic transformations of nitrate to
nitrite may contribute to an increased proportion of nitrous oxide production (Koskinen and
Keeney, 1982). Water temperature can cause temporal fluctuations in denitrification rates
(Ryden, 1983). Because the activation energy of nitrous oxide reduction is higher than that of
nitrous oxide production (Holtan-Hartwig et al., 2002), more nitrous oxide is produced at low
temperatures, resulting in increased production of nitrous oxide relative to dinitrogen gas
(Keeney et al., 1979; Avalakki et al., 1995).
The ratio of nitrous oxide to dinitrogen gas can also be influenced by biotic and abiotic
transformations of nitrogenous compounds that occur during denitrification. For example,
presence of suitable electron donors and denitrifying bacteria, possessing metabolic enzymes
capable of enhancing nitrous oxide production rates, can affect reaction rates and redox potential,
which facilitate nitrous oxide accumulation. As a result, nitrous oxide production is a complex
process that cannot be easily related to denitrification fluxes (Barnard et al., 2005). Meyer et al.
(2008) argued that the nitrous oxide produced by denitrification in deeper sediment layers is
consumed during its diffusion toward the sediment-water interface in nutrient-enriched
mangrove sediments.
In terrestrial systems, lower nitrate concentrations usually result in lower nitrous oxide to
dinitrogen gas ratios (Zaman et al., 2007). As nitrate concentration decreases, nitrous oxide acts
as a major electron acceptor, thereby decreasing nitrous oxide production (Swerts et al., 1996;
Dendooven et al., 1997). Increasing carbon availability tends to decrease the ratio of nitrous
20 oxide to dinitrogen gas (Dendooven et al., 1998). Senbayram et al. (2009), for example, observed
low nitrous oxide emissions following a large input of labile carbon in soils. This observation
was possibly due to microbial immobilization of some of the added N (Tiedje, 1988) or
dissimilatory nitrate reduction to ammonium (DNRA). Under anaerobic conditions, nitrate
concentration may control the denitrification product ratio of nitrous oxide to dinitrogen gas,
whereas labile carbon concentration may control the denitrification rate (Tiedje, 1988).
2.4. Anaerobic Ammonium Oxidation
Anammox (anaerobic ammonium oxidation) is the oxidation of ammonium coupled with
the reduction of nitrite under anaerobic conditions. This process is considered more
thermodynamically favorable than aerobic ammonium oxidation (Hu et al., 2011). Anammox is
an important pathway for N gas production in anaerobic marine sediments (Dalsgraad et al.,
2005). Anammox has been identified as an alternative microbial pathway of N gas production
and, from a biogeochemical perspective, can be considered a denitrifying process (Devol, 2008).
Anammox reduces nitrite to nitric oxide, which then reacts in a one-to-one stoichiometry with
ammonium to form hydrazine, and finally dinitrogen gas (Kartal et al., 2011). The overall
reaction is (Dalsgaard et al., 2005):
NH4+ + NO2- → N2 + 2 H2O
Although aerobic ammonia oxidation and anammox both act on ammonia, each process
involves different organisms. Anammox is mediated only by bacteria from the Planctomycetes
21 phylum; aerobic ammonia oxidation is carried out by both bacteria and archaea (Kirchman,
2012). The first described anammox bacterium was Brocadia anammoxidans (Dalsgaard et al.,
2005) Anammox bacteria oxidize ammonia by using nitrite as the electron acceptor to produce
gaseous N. Anammox bacteria were first discovered in anoxic bioreactors of wastewater
treatment plants, but have since been found in a variety of aquatic systems, including lowoxygen zones of the ocean, coastal and estuarine sediments, mangroves, and freshwater lakes.
The rate of anammox appears to be regulated by the availability of reduced substrates, nitrite
concentrations, and sediment organic content. However, in marine sediments, the reduction of
nitrate, and thus the production of nitrite, is fast enough to allow the anaerobic ammonium
oxidation (Dalsgraad et al., 2005).
In marine environments, temperature plays an important role in the anammox process.
Anammox is microbially catalyzed and therefore dependent on temperature. Ammonium
oxidizers have a temperature optimum from 12° to 15° C, whereas denitrifying bacteria have a
much higher temperature optimum between 25° and 30° C (Saad et al., 1993). The lower
temperature optimum of the anammox bacteria enables them to outcompete denitrifies in cold
environments.
In addition, organic matter content in sediments is a key factor for the anammox process.
The more organic matter mineralized, the more nitrite produced, and the higher the anaerobic
ammonium oxidation rate. Normally, organic matter loading of sediments decreases with
increasing water depth because of the long settling time of particles. The longer it takes for
organic matter to sink toward the ground, the higher the possibility of it being mineralized before
arrival. Thus, we can expect decreasing absolute anammox rates with increasing water depth
(Saad et al., 1993; Dalsgaard and Thamdrup, 2002; Dalsgaard et al., 2005; Schubert et al., 2006).
22 The importance of anammox in N-cycling processes across a range of aquatic ecosystems is
being increasingly recognized. In some areas of the ocean, anammox may be responsible for a
significant loss of N (Kuypers et al., 2005). However, Ward et al. (2009) argue that
denitrification, rather than anammox, is responsible for most N loss in other areas. Regardless of
whether anammox or denitrification is responsible for most N loss in the ocean, clearly,
anammox represents an important process in the global N cycle (Bernhard, 2012).
In wastewater treatment systems, the anammox process has several advantages. For
example, oxygen addition can be reduced by 60%. In turn, the energy required to perform the
oxygenation process can also be reduced. Anammox bacteria do not require organic carbon and
produce very little biomass, reducing the amount of sludge disposal (Fux et al., 2002; Egli et al.,
2001). In addition, greenhouse gas emissions are enormously reduced using the anammox
process. The overall carbon dioxide emission of a wastewater treatment plant can be reduced
with 95%, because the anammox process itself consumes carbon dioxide. Furthermore, the
production of nitrous oxide, which easily forms as a by-product of conventional denitrification,
is eliminated using the anammox process (Van Loosdrecht et al., 2004; Van Dongen et al.,
2001). Overall, the anammox process provides a sustainable alternative for nitrification.
2.5. Impacts of Nitrate on Mercury Cycling
Hg contamination in the water column and sediments of aquatic ecosystems raises
concerns related to biological exposure, including potential health effects in humans and wildlife
caused by consumption of contaminated fish (Benoit et al., 2003). On a global scale, rates of
atmospheric Hg deposition have increased three-fold as a result of human post-industrial
23 activities. An estimated 5-20% of deposited Hg is delivered to aquatic ecosystems where it may
accumulate in aquatic biota (Swain et al., 2007). In the U.S., three-fourths of states currently
have blanket fish consumption advisories due to elevated Hg in fish tissue. Nearly half of the
total U.S. lake acres and over a third of U.S. river miles are subject to Hg-related advisories
(USEPA, 2009). Hg cycling in aquatic ecosystems is greatly influenced by metabolic activity,
including primary production, decomposition processes, and sulfate reduction. Of primary
concern is the production of methylmercury (MeHg) which is a neurotoxin that strongly
bioaccumulates in aquatic food webs. MeHg represents more than 95% of total Hg in fish
biomass (Tchounwou et al., 2003).
The accumulation of MeHg in the anaerobic bottom waters of lakes is a recognized
phenomenon (Watras, 2009). Hg typically enters a lake in the inorganic form via aerial
deposition and watershed runoff. Inorganic Hg can complex with dissolved organic carbon
(DOC), and co-precipate with settling particles such as metal oxides of iron (Fe) and manganese
(Mn) (Watras, 2009; Chadwick et al., 2006). MeHg is produced in anaerobic environments by
SRB, which use sulfate as the terminal electron acceptor during the oxidation of organic matter
(Benoit et al., 2003). Sulfate reduction proceeds in the absence of energetically favorable
electron acceptors such as oxygen, nitrate, and oxidized manganese and iron. MeHg can also
enter the water column from the sediment as it is released from metal oxide complexes that
become reduced after the onset of anoxia (Merritt and Amirbahman, 2008; Chadwick et al.,
2006).
Due to its potential effect on redox conditions at the sediment-water interface, increased
nitrate concentration should reduce MeHg efflux from sediments by both of the mechanisms
described above: MeHg production by SRB and MeHg release from reduced metal oxides. SRB
24 activity in lake sediments and related internal production of MeHg may be inhibited by
maintaining adequate concentrations of alternate, energetically favorable electron acceptors, such
as oxygen and/or nitrate (Stephan et al., 1988). As detailed by Golterman (2001), oxidationreduction potential at the sediment-water interface is particularly critical in regulating the flux of
reduced and oxidized substances into and out of profundal sediments. Depressed oxidationreduction potential at the sediment-water-interface has a profound impact on metal cycling, as
reduced Fe (II) and Mn (II) tend to accumulate in anaerobic waters (Bryant et al., 2011). The
presence of nitrate at the sediment-water interface may poise redox potential in surfacial
sediments and repress metal oxide reduction and associated efflux of MeHg sorbed to metal
oxides.
Recent studies confirm nitrate’s ability to repress MeHg accumulation in lakes (Mathews
et al., 2013; Todorova et al., 2009). For example, Matthews et al. (2013) conducted a two-year
pilot study of nitrate addition to Onondaga Lake, a Hg-contaminated lake in New York. Results
showed that maximum concentrations of MeHg and phosphate in bottom waters decreased by
94% and 95%, respectively. The research analysis suggested that increased sorption to Fe and
Mn oxyhydroxides in surficial sediments was the likely regulating mechanism. Clearly more
study is needed to assess the potential for oxidants such as nitrate to inhibit MeHg efflux from
lake sediment, as well as the mechanisms by which oxidants repress MeHg efflux.
2.6. Case Studies of Managed Nitrate Addition in Lakes and Reservoirs for Water Quality
Improvement
2.6.1. Nitrate as a Pollutant of Concern
25 Nitrate is recognized as a critical pollutant in surface and ground waters. The most
remarkable global impacts include the leaching of nitrate to groundwater, the eutrophication of
surface waters and resultant marine “dead zones,” atmospheric deposition that acidifies
ecosystems, and the emission of N oxides (NOx) that deplete stratospheric ozone (Keeney and
Hatfield, 2007; Beever et al., 2007; Foley et al., 2005). These widespread environmental changes
can also threaten human health (Galloway et al., 2008; Guillette and Edwards, 2005; Galloway et
al., 2004; Vitousek et al., 1997; Fan and Steinberg, 1996; Jordan and Weller, 1996). In surface
waters, nitrate can stimulate algal productivity leading to eutrophication and associated water
quality problems and impairment of aquatic biota.
Whereas phosphorus (P) is a critical limiting nutrient in aquatic ecosystems (Schindler,
2012), N limitation or co-limitation of N and P in freshwater lakes is more common than
generally recognized (Elser et al., 2007). Therefore, nitrate is a potentially significant driver of
eutrophication in many freshwater systems. N-limited coastal waters are especially vulnerable to
eutrophication and the development of hypoxic conditions (Howarth and Marino, 2006). Nitrate
pollution from agricultural activities is responsible for the hypoxic dead zone at the mouth of the
Mississippi River in the Gulf of Mexico (Mitsch et al., 2001).
Nitrate contamination also limits the use of groundwater resources for human
consumption. A recent study concluded that nitrate contamination of drinking water wells,
resulting from agricultural activities, poses a health risk to hundreds of thousands of people in
California's Central Valley (Lund, 2012). When present in large quantities, nitrate in drinking
water is a potential health hazard to infants due to methemoglobinema, also known as blue baby
syndrome (Knobeloch et al., 2000). In addition, scientific evidence suggests that ingested nitrates
may result in mutagenicity, teratogenicity, and birth defects (Camargo and Alonso, 2006).
26 Unsuitably high nitrate levels may also affect pregnant women and adults with hereditary
cytochrome b5 reductase deficiency. In addition, nitrate and nitrite ingestion in humans has been
linked to goitrogenic (anti-thyroid) actions on the thyroid, fatigue and reduced cognitive
functioning due to chronic hypoxia, maternal reproductive complications including spontaneous
abortion, and a variety of carcinogenic outcomes deriving from N-nitrosamines formed via
gastric nitrate conversion in the presence of amines (Ward et al., 2005).
The concentration of nitrate fluctuates with the rate of nitrification and denitrification
processes. Nitrification requires the presence of oxygen; as a result, nitrification occurs only in
oxygen-rich environments and the surface layers of soils and sediments. However, the process of
nitrification has some important consequences. For example, ammonium ions are positively
charged and therefore stick to negatively charged clay particles and soil organic matter. The
positive charge prevents ammonium N from being leached out of the soil by rainfall. In contrast,
the negatively charged nitrate ion is not held by soil particles and can be washed out of the soil,
leading to decreased soil fertility and nitrate enrichment of downstream surface and groundwater
(Ravishankara et al., 2009; Compton et al., 2011; Harrison, 2003).
During denitrification, oxidized forms of N such as nitrate and nitrite are converted to
dinitrogen and, to a lesser extent, nitrous oxide gas. However, nitric oxide and nitrous oxide are
gases that have environmental impacts. Both gases are known as contributors to the atmospheric
ozone destruction. For example, nitric oxide contributes to smog which is known to cause
respiratory illnesses like asthma in both children and adults, and nitrous oxide is an important
greenhouse gas that is contributing to global climate change. The Intergovernmental Panel on
Climate Change (IPCC) estimates that the microbial conversion of agriculturally derived N to
nitrous oxide in soils and aquatic ecosystems is the largest source of anthropogenic nitrous oxide
27 to the atmosphere (Forster et al., 2007). Therefore, microbial denitrification is a large source of
nitrous oxide emissions in terrestrial and aquatic ecosystems.
The proportion of denitrified nitrate converted to nitrous oxide, rather than dinitrogen
gas, partially controls how much nitrous oxide is produced via denitrification (Codispoti, 2010 ).
Although nitrous oxide emission rates have been reported for streams and rivers (Cole and
Caraco, 2001; Beaulieu et al., 2008), the nitrous oxide yield has been studied mostly in lentic
freshwater and marine ecosystems, where it generally ranges between 0.1 and 1.0%. However,
yields as high as 6% have been observed (Seitzinger and Kroeze, 1998). These nitrous oxide
yields are low compared with observations in soils (0–100%) (Schlesinger, 2009), which may be
a result of the relatively lower oxygen availability in the sediments of lakes and estuaries. Few
studies provide information on the nitrous oxide yield in streams and rivers because of the
difficulty of measuring gas production in these systems. Yet, we know that N enrichment
increases nitrous oxide production by stimulating nitrification and denitrification. Overall,
nitrous oxide is a byproduct of both of these microbially mediated transformations (Seitzinger et
al., 2006; Forster et al., 2007; Compton et al., 2011).
2.6.2. Nitrate as a Resource to Improve Water Quality
While nitrate is a pollutant, it can also have beneficial effects on water quality. A number
of studies have documented that nitrate at the sediment-water interface can inhibit the release of
problematic compounds such as P, reduced iron and manganese, and sulfide. P can exacerbate
eutrophication, reduced iron and manganese can complicate potable water treatment, and sulfide
is extremely toxic to aquatic biota (Beutel et al., 2008; Wauer et al., 2005; Jonas et al., 2003;
DeGasperi et al., 1993). Some studies have documented that phytoplankton communities shift
28 away from harmful N-fixing cyanobacteria as N to P ratios increase in surface waters due to
changes in nutrient loading (Schindler, 2012; Effler et al., 2008; Levich, 1996).
Managed nitrate addition to anoxic sediment has been used in lakes and reservoirs to
oxidize organic sediment and repress internal P loading (Søndergaard et al., 2000a; Foy, 1986;
Ripl, 1976). Recent examples of managed nitrate addition to repress MeHg production in
profundal lake sediments also exist (Austin et al., 2011; Nolan et al., 2012). Thus, in some cases,
nitrate may have a beneficial role to play in managing the quality of water resources.
The following case studies revisit the concept of managed nitrate addition to lakes and
reservoirs for water quality improvement. Specifically, we focus on the potential for reuse of
nitrate-rich treated wastewater to improve water quality in raw water reservoirs used as a
drinking water sources. This concept is not especially novel. Ripl and Lindmark (1978) first
highlighted the potential to use nitrate in treated wastewaters to improve profundal water and
sediment quality over 30 years ago. A renewed focus on the topic is merited as water quality and
availability continues to be stressed by growing populations and climate change.
In a recent review of N discharges to European freshwater ecosystems, Petzoldt and
Uhlmann (2006) asked, "Is there a need for nitrate elimination in all wastewater treatment
plants?" The authors noted that nitrate could be used as an ecotechnological strategy to limit the
release of redox-sensitive P from anaerobic sediment, thereby improving water quality in Plimited aquatic ecosystems. They further argued that, rather than being mandated in all cases,
nitrate reductions in discharge to surface waters should be evaluated on an ecosystem-byecosystem basis. The reuse of nitrate in wastewater to improve the quality of raw drinking water
sources is compelling because it enhances the sustainability of both treatment processes. Since
the added nitrate is potentially converted to harmless dinitrogen gas via biological denitrification,
29 negative impacts related to eutrophication could be minimal if nitrate addition is properly
managed. In addition, nitrate addition to freshwater bodies in which P loading is low should not
result in substantial enhancement of productivity.
2.6.3. Sequential Treatment of Sediment with Nitrate
Rilp (1976) was one of the first researchers to propose the in-lake addition of nitrate to
manage internal P loading. The sediment treatment scheme consisted of a sequence of three
chemical additions. First, ferric chloride is added. The formation of iron oxides and resulting
acidic conditions in sediment have the dual effect of enhancing sediment P binding potential
while promoting hydrogen sulfide volatilization, which enhances subsequent nitrate treatment
since sulfide is a potential sink for nitrate via abiotic reduction. Second, calcium hydroxide is
added to neutralize the pH of the sediment, a precondition for effective microbial denitrification.
Third, calcium nitrate is added to oxidize organic matter in sediment via microbial
denitrification. Treated sediment should have a higher P retention capacity and lower oxygen
demand, resulting in lower rates of internal P loading. In 1975, sediment treatment was evaluated
in Lake Lillesjön (surface area = 4.2 ha; maximum depth = 4.2 m; mean hydraulic retention time
= 0.3 yr), a eutrophic lake in Sweden that received sewage loading before wastewater inflow was
diverted in 1971. During the treatment, water column nitrate concentrations peaked to 7 mg-N/L
but declined to zero after two months. After the treatment sediment oxygen demand decreased by
50% and water column P dropped from 0.2-0.4 mg-P/L in the previous year to < 0.1 mg-P/L.
Informed by the early work of Ripl (1976), Wauer at al. (2005) recently evaluated the use
of a newly developed compound, consisting of a matrix of iron oxide imbedded with calcium and
nitrate ions, to limit P release from lake sediment. The compound was added to replicate
30 experimental enclosures (diameter = 10 m; depth = 8.5 m) in eutrophic Dagowsee (surface area =
0.3 km2; volume = 1.5 x 106 m3; maximum depth = 9.5 m) located 100 km north of Berlin. Three
months after treatment, phosphate in bottom waters in untreated enclosures was 706 µg-P/L
compared to 16 µg-P/L in treated enclosures. P efflux from sediment was 4-10 mg-P/m2/d in
untreated enclosures and negligible in treated enclosures. The repression of P efflux from treated
sediment was still observed 1 year after the sediment treatment.
2.6.4. Quarry Lakes Experiment, Habo, Sweden
Swedish researchers were also the first to propose the use of nitrate-rich effluent from
wastewater to improve water quality in lakes (Ripl et al., 1979). They suggested that nitrate
addition could have a number of positive water quality impacts, including a shift away from Nfixing cyanobacteria, lower rates of P release from sediment, and better oxygen conditions as
organic matter in sediment is oxidized by nitrate-consuming bacteria. During the summer and
fall of 1978, Ripl et al. (1979) performed multiple low level additions of sodium nitrate to three
quarry ponds near Habo, Sweden. Two ponds were treated (surface area = 0.25 and 0.72 ha,
maximum depth = 5.5 m and 2.8 m) and another pond was used as a control (surface area = 1.6
ha, maximum depth = 4.5 m). Nitrate levels in water in the treated ponds ranged from 1-6 mgN/L. Nitrate addition resulted in a shift from a cyanobacteria dominated system to a green algae
dominated system with a more robust large-bodied zooplankton population with relatively lower
levels of N fixation, primary production and chlorophyll.
31 2.6.5. Lake Lyng, Denmark
A number of in-lake restoration measures have been implemented in Danish lakes over
the past two decades to improve water quality including manipulation of fisheries, macrophyte
implantation, dredging, and aeration (Søndergaard et al., 2000b). In one case, nitrate was added
to oxidize sediment and inhibit internal P release. Lake Lyng (surface area = 10 ha; maximum
depth = 7.6 m) is a eutrophic lake with a history of sewage loading from a nearby municipality
(Søndergaard et al., 2000a). Summertime surface waters were highly turbid (Secchi depth < 2 m)
and commonly dominated by small colony-forming cyanobacteria. Bottom waters were anoxic
and accumulated P up to a concentration of 3 mg-P/L. In the 1950s sewage inputs were diverted
but the water quality in the lake remained degraded due to high rates of sediment P release. In
the summers of 1995 and 1996, 500-600 kg of N as calcium nitrate was added to bottom waters,
a relatively low dose compared to other lake treatments. Nitrate addition yielded a 50% drop in
summertime P levels. Nitrate enhanced the co-precipitation of iron and phosphate in the bottom
waters of the lake, with the resulting iron-phosphate precipitate settling to the sediment, resulting
in less P available to stimulate phytoplankton growth in the water column. Nitrate also indirectly
enhanced iron-phosphate precipitation by inhibiting the formation of sulfides. Sulfides can
sequester iron through the precipitation of iron sulfide, thereby lowering the amount of iron
available to bind with P.
2.6.6. Schlei Fjord, Germany
Researchers have also evaluated the effect of nitrate addition on P cycling in estuaries
and coastal waters (Petzoldt and Uhlmann, 2006; McAuliffe et al., 1998; Feibicke, 1997). In one
European study, Feibicke (1997) evaluated the impact of nitrate addition to estuary sediment on
32 water quality in the Schlei, a eutrophic fjord in northern Germany. During summer, highly
reduced conditions in sediment lead to sulfide formation. Sulfide abiotically reduces Pcontaining iron oxides in sediment, resulting in P efflux from estuary sediment ranging from 130 mg-P/m2/d. Released P exacerbated eutrophic conditions by further stimulating phytoplankton
productivity. Nitrate addition, which would favor biological denitrification over sulfate
reduction, was hypothesized to impeded sulfide production and lower sediment efflux of P. A
nitrate solution was injected into sediment in a 3.5 ha enclosure in the summer of 1994. As
expected, nitrate addition resulted in higher levels of sulfate and dissolved iron and lower levels
of phosphate in sediment pore water. Total P and nitrate in water overlaying sediment was 0.4
mg-P/L and < 1 mg-N/L in untreated enclosures and 0.1 mg-P/L and 4-10 mg-N/L in treated
enclosures. A drop in phytoplankton biomass in treated enclosures, comparable to that for total
P, was also observed. Phytoplankton composition also shifted from cyanobacteria to diatoms and
other non-blue-green algal species in the treated enclosures. An estimated 1% of added nitrate
remained in the sediment after 100 days.
2.6.7. Occoquan Reservoir, USA.
The Occoquan Reservoir is a large, run-of-the-river reservoir (surface area = 6.23 km2;
volume = 31.4 x 106 m3; maximum depth = 19.8 m; mean hydraulic retention time = 20 d) near
Manassas, Virginia (Randall and Grizzard, 1995). The reservoir, which serves as the water
source for 2 million people, has a history of excessive nutrient loading from wastewater
discharges and storm runoff. In the late 1960s, massive cyanobacteria blooms led to frequent
taste and odor problems, depletion of oxygen in bottom waters, and complications and increased
costs to treat reservoir water for potable use. To address water quality concerns, regional
33 authorities evaluated a range of wastewater management options including a moratorium on new
sewage connections and the export of wastewater from the basin. Seeing the value in utilizing
wastewater as a resource, managers ultimately developed the "Occoquan Policy" – a farsighted
program of indirect reuse via tertiary wastewater treatment.
The Upper Occoquan Service Authority discharges 360 x 106 m3/yr of highly treated, low
nutrient wastewater to one of the two main tributaries of the reservoir. During the summer,
treatment is modified so that discharged effluent is rich in nitrate. Much of the added nitrate
flows with the colder stream water to the profundal zone of the reservoir. Typical nitrate levels
are > 10 mg-N/L in discharge and < 2 mg/L in bottom waters near the reservoir outlet 26 km
downstream (Cubas et al., 2014). The Fairfax County Water Authority withdraws water near the
outlet of the reservoir and treats it for potable use. Maintaining elevated nitrate levels in effluent
discharge to the reservoir has multiple benefits: decreasing tertiary treatment costs by avoiding
the denitrification step during the summer, improving water quality by reducing internal loading
of P and other reduced compounds from sediment, and increasing water supply.
2.6.8. Lake Tegel, Germany
Lake Tegel (volume = 23.2 x 106 m3; maximum depth = 16 m; mean hydraulic retention
time = 0.2 yr) is a eutrophic lake in northwestern Berlin. The lake, which serves as a water
supply reservoir, has a long history of cultural eutrophication and has been impacted by polluted
urban runoff, sewer overflows, and sewage discharges. The lake provides a unique case study in
which reductions in external P and N loading were implemented sequentially, resulting in an
interim period when the lake was relatively rich in nitrate (Schauser et al., 2006).
34 To improve water quality, a P elimination plant was constructed to remove P in surface
inflow in the 1980s. Then in the 1990s, to meet European Union nutrient discharge requirements,
regional wastewater treatment plants implemented N removal. Nitrification, the conversion of
ammonia to nitrate, began in 1992 and denitrification, the further conversion of nitrate to
dinitrogen gas, began in 1997. Nitrate rich conditions repressed P release from profundal
sediment. Bottom water P concentrations in 1994-1996 were < 0.2 mg-P/L, while prior to and
after this period P concentrations were 0.5-1.5 mg-P/L. 1994-1996 also yielded the three lowest
consecutive years of P accumulation in the hypolimnion (< 2 mg-P/m2/d). Accumulation rates
after 1996 were 4-8 mg-P/m2/d. Even in years before and after nitrification, the onset of P
accumulation in bottom waters correlated with depletion of nitrate in bottom waters. Detailed
evaluation of sediment biogeochemistry showed that sediment in Lake Tegel was saturated with
P relative to iron. While nitrate inhibited release of P bound to iron oxides, sediment still
released mineralized P since sediment had low capacity to retain P. These results indicate that the
ratio of P to iron in lake sediment is an important parameter in predicting the effectiveness of
nitrate in limiting P release from sediment.
2.6.9. Lake Onondaga, USA
Lake Onondaga (volume = 131 x 106 m3; maximum depth = 19.5 m; mean hydraulic
retention time = 0.25 yr) is located in Syracuse, New York. Since the 1920s, the urban lake has
received effluent from the Metropolitan Syracuse Wastewater Treatment Plant, which now
accounts for 20% of annual lake inflow. The input of treated wastewater has had profound
impacts on water quality, with elevated ammonia and P levels resulting in high phytoplankton
productivity and low dissolved oxygen levels (Effler et al., 2010; Matthews and Effler, 2006).
35 One historical industrial pollutant, Hg, is extremely elevated in lake sediment and waters
(Todorova et al., 2009). In 2004, the wastewater treatment plant underwent a $130 million
treatment upgrade to improve effluent quality. Treatment enhancements included conversion of
ammonia to less toxic nitrate via biological nitrification, removal of P via chemical precipitation,
and a new ultraviolet disinfection system. A 80% decrease in P loading to the lake has lowered
primary production in surface waters (Effler and O’Donnell, 2010), and nitrification has resulted
in lower ammonia levels and increased dissolved oxygen in lake bottom waters (Effler et al.,
2010). An unanticipated effect of the shift from ammonia to nitrate in wastewater effluent was a
50% drop in MeHg in bottom waters. MeHg, a potent neurotoxin that bioaccumulates in aquatic
food webs, is produced by anaerobic bacteria in the profundal zone of lakes. As detailed in
Todorova et al. (2009), the presence of nitrate in bottom waters alleviated anaerobic conditions
and repressed MeHg accumulation. The New York State Department of Environmental
Conservation is implementing a $400 million remediation cleanup effort focused on managing
Hg contamination in Lake Onondaga, including a 3-year pilot test to assess the benefits of nitrate
addition to impede MeHg accumulation in bottom waters (Nolan et al., 2012).
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47 CHAPTER 3
ENVIRONMENTAL AND ECONOMIC POLICY STUDY IN CARTAGENA,
COLOMBIA: WATER SUPPLY AND ENVIRONMENTAL MANAGEMENT PROJECT
3.1. Background
For decades, water pollution and lack of sanitation have been serious concerns for the
citizens of Cartagena, Columbia. Cartagena was the historic city where the “1983 Cartagena
Convention for the Protection of the Caribbean” was signed. The primary aim of this agreement
was to protect the coastal marine environment and the public heath of the Caribbean's citizens
through improved wastewater management. However, many water bodies near urban areas are
still highly polluted (ACUACAR, 1995, 2009b). Pollution emanating from domestic and
industrial wastewater has contaminated marine waters and the highly productive estuaries and
bays that provide critical ecological connections to marine environments. Inadequate wastewater
management has also polluted urban beaches, threatening public health and undermining
tourism.
Cartagena is a city of almost 900,000 people on the northern coast of Colombia. The city
serves as the capital of Bolívar, one of the Columbia's departments (states), and is an economic
hub and popular tourist destination. Prior to 1995, Cartagena's water and wastewater services
were extremely unreliable; less than 70 percent of households had water connections and less
than 55 percent had sewage service (World Bank, 1999, 2009). A large portion of the population,
particularly those with lower incomes, obtained water from private vendors. The other portion,
with connections to the city's water system, frequently experienced water pressure so low that
48 service was practically nonexistent. The pre-existing system operated with substantial financial
losses and had insufficient funding for maintenance or improvements (UNEP, 2010).
Considerable political pressure from the citizens of Cartagena led to the creation of a
Public-Private Partnership (PPP) to repair and operate the water and sewage systems. In 1994,
under the leadership of the mayor, the city's municipal council approved the creation of an
“empresa mixta” (mixed enterprise), which combined the resources of the public works
department and a private Spanish water firm (World Bank, 2007, 2008). Together, these two
entities operate under the name Aguas de Cartagena, known as ACUACAR.
Despite ACUACAR's significant progress in improving the efficiency and quality of
water and sanitation services, Cartagena still faced many challenges, including: (i) insufficient
water supply coverage, especially in poor neighborhoods located in the vicinity of the Cienaga
de la Virgen; (ii) insufficient sewerage services, especially in the poorest areas around the
Cienaga and in zones adjacent to the Cartagena Bay where open sewage canals on streets was
common; (iii) inadequate domestic wastewater management system with untreated wastewater
discharging to the Cartagena Bay, the Cienaga, and the inner-city water courses; and (iv)
minimally treated industrial wastewater discharging from the Mamonal zone to the Cartagena
Bay. Together, these conditions generated sanitation, environmental, and public health problems
that reduced the quality of life for residents and posed serious restrictions for sustainable
economic development, especially the city's tourism sector (World Bank, 2005).
To address these significant water and sanitation challenges, the District of Cartagena and
ACUACAR developed a water and wastewater master plan and implementation strategy. This
strategy primarily relied on a project called "The Cartagena Water Supply, Sewerage, and
Environmental Management Project (“the Project”)", financed by the World Bank. The Project's
49 goals were to: (i) expand water supply coverage and production capacity; (ii) expand sewerage
networks and improve the sewerage system in the Cienaga de la Virgen basin; and (iii) enhance
wastewater management in Cartagena. The wastewater management plan used a conveyance
system that transported the city’s wastewater to a preliminary treatment plant and subsequently
discharged the treated wastewater through a submarine outfall into the Caribbean Sea.
This chapter summarizes Cartagena’s experience and environmental policy issues
associated with implementing a water and wastewater management program. Specifically, this
chapter discusses the main efforts to improve water quality in Cartagena, the mechanisms for
overcoming impediments, and the key factors for achieving success. The political and
environmental empowerment of Cartagena is analyzed relative to implementing an effective
water and sanitation management program. Finally, this chapter examines the key environmental
policy lessons revealed through the Cartagena project experience.
A portion of this chapter is based on the publication, Restoring the Coastal Environment
of Cartagena, Columbia, authored by Ricardi Duvil (Consultant, World Bank) and Greg
Browder (Lead Water and Sanitation Specialist, World Bank), in collaboration with the World
Bank, the Colombian Ministry of Environment and Sustainable Development, the Cartagena
Water Utility (ACUACAR), and the Global Partnership for the Oceans (GPO) (Box 1). Also I
would like to thank Dr. Steven Stehr, associate professor of politics, philosophy and public
affairs at Washington State University for helpful discussions and reviews of this chapter.
50 Box 1: Global Partnership for the Oceans
The Global Partnership for Oceans is a new and powerful approach to restoring ocean
health. It mobilizes finance and knowledge to activate proven solutions at an
unprecedented scale for the benefit of communities, countries and global well-being. The
GPO is a growing alliance of over 140 governments, international organizations, civil
society groups, and private sector interests committed to addressing the threats to the
health, productivity and resilience of the ocean. The Partnership aims to tackle
documented problems of overfishing, pollution, and habitat loss. Together these problems
are contributing to the depletion of a natural resource bank that provides nutrition,
livelihoods and vital ecosystem services.
3.2. Project Objectives
The objectives of the Project were to: (i) improve the water and sewerage services of
Cartagena and the sanitary conditions of the city's poorest population by expanding water and
sewerage coverage, particularly in the poor neighborhoods; (ii) facilitate the environmental
cleanup of water bodies surrounding the city (Cartagena Bay, Caribbean beaches, and Cienaga
de la Virgen estuary) (Figure 1) by providing adequate collection, treatment, and disposal of the
city's wastewater; and (iii) improve the sustainability of water and sewerage services in
Cartagena by leveraging Bank support to shore up the private sector participation (PSP) model,
pioneered by ACUACAR, against the prospect of political interference.
A related objective of the city was to decrease the amount of non-revenue-water (NRW)
caused by leaks in the existing pipes. The city believed that 60 percent of the water was being
51 lost in this way and wished to reduce the loss to no more than 25 percent. Reducing NRW losses
would increase water pressure, water supply, and daily duration of water availability. In addition,
lowering NRW would improve the city's long-term financial situation because a larger portion of
water treatment and distribution costs could be recovered from consumers.
Another objective of ACUACAR was to transform the water and sewerage system into a
financially self-sustaining operation. This objective was to be achieved through increased
efficiency in the management of water supply and sewage services, paired with an increase in the
tariff collection ratio. The tariff structure incorporates a system of cross-subsidies based on the
1991 National Constitution. The structure divides customers into six categories, with the top two
categories subsidizing the lowest three and the remaining one paying cost-recovery tariffs
(World Bank, 2006). The municipal provider had over one thousand employees, but water
service was intermittent and tariff collection was less than 50 percent. Prior to the creation of
ACUACAR, the financial working ratio (the cash operating costs/cash collections) for the
municipal-owned provider was about 152 percent. The company was losing money and
providing inadequate service to customers. ACUACAR aimed to improve management, to
improve all-around service, and to eliminate the operating losses.
52 Figure 3-1. Map of project area.
53 3.3. Cartagena: A Dynamic Cultural and Industrial City
The District of Cartagena has experienced rapid growth with the population doubling
over the last two decades. Due to its history and spectacular natural scenery, Cartagena is
Colombia’s largest tourist area, with an annual influx of around 1 million visitors. In 1984,
UNESCO declared the key landmarks of Cartagena as a World Cultural Heritage Site. Cartagena
also has a thriving industrial sector with important petrochemical, beverage, and seafood
processing industries. In 2012, the Port of Cartagena was Colombia's main container port and
one of the busiest ports for transportation of grain in South America.
The socio-economic structure of Cartagena is complex and the city suffers from high
levels of poverty. Thirty-five percent of the population is classified as extremely poor, or “Strata
1” in the Colombian system, implying that they cannot meet their basic needs. Another 30% falls
under the “Strata 2” classification, implying that their basic needs are not met in a satisfactory
manner (Deloitte, 2009). Households classified as Strata 1 or 2 pay subsidized rates for utility
services, such as water and sanitation. A significant percentage of the population is composed of
recently arrived residents from rural areas seeking better economic opportunities and fleeing
drug trafficking violence in the countryside.
Cartagena is surrounded by water on all sides: the Caribbean Sea to the north, Cartagena
Bay to the west, and a lagoon known as the Ciénaga de la Virgen to the east (Figure 1).
Recognizing the importance of the marine environment for Cartagena, and its standing as a
leading city in the region, the Convention for the Protection and Development of the Marine
Environment in the Wider Caribbean Region, known as The Cartagena Convention, was signed
in 1983 in Cartagena (Box 2).
54 Box 2: Cartagena Convention
The Convention for the Protection and Development of the Marine Environment in the
Wider Caribbean Region (WCR) is a comprehensive, umbrella agreement for the
protection and development of the marine environment. This regional environmental
convention provides the legal framework for cooperative regional and national actions.
The Convention was adopted in Cartagena, Colombia on 24 March 1983 and entered into
force on 11 October 1986, for the legal implementation of the Action Plan for the
Caribbean Environment Programme. The Cartagena Convention has been ratified by 25
national governments in the Wider Caribbean Region. The Convention is supplemented
by three Protocols which entered into force with the following dates: i) Combating Oil
Spills (1986); ii) Protected Areas and Wildlife (2000); and iii) Pollution from Land-Based
Sources (2010).
3.4 Confronting Cartagena's Coastal Environmental Crisis
By the mid-1990s, rapid population growth, unplanned urban development, and poor
wastewater management had severely deteriorated Cartagena’s rich coastal resources and
generated a public health crisis. Less than half of the households had connections to a wastewater
disposal system and all wastewater was untreated. Untreated wastewater flowed into coastal
waters carrying organic waste and nutrients that destroyed rich fisheries in the Bay and the
Lagoon. Disease-causing bacteria and parasites from untreated wastewater flowed through
streets and into water courses, amplifying public health risks, particularly for the numerous low-
55 income and vulnerable communities living around the Lagoon. Cartagena’s famous beaches were
frequently closed due to microbiological contamination. In addition, industries further
contributed to the contamination of Cartagena’s environment by discharging wastewater with
impunity into the nearest water body.
The environmental crisis confronting Cartagena in the mid-1990s was undermining the
city’s long-term sustainability by threatening its world-class tourism industry, posing public
health risks, and lowering the quality of life. Furthermore, the rich ecosystems in the Bay, the
Lagoon, and along the coast—mangrove forests, highly productive and diverse fisheries, seagrasses and coral reefs—were in ecological decline. The prospects for recovery appeared bleak
in the absence of a comprehensive and effective water pollution control program.
3.5 Resetting the Institutional Framework for Water Supply and Management
Prior to 1995, Cartagena’s water and sanitation service suffered from an inefficient water
utility that was owned and controlled by the municipal government. Due to low tariffs and
inefficient operations, the lack of cost recovery on behalf of the water utility contributed to
financial problems. Ultimately, these problems resulted in inadequate service delivery to its
users. Major portions of the population, particularly those with lower incomes, were forced to
rely on water provided by private vendors. Even those with connections to the city water system
frequently experienced such low water pressure that their service was essentially nonexistent.
Unable to recover costs due to its inefficiency, the city's system operated with substantial
financial losses and had insufficient funding for maintenance or improvement of the system.
56 The situation in Cartagena was not an isolated case in Colombia during this time period.
As a result, in 1994, the Colombian government passed the Public Services Law that expanded
the role of specialized operators in public water utilities and created national tariff and service
regulatory agencies. To address its environmental and socio-economic problems, the District of
Cartagena was one of the first municipalities in Colombia to introduce private participation in the
water and sanitation sector.
In 1995, the Cartagena District Council and the City Mayor liquidated the municipal
utility and created a “mixed-capital” company for the operation of the water and wastewater
system, called ACUACAR. A private company from Barcelona, Spain (AGBAR) was selected
as the specialized operator and owns 46% of ACUACAR’s shares. Cartagena District retained
50% of the shares, with the remaining 4% owned by private shareholders. Although this model is
commonly used in Spain, this was the first time such an arrangement had been tried in Latin
America (BID, 2006).
The ACUACAR contract included various performance targets to improve the quality of
service and maintenance, reduce leakages in the distribution system, and improve the collection
rate. The District originally retained primary responsibility for financing future capital
investments, while ACUACAR was responsible for implementing the capital works program.
However, the contract has evolved over time and ACUACAR now self-finances a significant
share of capital works through its tariff revenue.
57 3.6 Developing New Environmental Policy and Management Institutions
In 1993, a new National Environmental Law created the Ministry of Environment and
formulated a framework for environmental management. The law also established autonomous
regional environmental authorities, which are responsible for implementing national
environmental policies at the local level. The regional environmental authorities in Colombia are
governed by a combination of national, regional, and local government representatives and
stakeholders from the business community, non-governmental groups, and indigenous peoples.
The regional environmental authorities have a steady funding stream, based mainly on a 15%
environmental property tax surcharge, to implement projects and cover operating costs.
However, these funds were often insufficient because the coastal regional environmental
authorities' jurisdiction now includes marine environments, which are an added responsibility
and contribute to higher operating costs.
The environmental authority for the region surrounding Cartagena is called CARDIQUE.
This institution is responsible for issuing environmental licenses and setting standards for water
quality and municipal and industrial discharge. The larger cities of Columbia, including
Cartagena, also have their own municipal agencies that are responsible for environmental
management and entitled to one-half of the city's environmental property tax surcharge.
3.7 Financing and Implementing a Phased Infrastructure Development Program
In the late 1990s, the financing needs for upgrading Cartagena’s water supply system,
and essentially constructing a new wastewater collection, treatment, and disposal system were
58 daunting. The upgrade would require over $150 million for wastewater infrastructure alone.
Fortunately, the Colombian national government and the District of Cartagena were committed
to providing most of the required financing. International development institutions, such as the
World Bank and the Inter-American Development Bank (IDB), also provided long-term lowinterest loans, coupled with technical assistance, to help Cartagena achieve its environmental
objectives (World Bank, 2004).
Tackling Cartagena’s severe coastal environmental degradation required a long-term and
incrementally phased approach, grouped into the following components:
1) Improving Water Supply Service: ACUACAR’s highest priority during the period from 1995
to 2000 was to improve water service to customers by rehabilitating and upgrading the existing
water supply system (Hazen and Sawyer, 2001). Improvements included increasing the treated
water supply, stabilizing water pressure, improving the electrical supply system, and improving
commercial practices such as billing, collection, and customer service. These interventions
allowed ACUACAR to enhance its financial stability and cost recovery and to secure the
confidence of its customers and the Cartagena District government, thus paving the way for
large-scale wastewater investments.
2) Improving Drainage in High-Value Economic Areas: Prior to 2000, key tourism and
commercial areas in Cartagena were subject to frequent flooding and sewer overflows into the
streets and surrounding beaches. For example, the low-lying Bocagrande zone of popular, highend tourist beaches and hotels is particularly vulnerable. With the help of an IDB loan of
59 US$40.5 million, ACUACAR improved the wastewater system draining into Cartagena Bay and
essentially eliminated sewage discharges in key tourism areas.
3) Improving Water Circulation in the Ciénaga de la Virgen: While ACUACAR focused on
water supply and high-priority drainage issues in the late 1990s and early 2000s, the District of
Cartagena and CARDIQUE developed an innovative project to improve water quality in the
Lagoon (Ciénaga de la Virgen). The Lagoon originally had numerous outlets to the sea and its
brackish waters supported a diverse and highly productive aquatic ecosystem. However, the
construction of a road along the coastline in the late 1980s sealed off most of the natural outlets,
thus severely curtailing tidal circulation in the estuary. By 2000, the Lagoon was receiving
around 60% of Cartagena’s untreated wastewater. The combination of high pollution loads with
poor water circulation had created an environmental disaster. The District of Cartagena and
CARDIQUE, with the financial and technical support of the Netherlands' government,
constructed a stabilized tidal inlet entrance known as “La Bocana Project” in the early 2000s.
The project consists of an enlarged canal entrance to the Lagoon, self-actuating tidal gates that
increase flow into the Lagoon, and a long seawall within the estuary that routes the tidal flows
and helps flush pollution out of the Lagoon.
4) Collecting, Treating, and Disposing of Cartagena’s Wastewater: By the early 2000s,
ACUACAR, with the support of a US$85 million loan from the World Bank, was poised to fully
implement its wastewater master plan. The plan consisted of the following components (Figure
2):
60 •
Convey the wastewater from the Cartagena Bay drainage area to the new central pump
station;
•
Complete the wastewater collection system in the Laguna drainage area, particularly in
the poor, low-lying squatter neighborhood along the Lagoon, and convey the wastewater
to the new central pump station;
•
Pump all of Cartagena’s collected wastewater from the central pump station to a new
treatment plant located 19 kilometers to the north; and
•
Treat the wastewater at a new wastewater treatment plant, and then discharge the effluent
through a 4.3-kilometer long, 2-meter diameter, submarine outfall into the Caribbean Sea
at a depth of approximately 20 meters.
A submarine outfall (Box 3), coupled with preliminary wastewater treatment was the
selected alternative due to its relatively low costs and the adequate natural conditions close to
Cartagena’s coast. The preliminary treatment plant was designed to remove floatable material
such as oils and plastic bags, as well as sand and grit particles. The submarine outfall technology
efficiently reduces organic material and suspended solids, producing up to a 99% reduction in
the concentration of these key pollutants outside the designated mixing zone around the outfall
diffusers. Extensive modeling was done with international experts to confirm the feasibility of a
submarine outfall and the plan was approved by CARDIQUE. The treatment plant layout also
allowed for upgrading the treatment process in the future, if so required.
61 Box 3: Submarine Outfalls
A marine outfall is a pipeline or tunnel that discharges wastewater under the sea’s
surface. For municipal wastewater, effluent is often discharged after only preliminary
treatment, with the intent of using the assimilative capacity of the sea for further
treatment (McCarthy, 2006). The siting and design of a submarine outfall through
sophisticated modeling and oceanographic studies is critical to ensuring that the receiving
seawater has sufficient dilution capacity. High quality construction is required to avoid
leakage in the outfall pipe or tunnel, and continuous monitoring under independent
regulatory supervision is indispensable to confirm that the outfall performs as planned.
Submarine outfalls, combined with preliminary treatment tend to be less expensive than
advanced wastewater treatment plants, using the natural assimilative capacity of the sea
instead of energy-intensive treatment processes in a plant. The costs of preliminary
treatment are about one tenth that of secondary treatment and preliminary treatment also
requires much less land than advanced wastewater treatment. Preliminary treatment,
however, typically does not remove nutrients such as nitrogen or phosphorous which are
a concern in many marine environments. Submarine outfalls are common throughout the
world and probably number in the thousands. More than 200 large outfalls have been
listed in a single international database maintained by the Institute for Hydromechanics at
Karlsruhe University for the International Association of Hydraulic Engineering and
Research (IAHR) / International Water Association (IWA) Committee on Marine Outfall
Systems.
62 3.8 Social Impact Concerns Raised by the Local Non-Governmental Organization
The selection of a submarine outfall option, combined with preliminary treatment, was
controversial (Neotropicos, 2002). The local non-governmental organization (LNGO), on its
behalf and on the behalf of 424 residents of the areas, was opposed to the project. As a result the
LNGO filled a number of claims to the World Health Organization (WHO), the United Nations
Environmental Programme (UNEP) and the Colombian government. These claims include:
•
The outfall would affect fishing village near the start of the 4-km long marine outfall
pipe.
•
Local property developers feared odor and seawater quality problems would affect real
estate value.
•
The Bank failed to identify affected communities as indigenous and failed to safeguard
their livelihood and reliance on fishing and farming as a subsistence living.
•
The Project would have serious impacts on the communities' culture and way of life, as
well as the impact of pollution on their health.
•
The Bank failed to perform proper supervision and insufficiently scrutinized the
economic investment and the environmental risk evaluations for the proposed submarine
outfall system.
•
The Project would place undue fiscal strain on the city of Cartagena and would violate
Colombia’s international obligations under the 1983 Cartagena Convention on LandBased Sources of Pollution and the related 1999 Aruba Protocol.
63 As a result, the marine outfall approach was contested in the Colombian environmental
regulatory system. In 2002, the Ministry of Environment ratified the environmental license
originally approved by CARDIQUE. Opponents of the outfall option also appealed to the World
Bank’s Inspection Panel in 2004. After a lengthy and extensive review, the Panel endorsed the
outfall alternative as an acceptable solution.
3.9 Addressing Concerns Raised by the LNGO
To address the LNGO’s concerns, the World Bank Panel conducted a Social Impact
Assessment (SIA) and an Environmental Impact Assessment (EIS) in the Punta Canoa areas. The
Bank found that the environmental impact and social impact assessment demonstrated the project
would not have significant adverse effects on the environment and the affected population. The
investigation concluded that the project would achieve significant public benefits with a high rate
of return, without significant environmental or social risks, and in a fiscally sustainable manner.
However, the Bank agreed that the project failed to address possible long-term environmental
and health effects on the coastal and marine environment as a result of multiple outfalls and
increased volumes of sewage and organic wastes. The Bank acknowledged that this issue needed
to be addressed in future. In addition, the bank agreed to support a program to strengthen fishing
activities and requested ACUACAR work with the communities to identify specific activities for
optimized fishing opportunities in Punta Canoa. Finally, the Bank included the affected
communities in the project decision-making process and ensured they would share in project
benefits, rather than compensating them for the project's existence. Later, the Bank and
64 ACUACAR compensated for the potential impacts, for example, a new school and community
center were built in the north zone area.
3.10 Implementations Metrics and Summary of ACUACAR’s Achievements
With the Bank's financial support, Cartagena water and sanitation services improved
significantly. Current customers saw an increase in the reliability of water and water pressure.
Existing pipelines and water meters were evaluated and repairs were made. A large percentage of
the loss of water was corrected, which increased water pressure to customers. Service was
extended into poor areas of the city and management improvements were identified and
implemented. Even before new pipelines were constructed, water trucks were sent into poor
areas to supply those areas with less expensive water.
Another aspect of Colombia’s Law No. 142 was that it required a survey of all residential
buildings, including location, form, state, and quality of housing. This information was used by
local governments to classify all residences on a six-level scale. Residences assigned a
classification of 1-3 were considered poor, residences categorized as 4 were middle-class, and
residences assigned a 5 or 6 were deemed upper class. Using this information, ACUACAR
created a new tariff system that included a cross-subsidy to benefit poor customers. Residences
were evaluated and classified per the six-category scale. For customers with high incomes, tariff
rates were set higher than the actual cost of service. This extra revenue then went to subsidize
service to poor residents, whose rates were set below the cost of service.
The new tariffs were intended to generate enough funds to cover operational and
investment costs. ACUACAR proposed the tariff rates and the Drinking Water Regulation
65 Commission reviewed and approved them before rates were assigned to customers. Residents
living in level 4 houses (middle-class) were assigned a tariff equivalent to the actual cost of
service. Lower rates were given to people in levels 1-3 (lower income households), while higher
rates were given to people living in levels 5-6 (the upper income households). This practice of
cross subsidization was used to cover the deficit in tariff revenues that would have resulted from
a general subsidy to poor households (ACUACAR, 2009c, 2009e).
Efficiency within the company and in provision of services was another objective of
ACUACAR. To increase efficiency within the organization, ACUACAR only hired back about
400 of the former 1,500 unionized municipal employees (those let go received severance pay
based on years of service). This move improved productivity from about 5.5 employees per one
thousand accounts in 1995 to 3.4 employees per one thousand accounts in 2005.
To extend service, about 35,000 new connections were made, almost exclusively in poor
neighborhoods. Many poor citizens were skeptical and suspicious of ACUACAR because they
were not used to dealing with a water company. In response, the company hired social workers
and community relations specialists to explain to residents how water and sewage service works.
Local unskilled laborers were also hired to complete the construction of the new pipelines for the
new connections. These actions, coupled with visible service expansion, improved ACUACAR’s
relationship with the community.
Another achievement was the decrease in the level of NRW. Originally, losses were was
estimated at about 40 percent of system volume, leading to an objective of lowering NRW to 25
percent (a 15 percent net reduction) over a period of ten years. However, reevaluation
determined that the NRW was actually closer to 60 percent. Nonetheless, with the improvements
initiated by ACUACAR, the NRW was reduced to about 40 percent by1999, a net reduction of
66 20 percent in only four years. In addition, water coverage went from less than 70 percent in the
early 1990s to close to 99 percent in 2005. The hours of water service increased from about 17
hours a day to 24. Sewage service availability increased from 55 percent of households in the
early 1990s to over 75 percent in 2005. ACUACAR has added an additional 260 km of pipeline,
bringing the water and sewage network up to 700 km in total, an increase of 48 percent (World
Bank, 2009; ACUACAR, 2009a, 2009d).
Equally important was the objective of increasing the collection fee ratio from less than
50 percent to 95 percent. Many households lacked a water meter. In other households, meters
were frequently broken. This situation made tariff collection difficult for the municipal-owned
provider and contributed to the negative cash flows. Within the first few months of their
operations, ACUACAR replaced or added water meters to existing connections and added water
meters to all new connections. By 2005, over 99 percent of connected households had water
meters. The improved billing and tariff collection has improved the ratio to just over 90 percent
in 2005. As a result, the financial working ratio dropped to less than 70 percent, from the 152
percent historical figure. ACUACAR also experienced an increase in the ratio of operating
revenue/operating costs to about 120 percent in 2005. Increases in efficiency, coupled with a
better collection record, gave ACUACAR positive cash flows.
As a result of improvements in utility management, moderate increases in tariffs, and
infrastructure investments, ACUACAR has been able to significantly improve the quality of
water and sanitation services and achieve sustainability. Potable water, which meets the national
quality standards, is provided on a continuous basis to all households—even to the poorest
neighborhoods. Based on its tariff revenues, ACUACAR is able to cover all operating and
maintenance costs, and help contribute to infrastructure investments. ACUACAR’s mixed capital
67 model has become institutionalized in Cartagena, and the combination of local political control
of the company combined with professional private sector management has performed
remarkably well (Table 1).
Table 3.1. Implementation Metrics Water Supply and Wastewater Treatment.
Year
Water
Supply
Coverage
Water
Supply
Customers
Length Water
Network
Continuity of
Water
Service
Unaccounted
for Water
1995
73%
92,572
700 kms
14 hours
45%
2012
99.9%
225,815
1,480 kms
24 hours
34%
Year
Sewerage
Coverage
Sewerage
Customers
Length
Sewage
Network
Wastewater
Treated
Revenues
COP-Million
1995
61%
77,553
500 kms
0%
25,592
2012
89%
200,150
1,062 kms
>90%
165,889
Source: Table based on ACUACAR water data information system.
3.11 ACUACAR’s Methods for Overcoming Impediments
Prior to the decision in 1994 to create the public-private partnership ACUACAR, the
unionized water service workers expressed strong opposition to the project because of potential
job loss. Many of these workers had been hired based on their relationship to the mayor and
managers of the municipal-owned provider, rather than on job-related qualifications. Under
ACUACAR’s management, many of these employees lost their jobs, but were provided
severance pay based on years of service (Gómez Torres, 2003).
ACUACAR also faced the immediate challenge of many complaints from consumers
based on their experiences with the previous service provider. Citizens voicing their opinions
68 about poor service helped bring about the creation of ACUACAR, but once the company took
over the operations and management, some of the vocal criticism continued. Only through
continuous efforts at improving services and public outreach was ACUACAR ultimately able to
turn around public opinion about Cartagena’s water and sewage services.
Another problem ACUACAR faced was that providing services to poor areas would take
time, because of the need to construct new pipelines. Many poor people had been purchasing
their water from independent vendors at a higher price than that paid by households with
municipal connections. Once ACUACAR was operational, the PPP began trucking water to
customers in these poorer neighborhoods. Even though these customers would have to wait for a
connection, they saw an immediate increase in service and were willing to support ACUACAR
during construction.
ACUACAR worked hard to earn and keep the support of the community, which is a
necessary step in the success of a PPP. The majority of the new connections were created in poor
areas where residents were not used to dealing with a water company. Many residents were
confused about the construction process and suspicious of, even hostile towards, the workers.
Community relations specialists and social workers were hired by ACUACAR to overcome this
obstacle. Public meetings were held explaining how water and sewage service worked and
ACUACAR’s methods. The social workers also helped set up payment schedules for customers
with irregular incomes. To further show dedication to the project and the poor areas, ACUACAR
hired local unskilled laborers to construct the pipelines and other necessary parts of the network.
69 3.12 Commissioning of Cartagena’s Wastewater System
Recognizing the importance of this achievement, President Juan Manuel Santos of
Colombia inaugurated the wastewater treatment system on March 20, 2013; his translated
comments are presented below:
“This is an extremely important project, a project that has great significance for the development
of the country, Cartagena, and the thousands of people which in one way or another will benefit
from no longer having a contaminated Bay or a contaminated Ciénaga de la Virgen. With this
project, Cartagena becomes the first city in the country to have a comprehensive sanitation
program with 100% wastewater treatment. This is the project that you have been waiting for 20
years, and after 20 years of much effort and challenges, finally today it is a reality.
3.13 Coastal Environmental Improvements
In addition to improved water services, the completion of the wastewater system has
generated environmental improvements that were hardly imaginable two decades ago. Although
there is less than one year of extensive water quality monitoring to date after the commissioning
of the new wastewater treatment system, the results are evident. Initial opposition to the
preliminary treatment with the marine outfall solution has dissipated and there is a consensus that
the approach is working well.
Figure 2 provides a summary of key water quality indicators “before” and “after” the
commissioning of the wastewater treatment system in mid-2013. This data was provided as a
70 final data set by ACUACAR based on their monitoring data collection. Figure 2 also compares
the water quality data with the regulatory requirements contained in ACUACAR’s
environmental license. For example, Figure 2 illustrates the total coliforms along the beaches of
Cartagena. “Before” reflects data analysis completed prior to the start of the wastewater
treatment system, and the average concentration was 920 MPN/100 mL and was calculated based
on a historical data analysis performed by ACUACAR. Data were collected between 2005 and
2012 from different regions along the beaches, with 13 sampling stations from Castillo Grande to
Punta Canoa. However, there was high variability in the data set. The standard deviation was 672
MPN/100 mL (sample size = 55) and the peak value was > 34,000 MPN/100 mL. Data indicated
that beaches were unsafe at certain times. “After” reflects data analysis completed after the start
of the wastewater treatment system, and the average concentration is 31 MPN/100 mL and was
calculated based on a monthly monitoring system that utilized 11 sampling stations to collect
data from April 2013 through August 2013. All water quality sampling and monitoring was also
completed by ACUACAR. As a result, the beaches are suitable for swimming, fishing and
recreational activities. As indicated in Figure 2, both average values before and after the project
are below the standard value (1,000 MPN/100 mL) for primary contact, but the total coliform
concentration shows a 97% decrease from 2005 to 2013.
Furthermore, Figure 2 shows Cienaga de la Virgen water quality improvements, the
average concentration of total coliforms before was 8,667 MPN/100 mL and the average
concentration “after” was 885 MPN/100 mL. The average value before the projects (8,667
MPN/100 mL) was calculated based on a historical data analysis performed by Hazen and
Sawyer (2001). Data were collected in the 1990s and 2000s from different regions of the lagoon,
and a peak value of > 20,000 MPN/100 mL was recorded close to the point of domestic
71 wastewater discharge. The average value after the projects (885 MPN/100 mL) was calculated
based on a monthly monitoring system that utilized 7 sampling stations to collect data from
January 2013 through July 2013. This water quality sampling and monitoring was completed by
ACUACAR. As indicated in this graphical representation, the average value in 2013 (885
MPN/100 mL) is below the standard value (1,000 MPN/100 mL) for primary contact. In this data
set, the highest concentrations were observed close to the urban areas. Monitoring indicated a
90% decrease in total coliform concentration from the 1990s to 2013.
In addition, Figure 2 shows the average concentration of total nitrogen, total phosphorus
and dissolved oxygen in the lagoon before (25 mg/L, 15 mg/L, and 1.7 mg/L) and after (0.4
mg/L, 0.2 mg/L, and 5.54 mg/L) the outfall project. Poor water quality before the project was the
result of human activity, which greatly increased nitrogen and phosphorus pollution. These
nutrients trigger the growth of algae that block light to seagrass and reduce dissolved oxygen
concentration through the decay of sea grass and algae. This in turn can result in fish kills. The
average values before the projects were calculated based on a historical data analysis completed
by Hazen and Sawyer (2001) as well as Grupo Ambiemtal de Cartagena, which utilized 5
sampling stations (as highlighted in the 2008 water quality survey). The average value after the
project was calculated based on a monthly monitoring system that utilized 7 sampling stations to
collect data from January 2013 through July 2013. This water quality sampling and monitoring
was completed by ACUACAR. The results of 2013 show that Cienaga de La Virgen meets the
water quality standards for a healthy lagoon with an overall decrease of 98% in total nitrogen,
97% in total phosphorus, and an increase of 69% in dissolved oxygen.
Finally, Figure 2 illustrates the total coliform concentration in Cartagena Bay. The
average concentration before was 6,200 MPN/100 mL and the average concentration after was
72 500 MPN/100 mL. The average value before the completion of the projects (6,200 MPN/100
mL) was calculated based on a historical data analysis performed by ACUACAR. Data were
collected in the 1990s and 2000s from different regions of the bay, and a peak value > 24,000
MPN/100 mL was found close to the industrial discharges. This peak value was detected during
high rain intensity; prior studies by Kress and Gifford (1984) have confirmed that high rainfall
intensity has an effect on peak coliform levels. The average after value (500 MPN/100 mL) was
calculated based on a monthly monitoring system that utilized 11 sampling stations to collect
data from January 2013 through July 2013. This water quality sampling and monitoring was also
completed by ACUACAR using the surface water sampling rules as guidelines. As indicated in
this graphical representation, the average value in 2013 (500 MPN/100 mL) is below the
standard value (1,000 MPN/100 mL) for primary contact. The highest concentrations were
observed at the surface waters of the Bay because coliforms are largely retained near the surface
and the Bay received continuous fecal contamination from a variety of sources, including
industrial wastewater and urban runoff with historically high coliform levels. The data shows a
92% decrease in total coliform concentration from the 1990s to 2013.
In addition, the level of dissolved oxygen concentration in the Cartagena Bay before was
0.5 mg/L and the after average concentration was 6.2 mg/L. In the Bay area, the dissolved
oxygen concentration varies from region to region, but in Colombia environmental regulations,
healthy water quality requires that the dissolved oxygen concentration should fall between 6 to 8
mg/L (IDEAM, 2000). In this data analysis, the average value before the project (0.6 mg/L) was
calculated based on an historical data analysis performed by the consulting firm Hazen and
Sawyer (2001) in 1998 and was used in the 1999 Environmental Impact Assessment for
Wastewater Management in Cartagena. The results indicate that dissolved oxygen concentration
73 was significantly influenced by pollution discharge from domestic and industrial waste.
Furthermore, the average value after the project (6.2 mg/L) was calculated based on a monthly
monitoring system that utilized 8 sampling stations to collect data from January 2013 through
July 2013 in varied weather conditions. During this sampling event, the Bay maintained an
average temperature of 26.5 °C and pH of 8.5. This data analysis was reported by ACUACAR in
September 2013. As indicated in this figure, the average value in 2013 (6.2 mg/L) was between
the standard value (6-8 mg/L) for healthy water quality. The lowest concentrations were
observed at the bottom of the Bay because of high level of organic matters. The data shows a
90% increase in dissolved oxygen concentration from the 1990s to 2013.
Since the wastewater treatment system was only commissioned in May 2013, the results
are preliminary, but encouraging. ACUACAR and CARDIQUE have a comprehensive and
continuous water quality monitoring program. The impacts can be qualitatively described as
follows:
Pollution Free Coastal Beaches: Cartagena’s Caribbean beaches are essentially free of
contamination from sewage and the “red-flag” days of beach closures are now history. Coliforms
are a broad class of bacteria found in our environment, including the feces of human and other
warm-blooded animals. The presence of coliform bacteria in water indicates a possible presence
of harmful, disease-causing organisms associated with human waste. The Colombian standard
for safe bathing is 1000 MPN/100 mL and was frequently exceeded in the past. Total coliform
concentrations have declined dramatically after the wastewater treatment system was
commissioned (Figure 2).
74 Significant Improvements in Cartagena Bay Water Quality: The wastewater generated in the
western part of the city, accounting for approximately 35% of the total pollution load, is now
conveyed to the wastewater treatment plant and disposed through the submarine outfall.
Pollution levels in the Bay, particularly alongside Cartagena in the “Inner Bay,” are now
significantly reduced. There are still challenges, however, in achieving a full Bay clean-up
including contaminated stormwater run-off, industrial pollution, and pollution from the “Canal
del Dique” in the southern part of the Bay which discharges untreated effluent from numerous
small municipalities.
Restoration of the Ciénaga de la Virgen: The removal of wastewater discharges into the Lagoon,
coupled with improved water circulation produced by the La Bocana project, has transformed the
estuary. All key parameters, including coliforms, dissolved oxygen, biochemical oxygen
demand, and suspended solids are now within regulatory standards and odor problems have been
eliminated. Cartagena residents can now enjoy the Lagoon through boating and fishing activities,
and CARDIQUE is planning to transform the Lagoon and its surrounding area into a protected
ecological park.
Water Quality around the Submarine Outfall: ACUACAR, under the regulatory supervision of
CARDIQUE, has undertaken extensive water quality monitoring around the outfall discharge
area. The monitoring program follows international standards and indicates that outside of the
prescribed mixing zone there is no discernible impact on the seawater quality. Outside of the 500
meter mixing zone around the outfall diffuser, biochemical oxygen demand and suspended solid
concentrations are equal to ambient seawater quality levels, and total coliform levels are less
than 5 MPN/100 mL—a low level and considered suitable for human contact.
75 IBRD 40434
Preliminary and Indicative Water Quality Monitoring Results in Cartagena
Beaches
Total Coliforms
(CT) (NPM/100mL)
1200
Average Values
Before Treatment
(2012 and earlier)
1000
CT
800
NPM/100mL
Average Values
After Treatment (2013)
Water Quality Standard
600
200
che
s
400
0
Before
Bea
CT=31.4
NPM/
100mL
After
Cienaga de la Virgen
Dissolved Oxygen (DO)
(mg/L)
7000
7
6000
6
5000
5
mg/L
NPM/100mL
8000
8
4000
3
2000
2
0
25
s
20
DO
P
15
10
DO
5
CT
1
NT=0.4
mg/L
0
0
Before
NT
4
3000
1000
30
9
ach e
CT
Total Nitrogen (NT),
Total Phosphorous (P)(mg/L)
Be
9000
Total Coliforms (CT)
(NPM/100mL)
mg/L
10000
After
Before
B
f
After
Aft
Before
After
P=0.2
mg/L
Before
After
Cienaga
de la
Virgen
Cartagena Inner Bay
7000
Total Coliforms (CT)
(NPM/100mL)
9
CT
Dissolved Oxygen (DO)
(mg/L)
8
6000
7
DO
5000
4000
5
mg/l
NPM/100mL
6
3000
Ceballos Pumping Station
Estación de Bombeo Ceballos
4
3
Cartagena
Bay
2000
2
1000
0
CT
Before
After
1
0
DO
Before
After
GSDPM
Map Design Unit
This map was produced by the Map Design Unit of The World Bank.
The boundaries, colors, denominations and any other information
shown on this map do not imply, on the part of The World Bank
Group, any judgment on the legal status of any territory, or any
endorsement or acceptance of such boundaries.
NOVEMBER 2013
Figure 3-2. Preliminary Water Quality Monitoring in Cartagena.
76 3.14 Future Challenges in Coastal Environmental Restoration
Cartagena has passed through the first and perhaps most important stage: the collection,
treatment, and safe disposal of its wastewater. International experience has shown, however, that
this first phase must be complemented with additional initiatives to protect the coastal
environment.
•
Ensuring all households are connected to the sewerage system and that sewer overflows
are eliminated;
•
Achieving full compliance with industrial discharge standards;
•
Reducing stormwater related pollution;
•
Dealing with non-traditional pollutants including nutrients, heavy metals, pesticides, and
pharmaceutical waste;
•
Monitoring water quality in a continuous and extensive basis to assess progress in
pollution control and help determine the need to upgrade the level of wastewater
treatment;
•
Managing land-use, particular with respect to protecting riparian areas along beaches,
estuaries, and rivers; and
•
Managing fisheries and natural resource exploitation to ensure sustainable use of
resources.
In addition, the environmental condition of the Canal del Dique with heavy metals
contamination, which generates delta formations in the Bay of Cartagena, has raised serious
concerns. Current sedimentation along Canal del Dique has prevented inland water transport,
which is vital for the economic development of Cartagena. Besides, the canal provides multiple
77 environmental functions such as source of water and nutrients for the neighboring wetlands,
natural control of salt intrusion, and vital source of drinking water for the different settlements.
Thus, the canal will still require an intense review and planning to protect water quality in the
Bay (Menahem and Roberts, 2002).
Furthermore, Cartagena Bay is known to have problems involving sedimentation and
eutrophication. Urban planning and management will be required to organize settlements along
the canal. For example, in the absence adequate planning, the Cerro de la Popa could suffer from
landslides. La Mojana, a sub-region near the canal, could undergo an accelerated process of
environmental degradation including the desiccation of wetlands, alteration of the natural
hydrological regime, severe deforestation, critical habitat alterations, and remarkable loss of
biodiversity. Finally, agricultural and fishing production could be affected by negative effects of
illegal mining practices, which are responsible for the sedimentation and contamination with
mercury and other highly toxic heavy metals.
3.15 Analysis of the Cartagena Water Supply and Environmental Management Project
Overall, the Cartagena Water Supply, Sewage, and Environmental Management Project
was successful. Service was extended into poorer neighborhoods and the existing service was
greatly improved. The system has continuous water pressure and leakage has been reduced.
Residents are happy with the service and vocal in supporting ACUACAR. Each time that a new
mayor is elected, residents want to make sure the new mayor does not do anything to change the
service.
78 Several of the keys to success for a PPP are evident in the case of ACUACAR. The initial
investment by the private sector to organize ACUACAR required an income stream for
repayment of this investment. Not only did the Spanish water firm receive dividends for being a
shareholder in ACUACAR, but it also received a percentage of the profits. This arrangement
gave the Spanish firm two reasons to operate all aspects of its business efficiently, including
operations and maintenance, infrastructure investments, long-term debt management, and
collected tariffs.
ACUACAR also sought out and obtained the support of the community, which made
success possible. Hiring social workers, community relations specialists, and local unskilled
laborers for the construction of pipelines helped show customers that the company was different
from the previous municipal provider.
3.16 Environmental Policy Issues Revealed in this Project
Although the District of Colombia has extensive environmental regulations, they are
inadequate for a number of reasons. First, in many cases, urgently needed regulation simply does
not exist. Second, some regulations are incomplete and lack critical details. For example, the lack
of regulations regarding the scope and applicability of environmental and social impact analysis
has made the use of public participation virtually incoherent. Third, some regulations are overly
prescriptive and potentially inappropriate for local economic and social circumstances. These
inadequacies in Colombia’s regulations lead to many problems. They contribute to poor
coordination between the Ministry of Ministry of Environment, ACUACAR, and CARDIQUE,
causing difficulties for carrying out one of their basic functions—implementing regulations
established at the national level. The inadequacies also make it difficult for other institutions to
79 perform their assigned roles. For example, lack of regulation—from constitutional precepts to
specific information standards—limits advancement of the Colombian System of Environmental
Information. As previously noted, incomplete licensing and permitting regulations lead to
inconsistent requirements and enforcement across the water and wastewater sector and create
opportunities for corruption.
In addition, environmental management capacity varies markedly across the
water/wastewater sector (Galán Santos, 1998). This marked variability in regulatory capacity is a
significant problem that has far-reaching consequences. For example, environmental regulations
are stringently enforced in some areas and virtually ignored in others. Also, locally generated
funds are efficiently collected and invested in some areas but are scarce and inefficiently
invested in others. At both the national and the regional levels, regulatory capture and corruption
are significant problems within the water/wastewater sector.
At the national level, private-sector interests have far more influence on environmental
policymaking than the organizations responsible for representing civil society—Nongovernmental organizations (NGOs). Private-Sector interest groups have a strong influence on
decision-making. Members of boards of directors with strong ties to the private sector include
not only two dedicated private-sector representatives, but often mayors and even NGO
representatives. The latter members sometimes represent spurious local organizations set up by,
or closely tied to, industry. Private-sector influence aside, water/wastewater project decisionmaking is often unduly influenced by political considerations. For example, environmental
investments such as water distribution systems are sometimes spatially targeted to maximize
political payoffs instead of environmental benefits.
80 Furthermore, inadequate data on environmental quality and institutional performance is a
critical contributor to the water/wastewater failings sector. For example, environmental
laboratories, measuring stations, documentation centers, and basic cartography services are
inadequate. In addition, lack of priority-setting across environmental subsectors and programs
contributes to imbalances in budgetary priorities. Budgetary allocations are apparently driven
more by institutional history than by environmental needs (Gómez Torres, 2003). For example, a
recent audit of the Ministry of Environment found that rural environmental issues accounted for
three-quarters of the ministry’s investment budget, even though more than 70% of Colombia’s
population is urban.
Finally, inadequate mechanisms for public participation have been an issue throughout
Colombia water/wastewater regulatory systems. Water/wastewater projects at the local level,
particularly in rural areas, confront problems from both lack of security and leadership. Thus,
Colombia environmental regulatory systems currently lack a consistent system of prior
notification of the government’s intention to take many major actions, such as promulgation of
major regulations. A systematic provision for public comment is also lacking.
3.17 Summary of Key Lessons in the Cartagena Project
Cartagena’s experience can be summarized into the following key points:
1) An efficient and sustainable water utility is crucial for effective wastewater management—
which is fundamental for coastal cities. The creation of ACUACAR in 1995 was a key
component of Cartagena’s success.
81 2) Partnerships at the local, national, and international levels can facilitate and expedite
environmental improvements. The Colombian national government formulated the policy
framework for private participation in the water sector and enhanced environmental
management, as well as providing significant financial support for Cartagena’s infrastructure.
Strong partnerships at the local level between CARDIQUE, the District of Cartagena, and
ACUACAR were indispensable to program continuity and coherence. International institutions,
such as the World Bank, provided financial support and technical assistance.
3) Efficient environmental regulatory policies are very important to address water and sanitation
problems. Recently, a key role of the Colombian government in the emerging field of water
sustainability management was the development of effective and efficient environmental
policies. Although Cartagena water and sanitation services have improved significantly, such
new policies have helped the Ministry of environment, CARDIQUE, and ACUACAR to better
understand the causes of environmental problems.
4) Public relations, community outreach and building consensus among local stakeholders is
critical to planning and implementing wastewater programs. There is usually no obvious “best
technical solution,” rather in consultation with all stakeholders and taking into account all
dimensions a “preferred alternative” often emerges. The challenge is then to implement this
approach in an expeditious manner to avoid further environmental degradation.
82 5) Comprehensive wastewater management is long-term process and can take a decade or longer
to implement. In Cartagena, it took around five years (1995-2000) to ensure the proper policy,
institutional planning and financial arrangements were in place before the construction could
commence, and then over ten years (2000-2013) before the wastewater treatment system could
be fully constructed and commissioned.
6) Long-term, incrementally phased and prioritized programs are necessary for water pollution
control and environmental restoration. With the commissioning of the wastewater management
system, Cartagena has completed the first phase of its long-term program to restore the coastal
environment.
7) Submarine outfalls, combined with preliminary treatment, can be an appropriate solution for
protecting coastal areas such as beaches, bays, and estuaries, while providing flexibility for
future upgrades as necessary and when affordable. The feasibility of an outfall approach
depends, of course, on the capacity of the receiving water body to assimilate the discharges and
must be accompanied by extensive environmental, engineering, and social studies.
83 3.18 References
ACUACAR. 2009a. Establecimiento de la línea base primera fase del plan de monitoreo
ambiental del proyecto para el tratamiento y disposición final de las aguas residuales de
Cartagena de Indias. Revista # 9.
ACUACAR. 2009b. Información Técnica, Económica y Financiera suministrada por staff del
PIU# 234.
ACUACAR. 2009c. Presentaciones Técnica y Financiera. Presentacion # 15.
ACUACAR. 2009d. Evaluación Expost Plan Abreviado de Reasentamiento Barrios Isla de León
y Valla Luna. Coreo # 315.
ACUACAR. 2009e. Encuesta de usuarios. Datos de Laboratorio Basico #12.
ACUACAR. 1995. Historia ambiental de Cartagena y Indicadores del Proyecto. Reporte 1-15.
BID. Inter-American Development Bank. 2006. Informe de Terminación de Proyecto, PCR-IT #
1010.
Deloitte. 2009. Auditoria de los Recursos Administrados por Aguas de Cartagena S.A.# 001.
Galán Santos, A. 1998. “Rationalization of Spending in SINA: Institutional Firming towards
Decentralization.” Contract DNP-075-98, Final Report, Executive Summary. December
1998. Bogotá, Columbia.
Gómez Torres, M. 2003. Fiscal Policy for Environmental Management in Colombia. XV
Regional Seminar on Fiscal Policy, Economic Commission for Latin America and the
Caribbean.” January 2003. Santiago, Chile.
Hazen and Sawyer, Inc. 2001. Estudio de Factibilidad Para el Tratamiento de las Aguas
Residuales de Cartagena y Para la Disposición Final del Efluente Al Mar Adyacente a
Través de Un Emisario Submarino- Informe Final # 003.
Kress, M. and G.F. Gifford. 1984. Fecal coliform release from cattle fecal deposits.Water Res.
Bull. 20:61–66.
IDEAM (El Instituto de Hidrología, Meteorología y Estudios Ambientales). 2000. El proyecto
Visión Amazonia 2020. Grupo de Comunicaciones IDEAM: 3527160.
Menahem, L.B, Roberts, J.W. 2002. Social aspects of wastewater submarine outfalls in
developing countries: the case of Cartagena, Colombia. CO# 112-RE-145.
84 McCarthy, E. 2006. A Public-Private Partnership for the Provision of Water and Sewerage
Services in Cartagena, Colombia: Looking back at the first Seven years. Report # CA2345.
Neotropicos. 2002. Estudio de Impacto Ambiental Emisario Submarino de Cartagena. Reporte
final # AGUA-122.
UNEP. 2010. Urban Water Expansion, Cartagena, Colombia: Examples of Successful PublicPrivate-Partnerships Volume 15, 135-143.
World Bank. 2009. Project appraisal document on a proposed loan in the amount of US$85
million to the district of Cartagena. Report No. 3987634-CO.
World Bank. 2008. Project Management Reports (PMR) CR#200934.
World Bank. 2007. Implementation completion and results report-Guidelines Report No. 32034CO.World Bank. 2006. Approaches to Private Participation in Water Services: Toolkit /
Public-Private Infrastructure Advisory Facility & the World Bank. EAN # 978-0-82136111-5.
World Bank. 2005. Management Report and Recommendation in Response to the Inspection
Panel Investigation Report. Report No. INSP/R2005-000-CO.
World Bank. 2004. Cartagena Loan Amendment documents Report No. 323674-CO.
World Bank. 1999. Loan agreement documents Report No. 367934-CO.
85 Chapter 4
Effects of Oxygen, Nitrate and Aluminum Hydroxide on Methylmercury Efflux from MineContaminated Profundal Lake Sediments
ABSTRACT
Experimental and field-based studies have evaluated remedial actions to mitigate mercury (Hg)
bioaccumulation in lake ecosystems. This chapter discusses results of chamber studies evaluating
the effects of oxygen, nitrate and sodium aluminate addition on methylmercury (MeHg) efflux
from profundal reservoir sediments from Guadalupe Reservoirs and Almaden Lake near San Jose,
CA. These reservoirs' sediments were highly contaminated as a result of historical Hg mining in
their watershed (Guadalupe total Hg ~3 mg/kg dry wt.; Almaden total Hg 6-20 mg/kg dry wt.).
Aerobic conditions in the overlying water column corresponded with a decrease in MeHg efflux.
In Guadalupe Reservoir, MeHg efflux was 5.5 ± 2.2 ng/m2/d (average plus/minus standard
deviation, n = 4) under aerobic conditions and 22 ± 11 ng/m2/d under anaerobic conditions. In
Almaden Lake, MeHg efflux was -2.3 ± 0.8 ng/m2/d (loss from water column) under aerobic
conditions and 11 ± 4 ng/m2·d under anaerobic conditions. MeHg efflux corresponded with the
release of other redox-sensitive compounds at the sediment-water interface including iron,
manganese, ammonia and phosphate. Results suggest that the dissolution of reduced metal
oxides in surficial sediments may be an important source of MeHg efflux from sediments.
Furthermore, additional chamber experiments were conducted in Almaden Lake examining the
effects of nitrate and sodium aluminate on MeHg accumulation. Like oxygen, nitrate repressed
86 MeHg efflux to less than 0.2 ng/m2/d by maintaining a higher redox condition at the sedimentwater interface. However, sodium aluminate, a common lake management additive that enhances
the sorption capacity of surficial sediments mainly for phosphorus control, did not appear to
impede MeHg efflux from profundal sediment.
4.1. Introduction
Methylmercury (MeHg), a known neurotoxin, is one of the most widespread waterborne
contaminants in the U.S. (USGS, 2001a; UNEP, 2002; USEPA, 2004) and poses major public
health concerns (WHO, 2009). During the 20th century, mercury (Hg) concentrations in lake
sediments across Europe, Asia, and North America have increased three-fold, with the highest
contamination in lower-latitude lakes near major emission sources (UNEP, 2008). In the U.S, Hg
pollution has contaminated 18 million acres of lakes, estuaries, and wetlands, and 1.4 million
river miles (NRDC, 2014). Exposure to toxic levels of MeHg affects more than 630,000
newborns each year, interfering with brain and nervous system development (Babal, 2005).
Studies have established a link between children exposed to MeHg and developmental deficits in
language, memory, cognitive thinking, and visual spatial skills (Babal, 2005; Bose-O’Reilly et
al., 2010).
Hg found in aquatic systems originates from both natural and anthropogenic sources.
Natural sources include volcano eruptions, geologic Hg deposits, and volatilization from the
ocean. The primary anthropogenic sources include coal combustion, chlorine alkali processing,
gold mining operations, and metal processing (Bose-O’Reilly et al., 2010). In addition,
atmospheric deposition of Hg from fossil fuel combustion and other sources associated with
87 atmospheric particulate matter has led to increased concentrations in freshwater systems and
biota, even in remote areas free from direct anthropogenic influences (Cohen et al., 2004; Rada,
et al., 1989). Once Hg is discharged from paper mills, coal power plants, and mining operations
and settles into bottom waters, the potential for ionic Hg, Hg(II), to be methylated increases at
the sediment-water-interface under anoxic conditions (Wang et al., 2004). Primarily taken up by
pelagic phytoplankton, MeHg then accumulates through aquatic food webs and can cause a range
of toxic and reproductive effects in fish and wildlife (Gilmour et al., 1998). The strong reactivity
of MeHg with sulfhydryl groups of proteins is responsible for its high degree of bioaccumulation
(Beckvar et al., 1996). Phytoplankton can concentrate dissolved MeHg in the water column
approximately 100,000 times, making this a critical step in the bioaccumulation process (Watras
et al., 1994). Klasing and Brodberg (2008) summarized human exposure and toxicity of MeHg.
The primary route of MeHg exposure in humans in the U.S. is fish consumption. In turn, human
consumption of contaminated fish and wildlife poses significant health risks (USEPA, 2009).
MeHg forms via the methylation of dissolved inorganic Hg by anaerobic bacteria in
reduced lake sediments and bottom waters (Benoit et al., 2003). Specifically, sulfate-reducing
bacteria (SRB) become active using sulfate as the terminal electron acceptor during organic
matter oxidation, which leads not only to MeHg production, but also hydrogen sulfide (Slotton et
al., 1995). This production of MeHg represents an important pathway for bioaccumulation of
MeHg in the food chain due to the availability of organic matter and the sulfate-rich environment
in lakes (Todorova et al., 2009). Furthermore, hydrogen sulfide is extremely toxic to aquatic
biota (Reiffenstein et al., 1992) and can degrade the water quality of lakes and reservoirs so that
potable water use becomes problematic (Levine et al., 2004).
88 Another pathway leading to MeHg production and accumulation, under anaerobic
conditions, is the release of highly reduced compounds such as iron (Fe) and manganese (Mn)
from lake sediments to overlaying waters (Chadwick et al., 2006). These redox-sensitive metal
oxides undergo reductive dissolution in the methylation zone. This dissolution, in turn, promotes
the release of previously sorbed or complexed MeHg and Hg(II). This additional source of
Hg(II) in water bodies adds to the background Hg concentrations, deposited from the atmosphere
and watershed runoff, further enhancing the potential for MeHg production (Chadwick et al.,
2006; Regnell et al., 2001)
As noted earlier, human exposure to MeHg is mainly through consumption of
contaminated fish and shellfish (Al-Saleh, 2009). As a result, MeHg consumption has caused
serious health problems in humans and wildlife worldwide. For example, Japanese women who
consumed MeHg-contaminated fish (50 µg/g on average), from Minamata Bay, during
pregnancy gave birth to children who developed severe neurological defects, including cerebral
palsy, microcephaly, blindness, and seizures (Tsubaki, 1977; Harada, 1995; Goto, 2000). In
China, in the Guizhou Province near heavy coal-based industries, surface water samples
contained total Hg (THg) concentrations up to 12,000 ng/L (Horvat et al., 2003; Zhang et al.,
2010), which bioaccumulated in the aquatic food chains leading to serious health concerns
(Cheng et al., 2009). Similar results and effects were reported in Philippines (Gray et al., 2003),
and India (Sharma, 2003). In Canada, MeHg contamination in the Arctic ecosystem has been
linked to both autism and Alzheimer disease in humans (El-Hayek, 2007).
In the U.S., freshwater fish contaminated with MeHg has affected the livelihood and
health of citizens in California, New York, Wisconsin, and Minnesota (Genuis, 2009; Hightower,
2009; Krabbenhoft, 2004; WDNR, 1999). For example, in Northern California, several studies
89 have reported an increasing trend in cases of Hg exposure that negatively impact fetus
development (EHHA, 2003; CDPH, 1987). The National Research Council, in its 2000 report on
the toxicological effects of MeHg, pointed out that the population at highest risk is the offspring
of women who consume large amounts of fish and seafood (NRC, 2000). The report went on to
estimate that more than 60,000 children are born each year at risk for adverse
neurodevelopmental effects due to in utero exposure to MeHg.
Owing to in-situ production of MeHg in anaerobic zones of lakes and reservoirs,
mechanisms regulating MeHg formation, accumulation, and biomagnification are poorly
understood (Eckley et al., 2005). Despite much literature on the topic, the production and
transformation mechanisms of MeHg in aquatic environment are still somewhat unclear
(USEPA, 2006; Ullrich et al., 2001). Knowledge is also lacking about the many chemicals
(sulfides and dissolved organic carbon) and biological processes (bacterial activity levels) that
control Hg methylation and bioaccumulation in aquatic environment, particularly at the
sediment-water-interface (de Wit et al., 2012; Ullrich et al., 2001). A better understanding of the
environmental factors affecting the formation of MeHg in lakes is also needed (Winfrey at al.,
2009; Langston et al., 1998).
To address MeHg contamination issues, USEPA has increased fish consumption
advisories for MeHg, which now account for more than three-quarters of all fish consumption
advisories in the U.S. These new advisories illustrate the severity of the problem (USEPA 2010).
In addition, the Food and Drug Administration (FDA) released a new advisory warning the
public about eating Hg-contaminated fish (FDA, 2004). From 2006 to 2008, the number of lake
acres under advisory increased by 18%, and the number of river miles increased by 52%. In
90 2008, all 50 U.S. states issued fish consumption advisories, warning citizens to limit how often
they eat certain types of fish because of known Hg contamination (NRDC, 2014).
Growing concerns about MeHg accumulation related to anoxic profundal sediment in lakes and
reservoirs in the United States have stimulated interest in remediation alternatives for the
inhibition of MeHg production in aquatic environments. USEPA has proposed to strengthen
efforts to control Hg (USEPA, 2010). Federal and State agencies are looking for management
strategies and remedial techniques for these polluted sediments that protect the public
adequately. For example, the State of California Water Board has initiated a statewide control
program for Hg that will focus first on Hg in California’s reservoirs. Fish containing potentially
harmful amounts of Hg are found in numerous reservoirs across the state. The program’s
objective is to reduce concentrations of inorganic Hg in reservoir sediment to minimize MeHg
production and its subsequent bioaccumulation in fish. Potential source controls include
remediation of historic gold and Hg mines upstream of reservoirs. In addition, reservoir
management practices and management of fish species are being implemented which are
important tools in addressing Hg impairments (CWB, 2013). Moreover, the California Water
Board has determined that total maximum daily load (TMDL) is the primary regulatory vehicle
for achieving water quality goals. As a result, a number of TMDL programs are being
implemented. For example, a single TMDL for the Guadalupe River Watershed near San Jose,
California, that considers all Hg sources in the watershed has been established. This program
includes all tributaries of the Guadalupe River upstream of tidal influence (CWB, 2013).
The objective of the research in this study was to provide a more comprehensive
understanding in MeHg accumulation at the sediment-water-interface by analyzing three
different lake remediation alternatives. This study presents a clear and logical approach to
91 evaluate how Hg efflux from sediment behaves under: (1) aerobic versus anaerobic conditions,
(2) anoxic conditions with nitrate addition, and (3) anaerobic conditions with sodium aluminate
addition. In contrast to many studies that have focused on anaerobic sediment release of
phosphorus (Nurnberg, 1988; Golterman, 2001; Schauser et al., 2006), this study analyzed
replicate experimental chamber incubations that quantify the effects of oxygen, nitrate, and
aluminum hydroxide on MeHg efflux from contaminated profundal sediments. A few recent
studies have correlated the presence of oxygen and nitrate with suppression of MeHg
accumulation in bottom water (Matthews et al., 2013; Auer et al., 2013; Austin, 2011; Todorova
et al., 2009). For our study, it was hypothesized that nitrate and oxygen would repress MeHg
accumulation by inhibiting the activity of SRB and/or repressing the release of MeHg from Fe
and Mn oxides in surficial sediments while maintaining an adequate concentration of the oxygen
and nitrate in the hypolimnion (Todorova et al., 2009; Benoit et al., 2003). Further, it was
hypothesized that aluminum hydroxide floc unlike Fe would not be reduced, but rather would act
as a potential sorption sink for MeHg, even under anaerobic conditions (DeGaspari et al., 1993).
A unique focus of this paper was the evaluation of correlation between aluminum hydroxide and
MeHg efflux at the sediment-water-interface using undisturbed sediment cores. This allowed for
a conclusive evaluation of the ultimate fate Hg methylation in replicate experimental profundal
sediments under anaerobic conditions.
92 4.2. Materials and Methods
4.2.1. Study Sites
Almaden Lake and Guadalupe Reservoir, part of the Guadalupe River Watershed (Fig. 41), suffer the impacts of Hg mining, particularly during large runoff events when Hg-laden
sediments from mine wastes are discharged into the surface waters and accumulate downstream
(Tetra Tech, 2005). Both water bodies are highly productive, with abundant phytoplankton.
Almaden Lake (maximum depth of 12.5 m; surface area of 3.2 ha) is a eutrophic quarry pond
located in an urban park in San Jose, California. Guadalupe Reservoir (mean depth of 12 m;
maximum depth of 17; surface area of 30 ha; storage capacity of 4,212,000 m3) is a eutrophic
reservoir located east of San Jose. For decades, the Guadalupe Watershed has experienced
cultural eutrophication and Hg contamination. Historical Hg mining in the region fed the demand
of gold rush miners who used Hg to separate gold from crushed rock (SFBRWQCB, 2008; Tetra
Tech, 2006). The watershed, managed by the Santa Clara Valley Water District (SCWVD), is
used as a source for drinking water for Santa Clara County's 1.8 million residents. The water is
captured and then percolated into the regional groundwater basin for later withdrawal for potable
use.
The low dissolved oxygen concentrations in bottom waters during the summer and fall
dry season create conditions that enhance MeHg production, as demonstrated by the sampling
results from Almaden Lake and Guadalupe Reservoirs. As a result of these conditions and
processes, the water bodies facilitate the production and downstream export of MeHg, the form
of Hg that most readily bioaccumulates. Average total MeHg concentrations in deep water
discharges measured during a 2004 study were 12.8 ng/L in Almaden Lake and 23.6 ng/L in
93 Guadalupe Reservoir (SFBRWQCB, 2008; Tetra Tech, 2005). Water quality profiles collected in
2011 show that MeHg accumulation is associated with anaerobic/anoxic conditions. Field
monitoring of temperature and DO showed the reservoirs were thermally stratified and that DO
in bottom water was depressed (DO < 1 mg/L) (Fig. 4-2). Data collected by the SCVWD also
show a tight linkage between dissolved oxygen, nitrate and MeHg in bottom waters (Fig. 4-3).
Low dissolved oxygen concentrations correlated with an increase in MeHg efflux and depletion
of nitrate concentrations during anaerobic conditions also associated with an increase in MeHg
efflux.
4.2.2. Collection of Sediment Cores
This study included two sampling events. Sediment-water interface samples were
collected using a 15 x 15 cm Wilco standard Ekman dredge and cylindrical 1.8-L polycarbonate
chambers. The dredge was brought to the surface with minimal disturbance. The collected
sediment was sub-sampled out of the dredge and placed into the chambers. Chambers have a
diameter of 9.5 cm, a height of 14 cm and a surface water of 71 cm2 (Beutel et al., 2007, 2006;
Beutel, 2003). The sediment-water interface samples were 4-10 cm thick cores.
During the first sampling event on November 16, 2012, replicate sediment-water
interface chambers were collected from Almaden Lake (5 chambers) and Guadalupe Reservoir (4
chambers). For this experiment, quadruplicate chambers were incubated under aerobic followed
by anaerobic conditions. For the second event on May 3, 2013, sixteen profundal sediment-water
interface samples were collected at the deepest station in Almaden Lake. In this experiment,
quadruplicate sediment chambers were incubated under four conditions: anaerobic control,
oxygen addition (air), nitrate addition (calcium nitrate), and aluminum floc (sodium aluminate).
94 4.2.3 Sediment Incubations
The nine chambers from the first sampling event were stored in an incubator at 14 °C.
Chambers were incubated over two phases: an aerobic phase for 10 days by bubbling with air,
and an anaerobic phase for 18 days by bubbling with nitrogen gas. Chamber water was tracked
over time for MeHg, THg, nitrate, ammonia, phosphate, sulfate, Fe and Mn. One chamber from
Almaden Lake was monitored every few days during the anaerobic phase for a limited number of
parameters (Fe, Mn, MeHg) to develop a more complete time series of accumulation of redoxsensitive compounds in chamber water.
The sixteen chambers from the second event were stored in an incubator at 14 °C.
Quadruplicate chambers were incubated as an anaerobic control (bubbled with nitrogen gas) and
three treatments: oxic (bubbled with air), anoxic (bubbled with nitrogen gas and enriched with
nitrate), and aluminum hydroxide floc (bubbled with nitrogen gas and enriched with sodium
aluminate). For nitrate chambers, 4.215g of calcium nitrate (CaNO3) were diluted with 250 mL
of lake water and chambers were spiked with a mean of 20 mg N/L of nitrate. For sodium
aluminate (NaAlO2) chambers, a primary solution was made with12.5 g NaAlO2 which were
diluted with 250 mL of lake water and neutralized to circumneutral pH with 2N HCl. Chambers
were then spiked with 12 mL of NaAlO2 solution resulting in an areal Al dose of 30 g/m2, which
is a typical dose range (10 to 40g/ m2) used in lake applications (DeGasperi et al., 1993; Cooke et
al., 1993; Welch and Cooke, 1999). Chambers were incubated for 24 days.
95 4.2.4. Water Quality Analyses
THg samples were collected from the chambers in clear 250 ml glass bottles and
preserved with 1% bromine monochloride solution; MeHg samples were collected in 250 ml
amber glass bottles to reduce photodegradation and preserved with 0.5 % HCl. Trace metals (Fe
and Mn) and nutrients were collected in 60 ml pre-wash plastic bottles. Metals were preserved
with 0.5% nitric acid and nutrients were filtered through 0.45 micrometer filters and frozen for
analysis. Each sample was drawn from chambers with clean syringes and Teflon tubing. All Hg
sampling protocol followed USEPA methods 1630 and 1631(USEPA, 2001, 2002). MeHg and
THg samples were analyzed using the Brooks Rand MERX-M Auto Analyzer with EPA method
1630 modular analytical technique, which includes sample distillation, ethylation, purge and
trap, thermal desorption, gas chromatography separation, pyrolyzation, and subsequent detection
via a Tekran 2500 CVAFS (USEPA, 2001). Standard quality control procedures for MeHg
included duplicates (< 25% relative percent difference), matrix spikes (77-125% recovery), and
method blanks. Method detection limits were 0.2 ng/L for THg and 0.02 ng/L for MeHg.
Nutrient samples were analyzed for soluble reactive phosphorus (SRP), nitrate, and ammonia on
a Lachat QuikChem 8500 nutrient auto-analyzer using standard colorimetric method (APHA,
1998). Nutrient detection limits were 0.01 mg/L. Samples for Fe and Mn analyses measured
using inductively coupled plasma mass spectrometry (APHA, 1998). Method detection limits
were 50 µg/L for Fe and 0.1 µg/L for Mn. Sulfate samples were analyzed using ion
chromatography with EPA method 300 (USEPA, 1993). Due to the high concentration of sulfate,
the matrix spike/matrix spike duplicate (MS/MSD) were reported above the calibration range.
The associated laboratory control sample (LCS) met acceptance criteria. Method detection limit
was 20 µg/L.
96 4.2.5. Sediment Quality Analyses
Sediment was analyzed for a number of standard characteristics. Subsamples were placed
in weighed crucibles and weighed. Weight loss was measured after heating at 100 oC overnight
to remove water, at 550 oC for four hours to remove organic matter, and at 1000 oC for two hours
to remove carbonates. Weight losses associated with water and carbon dioxide evolutions were
quantified by recording sample weights before and after controlled heating (ignition at 550 and
1000 oC) (Heiri et al., 2001; Santisteban et al., 2004). From these data the percentages by dry
weight of sand, silt, and clay were calculated for each sample and classified according to the
nomenclature of Shepard (1954). Sediment cores were analyzed using the ultrasensitive DMA-80
analyzer to determine THg following EPA method 7473 (USEPA, 2007). Sediments were also
analyzed for total nitrogen (N), total carbon, total phosphorus and total sulfur (S), iron, and Mn
using inductively coupled plasma mass spectrometry (ICP-MS) (APHA, 1998).
4.3. Results
4.3.1. Sediment Quality
In Almaden Lake, sediment cores were collected from 4 different stations (depth = 12.5
m, 10.5 m, 8.5 m and 8 m). Sediment compositions were reported with clay (16-24%), silt (3239%), sand (35-53%), loss on ignition (83-88%), total carbon (2.4%), total nitrogen (0.27), total
Hg (7-35 mg/kg dw), total Fe (51 mg/g dw), total phosphorus (0.84 mg/g dw) and total sulfur
(2.3 mg/g dw) (Table 1). Guadalupe sediment cores were collected from a single station at depth
15 m. Sediment compositions included clay (38%), silt (30%), sand (32%), loss on ignition
97 (91%), total carbon (5.9%), total nitrogen (5.9%), total Hg (6 mg/kg dw), total iron (40 mg/g
dw), total phosphorus (1 mg/g dw), and total Sulfur (7 mg/g dw) (Table 4-1).
4.3.2 Study 1: Total Mercury, Sulfate and Methylmercury Effluxes under Aerobic and Anaerobic
Conditions
In Guadalupe Reservoir, aerobic conditions yielded high rates of THg efflux (119 ± 110
ng/m2/d) and sulfate efflux (66.9 ± 21.7 ng/m2/d). But in Almaden Lake, aerobic conditions
substantially repressed THg efflux (-207 ± 98 ng/m2/d) and lowered sulfate efflux (-90.3 ± 53.1
ng/m2/d) (Table 4-2; Fig. 4-4). Note negative values indicate loss from the water column. Data
sets were statically significantly different (two-tailed paired t-test, p < 0.05).
Under anaerobic conditions, Guadalupe showed THg efflux (45.9 ± 74.4 ng/m2/d)
slightly lower than levels Almaden (51 ± 35 ng/m2/d), but sulfate efflux (-167 ± 17.6 ng/m2/d) in
Almaden was substantially lower than levels in Guadalupe (-56.3 ± 37.3 ng/m2/d) (Table 4-2;
Fig. 4-4). In Almaden Lake, aerobic conditions decreased MeHg efflux (-2.4 ± 0.8 ng/m2/d) but
Guadalupe showed positive MeHg efflux (5.5 ± 2.2 ng/m2/d) at the sediment-water interface.
Under anaerobic conditions, Almaden MeHg efflux (11 ± 4 ng/m2/d) was slightly lower than
Guadalupe MeHg efflux (22 ± 11 ng/m2/d) (Table 4-2; Fig. 4-4). Data sets were nearly
significantly different (two-tailed paired t-test, p = 0.07). The ratio of MeHg to THg in Almaden
was 0.04 ± 0.2 in aerobic chamber water but increased to 0.17 ± 0.06 in anaerobic chambers. In
Guadalupe Reservoir, the ratio of MeHg to THg was 0.09 ± 0.01 in aerobic chamber water but
increased to 0.25 ± 0.07 in anaerobic chambers.
98 4.3.3 Study 1: Iron and Manganese in Experimental Chambers
In Almaden Lake, both Fe (-2.2 ± 0.9 mg/m2/d) and Mn (-1.1 ± 0.9 mg/m2/d) efflux rates
were negative under aerobic conditions. However, in Guadalupe, only Mn (-4.2 ± 0.5 mg/m2/d)
efflux showed a negative efflux while Fe (2.5 ± 0.5 mg/m2/d) showed a positive efflux under
aerobic conditions (Table 4-2; Fig 4-4). During the anaerobic phase, Fe (6.7 ± 5.6 mg/m2/d; 4.1 ±
2.3 mg/m2/d) and Mn (17.7± 6.7 mg/m2/d; 20.6 ± 7.0 mg/m2/d) efflux increased at both sites,
with substantial increases in Mn. During all incubation phases, Fe and Mn showed a strong
correlation revealing the common cycling mechanisms shared by these two metals at the
sediment-water interface.
4.3.4 Study 1: Cumulative Mass of Methylmercury, Manganese and Iron in Anaerobic Phase
One chamber from Almaden Lake was monitored every few days for a period of 17 days
under anaerobic conditions. A limited number of parameters (Fe, Mn, MeHg) were evaluated to
develop a more complete time series of accumulation of redox-sensitive compounds in chamber
water. Results showed at day 0, mass of MeHg was 0.41 ng while Fe and Mn were 0.05 and 0.01
mg respectively. At day 3, mass of MeHg slightly increased to 0.98 ng but mass of Fe showed no
change (0.05 mg) while Mn increased to 0.77 mg. At day 8, mass of MeHg substantially
increased to 4.79 ng while Fe and Mn increased to 2.81 and 1.35 mg (Fig. 4-5). At day 12, mass
of MeHg slightly increased to 5.62 ng while Fe and Mn continued to increase to 3.05 and 1.89
mg respectively. At day 17, mass of MeHg showed an increase from 5.62 to 6.03 ng but Mn
dropped to 3 mg while mass of Fe increased slowly from 1.89 to 1.92 mg (Fig. 4-5).
99 4.3.5 Study 2: Efflux Rates for Methylmercury, Manganese, Iron and Ammonia
Four alternative treatments were evaluated: anaerobic, aerobic, anoxic with nitrate
addition, and anaerobic with sodium aluminate addition. In the anaerobic control, high level of
MeHg and ammonia efflux were observed at 10.8 ng/m2/d and 3.2 mg/m2/d, respectively, but Fe
efflux decreased negatively to -0.56 mg/m2/d while Mn efflux was 1.7 mg/m2/d. Data sets were
statically significantly different (two-tailed paired t-test, p < 0.05). In aerobic control, all
parameters showed a decrease in efflux rates. MeHg and ammonia efflux rate dropped to 0.5
ng/m2/d and 1.6 mg/m2/d while Mn and Fe decreased to -1.05 and -1.95 mg/m2/d respectively
(Fig. 4-6). In anoxic with nitrate treatment, MeHg and Fe efflux continued to decrease with
efflux rates of 0.2 ng/m2/d and -0.32 mg/m2/d, respectively, while ammonia and Mn were 3.2 and
1.4 mg/m2/d, respectively. In sodium aluminate/anaerobic treatment, MeHg efflux substantially
increased to 10.5 ng/m2/d and ammonia also increased to 3.7 mg/m2/d. Fe and Mn effluxes were
0.2 and 1.0 mg/m2/d (Fig. 4-6).
4.4. Discussion
4.4.1. Comparison of Release Rates of Methylmercury Efflux
In our experimental study, MeHg efflux ranged from -2.3 to 22 ng/m2/d which was
similar to MeHg efflux rates found in other Hg-contaminated freshwater lakes and reservoirs
(Table 2). Kuwabara et al. (2003) measured MeHg efflux rates of 1-3 ng/m2/d from aerobic
profundal sediments from Camp Far West Reservoir, a mesotrophic reservoir impacted by
historic Hg mining activity in the Sierra Nevada Mountains of Northern California. Kuwabara et
al. (2002) in collaboration with USGS examined sediment-water core incubations in Lahontan
100 Reservoir, NV, a mining impacted reservoir, to determine the benthic flux from the sediment into
overlaying water. MeHg efflux rates ranged from -3.9 to 27.4 ng/m2/d. Our rates were generally
near or higher than levels measured in uncontaminated systems, though there are few studies of
non-contaminated systems. Efflux rates of MeHg in a non Hg-impacted Deer Lake were < 0.4
ng/m2/d (Cox, 2011). However, other non-chamber studies that evaluate MeHg accumulation via
seasonal mass balance have shown relatively high net accumulation rates. Watras et al. (1996)
reported MeHg production rates in temperate freshwater lakes ranging from 1.4 to 14 ng/m2/d.
Specifically in north-central Wisconsin, Watras et al. (1994) examined two non-Hgcontaminated lakes and found MeHg efflux rates (>50 ng/m2/d) in Little Rock Lake and (>54
ng/m2/d) in Mary Lake due to atmospheric deposition of Hg. Both Lakes are dimictic that stratify
thermally soon after ice-out in spring and remain stratified until autumn. As noted above,
chamber studies from Hg contaminated lakes showed higher MeHg fluxes. This correlated with
the growing evidence that sedimentary production of MeHg is influenced strongly by the
availability of dissolved inorganic Hg to methylating bacteria. As inorganic Hg(II) is the
substrate for the production of MeHg, abundant availability of Hg can promote higher potential
for MeHg efflux (Benoit et al., 2003). Several studies have found a positive relation between the
methylation rate and the Hg(II) concentration (Hammerschmidt and Fitzgerald 2004, 2006;
Fitzgerald et al., 2007; Benoit et al., 2003).
Most of the research quantifying Hg efflux from sediments has focused on coastal marine
systems. Release rates measured in this study were lower than those reported for marine systems.
Benoit et al (2009) reported MeHg efflux rates range from -0.9 to 41 ng/m2/d for coastal marine
sediments. Point et al. (2007) examined the Thau Lagoon in France. This study found MeHg
efflux rates from -32 to 68 ng/m2/d. Hammerschmidt and Fitzgerald (2008) examined MeHg
101 efflux from marine sediment using shipboard benthic flux chambers and reported MeHg efflux
from 15 to 19 ng/m2/d. Additional studies from Hg-contaminated marine systems that measured
MeHg efflux rates in excess of 100 ng/m2/d (Choe et al., 2004; Covelli et al., 2008). Marine
systems have a tendency to have very high levels of biological sulfate reduction that result in
elevated rates of Hg methylation and enhanced MeHg efflux rates from sediments
(Hammerschmidt et al., 2004). As noted by Lawson and Mason (1998), in marine ecosystems,
the chemical speciation of MeHg and its bioavailability are enhanced by an abundance of sulfur.
Similar observation were made by a number of studies (Han et al., 2007; Benoit et al., 2003),
reinforcing the contention that SRB in anoxic waters and sediments are the major producers of
MeHg. These systems are therefore considered to be hot spots for MeHg production. It is
important to note that many field studies have found that MeHg efflux varies substantially with
different type of systems. Marine systems tend to have higher MeHg fluxes compared to
freshwater systems (Table 4-3) (Benoit et al., 2009; Covelli et al., 2008; Lambertsson and
Nilsson 2006).
4.4.2. Efflux Rates in Almaden Lake versus Guadalupe Reservoir
Under oxic conditions Almaden showed uptake of both THg and MeHg, as expected
(Table 4-2). But an unexpected result of this study was that Guadalupe showed continuous
release of these compounds. In Guadalupe chambers, aerobic release of THg was twice the
anaerobic release rate, and aerobic release of MeHg did decrease but was still positive. Aerobic
release of MeHg was one fourth of anaerobic release (Table 4-2).
Almaden responded conventionally to oxygenation. MeHg efflux was repressed, and this
repression corresponded with substantial drops in total Fe, Mn, ammonia, and SRP under oxic
versus anoxic conditions. As discussed by Watras (2009), lakes with anoxic hypolimnia tend to
102 accumulate MeHg at concentrations that are orders of magnitude higher than lakes with oxic
bottom waters. Anoxic conditions support the activity of anaerobic microorganisms including
sulfate-reducing bacteria that methylate ionic Hg (Fleming et al., 2006; Gilmour et al., 1992). In
addition, anoxic conditions can lead to the production of sulfide, which at moderate levels can
enhance bioavailability of ionic Hg for methylation (Benoit et al., 2003) and MeHg for uptake
into phytoplankton through the formation of neutral lipophilic Hg-S complexes (Masson, 2002).
Sulfide may also increase the diffusive efflux of MeHg from anaerobic sediment (Holloweg et
al., 2010) and can strip Hg from settling metal oxides before it settles onto the sediments (Morel
et al., 1998). Anoxic conditions may also result in reduction and dissolution of Mn and Fe
oxides in surficial sediments with corelease of Hg/MeHg-DOC complexes (Feyte et al., 2010;
Chadwick et al., 2006; Ullrich et al., 2001).
Mn and Fe are highly sensitive to redox conditions and oxic conditions are typically
effective in repressing the release of reduced Mn (II) and Fe (II) from reservoir sediments
(Bryant et al., 2011). The observed difference in MeHg efflux rates between oxic and anoxic
conditions in Almaden may also be the result of sorption with Fe (III) in surficial sediments
during the oxic phase . Under oxic conditions, Fe and Mn oxides have the ability to sorb MeHg,
but in contrast to the anoxic phase, the reduction of these oxides can enhance the release of
MeHg to overlaying water (Regnell et al., 2001). A handful of studies have reported that Fe and
Mn oxides play important roles in MeHg accumulation at the sediment water interface through
impacts on Hg complexation (Chadwick et al., 2006; Merritt et al., 2008). Sediments can be a
source of both Hg (II) and MeHg through the reduction of oxyhydroxides of Fe (III) and Mn (IV)
with subsequent diffusion of new dissolved Hg species to the overlying water (Chadwick et al.,
2006) .Fe (II) and Mn (II) have been detected in anaerobic bottom water as a result of the
103 reduction of metal oxides in profundal surface sediments and sinking particulates from
overlaying surface waters (Davison, 1993; Balistrieri et al., 1992). As observed in our study, Fe
and Mn in Almaden indicated a strong redox response to anoxic conditions. Increases in Fe and
Mn suggested diffusive flux from sediments into the overlying water. It is equally important to
observe that MeHg accumulates only after Fe starts to accumulate, even though Mn
accumulation started earlier. Mn increased first in anaerobic chamber water because Mn(IV) is
biologically reduced at higher redox potentials than Fe(III). It appears that anaerobic conditions
and associated Fe oxide reduction were mainly the source of MeHg in chamber water. Similar
responses were observed by a number of researchers (Chadwick et al., 2006; Merritt et al., 2008;
Meili, 1997; Davison, 1993; Balistrieri et a., 1992). For example, Chadwick et al. (2006)
evaluated Hg cycling in the Experimental Lakes in Ontario, Canada. They found Fe and Mn
oxides had an indirect influence on Hg accumulation at the sediment water interface. Under
reducing conditions, these metal oxides appeared to release sorbed dissolve organic, which likely
contained Hg. As a result, metal oxides can also be a sink for Hg. Meili (1997) also discussed the
importance of adsorption of Hg to oxyhydroxides in lakes. His research suggested the
enrichment of MeHg in anoxic waters may result from the sedimentation of Hg-laden
oxyhydroxides of Fe and Mn from the surface and their dissolution in the anoxic bottom water.
Differences in total ammonia and dissolved orthophosphate between oxic and anoxic
conditions in Almaden suggest that oxygen inhibited sediment orthophosphate release, likely by
repressing the biological reduction of phosphate-containing iron-oxide complexes in profundal
surficial sediments (Golterman, 2001). Oxygen also inhibited sediment release of ammonia,
likely by enhancing nitrification of ammonia to nitrate and ammonia assimilation into bacterial
biomass (Beutel, 2006).
104 Guadalupe chambers responded in a surprising way in that THg efflux was high and
MeHg efflux was not inhibited under aerobic conditions. For this scenario, we noticed a
correlation between the efflux of ionic Hg, Fe and sulfate under oxic conditions. Our results
showed that Fe efflux under oxic conditions was still positive; meanwhile the efflux of Mn,
which is more resistant to rapid abiotic oxidation, was negative (Table 4-2). This suggests that
there was some source of Fe release. Also, in contrast to Almaden where sulfate was negative
under both oxic and anoxic conditions, we observed substantial release of sulfate in experimental
sediments (Table 4-1) leading to high sulfate efflux rates (Table 4-2) at the sediment-water
interface. As discussed by Kusel et al. (2002), lakes near mining areas can be characterized by
low pH and nutrient status, and high Hg, Fe and sulfate ions due to the oxidation sulfide minerals
in the surrounding mine tailings.
In Guadalupe chambers, these observations were most likely due to the oxidation and
dissolution of cinnabar (HgS) and pyrite (FeS2) in profundal sediments. Under oxic conditions,
the formation of oxidation products from cinnabar and pyrite can cause liberation of Fe, Hg and
sulfur as sulfate ions. This process involves the transfer of several electrons from each sulfur
atom in the minerals to an aqueous oxidant. Since cinnabar is unstable in the presence of oxygen,
Hg may be released by simple oxidative dissolution (Eq. 1). Likewise, the oxidation of pyrite
leads to the oxidation of sulfide to sulfate with the co-release of dissolved ferrous iron. After
dissolution the ferrous iron undergoes oxidation to ferric iron, which then hydrolyzes to form
insoluble ferric hydroxide (Eq. 2). The overall reactions are commonly written as (Barnett et al.,
2001; Rimstidt et al., 2003):
105 HgS + 2O2 + 2 H2O => Hg(OH)20 + SO42- + 2H+ (cinnabar)
Eq. 1
FeS2 + 15/4 O2 + 7/2H2O => Fe(OH)3 + 2SO42- + 2H+ (pyrite)
Eq. 2
This phenomenon has been observed in other Hg-contaminated sites (Barnett et al., 2001;
Monterroso et al., 1998; Holley et al., 2007). Barnett et al. (2001) investigated cinnabar with
respect to oxidative dissolution by dissolved oxygen focusing on Hg cycling in contaminated
environments. Results showed cinnabar, owing to the presence of reduced sulfur, is
thermodynamically unstable in the presence of oxygen, causing Hg to be released. Therefore,
dissolution will have a much larger thermodynamic driving force in the presence of oxygen
(Barnett et al., 2001). Monterroso et al. (1998) evaluated the quality of drainage waters from
Puentes Lignite Mine Dump in Galicia, Spain, containing sediments with different amounts of
carbonaceous material, with pyrite often being present. Results showed the presence of elevated
concentrations of Fe, sulfate, and H+, liberated from the oxidation of Fe-S (Monterroso et al.,
1998).
In addition to Hg and Fe release during oxidation of cinnabar and pyrite, there is growing
evidence that for many sulfide minerals such as cinnabar and pyrite, the first step in oxidation
produces a sulfate-enriched zone in natural waters (Buckley and Woods, 1987; Kartio et al.,
1998; Rimstidt et al., 2003). Since sulfide minerals were likely the dominant source of sulfate
release in our chamber, it is reasonable to conclude that this evidence confirmed the high levels
of sulfate observed in our experiment (Table 4-2).
Holley et al. (2007) discussed Hg mobilization by oxidative dissolution of cinnabar
(alpha-HgS) and metacinnabar (beta-HgS) in oxygenated slurry experiments. Results showed
106 that a great amount of sulfur is oxidized to thiols, which tend to be liberated over time as sulfate;
therefore, the initial observation of sulfate production via oxidative dissolution underestimates
the actual amount of mineral that has been oxidized.
Much of the Hg that is liberated by oxidative dissolution of cinnabar ends up sorbing to
remaining mineral surfaces, and as a result Hg fluxes tend to be 1 to 2 orders of magnitude lower
that sulfate fluxes (Holley et al., 2007). Presumably, this scenario was observed on the macro
scale in our chambers. Hg fluxes were approximately 1.2 orders of magnitude less than sulfate
fluxes. It is reasonable to conclude that source of sulfate and Hg in our chambers was due to
oxidative dissolution.
It is important to note that these oxidation reactions, however, are extraordinarily
simplistic compared to the realities of chemistry and the actual processes occurring in the
environment (Nordstrom, 1997). Because so many electrons must be removed from each sulfur
atom to oxidize it to sulfate, this oxidation process is one of the most complex processes in
sulfide mineral oxidation. It is unlikely that more than one electron is removed from the sulfur
atom at a time, therefore, there must be several steps to this overall process (Rimstidt et al.,
2003; Williamson and Rimstidt, 1992; Basolo and Pearson, 1967).
Furthermore, in Guadalupe chambers, MeHg efflux was not reversed even under oxic
conditions. The high availability of ionic Hg and sulfate at the sediment-water interface could
have enhanced methylation and could account for MeHg release under oxic conditions. As
observed in Guadalupe chambers, sulfate, a key substrate for methylation bacteria, was very high
in chamber water. Gilmour et al. (1992) reported that high levels of sulfate above intact sediment
cores resulted in increased microbial production of MeHg from bioavailability of inorganic Hg.
Thus, even though chamber water was oxic, high sulfate levels combined with elevated ionic Hg,
107 both liberated via oxidative dissolution of Hg-containing minerals, likely stimulated Hg
methylation in anoxic sediments. And this MeHg diffused upwards into oxic chamber waters.
4.4.3. Effects of Oxygen and Nitrate in Experimental Chambers
Our experimental efforts confirm that MeHg efflux rates from both oxic and nitrate
addition chambers were repressed at the sediment water interface (Fig. 4-6), which correlate well
with reported relationship between oxygen and nitrate and sediment MeHg release (Matthews et
al., 2013; Austin et al., 2011; Todorova et al., 2009; Effler et al., 2008; Eckley et al., 2006;
Benoit et al., 2006; Golterman, 2001). Our study demonstrated oxygen inhibited the release of
MeHg from lake sediments in chamber water (Fig. 4-6). Similar dynamics were observed in
Benoit et al. (2006). The study reported the penetration of oxygen into sediments with the net
result of pushing the site of Hg methylation deeper into sediments and eventually inhibiting
MeHg release. In addition, Gilmour et al. (1992) amended intact lake sediment cores with Hg
and incubated them under water with various sulfate concentrations. High levels of MeHg in
surface sediment correlated with intermediate levels of sulfate but dropped significantly in welloxygenated sediments. DeLaune et al. (2004) showed that Hg-enriched surface sediments from
three Louisiana lakes accumulated less MeHg when water above the sediments was aerated
(oxygen). Klein (2006) reported a number of studies in which MeHg production was inhibited
from lake sediments by maintaining an oxidized sediment-water interface. Golterman (2001)
suggested that oxygen levels and oxidation-reduction potential at the sediment-water interface
are particularly critical in regulating the flux of reduced and oxidized substances into and out of
profundal sediments.
108 Our experimental study also demonstrated the presence of nitrate overlaying the
sediment-water interface was effective in repressing the release of MeHg in chamber water (Fig.
4-6). Some researchers even argue that nitrate addition may be more effective and ecologically
benign than aeration in inhibiting MeHg production in the profundal zone (Effler and Matthews,
2008, Matthews et al. 2013). An interesting potential advantage of nitrate is that at sites with
contaminated sediments, nitrate addition can elevate redox without relieving hypoxia, thereby
limiting the exposure of aerobic macrobenthos and connected food webs to sediment-based Hg
contamination. Similar studies confirmed nitrate’s ability to repress MeHg accumulation in lakes
and reservoirs (Mathews et al., 2013; Todorova et al., 2009). Sulfate-reducing bacteria activity in
lake sediments and related internal production of MeHg may be inhibited by maintaining
adequate concentrations of alternate, energetically favorable electron acceptors, such as oxygen
and/or nitrate (Stephan et al., 1988). Due to its effect on redox conditions in the hypolimnion,
increased nitrate concentration can reduce the process of methylation. For example, Matthews et
al. (2013) conducted a two-year pilot study of nitrate addition to Onondaga Lake, a Hgcontaminated lake in New York. Results showed that maximum concentrations of MeHg and
soluble reactive phosphorus in bottom waters decreased by 94% and 95%, respectively. Austin
(2011) documented nitrate repression of MeHg accumulation in bottom waters, including one by
the consulting firm CH2MHill that added liquid calcium nitrate to a number of small lakes in
Minnesota. One potential downside of nitrate addition is the production nitrous oxide. During
denitrification, oxidized forms of N such as nitrate and nitrite are converted to dinitrogen and, to
a lesser extent, nitrous oxide gas. Nitrous oxide is a major contributor to the atmospheric ozone
destruction and a potent greenhouse gas. However, the nitrous oxide yield in fresh water systems
109 is low, generally ranging between 0.1 and 1.0% of total N gas production (Seitzinger and
Kroeze, 1998).
4.4.4. Sediment Ammonia Release
In our research, our first study showed ammonia release under aerobic conditions was
lower compared to anaerobic conditions in both Almaden lake (aerobic: 14.9 ± 13.9 mg/m2/d;
anaerobic: 29.0 ± 4.1 mg/m2/d) and Guadalupe reservoir (aerobic: -1.5 ± 5.2 mg/m2/d; anaerobic:
16.6 ± 1.5 mg/m2/d) (Table 4-2). In our second study, ammonia release under aerobic treatment
was low (1.7 mg/m2/d) in comparison to other treatments (Fig. 4-6). Under both nitrate-rich and
nitrate-free anoxic (anaerobic) conditions, sediment ammonia release rates were relatively high
(3.2 mg/m2/d). This indicated that when oxygen is present, biological nitrification of ammonia to
nitrate could proceed, but in the absence of oxygen nitrification in sediments was inhibited.
Anaerobic microorganisms also have much lower cell yields and growth rates compared to
aerobic microorganisms. Thus, anaerobic bacteria take up ammonia liberated from decaying
matter at much slower rates than aerobic bacteria (Snoeyink and Jenkins, 1980). This could
exacerbate ammonia release under anoxic conditions. Nitrate-rich anoxic conditions can also
result in enhanced biological ammonia removal. But this did occur in our chambers. As
mentioned by Burgin et al. (2007), nitrate may be reduced either to ammonium, as a form of
dissimilatory reduction of nitrate to ammonium (DNRA), or to N2, as a form of denitrification. In
this process, the predominant fate of the reduced nitrate may be determined by the ambient
concentration of free sulfide, which is known to inhibit the final two reduction steps in the
110 denitrification sequence. As a result, sulfide inhibition of these terminal steps may drive the
reduction to ammonium rather than N2O and N2 (Burgin et al., 2007).
Similar dynamics between ammonia, oxic, nitrate-rich and nitrate-free anoxic (anaerobic)
conditions were observed in a number of studies (Beutel et al., 2006; Rysgaard et al., 1994;
Mengis et al., 1997; Graetz et al., 1973). For example Beutel et al. (2006) evaluated how oxygen
and nitrate in overlaying water affect nutrient release from profundal sediments in a reservoir
(Lake Mathews, CA) and as well as in experimental sediment-water chamber incubations. The
study found that under nitrate-free anoxic conditions, ammonia release from sediments in
experiment incubations was 2.8 mg/m2/d; this compared to -3.6 mg/m2/d (uptake) under oxic
conditions. As observed in our study, only oxic phase and not nitrate-rich and nitrate-free anoxic
(anaerobic) phase was successful in repressing sediment ammonia release, perhaps by enhancing
biological nitrification and assimilation in surficial sediment under oxic conditions. Cubas (2012)
performed extensive experimental and field studies of nitrogen cycling in the Occoquan
Reservoir, Virginia. Results showed that ammonia release was higher under anoxic conditions
versus oxic conditions. But when nitrate input to the reservoir was increased, sediment
ammonium release was decreased. As a result, nitrate was converted to nitrogen gas as it moved
through the reservoir. On the other hand, Austin et al. (2015) reported substantial DNRA driven
by high sulfide levels in a shallow lake in Minnesota where nitrate was added to repress Hg
accumulation. The main nitrate sink was anaerobic oxidation of sulfidic compounds, but some
dissimilatory nitrate reduction cannot be ruled out. Sulfate at the time of nitrate addition was
0.99 mg/L, rose to 4.0 mg/L at depletion, and fell to 0.89 mg/L at turnover in early October. This
showed sulfide is much more of a sink for nitrate in natural system than organic carbon and plays
a critical role in determining the dominant pathway of nitrate reduction.
111 4.4.5. Effect of Sodium Aluminate in Chamber Water
There is growing interest in the use of sorbents for the in situ treatment of Hgcontaminated sediment. In situ sorbent amendments have attracted recent attention as a low-cost,
low-impact approach for remediation of organic-contaminated sediments (Ghosh et al. 2011).
Conventional remedial methods for Hg-contaminated aquatic sediment, including dredging,
capping and monitoring natural recovery, have shortcomings (Ghosh et al. 2011, Wang et al.
2004).
Dredging and disposal are expensive and inappropriate for large areas of contamination.
Capping with clean material can inadvertently enhance MeHg production by promoting
conditions that accelerate methylation (Johnson et al. 2010). Recent studies have focused on
sorbent amendments such as activated carbon and biochar (Gilmour et al. 2013, Ghosh et al.
2011). These amendments act as a sorption barrier to the efflux of hydrophobic organic
compounds (e.g., PCBs, dioxins, pesticides) from surficial sediment, thereby repressing uptake
of pollutants into the aquatic food web. Over time the barrier is mixed into sediments via
bioturbation and/or covered with clean sediment via deposition. A recent study by Gilmore et al.
(2013) showed activated carbon amendments to Hg-contaminated sediments were highly
effective in reducing methyl Hg bioaccumulation by benthic invertebrates. In these studies, the
effectiveness of activated carbon and other amendments in reducing invertebrate
bioaccumulation was well-correlated with their effectiveness in decreasing pore water MeHg
concentrations, and increasing sediment: water partition coefficients of MeHg. The mechanism
of activated carbon remediation appears to be mainly by reductions in MeHg bioavailability to
112 worms, rather than by reductions in sediment MeHg production or bulk sediment methyl Hg
concentration.
In our experimental chamber study, we evaluated sodium aluminate, a potential sorption
sink for MeHg. Aluminum treatment via alum (aluminum sulfate) and/or sodium aluminate is a
common lake management strategy to control internal P cycling (Cooke et al., 1993). Aluminum
addition results in a floc of aluminum hydroxide overlaying sediment, which is a highly effective
technique to reduce internal loading of phosphate in both stratified and unstratified lakes and
reservoirs (Welch and Cooke, 1999; Rydin et al. 2000). As noted by a number of authors, native
aluminum and manganese oxides have been implicated in the retention of MeHg in surficial
sediments (Chadwick et al., 2006; Merritt and Amirbahman, 2008). However, in our study
aluminum addition did not impede sediment release of MeHg efflux rates (Fig. 4-6). MeHg
efflux rates (> 12 ng/m2/d) appeared to be as high as in anaerobic conditions. The question arises
as to why aluminum addition did not repress MeHg release. Presumably, strong ligands such as
sulfide and DOC have an effect on the effectiveness of aluminum as a potential sorption to
suppress MeHg release. The binding of DOC and sulfide with ionic Hg and MeHg has a major
effect on Hg cycling in surface waters (Ullrich et al., 2001). As with inorganic Hg (II), MeHg
can form highly stable complexes with sulfide (CH3HgS-) and other ligand (DOC), depending on
the concentration of various ligands (More et al., 1998; Ullrich et al., 2001). Sulfide
complexation dominated MeHg speciation in anoxic conditions (Han et al., 2007; Zhan et al,
2004). In our laboratory experiment, we observed strong sulfide odor in our chambers during
initial and medium points during incubation. Apparently, the availability of sulfide may have
controlled metal distribution and solubility. Perhaps in our systems, strong ligand like sulfide
may have repressed the effectiveness of aluminum floc. Other studies have reported the binding
113 of MeHg to other ligand such as DOC may be affected by the presence of other cationic metals.
Driscoll et al. (1995) found that MeHg levels in fish were positively correlated with
concentration of monomeric aluminum and suggested that this was because aluminum competes
with MeHg for binding sites on DOC, thereby increasing the concentration of unbound MeHg.
Our results suggest that aluminum hydroxide, whether it comes from aluminum sulfate
(alum) or sodium aluminate, does not impede MeHg efflux. In fact, addition of alum could
enhance methylation due to the addition of sulfate. A number of studies have reported addition of
sulfate to anoxic sediment above intact sediment cores resulted in increased production of MeHg
by SRB activity (Gilmour et al., 1992; Gilmour and Henry, 1991). SRB are major producers of
MeHg in anoxic sediments and typically inhabit the anoxic zones of sediment where sulfate is
abundant (Gilmour and Henry, 1991). SRB are obligate anaerobes that obtain energy for growth
by oxidation of organic substrates. They use sulfate as the terminal electron acceptor and
consequently convert sulfate to sulfide (Harmon et al., 2007). Other studies have investigated
sulfate limitations in fluvial sediment and wetland systems by testing whether sulfate addition
alone can stimulate Hg methylation (Braunfireun et al. 1999; Gilmour et al. 1992; Jeremiason et
al. 2006; Harmon, 2004). Since SRB are the primary methylators of Hg in many environmental
systems, sulfate addition to an anoxic system starved for sulfate often increases SRB activity,
thereby enhancing MeHg production (Benoit et al., 2003). In addition, excess sulfate can also
trigger SRB to produce hydrogen sulfide, which can make water undesirable as a raw water
source for drinking water (Levine et al., 2004) and can be highly toxic to aquatic biota
(Reiffenstein et al., 1992).
In some cases, there could be benefit of using a combination of alum/sodium aluminate
instead of alum with buffer, since less sulfate would be available to enhance MeHg production.
114 This is because in the alum/sodium aluminate application, some of the aluminum is added with
sodium rather than sulfate to stimulate methylation. A number of studies have reported a limited
availability of sulfate can repress MeHg production at the sediment-water interface because SRB
are the dominant methylators of inorganic Hg in freshwater sediments (Benoit et al., 2003;
Ullrich et al., 2001; Morel et al., 1998; Gilmour et al., 1992). In this scenario, SRB are unable to
use sulfate as the terminal electron acceptor during the oxidation of organic matter to produce
hydrogen sulfide, reinforcing the idea that the activity SRB, and consequently the production of
MeHg, might be inhibited by adequate limitation of sulfate.
4.4.6. Implications for Lake Management
Best management practices that address MeHg production in Hg-contaminated
freshwater and marine sediments need to be implemented toward sustainable remediation
alternatives. The input of Hg into the environment and its distribution, behavior and fate in the
different environmental compartments need to be further investigated (Ullrich et al., 2001).
Furthermore, a detailed understanding of the processes of bioaccumulation and biomagnification
of MeHg in the food chain that lead to different levels of MeHg in different fish species and
different areas is crucial to gaining a fuller picture of the whole Hg pathway (Benoit et al., 2003).
Therefore, it is important to understand Hg chemistry and the corresponding health risks to the
entire ecosystem. In this manner, MeHg contamination can be pinpointed and addressed so that a
given remediation project can be both effective in reducing health risks and feasible in an
economic and technological sense. In Guadalupe Reservoir and Almaden Lake, remediation of
MeHg within watersheds is currently being addressed as a major priority in water quality
management, but there are several legal and technical obstacles to Hg clean up.
115 This experimental study can advance lake management practices with respect to the
MeHg efflux from sediments. One approach is to increase oxygen in anoxic bottom waters of
lakes to inhibit MeHg production. Hypolimnetic oxygenation is a relatively new management
strategy that uses pure oxygen gas as oxygen source to release soluble oxygen into the
hypolimnion during summer stratification to maintain a well-oxygenated condition in bottom
waters while maintaining thermal stratification. Hypolimnetic oxygenation has the potential to
decrease sediment release of MeHg, as suggested by Beutel and Horne (1999). In addition,
oxygenation systems have successfully oxygenated large multi-purpose reservoirs (Speece,
1994), moderately sized drinking water reservoirs (Jung et al., 2003), and relatively small natural
lakes (Moore et al., 1996). Higher oxygen levels at the sediment-water interface could result in
lower bioaccumulation in benthic biota by two key mechanisms: (1) deeper oxygen penetration
into the sediments could push the site of methylation by SRB downwards and away from the
sediment-water interface where benthic biota primarily reside; and (2) deeper oxygen penetration
can maintain an well-oxygenated surficial sediment layer rich in metal oxides which may act as a
sorption sink for ionic Hg and MeHg, thereby making ionic Hg less bioavailable to get
methylated while limiting the diffusion of MeHg out of the sediment and into overlaying water.
Nitrate addition is another attractive management alternative for control of MeHg
mobilization as shown in this study. It may represent a viable remediation alternative to
dredging, capping, and oxygenation at Hg-contaminated sites (Matthews et al., 2013). For
instance, nitrate can inhibit sulfur-reducing bacteria and MeHg efflux from sediment, without
providing an oxygen-rich environment for biota. In this experimental study, nitrate addition
repressed MeHg efflux rates even lower compared to oxygenated chamber water. Todorova et al.
(2009) reported the maintenance of nitrate at concentrations greater than 1 mg-N/L in the water
116 column overlying the sediments was sufficient to prevent mobilization of MeHg in Onondaga
Lake. Similar dynamics were observed at Round Lake, Minnesota, where the addition of nitrate
poised the oxidation-reduction potential and repressed MeHg accumulation in bottom waters
(Austin, 2013). Once nitrate was exhausted, MeHg concentrations rapidly returned to pretreatment levels. A recent study by Mathews et al. (2013) also found in anaerobic bottom waters
of Onondaga Lake, nitrate addition inhibited MeHg efflux rates. The reuse of nitrate in
wastewater to improve surface water quality is compelling because it lowers treatment costs and
complexity while enhancing environmental quality. The case study of Occoquan Reservoir,
described in Chapter 2 and 5, was a good example of such an application. Because the added
nitrate is mainly converted to harmless dinitrogen gas via biological denitrification, negative
impacts related to eutrophication should be minimal if nitrate addition is properly managed. In
addition, nitrate addition to freshwater bodies in which P loading is low may not substantially
enhance productivity. To ensure minimal impacts, N could be added at stoichiometric amounts
relative to environmental levels of P (N:P < 10:1) so that little or no residual N is available to
stimulate phytoplankton growth.
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129 Figure 4-1.Map of the Guadalupe River watershed in San Jose, California. Our study sites
include Guadalupe Reservoir in the upper center of the watershed, and Almaden Lake which is
downstream of the reservoir. Figure is from the Draft Final Conceptual Model Report, June 2004
by Tetra Tech, Inc.
130 Figure 4-2. Profiles of temperature, dissolved oxygen (DO), and methylmercury (MeHg). A.
Lake Almaden sampled September 13, 2011. B. Guadalupe Reservoir sampled August 16, 2011.
131 Figure 4-3. 2008 dissolved oxygen (DO), nitrate and methylmercury (MeHg) concentrations in
bottom water near the sediment-water interface of Lake Almaden in 2008. Note high levels of
MeHg in bottom water are associated with depletion of both DO and nitrate.
132 Figure 4-4. Total mercury and sulfate efflux from Lake Almaden (left) and Guadalupe Reservoir
(right) to aerobic conditions (clear bars) and anaerobic conditions (black bars). Error bars are one
standard deviation (n = 4).
133 Figure 4-5. Cumulative mass of methylmercury, manganese and iron in chamber water from
anaerobic incubation of Lake Almaden sediment.
134 Figure 4-6. Efflux rates from Lake Almaden sediment for methylmercury (ng/m2·d), manganese
(mg/m2·d), iron (mg/m2·d), and ammonia (mg/m2·d) in anaerobic control and three experimental
treatments: aerobic, anoxic with nitrate addition, and anaerobic with sodium aluminate addition.
Error bars are one standard error (n = 3-4). Asterisk (*) signifies that efflux rate in experimental
treatment is significantly different from control (two-tailed t-test, p < 0.05).
135 Table 4-1. Sediment Quality in Lake Almaden and Guadalupe Reservoir
Site
Loss
on
Total Total Total
ignition, C,
N,
Hg,
%
%
%
µg/g
Total
Mn,
µg/g
Total
Fe,
µg/g
Total
P,
µg/g
Total
S,
µg/g
pH
Lake Almaden
Station 1
(12.5 m
deep)
Station 2
(10.5 m
deep)
Station 3
(8 m deep)
Station 5
(8 m deep)
84.2
± 0.1
2.4
0.27
7.6
± 0.6
85.3
±0.1
nm
nm
13.7
± 0.2
nm
nm
nm
nm
5.9
0.70
83.1
± 0.3
87.8
± 0.1
7.8
± 1.7
34.6
± 6.2
1,100 51,000
840
2,300 7.3
nm
nm
nm
nm
nm
nm
nm
nm
nm
nm
nm
nm
nm
nm
nm
Guadalupe Reservoir
Deep
Station
(15 m deep)
90.8
± 0.3
6.3
± 0.3
1,100 40,000 1,000 7,000 7.2
Values are average plus/minus standard deviation (n = 3-6); concentrations are µg/g dry
weight; nm = not measured.
136 Table 4-2. Fluxes measured in experimental chambers under aerobic and anaerobic conditions in
Lake Almaden and Guadalupe Reservoir.
Lake Almaden
Constituent
Guadalupe Reservoir
Aerobic
Anaerobic
Total Hg, ng/m2·d
-207 ± 98*
51 ± 35*
119 ± 110 45.9 ± 74.4
Methyl Hg, ng/m2·d
-2.3 ± 0.8†
11 ± 4.1†
5.5 ± 2.2* 22.3 ± 11.3*
Total Manganese, mg/m2·d
-1.1 ± 0.9* 17.7 ± 6.7*
-4.2 ± 0.5* 20.6 ± 7.0*
Total Iron, mg/m2·d
-2.2 ± 0.9
6.7 ± 5.6
2.5 ± 0.5*
Nitrate, mg-N/m2·d
8.3 ± 3.0*
-2.4 ± 1.6*
12.7 ± 3.2* -2.6 ± 1.2*
Ammonia, mg-N/m2·d
Dissolved Orthophosphate, mgP/m2·d
14.9 ± 13.9 29.0 ± 4.1
-1.5 ± 5.2 16.6 ± 1.5
Sulfate, mg/m2·d
-90.3 ± 53.1 -167 ± 17.6
1.0 ± 0.4*
17.6 ± 5.0*
Aerobic
0.3 ± 0.1*
Anaerobic
4.1 ± 2.3*
5.1 ± 1.1*
66.9 ± 22* -56.3 ± 37.3*
Values are mean plus/minus standard deviation (n = 4).
*Aerobic and anaerobic fluxes are significantly different (two-tailed paired t-test, p < 0.05)
*Aerobic and anaerobic fluxes were nearly significantly different (two-tailed paired t-test, p =
0.07)
137 Table 4-3. Comparison of net flux of methylmercury rates to similar studies in freshwater and
marine systems.
Site
Almaden Lake, CA
Guadalupe Res., CA
Deer Lake, WA
Camp Far West, CA
Boston Harbor, MA
Lahontan Res., CA
Little Rock Lake, WI
Mary Lake, WI
NY/NJ Harbor, NY
Thau Lagoon, France
San Francisco Bay, CA
Gulf of Trieste, Italy
Grado Lagoon, Italy
Net Flux (ng/m2/d)
-2.3 ~ 11
5.5 ~ 22.3
0.7 ~ 0.4
1.0 ~ 3.0
-0.9 ~ 41
-3.9 ~ 27.4
~ 50.0
~ 54.0
15~19
-32 ~ 68
-20 ~ 183
-12~ 238
34 ~ 498
138 Reference
This research
This research
Beutel et al., 2014
Kuwabara et al., 2003a
Benoit et al., 2009
Kuwabara et al., 2002
Watras et al., 1994
Watras et al., 1994
Hammerschmildt et al., 2008
Point et al., 2007
Choe et al., 2004
Covelli et al., 2008
Covelli et al., 2008
CHAPTER 5
Spatial and Temporal Patterns of Nitrate, Oxygen and Methylmercury during Summer
Stratification in Occoquan Reservoir, Virginia.
5.1. Introduction
Freshwater quality studies and management have become increasingly important over the
past decade (USEPA, 2000). As the global population continues to expand, the demand for
freshwater increases yearly by approximately 64 billion m3 (United Nations, 2010). As a result,
the global water consumption rate doubles every twenty years, a pace that is twice the rate of
population growth (World Water Organizations, 2010). Due to this increased stress on water
sources, water quality in lakes and reservoirs has become a chronic problem for some water
utilities and has forced many others to embrace best management practices as possible strategies
for improving raw water quality and treatability (Wetzel, 2001; USEPA, 1997). Consequently,
federal, state and local government agencies have developed plans and laws that protect and
preserve the quality of drinking water sources (USGS, 2005). However, USEPA (2004) reported
64% of the 16 million acres of lakes in the nation was found to be impaired and not suitable to
support the use for primary contact (aquatic life, fish consumption, and drinking water supply)
and secondary contact (swimming and recreation). Nutrients, sediment and metals, including
mercury (Hg), were listed as leading pollutants and caused dramatic impact on water quality.
A factor that contributes to poor water quality in productive, moderately deep lakes and
reservoirs is stratification in the summertime. During the period of thermal stratification, the
hypolimnion is essentially cut off from oxygen exchange with the atmosphere and is often too
139 dark for plants and algae to grow and produce oxygen by photosynthesis. As a result, bottom
water become anaerobic as the summer progresses. This occurs as supply of oxygen is consumed
by bacteria and other bottom-dwelling organisms (Beutel et al., 2008). As anaerobic conditions
at the sediment-water interface continue, bottom waters accumulate a range of compounds that
exacerbate eutrophication, are toxic to aquatic biota, and limit the potential for potable use.
A critical toxic compound that tends to be released from anaerobic sediments is
methylmercury (MeHg). MeHg production is primarily mediated by anaerobic bacteria in surface
sediments, particularly sulfate reducing bacteria (SRB) in organic-rich sediments. SRB are the
key Hg-methylating organisms in nature (King et al., 2001). SRB typically inhabits the anoxic
transition zones of sediment where sulfate is present, such as subsurface zones of lakes and
anoxic sediments (Gilmour and Henry, 1991). SRB are obligate anaerobes that obtain energy for
growth by oxidation of organic substrates. In many cases, they use sulfate as the terminal
electron acceptor and consequently convert sulfate to sulfide (Harmon et al., 2007). Methylation
of Hg occurs in SRB as an accidental co-process of sulfate reduction (Gilmour et al., 1992). The
bioavailability of ionic Hg to get methylated appears to be enhanced at moderate sulfide levels,
at which neutral Hg-S complexes are formed, which more easily diffuse across the polar cell
walls of SRB and into the cell where methylation occurs (Benoit et al., 2003).
Once released, MeHg can accumulate in the surface sediments since a concentration
gradient between sediment pore waters and overlaying waters develops, which drives the
diffusive flux of MeHg from the sediment to the overlaying water (Rolfhus et al., 2003). As a
result, MeHg, a potent neurotoxin, can quickly bioaccumulates in phytoplankton, which is
consumed by the next level in the food chain (Farina et al., 2011). The National Research
Council, in its 2010 report on the toxicological effects of MeHg, pointed out that the population
140 at highest risk is the offspring of women who consume large amounts of fish and seafood (NRC,
2010). MeHg affects the immune system, alters genetic and enzyme systems, and damages the
nervous system, including coordination and the senses of touch, taste, and sight (Clifton, 2007).
MeHg is particularly damaging to developing embryos, which are five to ten times more
sensitive than adults (Clarkson, 1997).
As redox potential continues to decrease, sediments release manganese (Mn (II)) and iron
(Fe (II)) via microbiological and chemical reduction of metal oxides (Davison, 1993). Iron and
manganese are chemically similar and cause similar problems. Elevated Fe and Mn in raw water
reservoirs complicate drinking water treatment and cause build up in pipelines, pressure tanks,
water heaters, and water softeners. This reduces the available quantity and pressure of the water
supply (Logsdon et al., 1999; USEPA, 2000). In addition, Fe and Mn accumulation become an
economic problem when water supply and water softening equipment must be replaced. There
also are associated increases in energy costs from pumping water through constricted pipes and
heating water with electric heating rods coated with iron and manganese mineral deposits
(Gaskill, 1996; O’Connor, 1971). Although Fe and Mn have been considered non-hazardous,
excessive concentrations in water supply can cause several health problems. Kawamura et al.
(1981) reported health effects resulting from the ingestion of Fe and Mn-contaminated well
water for an estimated 2-3 months by 25 individuals. Results showed heath effects included
lethargy, increased muscle tonus, tremor and mental disturbances.
The goal of this study was to perform an evaluation of spatial and temporal patterns of
nitrate and oxygen on MeHg, as well as the redox sensitive metals Fe and Mn, during the
summer stratification period in Occoquan Reservoir, VA. My hypothesis was that MeHg would
only be observed in substantial levels in bottom waters if oxygen and nitrate were not present at
141 the sediment-water interface, that is, when or where water conditions shift from anoxic (no
oxygen but some nitrate) to anaerobic (no oxygen or nitrate). The rationale for this hypothesis
was that the production of MeHg is usually associated with anoxic conditions and the
enhancement of the activity of SRB. These bacteria are important Hg methylators in sediments
and are regulated by the flux of organic matter to the sediment and the availability of oxygen or
nitrate for oxidation of organic matter. In the absence of oxygen or nitrate, SRB produce
hydrogen sulfide, which facilitates the methylation process (Benoit et al., 2003). Our study
evaluated both nitrate and oxygen, and more specifically how MeHg accumulation correlates
with the accumulation of other redox-sensitive metals such as Fe ad Mn. While previous studies
have focus on the effect of oxygen addition on hypolimnetic accumulation of nutrients and
conventional metals, this study is novel in that it also evaluated the fate of MeHg, a potent and
widespread bioaccumulatory toxin in aquatic ecosystems.
5.2. Materials and Methods
5.2.1. Study Site
The Occoquan Reservoir is a large, run-of-the-river reservoir (volume = 31.4 x 106 m3;
maximum depth = 19.8 m; mean hydraulic retention time = 20 d) located in northern Virginia,
southwest of Washington D.C. (Fig.5-1) (Randall and Grizzard, 1995). The reservoir ultimately
discharges into the Potomac River, a major tributary to the Chesapeake Bay. The reservoir,
which serves as a water source for 2 million people, has a history of excessive nutrient loading
from wastewater discharges and storm runoff. In the late 1960s, massive cyanobacteria blooms
led to frequent taste and odor problems, depletion of oxygen in bottom waters, and complications
142 and increased costs to treat reservoir water for potable use. To address water quality concerns,
regional authorities evaluated a range of wastewater management options including a
moratorium on new sewage connections and the export of wastewater from the basin. Seeing the
value in utilizing wastewater as a resource, managers ultimately developed the "Occoquan
Policy" – a farsighted program of indirect potable reuse via advanced wastewater reclamation
practices. The Upper Occoquan Service Authority (UOSA) discharges 360 x 106 m3/yr of highly
treated, low nutrient, reclaimed water to one of the two main tributaries of the reservoir. During
the summer, treatment is modified so that the discharge is rich in nitrate. Much of the added
nitrate (approximately 80%) flows with the colder stream water to the profundal zone of the
reservoir due to density gradients that form at the reservoir inlet. Typical nitrate levels are > 10
mg-N/L in the discharge and < 2 mg/L in bottom waters near the reservoir outlet 26 km
downstream. Fairfax Water, the drinking water authority, withdraws water near the outlet of the
reservoir and treats it for potable use.
The Occoquan Reservoir water reuse program, combined with UOSA seasonal nitraterich product water, has the multiple benefits (Cubas 2012, Cubas et al. 2014). Treatment costs
are reduced by avoiding a denitrification treatment step during the late spring and summer. The
water supply yield of the reservoir has also increased. From a water quality perspective, nitrate
addition has reduced internal loading of P and other reduced compounds from sediments. In the
reservoir, nitrate concentrations greater than 1 mg-N/L above the sediment-water interface
prevented the release of phosphate from sediments during periods of hypolimnetic anoxia (Fig.52). When nitrate concentration in water overlaying sediment was higher than 1 mg-N/L,
phosphate seldom exceeded 0.03 mg-P/L. At the same nitrate levels, sediment release of total
organic carbon was also reduced. Organic carbon is a precursor to the formation of cancer-
143 causing disinfection byproducts during drinking water treatment. Both field and experimental
sediment-water interface chamber work showed that nitrate in the bottom layers of the reservoir
also prevented sediment release of Fe and Mn and delayed ammonia release. As detailed in
Cubas et al. (2014), decreasing the nitrate load from UOSA to comply with new stringent
regional N regulations could inadvertently result in enhanced internal release of undesirable
reduced compounds into reservoir waters, thereby impacting water quality in the reservoir and in
downstream waters.
5.2.2. Collection of Field Sampling
This study included two summers of sampling events. During the first summer (2012),
water samples were collected once a month from July through mid-October at 2-m intervals from
stations RE02 (near the dam), RE05, RE10, RE15, RE20 (confluence of Bull Run and Occoquan
Creek), RE30 (upstream shallow arm, nitrate-rich effluent), and RE 35 (downstream shallow
arm, nitrate-poor effluent) (Fig.5-1).During the second summer (2013), water samples were
collected every ten days on average from June through September at 1-m intervals from stations
RE20, RE30, and RE 35. For both summer sampling events, vertical profiles of dissolved oxygen
and temperature were measured by taking reading in the water column at 1-m intervals using a
YSI 600 XL Sonde probe. Nitrate profiles were measured at 1-m vertical intervals with a
Satlantic SUNA optical nitrate probe. Surface and bottom samples were collected and analyzed
for nitrate in accordance with Standard Methods (APHA, 2005). For summer 2012, MeHg, Fe
and Mn samples were collected at 2-m intervals. However, summer 2013, these samples were
collected at 1-m intervals. A Teflon Kemmerer sampler was used for sample collection.
144 5.2.3. Water Quality Analysis
MeHg samples were stored in acid washed 250 ml amber glass bottles with Teflon lined
caps. These samples were preserved with trace metal grade hydrochloric acid then were stored
4°C until analysis. MeHg were analyzed using a MERX automated modular Hg system (Brooks
Rand Labs, Seattle, WA, USA) using cold vapor atomic fluorescence spectroscopy (CVAFS)
based on USEPA method 1630 for MeHg (USEPA, 2001). The method detection limit was 0.02
ng/L for MeHg. Standard quality control procedure were followed including calibration blanks,
matrix spikes samples (acceptable range of 70-130 %), and ongoing precision recovery samples
(acceptable range of 75-125%).
Fe and Mn samples were stored in pre-washed 60 ml wide neck bottles from Thermo
Scientific Nalgene. These samples were preserved with nitric acid and analyzed using
inductively coupled plasma mass spectrometry (ICP-MS) (American Public Health Association,
2005). Detection limits were 0.001 mg/L for Mn and 0.05 mg/l for Fe.
5.3. Results
5.3.1. Water Quality
Occoquan Reservoir exhibited spatial and temporal patterns of water quality indicative of
summer thermal stratification and oxygen consumption in the hypolimnion (Fig.5-2). In 2012
and 2013, DO was < 1 mg/l near the sediment-water interface during much of the summer.
However, nitrate varied according to the outlet of the reservoir and ranged from (0-8 mg/l). DO
levels tended to drop as the stratified period progressed. While anaerobic, the hypolimnetic water
column of the reservoir accumulated MeHg, Mn and Fe during the stratified period.
145 In 2012, a peak of MeHg was observed near the sediment-water interface at station RE35, the upstream tributary that does not receive nitrate rich discharge (Fig.5-1), early in the
summer. The mean hypolimnetic MeHg concentration was > 1 ng/L where oxygen was depleted
(DO < 0.3 mg/L), but MeHg levels were 0.04-0.12 ng/L in the latter part of the stratified period
in conjunction with increased oxygen levels (DO < 5 mg/L) (Fig.5-3). Mn and Fe tended to be
fairly well correlated with MeHg accumulation throughout the hypolimnion, though mean peak
levels were observed near the sediment-water interface. Hypolimnetic concentrations ranged
from 0.15-1.2 mg/L for Mn and 0.5 -2.2 mg/L for Fe (Fig.5-3). Obviously, nitrate concentration
was low throughout the hypolimnetic water column of the reservoir and ranged from 0 – 0.6 mgN/L) (Fig.5-3).
At Station RE-30, the upstream tributary that does receive nitrate-rich discharge (Fig.51), mean hypolimnetic MeHg concentration ranged from 0.02-0.1 ng/L, and was lower compared
to R35; it also showed little variation. The hypolimnion accumulated little Mn and Fe and
showed the same trend with MeHg. Values peaked near the sediment-water interface at 0.05-1.2
mg/L for Mn and 0.2-0.4 mg/L for Fe (Fig.5-3). DO and nitrate tended to be higher, and MeHg
tended to be lower, at RE-30 relative to RE-35. Mean hypolimnetic DO ranged from 0.4-9 mg/L
and 4-6 mg/L for nitrate (Fig.5-3). Regarding mean hypolimnetic concentrations of Mn and Fe,
the only difference was that Mn levels were statistically significant compared to Fe. Mean
hypolimnetic Mn was, on average, three times higher than Fe.
At Station RE-20, downstream of the confluence of the two tributaries in the upper
portion of the reservoir (Fig.5-1), MeHg was consistently lower and showed little to no variation
at the sediment-water interface. Mean hypolimnetic MeHg ranged from non-detect to 0.09 ng/L
with peak value early during the stratified period (Fig.5-3). In mid-June, while DO and nitrate
146 were near zero, the hypolimnion accumulated Mn and Fe. Hypolimnetic concentrations ranged
from 0.07-3.5 mg/L for Mn and 0.2-1.3 mg/L for Fe (Fig.5-3). Peak concentrations of Mn and Fe
displayed similar patterns as the mean hypolimnetic concentrations. DO and nitrate started lower
but increased dramatically early August. Mean hypolimnetic concentrations ranged from 0.2-8
mg/L for DO and 0.1-4 mg/L for nitrate (Fig.5-3). Similar trends were observed for DO and
nitrate.
At Station RE-15, downstream of the reservoir near segment 3 (Fig.5-1), mean
hypolimnetic MeHg concentrations ranged from non-detect to 0.3 ng/L, but levels consistently
dropped in the latter part of the stratified period coincidental with a substantial increase of DO
(0.2-6 mg/L) and slight increases of nitrate (0.06-2 mg/L) (Fig.5-4). Mn and Fe showed similar
trends with MeHg accumulation throughout the hypolimnion, though peak levels were observed
near the sediment-water interface. Hypolimnetic concentrations ranged from 0.06-5 mg/L for Mn
and 0.1-3 mg/L for Fe (Fig.5-4). Peak hypolimnetic concentration of Mn for RE-15 was, on
average, the highest compared to other stations throughout the reservoir.
At Station RE-10, downstream of the reservoir near segment 2 (Fig.5-1), mean
hypolimnetic concentrations of MeHg ranged from 0.02-0.2 ng/L with a peak above the sediment
of 0.3 ng/L (Fig.5-4). Mn and Fe, which correlated to MeHg, showed a similar seasonal trend.
Values peaked near the sediment-water interface at 0.08-1.5 mg/L for Mn and 0.05-1.6 mg /L for
Fe (Fig.5-4). DO and nitrate exhibited similar seasonal patterns and showed little variation at the
end of the stratified period. Mean values ranged from 0.1-5 mg/L for DO and 0-1.2 mg/L for
nitrate (Fig.5-4.).
At Station RE-05, downstream of the reservoir near the Dam (Fig.5-1), accumulation of
MeHg showed no variation for the month of June and July, but levels decreased in the
147 hypolimnion as thermal stratification progressed. Mean hypolimnetic MeHg ranged from nondetect to 0.16 ng/L (Fig.5-4). Regarding mean hypolimnetic concentration of Mn and Fe, the
only difference was that Mn levels were statistically significant relative to Fe at the sediment
water interface. Mean values ranged from 0.4-2.3 mg/L for Mn and 0.2-1.4 mg/L for Fe. DO and
nitrate displayed similar trends and mean hypolimnetic concentrations ranged from 0.1-5 mg/L
for DO and 0-1.4 mg/L for nitrate (Fig.5-4).
At Station RE-02, downstream of the reservoir near the Dam (Fig.5-1), mean
hypolimnetic concentrations of MeHg ranged from non-detect to 0.2 ng/L with peak values near
the sediment-water interface (Fig.5-4). Mn and Fe, which correlated with MeHg, showed similar
patterns. Fe peak concentration was lower compared to Mn. Values peaked near the sedimentwater interface at 0.3-1.2 mg/L for Mn and 0.2-0.8 mg /L for Fe (Fig.5-4). DO and nitrate
exhibited similar seasonal patterns and showed little variation at the beginning and end of the
stratified period. Mean values ranged from 0.1-4 mg/L for DO and 0-1.5 mg/L for nitrate (Fig.54).
In 2013, Station RE-35, the upstream tributary that does not receive nitrate rich discharge
(Fig.5-1), peak hypolimnetic MeHg was observed between July and September near the
sediment-water interface. Mean hypolimnetic MeHg was > 0.5 ng/L where oxygen was depleted
(DO < 0.6 mg/L), but levels were 0.03-0.1 ng/L in the latter part of the stratified period when DO
increased (DO < 6 mg/L) (Fig.5-5). Clearly, nitrate accumulation was near zero throughout the
hypolimnetic water column of the reservoir. Mn and Fe tended to be fairly well correlated with
MeHg accumulation throughout the hypolimnion, though mean peak levels were observed near
the sediment-water interface. Hypolimnetic concentrations ranged from 0.2-2.7 mg/L for Mn and
0.5-2.3 mg/L for Fe (Fig.5-5).
148 At Station RE-30, the upstream tributary that does receive nitrate rich discharge (Fig.51), mean hypolimnetic for MeHg was ranged from (0.02 – 0.08 ng/L) much lower compared to
R-35 and showed little variation. The hypolimnion accumulated little Mn and Fe and showed the
same trend with MeHg. Values peaked near the sediment-water interface at 0.04 – 0.6 mg/L for
Mn and 0.1 - 0.7 mg/L for Fe (Fig.5-5). DO and nitrate tended to be higher and MeHg tended to
be lower in RE-30 relative to RE-35. Mean hypolimnetic DO was ranged from 2 to 8 mg/L and 2
to 9 mg/L for nitrate (Fig.5-5). Regarding peak mean hypolimnetic concentrations of MeHg
between RE-30 and RE-35, the only difference was that MeHg levels in RE-35 were statistically
significant compared to RE-30. Peak mean hypolimnetic MeHg was, on average, six times
higher in RE-35 relative to RE-30.
At Station RE-20, downstream of the confluence of the two tributaries in the upper
portion of the reservoir (Fig.5-1), MeHg was consistently lower near the sediment-water
interface compared to RE-35 and showed little variation. Mean hypolimnetic concentrations
ranged from 0.02-0.2 ng/L with peak value around mid-August (Fig.5-5). In mid-June, while DO
and nitrate were near zero, the hypolimnion accumulated Mn and Fe. Hypolimnetic
concentrations ranged from 0.2-2 mg/L for Mn and 0.2-3 mg/L for Fe (Fig.5-5). Peak
concentrations of Mn and Fe displayed similar patterns. Similar trends were also observed for
DO and nitrate throughout the hypolimnion. Mean hypolimnetic concentrations ranged from 0.212 mg/L for DO and 0.07-7 mg/L for nitrate (Fig.5-5).
5.3.2 Correlated Variation between Methylmercury, Oxygen, Nitrate, Manganese and Iron
Correlations were observed between DO, nitrate and MeHg in 2013 from stations RE-35,
RE-30 and RE-20 (Fig.5-6). At station RE-35, where nitrate and DO were consistently near zero
149 throughout the hypolimnetic water column of the reservoir, MeHg accumulated up to 0.75 ng/L.
However, at the end of the stratified period, DO increased rapidly to 6 mg/L and MeHg
decreased to the low levels (0.03 ng/L), even though nitrate was not present. In contrast to RE30, where nitrate was moderately present near the sediment-water interface and DO was low,
MeHg showed little to no accumulation with a peak concentration of 0.07 ng/L. Interestingly, no
MeHg release was observed in the presence of nitrate or DO (Fig.5- 6), suggesting that DO and
nitrate are interrelated. Similar observation continued in RE-20. Nitrate and DO levels above the
sediment were higher relative to RE-35 but lower than RE-30. MeHg peak accumulation was
between RE-30 and RE-35 [(RE-30) 0.07 ng/L < (RE-20) 0.23 ng/L < (RE-35) 0.7 5ng/L)]
(Fig.5-6). MeHg were observed in bottom waters in mid-July, and these levels increased to
around 0.23 ng/L by August while DO and nitrate was very low.
The distribution of MeHg was correlated closely with the distribution Mn and Fe in all
three stations (Fig.5-7). In RE-35, concurrent peaks of mean hypolimnetic concentrations of Mn
(3 mg/L) and Fe (2.2 mg/L) were observed at the same time of MeHg (0.75 ng/L). Clearly,
MeHg tended to display comparable spatial patterns with Mn and Fe. Similar observation
continued in RE-30 and RE-20 (Fig.5-7).
5.4. Discussion
5.4.1 Comparison between Stations RE-30 and RE-35, 2012 and 2013
MeHg production in bottom waters is primarily mediated by anaerobic bacteria. The
efficiency of this microbial process is dependent on the activity of the methylating bacteria and
the bioavailability of Hg(II) associated with sulfides and dissolved organic matter (Benoit et al.,
150 2003; Ullrich et al., 2001). Microbial respiration in bottom waters combined with thermal
stratification can lead to elevated levels of hydrogen sulfide (Benoit et al., 1999). The sulfide is
produced by sulfate reducing bacteria, which have also been shown to be the principal
methylators of Hg to MeHg (Gilmour et al., 1992). Surface water concentration of MeHg is
generally much lower than in anoxic hypolimnetic water, and since water discharged from most
lakes is surface water, it is also generally low in MeHg. However, water discharged from
reservoirs frequently comes from lower in the water column. In some cases, this includes water
coming directly off the bottom where there is a greater potential for MeHg to exist. The MeHg
enriched discharge water can then be transported downstream where it would be bioavailable to
accumulate in the aquatic food web below the reservoir (Canavan et al., 2000).
In the case of Occoquan Reservoir, discharges of nitrate rich water resulted in decreased
buildup of MeHg in bottom waters in station RE-30 (the upstream tributary that does receive
nitrate rich discharge) relative to RE-35 (the upstream tributary that does not receive nitrate rich
discharge). This study provided a rare opportunity to further our understanding of Hg cycling in
a reservoir and advance management and remediation to improve water quality. Based on the
two-year data set, mean hypolimnetic MeHg was lower in RE-30. There was also a modest
negative correlation between mean hypolimnetic MeHg and DO/nitrate (2013) in RE-30 (Fig.56), reinforcing the contention that higher DO or the presence of nitrate resulted in lower MeHg
buildup. Also supporting this assessment was the observation that years with higher DO or the
presence of nitrate in RE-30 (2012 and 2013) had lower MeHg. Mean hypolimnetic nitrate in
2012 and 2013 on average remained above 5 mg/L; mean hypolimnetic MeHg remained below
0.1 ng/L in 2012 and 0.08 ng/L in 2013. In 2012, mean hypolimnetic DO stayed below 3 mg/L
until the end of the stratified period when it increased to 8 mg/L compared to 2013, when it
151 remained on average above 5 mg/L. Mean hypolimnetic MeHg did not accumulate even though
DO was low at the beginning at the stratification period due to the presence of nitrate. In contrast
to RE-35, mean hypolimnetic nitrate (2012 and 2013) remained below 0.6 mg/L and mean
hypolimnetic MeHg showed a steady buildup to 1.10 ng/L in 2012 and 0.75 ng/L in 2013.
Hypolimnetic buildup of MeHg was clearly associated with anoxia and the absence of nitrate. In
both years, mean hypolimnetic MeHg began to increase once hypolimnetic mean DO dropped
near zero, and MeHg tended to increase through the stratified period (Fig.5-3; Fig.5-5). A highly
significant negative correlation was observed between mean hypolimnetic MeHg and nitrate plus
DO for 2013 data set (Fig.5-6). In RE-35, elevated levels of MeHg corresponded with nitrate
and DO below 1 mg/L on a mean hypolimnetic basis. In addition, mean hypolimnetic MeHg
concentrations were typically higher in RE-35 compared to RE-30.
This is one the few studies to examine the effects of nitrate and DO on the hypolimnion
Hg concentrations in a reservoir. Our observation confirmed that the presence of nitrate or DO
repressed MeHg accumulation during extended period of hypoxia. This observation correlated
well with reported relationships between nitrate/DO and MeHg release at the sediment-water
interface (Matthews et al., 2013; Austin et al. 2015; Todorova et al., 2009; Effler et al., 2008;
Eckley et al., 2006; Benoit et al., 2006; Golterman, 2001). Similar dynamics were observed by
Austin et al. (2015). The study reported the effectiveness of nitrate addition to urban lakes in
Saint Paul/Minneapolis, Minnesota, USA in an effort to lower Hg bioaccumulation. A wholelake pilot study consisting of a slug dose of liquid calcium nitrate in Round Lake, Mn (area= 13
ha; maximum depth = 11 m) was conducted in 2010. Nitrate in bottom water peaked at 4 mg-N/L
then dropped to zero over a 2-month period. Bottom water MeHg remained steady at 0.3 ng/L
152 when nitrate was present, then rapidly increased to 1 ng/L once nitrate disappeared (Austin et al.,
2015).
Matthews et al. (2013) reported results from a whole-lake pilot test to assess the effect of
nitrate addition on mobilization of MeHg from the sediments of Onondaga Lake (area= 1 km2;
maximum depth = 20 m), a Hg-contaminated urban lake in New York, USA. Eighty-four
thousand kilograms of nitrate-nitrogen was added as liquid calcium nitrate to bottom waters from
a barge during thermal stratification. Bottom waters were maintained at a nitrate concentration of
2-4 mg-N/L and MeHg, which typically ranged from 2 to 6 ng/L in prior years, was negligible.
The probable regulating mechanism was enhanced sorption of MeHg to Mn and Fe in surficial
sediment. Recent incubations with sediment from Almaden lake in San Jose, California
(maximum depth of 12.5 m; surface area of 3.2 ha), a highly Hg polluted lake from gold mining,
showed that while both nitrate and oxygen addition repressed MeHg efflux, nitrate was even
more effective than oxygen. Results showed that in the nitrate added chamber, MeHg efflux
decreased to 0.2 ng/m2/d compared to 0.5 ng/m2/d in the oxic chamber (Duvil et al., 2013). A
potential benefit of nitrate addition compared to oxygenation is exclusion of aerobic
macroinvertebrates from polluted profundal sediment (Beutel et al., 2014). Effler and Matthews
(2008) suggested that macroinvertebrates could enhance MeHg efflux from Hg-contaminated
sediment as a result of bioirrigation or could act as a vector for Hg uptake into the aquatic food
web. This potential benefit is likely less of an issue in oxygenation of non-Hg contaminated lakes
and reservoirs like Occoquan Reservoir. On the contrary, oxygen addition is more effective than
nitrate in repressing Mn buildup in bottom waters, a compound that can greatly complicate
potable water treatment (Bryant et al., 2011; Betancourt et al., 2010) while repressing MeHg
release at the sediment-water interface.
153 The magnitude of MeHg release also correlated closely with the distribution Mn and Fe
in both stations near the sediment-water interface (2012 and 2013) (Fig.5-4; Fig.5-7). Mean
hypolimnetic Mn and Fe was lower in RE-30 compared to RE-35 in 2012 and 2013. In both 2012
and 2013, peak mean hypolimnetic concentrations of Mn and Fe were observed at the same time
as MeHg peaks, suggesting that the reduction Mn and Fe oxides were associated with MeHg
accumulation and reductive mobilization occurred at the sediment-water interface. Observations
of simultaneous increase in hypolimnetic concentrations of MeHg, Fe, and Mn under anoxic
conditions suggest that mobilization of MeHg is linked to the dissolution of Fe and Mn
oxyhydroxides in surficial sediments (Jacobs et al., 1995; Regnell et al., 2001; Chadwick et al.,
2006). As observed in RE-35, there was a rapid and synchronous increase in Fe and MeHg
following DO depletion and the absence of nitrate. On the contrary, in RE-30, Fe and MeHg
remained low and showed little to no variation, suggesting the presence of nitrate inhibited the
reduction of Fe and Mn at the sediment-water interface, hence repressing MeHg release.
Todorova et al. (2009) also reported nitrate control MeHg accumulation in Onondaga Lake may
result from direct inhibition of dissolution of hydrous ferric and manganese oxides from anoxic
sediments.
5.4.2 Variation of Methylmercury, Manganese and Iron in Downstream Region of Reservoir
Mean hypolimnetic MeHg, Mn and Fe concentrations downstream of the reservoir from
RE-20 to RE-02 were lower compared to upstream stations (RE-30 and RE-35), except for the
month of July in RE-15 where Mn concentration was around 5 mg/L and Fe concentration was
around 3 mg/l, the highest throughout the whole reservoir (Fig.5-4). During that month, MeHg
154 increased to around 0.3 ng/L in RE-15, the second highest MeHg accumulation observed behind
RE-35. This also suggested that Mn and Fe oxides were present within the sediments prior to the
onset of anoxia. Most of the other stations showed little to no variation at the sediment-water
interface and minimal MeHg accumulation. Mean hypolimnetic MeHg in RE-20 was
approximately the same in 2012 and 2013, even though mean hypolimnetic Mn and Fe was
higher in 2012 relative to 2013 (Figs.5-4; Fig.5-5). RE-02 and RE-05 displayed similar patterns
and showed no hotspot of MeHg release compared to RE-10 and RE-15. Mean hypolimnetic
MeHg was two times higher in RE-10 and RE-15 relative to RE-02 and RE-05.
The relationships observed between the downstream stations in regard to low levels of
MeHg release were surprising. Since most of the nitrate did not reach the downstream of the
reservoir, it was expected that some MeHg accumulation would occur in the water column, but
the transition zone showed little to no buildup of MeHg release. As discussed by Cubas et al.
(2014), one of the unique features of the reservoir is its capability to consume most of the nitrate
that entered to the system. The mean nitrate concentration measured in downstream stations of
the reservoir never exceeded 2.5 mg-N/L, even though the nitrate concentration in the input
exceeded 10 mg-N/L. The nitrate depletion along the length of the reservoir is attributed to high
denitrification rates favored by high concentrations of organic matter and low DO concentration
at the sediment-water interface (Randall and Grizzard, 1995). At station RE-15, RE-10, RE-05
and RE-02, DO was depleted by the end of June and remained low until early September (Fig.54). As a result, most of the nitrate was consumed in the upper reaches of the reservoir during the
stratified period. Clearly, it can be concluded downstream stations were not hotspot for MeHg
accumulation since methylation process showed little activity above the sediment in the absence
of nitrate and DO.
155 Due to the differences in standard potential, rates of diffusion, and kinetics of oxidation,
differences in the water column distribution of Mn and Fe were expected downstream of the
reservoir. Because of their differing responses to redox conditions (Davison, 1993), our results
showed strong variation of Mn and Fe between stations throughout the stratified period,
indicating that Mn and Fe are good indicators of the redox status of surficial sediments. The
highest peak of Mn and Fe was observed at RE-15 in 2012, suggesting that both were very
persistent in the anoxic portion of the hypolimnion at this station and possibly Mn and Fe oxides
were present within the sediment prior to the onset of anoxia. Mean hypolimnetic of Mn was
higher than Fe throughout the reservoir, except for the month of July in some stations.
Apparently, the rate of reduction of Fe is faster than the rate of reduction of Mn, suggesting that
Mn is more resistant to rapid abiotic oxidation (Chadwick et al., 2006). Since Mn does not
readily oxidize as quickly as Fe, Mn oxidation and subsequent settling cannot account for lower
Mn in the reservoir. The probable sink for Mn in the reservoir was coprecipitation with Fe
oxides, a phenomenon that has been observed in other lakes and reservoirs (Hongve, 1997;
Davison, 1993). Our results were consistent with this observation in RE-15, RE-20 in 2012
(July-August) and RE-35 in 2013 (September-October)(Fig.5-4; Fig.5-5). In this case, Fe(II) and
Mn(II) were released from anoxic surficial sediment and mixed into the upper hypolimnion
where Fe(III) coprecipitated with Mn(II). Fe(III) oxides enriched with Mn(II) then settled out of
the water column with some of the Fe(III) oxides redissolving at the anoxic sediment-water
interface, thereby recycling Fe(II) and Mn(II) back into the water column (Beutel et al., 2014).
Dent et al. (2014), in his study at Twin Lakes in Washington, USA, also discussed how Mn(II)
coprecipitation with Fe(III) was also the mechanism proposed for a rapid and massive drop in
Fe(II) and Mn(II) from the hypolimnetic water during an oxygenation system test. In contrast to
156 coprecipitation mechanisms observed in RE-15, RE-20 and RE-35, RE-05 in 2012 and RE-20 in
2013 exhibited different patterns. Mn oxidized slower than Fe, reinforcing the contention that
Mn is biologically reduced at higher redox potentials than Fe. An example of this phenomenon is
the temporal disparity between the accumulation of Mn and Fe in bottom waters of some lakes.
Mn accumulation in eutrophic Lake Sammamish, Washington, started around 10 weeks before
Fe accumulation (Balistrieri et al., 1992). Redox potential in seasonally anoxic Rostherne Mere,
UK never dropped low enough to induce Fe accumulation; only Mn was observed at the
sediment-water interface (Davison and Woof, 1984).
5.5. Conclusion
Reservoir water quality is a complex problem and needs advance management and
sustainable remediation alternatives (Holland, 1993). Rational planning and operation of water
supply systems requires recognition of the cause-effect relationships that influence water quality
and, therefore, influence the feasibility and costs of supplying water that meets state and federal
standards and criteria. The trend toward increasingly restrictive drinking water standards can be
attributed to increased scientific understanding of the relationships between drinking water
quality and health and to vast improvements in analytical capabilities, particularly with respect to
nutrient and metals (USEPA, 2004). Given this trend and the current and projected waters supply
shortages in many areas of the country, the protection or enhancement of source water quality
has become increasingly critical.
In the case of Occoquan Reservoir, depletion of oxygen at the sediment-water interface
caused the release of Fe, Mn, P and ammonia into the water column, and likely MeHg. However,
157 the nitrified effluent from the Upper Occoquan Service Authority (UOSA) has had clear
beneficial effects on water quality in the Occoquan Reservoir by repressing the release of these
redox-sensitive compounds (Cubas, 2012). Regarding MeHg, nitrate addition affected the upper
portion of the reservoir, and still kept downstream portion from going severely anaerobic. As a
result, MeHg release downstream was lower than expected. This suggested that perhaps if nitrate
was not added, downstream would exhibit higher MeHg release at the sediment-water interface.
Because this distinctive approach to nutrient and metal cycling treatment, several concerns have
been raised about the effects of implementing stringent nitrogen removal regulations in
wastewater treatment plant effluents in order to minimize excessive nitrogen to estuarine
tributaries of the Chesapeake Bay (Cubas et al., 2014). But data collected by the Occoquan
Water Quality Monitoring Laboratory over the last 40 years have demonstrated that there is a
well-understood water quality benefit from the maintenance of oxidized nitrogen in the
hypolimnion of the reservoir (Randall and Grizzard, 1995; Banchuen, 2003; Cubas et al., 2014).
In fact, an important observation is the reservoir’s capacity to consume most the nitrate that
enters in the system from the UOSA. As a result, effluents from the UOSA served as a
significant supplement to the safe drinking water supply yield of the Occoquan Reservoir, and
nitrate was an attractive management alternative for lakes and reservoirs that are subject to
extended period of hypoxia.
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163 Figure 5-1. Map of Occoquan Reservoir complements of Francisco Cubas and Tom Grizzard of
Virginia Tech. Stations monitored in 2012 include Stations 1 (RE30), 2 (RE35), 3 (RE20), 4
(RE15), 5 (RE10), 6 (RE5) and 7 (RE02). Stations monitored in 2013 include 1 (RE30), 2
(RE35), 3 (RE20). The nitrate-rich effluent from the WWTP is input upstream of Station 1.
164 DO
Nitrate
0.07
Phosphate
0.06
8
0.05
6
0.04
0.03
4
0.02
2
0
Feb
Phosphate, mg-P/L
DO, mg/L and nitrate, mg-N/L
10
0.01
Mar
Apr
May
Jun
Jul
Aug
Sep
Oct
0.00
Dec
Figure 5-2. 2009 dissolved oxygen (DO), nitrate and phosphate in the bottom waters of the
Occoquan Reservoir, Virginia. Depletion of DO and nitrate in June and July corresponded with
elevated phosphate. Elevated nitrate in August and September, a result of inflow of nitrate-rich
tertiary treated wastewater, corresponded with lower levels of phosphate. Data from Cubas
(2012).
165 Figure 5-3. Bottom water quality in Occoquan Reservoir in 2012. RE 35 is the upstream tributary
that does not receive nitrate rich discharge. RE 30 is the upstream tributary that does receive
nitrate rich discharge. RE 20 is downstream of the confluence of the two tributaries in the upper
portion of the reservoir. Stations continuing downstream are shown in the following figure.
Values are average of bottom water samples collected below the thermocline (n = 1-5).
166 Figure 5-4. Bottom water quality in Occoquan Reservoir in 2012. Stations are moving
downstream through the reservoir from RE 20 to RE 02 near the dam. This is a continuation of
the previous figure. Values are average of bottom water samples collected below the thermocline
(n = 1-5).
167 Figure 5-5. Bottom water quality in Occoquan Reservoir in 2013. RE 35 is the upstream tributary
that does not receive nitrate rich discharge. RE 30 is the upstream tributary that does receive
nitrate rich discharge. RE 20 is downstream of the confluence of the two tributaries in the upper
portion of the reservoir. Values are average of bottom water samples collected below the
thermocline (n = 1-5).
168 Figure 5-6. Correlations between DO/nitrate and methylmercury (MeHg) in bottom waters of
station RE-35, RE-30 and RE-20 for 2013. Values are average of bottom water samples collected
below the thermocline (n = 1-5).
169 Figure 5-7. Correlations between iron/manganese and methylmercury (MeHg) in bottom waters
of station RE-35, RE-30 and RE-20 for 2013. Values are average of bottom water samples
collected below the thermocline (n = 1-5).
170