U.S. Department of Interior 

U.S. Geological Survey 

Flux of Dissolved Forms of Mercury 

Across the Sediment-water Interface in 

Lahontan Reservoir, Nevada 

By James S. Kuwabara, 

Mark Marvin-Dipasquale, 

Wayne Praskins, 

Earl Byron, 

Brent R. Topping, 

James L. Carter, 

Steven V. Fend, 

Francis Parchaso, 

David P. Krabbenhoft,

Water Resources 

Investigations Report 02-4138 

and Mae S. Gustin
Flux of Dissolved Forms of Mercury Across the Sediment-water 

Interface in Lahontan Reservoir, Nevada 

 

By James S. Kuwabara1, Mark Marvin-Dipasquale2, Wayne Praskins3, Earl Byron4, Brent 

R.Topping5

, James L. Carter6, Steven V. Fend7, Francis Parchaso8, 

David P. Krabbenhoft9, and Mae S. Gustin10 

______________________________________________________________________ 

U.S. GEOLOGICAL SURVEY 

Water Resources Investigations Report 02-4138 

1 kuwabara@usgs.gov, U.S. Geological Survey, Menlo Park, CA 

2 

mmarvin@usgs.gov, U.S. Geological Survey, Menlo Park, CA 

3 Praskins.Wayne@epamail.epa.gov, U.S. Environmental Protection Agency, San Francisco, CA 

4 ebyron@ch2m.com, CH2M HILL, Sacramento, CA 

5 btopping@usgs.gov, U.S. Geological Survey, Menlo Park, CA 

6 jlcarter@usgs.gov, U.S. Geological Survey, Menlo Park, CA 

7 svfend@usgs.gov, U.S. Geological Survey, Menlo Park, CA 

8 parchaso@usgs.gov, U.S. Geological Survey, Menlo Park, CA 

9 dpkrabbe@usgs.gov, U.S. Geological Survey, Middleton, WI

10 msg@unr.nevada.edu, University of Nevada, Reno, NV 

1
U.S. DEPARTMENT OF THE INTERIOR 

Gale Norton, Secretary 

U.S. GEOLOGICAL SURVEY 

Charles G. Groat, Director 

Any use of trade, product, or firm names is for descriptive purposes only and does not imply 

endorsement by the U.S. Geological Survey. 

______________________________________________________________________ 

For addition information contact: Copies of this report may be purchased from: 

James S. Kuwabara U.S. Geological Survey 

U.S. Geological Survey Branch of Information Services 

345 Middlefield Road, MS 439 Box 25286 

Menlo Park, CA 94025 Denver, CO 80225 

2
 TABLE OF CONTENTS 

Executive Summary -------------------------------------------------------------------- 4 

Background ------------------------------------------------------------------------------ 8 

Results and Discussion ----------------------------------------------------------------- 9 

Study Design and Methods ------------------------------------------------------------ 13 

References Cited ------------------------------------------------------------------------ 15 

Acknowledgements --------------------------------------------------------------------- 17 

Appendix 1: Comments on the Report Structure ----------------------------------- 18 

Appendix 2: List of Figures ---------------------------------------------------------- 19 

Appendix 3: List of Tables ------------------------------------------------------------ 20 

 

 

3
Flux of Dissolved Forms of Mercury Across the Sediment-water 

Interface in Lahontan Reservoir, Nevada 

by James S. Kuwabara, Mark Marvin-Dipasquale, Wayne Praskins, Earl Byron, 

Brent R.Topping, James L. Carter, Steven V. Fend, Francis Parchaso, 

David P. Krabbenhoft, and Mae S. Gustin

Executive Summary 

Field and laboratory studies were conducted between April 30, 2001 and July 19, 2001 to provide the first direct

measurements of the benthic flux of dissolved (0.2-micrometer filtered) mercury species (total and methylated

forms) between the bottom sediment and water column at three sites: two in the southern lobe and one in the 

northern lobe of Lahontan Reservoir, Nevada (Background, Fig. 1). Dissolved-mercury species and predominant 

ligands (represented by dissolved organic carbon, and sulfides) were the solutes of primary interest. Benthic flux, 

sometimes referred to as internal recycling, represents the transport of dissolved chemical species between the water 

column and the underlying sediment. 

Water-quality managers often assess and prioritize remediation strategies for aquatic systems, in particular Super 

Fund sites that have been adversely affected by anthropogenic activities. In the case of the Lahontan Reservoir 

along the Carson River, mercury associated with historic gold and silver extraction has been fluvially transported 

and accumulated in the bottom sediments. Frequent demands have been made by Super Fund site managers and the 

general public to quantify the connections between fluxes of contaminants and the health, abundance, and 

distribution of biological resources (Kuwabara and others, 1999). As part of a comprehensive examination of 

transport processes affecting mercury dynamics in Lahontan Reservoir, this study focuses on a poorly understood, 

yet potentially predominant, source of mercury to the reservoir water column, which is internal recycling, or benthic 

flux of mercury species and associated ligands. Mobilization, flux, and biological availability of mercury into the 

water column of the reservoir are affected by physical (e.g., advection and diffusion), chemical (i.e., oxidationreduction reactions, complexation and repartitioning) and biological processes (Flegal and others, 1991; Kuwabara 

and others, 1996; Grenz and others, 2000, Topping and others, 2001).

The results described herein followed from the integration of current project studies with information needs 

identified by the U.S. Environmental Protection Agency, Region 9 (USEPA) to provide initial determinations of

dissolved total and methyl-mercury fluxes from the sediments into the water column of Lahontan Reservoir. Recent 

mercury distribution and transformation studies in the Carson River system by Marvin-Dipasquale and others (2001) 

indicated the potential importance of sediment-water interactions in describing mercury speciation, its sources and 

sinks. Quantifying and understanding the magnitude and variability of these interactions are critical to the accurate 

assessment of contaminant sources and loads as well as to the development of process-integrated water-quality 

models for this mining-affected system. 

With a variety of strategies under consideration to determine how to most efficiently improve the water quality 

in the Carson River system, the primary question posed in this study was, “What processes regulate the fate and 

transport of mercury species in Lahontan Reservoir? In particular, are sources and sinks of dissolved total and 

methyl mercury associated with the bottom sediment within Lahontan Reservoir significant relative to major 

surface-water inputs from the Carson River?” The question was motivated by a number of factors. First, extraction 

of precious metals up gradient of the reservoir provide a historic source of elemental mercury that continues to be 

fluvially transported in dissolved and particulate phases (Hoffman and Taylor, 1998; Carroll and others, 2000;

Carroll and Warwick, 2001). Elevated total and methyl mercury concentrations into the reservoir have been well

documented (Priessler and others, 1999; Jones and others, 1999; Allander and others, 2001, Marvin-Dipasquale and 

others, 2001). Therefore, determining whether some fraction of this sediment-associated mercury can remobilize for 

transport to the overlying water and subsequently to down-stream portions of the Carson River is necessary. 

4
Second, elevated concentrations of mercury species in reservoir water, sediment, and fish have prompted a 

compilation, and comparison of dominant contaminant sources so that appropriate remedial strategies can be 

designed and implemented. Third, changes in oxidizing or reducing (redox) conditions and nutrient availability near 

the sediment-water interface (e.g., during phytoplankton blooms) can dramatically alter the mobility of metals and 

ligands associated with the bottom sediment as episodic sources of carbon settle out and accumulate (Thompson and 

others, 1981). Finally, there is a growing body of evidence from other aquatic systems that benthic flux or internal

recycling of contaminants and nutrients is an important process to consider in developing appropriate ecosystem

water-quality models (Wood and others, 1995; Kuwabara and others, 2000). The need for more refined numerical 

and conceptual models for mercury dynamics within the Carson River system thus clearly exists. 

This report is formatted unconventionally in a pyramid-like structure to effectively serve the needs of diverse 

parties interested in reviewing or acquiring information at various levels of detail (Appendix 1). The report enables 

quick transitions between the initial summary information (figuratively at the top of the pyramid) and later details of 

methods or results (that is, figuratively towards the base) using hyperlinks to supporting figures and tables, and an 

electronically linked Table of Contents. 

During two sampling events, two replicate sediment cores were collected from three reservoir locations (Fig. 1) 

for incubation experiments to provide flux estimates and benthic biological characterizations (Fig. 2). Ancillary 

data, including nutrient and ligand fluxes, were determined to provide a water-quality framework from which to

compare the results for mercury. Because the first sampling event spanned the last day of April and the first days of 

May 2001, the results for that experimental series will be referred to as "April/May 2001". The following major 

observations from interdependent physical, biological, and chemical data were made: 

Physical and Biological Characterizations

1. Hydrologic Conditions: The sampling associated with this study occurred during one of the driest years 

on record for the area as depicted by the discrepancy between the daily average streamflow into 

the reservoir and the historical average daily streamflow since the beginning of the construction of 

the Lahontan Dam in 1911 (Fig. 3; U.S. Geological Survey, 2001). These drought conditions 

caused a rapid draw down of the reservoir as indicated by reservoir-storage data between the two 

coring trips. The time of the second coring trip was greatly influenced by restricted boat-ramp

access. For example, the water-column depths at the coring sites in the southern and northern 

lobes on April 30, 2001 were approximately 6.7 and 21.3 meters, respectively, while on July 18, 

2001, during our second coring trip, those depths had decreased to 4.2 and 18.9 meters, 

respectively. 

Dissolved-oxygen and temperature profiles during the coring trips indicate that an increase in 

water-column temperatures between the trips coincides with a decrease in bottom-water dissolvedoxygen concentration (Fig. 4). However, the water column remained oxic through the summer, 

and in fact subsequent water-column sampling indicated that the bottom water did not become

anoxic that year. This may partially be explained in the northern lobe by vertical mixing of water 

from the extraction depth of Lahontan Dam and the augmented input of oxygenated water from

the Truckee Canal near the dam to meet agricultural demands down gradient along the Carson 

River (Fig. 1). 

2. Porosity: The surficial sediments at the three coring sites were observably different during both 

sampling trips. Cores from main-channel, southern-lobe Site 1, consistently exhibited the lowest 

porosities (0.65 to 0.70). Less than 0.5 kilometers away from Site 1, the near-shore site in the 

southern lobe (Site 2) consistently indicated higher porosities (0.80 to 0.85). The highest 

porosities at the northern-lobe site 3 (0.89 to 0.93) were indicative of the fine-grained

unconsolidated, viscous slurry at the sediment-water interface (Table 1). 

 

3. Macrobenthos: The presence of certain macroinvertebrate taxa may increase, by orders of magnitude, 

the solute flux across the sediment-water interface due to biologically enhanced advection 

5
(Caffrey and others, 1996; Kuwabara and others, 1999). Macroinvertebrate abundances from the 

Lahontan sites appear to be similar to published abundance values from around the world (Fig. 5). 

However, macroinvertebrate populations in the Lahontan samples were considered to be of low 

importance to benthic flux because they were dominated by very small immature tubificid worms 

and nematodes that typically pass through a conventional 500-micrometer mesh sieve (Tables 2 -

4; Macrobenthos discussion). Worm fragments, associated with particle aggregates, inflate the 

data depicted for Lahontan. The low biomass of the benthic community at the coring sites suggest

that there were insufficient biota to generate biologically enhanced advection through bioirrigation 

or bioturbation. This is consistent with low benthic- chlorophyll concentrations (see section 

below), and with sulfide-flux determinations. However, other lines of evidence such as bromide 

injections, radon emanation or silica fluxes would be required to confirm that diffusion controls 

benthic flux. 

4. Benthic chlorophyll: Chlorophyll-a concentrations in surficial sediments ranged from 0.8 to 2.1 

micrograms per square centimeter with highest concentrations observed at the near-shore Site 2 

during the spring sampling (April/May 2001), and the lowest concentrations at the northern-lobe 

site (Site 3) on both sampling dates (Table 5). A compilation of benthic biological data from our 

study and compilations from publications on 14 reservoirs and oligotrophic and mesotrophic lakes 

around the world (Fig. 5; Bass, 1992; Popp and Hoagland, 1995; Kuwabara and others, 2000) 

clearly indicated that the benthic algal biomass was sparse at all three coring sites relative to 

previously reported values. 

Chemical Characterizations 

Note: The dissolved-mercury concentrations discussed in this section refer to samples filtered

with 0.7-micrometer quartz-fiber filters pre-combusted at 500 oC. 

1. Water-column concentrations: Both dissolved total and methyl mercury concentrations in the bottom

water over the two coring dates were lowest at the northern-lobe Site 3 (0.3 to 0.5 picomolar 

methyl mercury and 373 to 787 picomolar total mercury) and highest at the main-channel

southern-lobe Site 1 (0.7 to 1.1 picomolar methyl mercury and 647 to 1952 picomolar total

mercury). Total-mercury concentrations in reservoir bottom waters were higher on July 18, 2001 

during low-flow conditions than during high-flow (April 30, 2001) by a factor of 2 to 3, but such 

temporal differences in methyl-mercury concentrations were not evident. 

Dissolved total and methyl mercury of riverine sources were 135 + 21 picomolar (n=56) and 5.4 +

1.2 picomolar (n=56), respectively (Priessler and others, 1999; Jones and others, 1999; Allander 

and others, 2001). Compared to the reservoir bottom waters (Table 7), the total-mercury 

concentrations of the riverine inputs were lower and the methyl-mercury riverine concentrations 

were higher. 

2. Benthic flux of dissolved forms of mercury: Dissolved methyl-mercury fluxes were positive except 

from two cores at the main-channel southern-lobe Site 1. The highest fluxes were at the northernlobe Site 3, and the lowest at the main-channel southern-lobe Site 1. The methyl-mercury fluxes 

were about three orders of magnitude less than the flux estimates for total mercury, that were 

roughly consistent with concentration differences between species. 

In contrast to methyl-mercury fluxes, dissolved total-mercury fluxes, based on end-point

concentrations, were greatest at the main-channel southern-lobe Site 1. Total-mercury fluxes at 

the other two sites (near-shore southern-lobe Site 2 and northern-lobe Site 3) were lower in July

than in April/May 2001. 

6
Initial riverine, atmospheric and benthic flux estimates do not indicate a dominant transport 

process regulating the fate of mercury species in the reservoir water column (Fig. 7). But, at 

times, the release of dissolved mercury from the benthos is comparable to, or greater than, the 

magnitude of riverine and atmospheric flux (Fig. 6). 

Remedial Implications 

Because the benthic flux of mercury species may represent a dominant transport process, it also 

suggests an important management implication. Remediation efforts in the Carson River have an 

objective of decreasing concentrations and loads to downgradient systems. In the reservoir, a 

subsequent decrease in dissolved-mercury concentrations in the water column would shift and

increase the vertical concentration gradient near the sediment-water interface. Consequently, 

solute flux into the water column from the sediments may increase as an unintentioned result of

upstream remediation efforts (Fig. 7). Therefore, when water-quality models are developed to

evaluate, justify and establish remediation strategies in the basin, the importance of reservoir 

sediment as mercury sources and sinks should be considered. 

7
Background 

What is benthic flux and why should it be considered?

Many fundamental processes affect the transport of dissolved chemical species (e.g., nutrients, metals, or 

ligands) through and within reservoirs and other aquatic systems. A conceptual model of these processes (Fig. 8) 

illustrates some physically based processes that have been examined and carefully quantified for several years (e.g., 

advective transport and point source inputs; Fisher and others, 1979). 

Conversely, there are flux terms in the conceptual model that have received little attention. A prime example is 

the benthic flux term where no direct measurements for dissolved mercury species have heretofore been available 

for the study area. Benthic flux (sometimes referred to as internal recycling) represents the transport of dissolved

chemical species between the water column and the underlying sediment. Flux of solutes can be either positive (into 

the water column from the sediment or atmosphere) or negative (out of the water column into the sediment or 

atmosphere) and can vary over multiple temporal and spatial scales (Kuwabara and others, 2000; Topping and 

others, 2001). 

As a result of physical, chemical, and biological changes in the vertical cross section of the sediment-water 

interface, geochemical gradients take on a variety of forms that have been previously reported (Kuwabara and 

others, 2000; Fig. 9). When the solute concentration above and below the sediment-water interface are equivalent; 

there is no discernable concentration gradient, and, consequently, no net transfer of that substance across the 

interface results in a condition of no benthic flux. In contrast, when solute concentrations in the water column are 

higher than those in the bottom sediment pore waters, a negative benthic flux results whereby the substance moves 

into the sediment. Dissolved oxygen is a typical example of such a solute where microbial respiration can create a 

sediment demand for oxygen. When concentrations in the water column are lower than those in sediment pore 

waters, the vertical concentration gradient can physically drive the release from the sediment of chemical species to 

the overlying water yielding a positive benthic flux. The remobilization or chemical transformation of sedimentassociated trace metals or ligands may exemplify this condition. 

When interdependent factors regulate the benthic flux of biologically reactive substances, the vertical gradient

for one dissolved species may be dependent on the gradient of another chemical species. For example, an 

attenutated release may occur when solute concentrations increase below the sediment-water interface only when 

another solute is depleted. Dissolved iron often behaves in this manner when suboxic conditions reduce it from

ferric to ferrous forms, increasing its solubility. Additionally, macroinvertebrates can biologically enhance the 

benthic flux by irrigating, or perturbing surficial sediment layers (bioirrigation, bioturbation, or biologically

enhanced advection). Certain productive benthic communities can enhance benthic flux beyond diffusive-control 

by orders of magnitude (Kuwabara and others, 1999). In summary, vertical chemical gradients generated by a 

variety of interdependent biogeochemical processes can induce the movement of dissolved-mercury species across 

the sediment-water interface. 

Scientists and water-quality managers are only beginning to appreciate the importance of benthic flux in many

aquatic environments. Within the past decade or two, researchers have gradually realized that there are nonhydrologic processes (for example, benthic flux) that must be incorporated into water-quality models in order to

generate physically meaningful information. Benthic-flux studies have lagged behind studies of other transport

processes (Flegal and others, 1991; Kuwabara and others, 1996) because quantifying benthic flux is instrument and 

manpower intensive. Each flux estimate requires a concentration time-series analysis or vertical-profile analysis. 

8
Results and Discussion 

Physical Data

There were visible differences in sediment texture among the three sampling sites that directly

affected the collection of cores for incubation. The main-channel site in the southern lobe (Site 1)

was consistently the lowest in porosity (Table 1) and the most difficult to core. This was because 

the finer surficial sediment of approximately 2 to 3 centimeter depth overlay sandy, 

unconsolidated material that was sometimes lost during core retrieval. The near-shore site in the 

southern lobe (Site 2), although less than 0.5 kilometers away from Site 1, was consistently higher 

in porosity and easier to core because of a thicker (4 to 5 centimeter) layer of fine-grained surficial 

material that allowed deeper core penetration. The deepest site in the northern lobe (Site 3)

yielded sediment of the highest porosity. 

Biological Data

1. Benthic chlorophyll-a: Benthic chlorophylls were lower in Lahontan Reservoir (Table 5) than reported 

for other oligotrophic/mesotrophic lakes and reservoirs (Fig. 5). 

 

2. Benthic macroinvertebrates: The benthic macroinvertebrates were dominated by nematodes and 

tubificid worms both spatially and temporally (Fig. 10). Although the number of individuals in

the benthic community was greater in July 2001 than in April/May 2001, the number of taxa 

remained similar. Typically, nematodes and immature tubificids are not quantitatively retained by

500-micrometer sieving. The effects of particle aggregation on sieving efficiency may therefore 

inflate the number of total individuals relative to other studies using a 500-micrometer mesh. 

Reible and others (1996) reported that the flux of hydrophobic organics was significantly

increased when large, laboratory-cultured tubificids were present at densities between 6,000 to 

27,000 per square meter. Not only were the densities of tubificids in Lahontan Reservoir cores 

typically much lower, but they were also dominated by smaller immature individuals (Tables 2 –

4). 

Taxonomic analyses of July 2001 cores were also performed using individuals retained on a 106–

micrometer sieve as well as those retained on a 500-micrometer sieve to check for the potential

importance of losing small individuals while using conventional-sized mesh. Data from these 

analyses indicate that the 500-micrometer mesh lost about 50% of the individuals that were 

retained by the smaller mesh, but those individuals were obviously smaller and insufficient in

number to alter the conclusion that the benthic macroinvertebrate community at all reservoir sites 

contributed little to enhance benthic flux (Word, 1980; Miller, 1984; Tables 3 and 4; Fig. 11). 

Chemical Data

For consistency with previous geochemical studies, mercury flux estimates were determined in 

mass-flux units, but were also tabulated in molar units. 

1. Dissolved oxygen (DO) consumption: Dissolved-oxygen fluxes were consistently negative, indicating 

benthic consumption. DO flux ranged from -483 micromoles per square meter per hour at Site 3 

in July 2001 (average of two measurements, n=8) to -1240 micromoles per square meter per hour 

at Site 2 on the same date (Table 6). A comparison of average dissolved-oxygen flux 

measurements for the three sites indicates that the southern-lobe Sites 1 and 2, in contrast to the 

northern-lobe Site 3, displayed a temporal increase in oxygen consumption in July 2001 relative to

April/May 2001. This observation is consistent with the lower DO level and higher bottom-water 

temperature in July 2001 (Fig. 4). Although the reason for contrasting temporal trends for Site 3 

9
are unclear, the augmented flow from the Truckee Canal and modified circulation patterns in the 

northern lobe during the drought conditions between the two coring dates could have affected the 

DO flux. It should be pointed out that these temporal differences, though notable, are not

statistically significant because of non-linearities in the DO time series (Table 6). 

2. Bottom-water mercury concentrations: Dissolved total-mercury concentrations in the reservoir 

bottom water over the two coring dates ranged from a minimum of 375 + 3 picomolar (n = 2) at

the northern-lob Site 3 on April 30, 2001, to a maximum of 1930 + 32 picomolar (n = 2) at the 

main-channel southern-lobe Site 1 on July 18, 2001 (Table 7). Similarly, dissolved methylmercury concentrations in the bottom water were lowest at Site 3, on both coring dates (0.4 

picomolar) and highest at Site 1 on April 30, 2001 (0.9 picomolar; Table 7). Furthermore in terms 

of spatial distributions, the lowest concentrations were consistently observed for both dissolved 

total and methyl-mercury concentrations at the northern-lobe Site 3 and highest at southern-lobe 

Sites 1 and 2. Temporally, the total-mercury concentrations in reservoir bottom waters were 

higher in July than April/May 2001 by a factor of 2 or 3, but similar differences in methylmercury concentrations were not evident. 

Although mercury analyses for filtered surface waters are not available, concentrations of filtered 

bottom-water samples may be compared to a time series of surface-water analyses taken at Weeks 

Bridge, near the riverine inflow to the reservoir (Priessler and others, 1999; Jones and others, 

1999; Allander and others, 2001). The total-mercury concentration range for reservoir bottom

waters determined in this study (373 to 2068 picomolar) were consistently higher than observed at

Weeks Bridge in Fort Churchill, near the Carson River inflow to the reservoir (135 + 21 

picomolar, n=56) between April, 1997 and May, 2001. In contrast, the methyl-mercury

concentration range in the bottom waters (0.3 to 1.1 picomolar) was lower than observed for the 

riverine source (5.4 + 1.2 picomolar, n=56). 

3. Mercury benthic flux: For both incubation experiments, dissolved methyl-mercury fluxes were positive 

except from two main-channel southern-lobe Site 1 cores (Table 8). The highest methyl-mercury 

fluxes were at the northern-lobe Site 3, and the lowest at the main-channel southern-lobe Site 1. 

For both sampling dates, the magnitude of the methyl-mercury fluxes were about three orders of 

magnitude less than the flux estimates for total mercury, roughly consistent with concentration 

differences between species. 

In contrast to methyl-mercury fluxes, dissolved total-mercury fluxes were greatest at the mainchannel southern-lobe Site 1. Total-mercury fluxes at the other two sites (near-shore southernlobe Site 2 and northern-lobe Site 3) were lower in July 2001 than in April/May 2001. In both 

series of incubation experiments, within-site variability in the flux estimates, typically obscured

between-site differences. That is, differences between fluxes for dissolved total-mercury were not

statistically significant (very low coefficients of determination for the regressions) when the 

intermediate sampling intervals at 2 and 6 hours of elapsed time were used for calculating flux. 

The total-mercury concentrations in all cores initially decreased before increasing between 6 and

12 hours. This type of non-linear temporal trend for the concentrations is atypical for incubation 

experiments of this kind, and may reflect a transient response to a perturbation at the beginning of 

the incubation. One possibility is that atmospheric exposure of the overlying water added at the 

beginning of the incubation caused initial mercury precipitation. This explanation might be tested

by incubating over longer durations (for example, a 24-hour end point) and evaluating linearity. 

As one might expect, the bias of omitting the initial concentrations in each time series (that is, the 

use of only the 2, 6 and 12 hour data in the flux calculations) would yield statistically significant

regressions and much larger flux values. Such values might represent maximum potential flux 

estimates. Because of volume constraints, methyl-mercury concentrations were measured only at 

the beginning and the end of the incubation period. 

A summary of mercury-flux data from a variety of processes (Fig. 7) illustrates that flux estimates 

not only vary over orders of magnitude, but also vary in direction where a positive flux represents 

10
solute transport into the water column, and a negative flux indicates solute moving from the water 

column to the sediment or to the atmosphere. The core-incubations were used to estimate benthicflux over the projected area of the reservoir for the two sampling dates (31.3 and 23.8 square 

kilometers on April 30 and July 18, 2001, respectively), and ranged between 0.8 to 5.7 and –78 to

7291 picomoles mercury per square meter per hour for methyl and total mercury, respectively,

after using the mean of two replicate cores per site. This corresponds to a mass range of –4 to 27 

and -374 to 35,100 nanograms mercury per square meter per day for methyl and total mercury,

respectively. Secondly, a comparison of flux magnitudes from these sources and sinks indicate no 

obvious dominant process. Riverine fluxes into the reservoir for the two sampling dates were 

based on USGS stream-flow data available on the internet (USGS, 2001) for the Carson River at

Weeks Bridge, just up gradient of the reservoir, and on storage data and surface-area estimates for 

Lahontan Reservoir (Katzer, 1971; USGS, 2001). In addition, dissolved total and methyl-mercury 

data from Weeks Bridge were used (Priessler and others, 1999; Jones and others, 1999; Allander 

and others, 2001). For April 30 and July 18, 2001, the riverine fluxes were estimated at 12.6 and 

0.1 picomoles methyl mercury per square meter per hour, and 316 and 1.9 picomoles total

dissolved mercury per square meter per hour, respectively. The estimate for daily Hg emission 

from the reservoir (21 picomoles mercury per square meter per hour or 100 nanograms mercury

per square meter per day) is based on a diel atmospheric-flux study conducted in July 2001. 

Based on our current comparative information, it would seem unwise to neglect any of these 

processes when trying to quantitatively describe mercury-transport processes within and through 

this system. 

4. Dissolved organic carbon (DOC): DOC concentrations serve as a model for the presence of organic 

ligands that can complex and enhance the solubility of trace metals like mercury. The partitioning 

and speciation of certain metals in aquatic systems are significantly affected by organic 

complexation (Mantoura and others, 1978; McKnight and others, 1983) and, in particular, for 

mercury. Ligands represented by dissolved organic carbon can compete with sulfide to regulate 

solubility (see section 5 below, Ravichandran and others, 1998). This solute interdependency is 

consistent with concentration trends in bottom-water DOC (Table 9). Similar to total dissolved

mercury in the bottom waters, DOC was: (1) significantly lowest on both dates at the northernlobe site 3, and (2) consistently higher at all sites in July 2001 by approximately 40% relative to

April/May 2001 concentrations. Application of a physical-transport model might clarify whether 

this temporal increase in the summer was primarily due to one or more of the following: snowmelt dilution (leading to lower DOC levels during April/May), a source of DOC available during

the summer within the watershed, temporal changes in mass transport through the reservoir, and/or 

evaporation. 

The benthic flux of DOC ranged from –45 micromoles per square meter per hour at Site 3 in July

2001 (average of two measurements, n=8) to 116 micromoles per square meter per hour at Site 2 

in April/May 2001 (Table 10). The fact that the highest measured DOC flux into the sediment at

Site 3 was coincident with the lowest average consumption rate for dissolved oxygen (Table 6) 

suggests an electron-donor (organic carbon) limitation at Site 3. Similarly in July 2001, the 

highest DOC flux out of the sediment at Site 2 was coincident with the highest average rate of 

benthic consumption for dissolved-oxygen. This is not surprising with elevated DO 

concentrations in the shallower depths of the southern lobe relative to those of northern lobe 

bottom water (Fig. 4). In April/May 2001, the highest DOC flux out of the sediment was also at 

Site 2, but dissolved-oxygen saturation in the bottom waters of all three coring sites generated

spatially comparable DO fluxes. This temporal shift in bottom-water DO concentrations may also

have contributed to the reversal in direction of DOC flux for Site 3 between the April/May and

July 2001 experiments. 

5. Dissolved Sulfides: Sulfide flux was consistently positive (that is, from the sediment to the overlying

water column) and ranged from 419 to 766 nanomoles per square meter per day (average of two 

11
replicates per site per date; Table 11). A spatial comparison of dissolved-sulfide benthic fluxes 

indicates enhanced flux at the northern-lobe Site 3 in April/May 2001 relative to the southern-lobe 

Sites 1 and 2, but not in July 2001. For both coring dates, the sulfide benthic fluxes from the 

southern-lobe sites were statistically similar. Temporally, all sites exhibited an increase in

dissolved-sulfide flux in July relative to April/May 2001, but this difference was most pronounced 

at the southern-lobe Sites 1 and 2 where flux nearly doubled between experiments. In contrast, the 

increase in average dissolved-sulfide flux for northern-lobe Site 3 was not statistically significant. 

Using dissolved-sulfide concentrations from the bottom water and surficial porewater, a diffusive 

flux from the benthos was calculated with a diffusion coefficient of 1x 10-5 square centimeters per 

second (Li and Gregory, 1974). With one exception, the calculated diffusive-flux estimates were 

lower than directly measured sulfide-fluxes and were of the same order of magnitude (Table 11). 

In contrast, diffusive sulfide fluxes were three orders of magnitude lower than measured sulfide 

benthic flux in San Francisco Bay where bioturbation effects have been well documented (Caffrey

and others, 1996; Kuwabara and others, 1999). This comparison supported the conclusion that the 

reservoir had insufficient biomass of critical macroinvertebrate species to impose bioturbation or 

bioirrigation effects observed in other environments. 

Dissolved sulfides represent a metastable ligand with high affinity to complex most divalent

metals, but particularly mercury that has a logarithmic mercuric-sulfide solubility product of 

approximately –50 (Hogfeldt, 1983). At thermodynamic equilibrium, the product of the molar 

concentrations of uncomplexed mercury and uncomplexed sulfide ions is expected to be very low 

(of the order of 10-50) due to the formation of mercuric sulfide. 

12
Study Design and Methods 

The protocol described in this section focuses on method applications in this series of two core-incubation 

experiments. Details (e.g., quality control specifications) for each analysis have been previously documented 

(Woods and others, 1999; Praskins and others, 2001). Sampling was performed at a main-channel and near-shore 

site in the southern lobe of the reservoir (Fig. 1) (39 degrees 26.6328 minutes north latitude by 119 degrees 10.1403 

minutes west longitude, and 39 degrees 22.9939 minutes N by 119 degrees 9.9412 minutes W, respectively), and 

one site up gradient of Gull Island in the northern lobe (39 degrees 26.3302 minutes N by 119 degrees 4.5980 

minutes W). Horizontal coordinate information is referenced to the North American Datum of 1983 (NAD 83). 

Cores were taken during the morning to avoid the effects of wind on core retrieval and sediment resuspension. The 

site map (Fig. 1) also indicates flow of the Carson River into the southern delta region and the out-flow from the 

northern lobe, regulated at the Lahontan Dam. Also depicted is the Truckee Canal, a regulated supplemental input

to the reservoir. 

Coring Operation

On both sampling days (April 30 and July 18, 2001), three cores were collected at each of three 

sites using a coring device fabricated from non-metallic parts (Fig. 2, Savillex Corporation,

Minnetonka, Minnesota). To avoid sample contamination, wetted surfaces of the coring device 

and core tubes were acid-washed polypropylene or fluoroethylene polymer. Because the bottom

water was aerobic on both sampling dates, and throughout that summer, the nine cores were 

aerated overnight to condition wetted surfaces before the incubation period (Topping and others, 

2001). Two of the three cores per site were selected for incubation based on visual inspection of 

the sediment-water interface for any disturbances. Water overlying the sediment in the selected 

cores was sampled at four intervals during a 12-hour incubation. Dissolved-mercury samples were 

processed in a Class-100 laminar-flow hood. The time series of solute concentrations was used to

determine the benthic flux from each of the incubated cores. 

Physical Data 

Sediment Porosity: After core-incubations, approximately 10 milliliters of surficial sediment was 

collected from each core. Wet weight and dry weight after lyophilization were measured to 

calculate porosity (Table 1). 

Biological Data

1. Benthic Macrofauna: After core incubations, each core was sieved (500-micrometer mesh). The 

sieved samples were fixed with 10% buffered formalin, transferred to 70% ethanol, then sorted 

and identified to the lowest practicable taxonomic level (Fig. 12). In July 2001, the sediments 

were also sieved at 106 micrometers to provide additional information about the scarcity of 

benthic invertebrates. Samples were stained with rose bengal to facilitate sorting.

2. Benthic Chlorophyll-a: Each incubated core was sub-sampled in triplicate for benthic chlorophyll-a. 

Surficial sediment was collected on a glass-fiber filter and buffered with magnesium carbonate. 

Samples were then frozen in darkness for preservation until spectrophotometrically analyzed by 

methods described in Thompson and others (1981) and Franson (1985) (Fig. 13). 

13
Chemical Parameters 

1. Bottom-water sampling: Prior to coring at each sampling site, bottom-water samples (approximately 1 

meter above the sediment-water interface) were collected for subsequent analysis of dissolvedmercury speciation by cold-vapor atomic fluorescence spectroscopy (CVAFS) using a highdisplacement peristaltic pump and a tethered length of rigid fluoroethylene polymer tubing. These 

samples were used to replace overlying water in cores collected from each site for incubation. 

2. Dissolved Oxygen (DO): Consumption of dissolved oxygen was monitored for each incubated core 

using a micro-Winkler titration method (Fig. 14; Kuwabara and others, 2000). 

3. Dissolved Mercury: Mercury samples from water-column sampling, and core incubations were filtered

with 0.7-micrometer quartz membranes that had been baked for 6 hours at 500 degrees centigrade 

to remove residual organics. Filtered samples were acidified with quartz-distilled nitric acid (pH 

less than 2), and refrigerated in darkness until analyzed by CVAFS (Fig. 15, Krabbenhoft and 

others, 1998). 

4. Dissolved organic carbon (DOC): Dissolved organic carbon was determined by high-temperature, noncatalytic combustion (Qian and Mopper, 1997). Potassium phthalate was used as the standard. 

Low-DOC water (blanks <40 micrograms-organic C per liter) was generated from a doubledeionization unit with additional ultraviolet treatment (Milli-Q Gradient, Millipore Corporation)

(Fig. 16). 

5. Sulfides: Dissolved sulfides in overlying-water samples were analyzed by square-wave voltammetry 

(Fig. 17; Kuwabara and Luther, 1993). 

14
References Cited 

Allander, Kip, Gortsema, Gary, Huchinson, Douglas, and Schwartzenberger, James, 2001, Water Resources Data, 

Nevada, Water Year 2000: USGS Water-Data Report NV-00-1, 570 p. 

Bass, D., 1992, Colonization and succession of benthic macroinvertebrates in Arcadia Lake, a South-Central USA 

reservoir: Hydrobiologia, v. 242, No. 2, p. 123-131. 

Caffrey, J., Hammond, D., Kuwabara, J., Miller, L. and Twilley, R., 1996, Benthic processes in San Francisco Bay: 

the role of organic inputs and bioturbation, in Hollibaugh, J.T., ed., San Francisco Bay: the Ecosystem:

American Association for the Advancement of Science, Pacific Division, San Francisco, p. 425-442. 

Carroll, R.W.H., and Warwick, J.J., 2001, Uncertainty analysis of the Carson River merury transport model:

Ecological Modelling, v. 137, p. 211-224. 

Carroll, R.W.H., Warwick, J.J., Heim, K.J., Bonzongo, J.C., Miller, J.R., and Lyons, W.B., 2000, Simulation of 

mercury transport and fate in the Carson River, Nevada: Ecological Modelling, v. 125, no. 2-3, p. 255-278. 

Fischer, H.B., List, E.J., Koh, R.C.Y., Imberger, J., and Brooks, N.H., 1979, Mixing in Inland and Coastal Waters:

Academic Press, Inc., Orlando, FL, 483, p. 

Flegal, A.R., Smith G.J., Gill, G.A., Sanudo-Wilhelmy, S., Scelfo, G., and Anderson, L.C.D., 1991, Dissolved trace 

element cycles in the San Francisco Bay estuary: Marine Chemistry, v. 36, p. 329-363. 

Franson, M.A.H., 1995, Standard Methods for the Examination of Water and Wastewater, Sixteenth Edition, 

Method 1003C.6: American Public Health Association, American Water Works Association, Water 

Pollution Control Federation, Washington, D.C., 1268 p. 

Grenz, C., Cloern, J.E., Hager, S.W., and Cole, B.E., 2000, Dynamics of nutrient cycling and related benthic nutrient

and oxygen fluxes during a spring phytoplankton bloom in South San Francisco Bay (USA): Marine 

Ecology Progress Series, v. 197, p. 67-80. 

Hoffman, R.J., and Taylor, R.L., 1998, Mercury and suspended sediment, Carson River Basin, Nevada, - Loads to

and from Lahontan Reservoir in flood year 1997 and deposition in Reservoir prior to 1983: U. S. 

Geological Survey FS-001-98. 

Hogfeldt, E., 1983, Stability constants of Metal-Ion Complexes. Part A, Inorganic Ligands. Pergamon Press, 

Oxford. 310 pp. 

Jones, C.Z., Rowe, T.G., Sexton, R.J., and Tanko, D.J., 1999, Water Resources Data, Nevada Water Year 1999:

USGS Water-Data Report NV-99-1, 569 p. 

Katzer, T.L., 1971, Reconnaisance bathymetric map and general hydrology of Lahontan Reservoir, Nevada: Water 

Resources - Information Series, Report 9, U.S. Geological Survey and U.S. Bureau of Reclamation. 

Krabbenhoft, D.P., Gilmour, C.C., Benoit, J.M., Babiarz, C.L., Andren, A.W., and Hurley, J.P., 1998, Methyl

mercury dynamics in littoral sediments of a temperate seepage lake: Canadian Journal of Fisheries and

Aquatic Sciences, v. 55, p. 835-844. 

Kuwabara, J.S. and Luther, G.W., III, 1993, Dissolved sulfides in the oxic water column of San Francisco Bay,

California: Estuaries, v. 16, p. 567-573. 

Kuwabara, J.S., Chang, C.C.Y., Khechfe, A.I. and Hunter, Y.R., 1996, Implications of dissolved sulfides and 

organic substances for the chemical speciation of trace contaminants in the water column of San Francisco 

Bay, California, in Hollibaugh, J.T., ed., San Francisco Bay: the Ecosystem: American Association for the 

Advancement of Science, Pacific Division, San Francisco, p. 157-172. 

Kuwabara, J.S., Topping, B.R. Coale, K.H. and Berelson, W.M., 1999, Processes Affecting the Benthic Flux of 

Trace Metals into the Water Column of San Francisco Bay, In Morganwalp, D.W., and Buxton, H.T., eds., 

U.S. Geological Survey Toxic Substances Hydrology Program--Proceedings of the Technical Meeting, 

Charleston, South Carolina, March 8-12, 1999--Volume 2--Contamination of Hydrologic Systems and 

Related Ecosystems: U.S. Geological Survey Water-Resources Investigations Report 99-4018B, p. 115-

119. 

15
Kuwabara, J.S., Berelson, W.M., Balistrieri, L.S., Woods, P.F., Topping, B.R., Steding, D. J., and Krabbenhoft,

D.P., 2000, Benthic flux of metals and nutrients into the water column of lake Coeur d'Alene, Idaho: 

Report of an August, 1999 Pilot Study: USGS Water-Resources Investigations Report 00-4132, 74 p. 

(Internet access at: http://water.usgs.gov/pubs/wri/wri004132/). 

Li, Y.H. and Gregory, S., 1974, Diffusion of ions in sea water and in deep-sea sediments: Geochimica et

Cosmochimica Act, v. 38, p. 703-714. 

Mantoura, R.F.C., Dixon, A., and Riley, J.P., 1978, The complexation of metals with humic materials in natural

waters: Estuarine and Coastal Marine Sciences, v. 6, p. 387-408. 

Marvin-DiPasquale, M., and Oremland, R.S., 2001, Methylmercury Formation and Degradation in Sediments of the 

Carson River System: U.S. Geological Survey Administrative Report to U.S. Enironmental Protection 

Agency, Region IX, 69 p. 

McKnight, D.M., Feder, G.L., Thurman, E.M., Wershaw, R.L. and Westall, J.C., 1983, Complexation of copper by

aquatic humic substances from different environments: Science of the Total Environment, v. 28, p. 65-76. 

Miller, D.C., 1984, Mechanical post-capture particle selection by suspension- and deposit-feeding Corophium: 

Journal of Experimental Biology and Ecology, v. 82, p. 59-76. 

Popp, A., and Hoagland, K.D., 1995, Changes in benthic community composition in response to reservoir aging: 

Hydrobiologia, v. 306, No. 1, p. 159-171. 

Praskins, W., Byron, E., Marvin-Dipasquale, M., Kuwabara, J., Diamond, M., and Gustin, M., 2001, Sampling and 

Analysis Plan: Mercury Dynamics in Lahontan Reservoir: U.S. Environmental Protection Agency, March 

22, 2001, 46 p. 

Preissler, A.M., Roach, G.A., Thomas, K.A., and Wilson, J.W., 1999, Water Resources Data, Nevada Water Year 

1998: USGS Water-Data Report NV-98-1, 598 p. 

Qian, J. and Mopper, K., 1996, Automated high-performance, high-temperature combustion total organic carbon 

analyzer: Analytical Chemistry, v. 68, p. 3090-3097. 

Ravichandran, M., Aiken, G.R., Reddy, M.M., and Ryan J.N., 1998, Enhanced dissolution of cinnabar (mercuric 

sulfide) by dissolved organic matter isolated from the Florida Everglades: Environmental Science and 

Technology, v. 32, p. 3305-3311. 

Reible, D.D., Popov, V., Valsaraj, K.T., Thibodeaux, L.J., Lin, F., Dikshit, M., Todaro, M.A. and Fleeger, J.W., 

1996, Contaminant fluxes from sediment due to tubificid oligochaete bioturbation: Water Research, v. 30, 

no. 3, p. 704-714. 

Thompson, J.K., Nichols, F.H., and Wienke, S.M., 1981, Distribution of benthic chlorophyll in San Francisco Bay,

California, February 1980 – February 1981: U.S. Geological Survey Open File Report 81-1134, 55 p. 

Topping, B.R., Kuwabara, J.S., Parchaso, F, Hager, S.W., Arnsberg, A.J., and Murphy, F., 2001, Benthic flux of 

dissolved nickel into the water column of South San Francisco Bay: U.S. Geological Survey Open-file 

Report 01-89, 50 p. (Internet access at: http://pubs.water.usgs.gov/ofr01089/). 

U.S. Geological Survey, 2001, National Water Information System (NWIS), Carson River at Fort Churchill, 

http://water.usgs.gov/nv/nwis (for Site 10312000). 

Wood, T.M., Baptista, A.M., Kuwabara, J.S. and Flegal, A.R., 1995, Diagnostic modeling of trace metal partitioning 

in south San Francisco Bay: Limnology and Oceanography, v. 40, p. 345-358. 

Woods, P.F., Nearman, M.J., and Barton, G.J., 1999, Quality Assurance Project Plan for U.S. Geological Survey 

Studies in Support of Spokane River Basin RI/FS.: U.S. Environmental Protection Agency, Seattle, 

Washington, and U.S. Geological Survey, Boise, Idaho, 153 p. 

Word, J.Q., 1980, Classification of benthic invertebrates into infaunal trophic index feeding groups, Bascom, W., 

ed., Biennial Report: Southern California Coastal Research Project, Long Beach, CA, p. 103-121. 

16
Acknowledgements 

 

Collaboration and interpretive discussions with M. Diamond and N. Gandhi are gratefully acknowledged. The 

authors also wish to thank M. Olson, J. DeWild, and S. Olund (USGS, Madison, WI) for mercury-speciation 

analyses, and A. Arnsberg and C. Lopez for benthic-chlorophyll analyses. Logistical and facilities support by 

Rangers K. Penner and P. Schmidt of the Nevada Division of State Parks, and C. Azevedo are much appreciated. 

We are also grateful for digital and analytical information from J. Gronberg, M. Lico, and L. Smith. Colleague 

reviews by R. Bouse and D. Vivit helped clarify a number of points made in this report. The U.S. Environmental

Protection Agency, Region 9 (Interagency Agreement DW14-95550801-0) is gratefully acknowledged for its

support of this work, as is the U.S. Geological Survey Toxic Substances Hydrology Program for long-term support

of this area of research. 

Product names are provided for identification purposes only and do not constitute endorsement by the U.S. 

Geological Survey. 

17
Appendix 1: Comments on the Report Structure 

A major objective of this electronic document is to provide a structure that is easily accessible to the wide range 

of interests in this work. Therefore, pathways within this document have been constructed to be both logical and 

intuitive. In addition to hyperlinks within the document to supporting figures and tables, links in Appendices 2 and 

3 provide a quick way to directly review and examine all figures and tables. 

Although hard copies of this report are available on request, the advantages of the electronic version relative to

the hard copy are substantial in many respects, but particularly in the rapid access of information at multiple levels 

of detail. 

Your comments about how to better serve readers with this evolving type of product are most welcome and may

be directed to any of the email addresses provided in the title page, but we request that at least a copy of all 

comments about the report and its format be addressed to the major author (kuwabara@usgs.gov) so that they may 

be compiled for future revisions and reports. 

18
Appendix 2: List of Figures 

Fig. 1 – Site map of the study area 

Fig. 2 – Photos of the coring operation 

Fig. 3 – Hydrological conditions during the study period 

Fig. 4 – Temperature and dissolved-oxygen trends in the water-column 

Fig. 5 – Comparison of the ranges for benthic indices 

Fig. 6 – Methyl-Hg and total-Hg flux comparisons 

Fig. 7 – Upstream remediation efforts potentially affect lake gradients 

Fig. 8 – Conceptual model of solute transport through a reservoir 

Fig. 9 – Processes regulating benthic flux of solutes 

Fig. 10 – Abundance of benthic macroinvertebrates 

Fig. 11 – Size variation of individuals retained by different sieve sizes 

Fig. 12 – Photo of macroinvertebrate taxonomy methods 

Fig. 13 – Photo of benthic-chlorophyll analyses 

Fig. 14 – Photo of dissolved-oxygen analyses 

Fig. 15 – Photo of dissolved-mercury analyses 

Fig. 16 – Photo of dissolved organic carbon (DOC) analyses 

Fig. 17 – Photo of dissolved-sulfide analyses 

19
 

Appendix 3: List of Tables 

Table 1 – Surficial-sediment porosities 

Table 2 – Benthic-invertebrate taxonomy for cores incubated May 1, 2001 (500-micrometer sieve) 

Table 3 – Benthic-invertebrate taxonomy for cores incubated July 19, 2001 (500-micrometer sieve) 

Table 4 – Benthic-invertebrate taxonomy for cores incubated July 19, 2001 (106-micrometer sieve) 

Table 5 – Benthic-chlorophyll analyses 

Table 6 – Dissolved-oxygen fluxes 

Table 7 – Bottom-water concentrations (picomolar, pM) for mercury species in Lahontan Reservoir. 

Table 8 – Dissolved-mercury fluxes 

Table 9 – Bottom-water dissolved organic carbon (DOC) concentrations (micromolar, uM) 

Table 10 – Dissolved organic carbon (DOC) fluxes 

Table 11 – Dissolved-sulfide fluxes 

 

20
Fig. 1. Fig. 1. Site map of the map of the study area study area 

including locations of sampling 

and coring sites in and coring sites in Lahontan Lahontan 

Reservoir.

Nevada
Fig. 2. Fig. 2. Photos of the coring op Photos of the coring operation eration showing the release 

showing the release 

metallic coring de tallic coring device (A), capp vice (A), capping of the core upo the core upoand rem and removal of the core tube (C core tube (C) in preparation fo

) in preparation fo

AB

Modif
Fig. 3. Fig. 3. Hydrologic conditio Hydrologic conditions during th ns during the study period, including 20

e study period, including 20

River, averaged historical dail River, averaged historical daily inflow since 1911, and rese

y inflow since 1911, and rese

(http://water. http://water.usgs.gov/nv/nwis).

. 

May 1 July 1 Sept. 1 

0

2040

Date

Flow (m Flow (m

3/s)

Storage

Average Inflow

2001Inflow

Boa

Sampling Trip
Fig. 4. Fig. 4. Temperature and disso ature and dissolved oxygen d oxygen (DO) concentration trend

(DO) concentration trend

between the May and July sam May and July sampling periods, with arrows re

periods, with arrows re

temporal change. oral change.

10.0

0.0

0

1020

DO (mg L-1)

Depth (m)

Northern LobeSouthern Lobe

Temperature (o

5 10 15 20

0

10

0
Fig. 5 Comparison of the ra of the ranges for bent es for benthic indices (benthic 

benthic 

macroin macroinvertebrate ertebrate abundance) de abundance) determined dur termined during th

ing th

to those reported for other to those reported for other oligotrophic trophic/mesotroph

mesotroph

0.1

10

1000

100000

Benthic Index

Lahontan Lahontan 2001

Published PublishedRanges

Chl-a (μg/cm2) Macroin

Macroi(individ
Fig. 6. Fig. 6. Methyl-Hg and Total Hg and Total-Hg

flux com flux comparisons between riverine riverine, atmospheric and heric and benthic sources and/or benthic sources and/or sinks. Positive flux sinks. Positive flux indicate indicates flux into flux into thewater column. water column.

1

100

10000

1

100

10000

Methyl-Hg

Total Hg

Riverine** Atmo

Source

Source

Flux (pmoles-Hg m

-2

h

-1

)

- 10000

- 100

- 1

Source: http://water.usgs.gov/nv/nwis

* Daily emission estimate

**

*** Source: Estimate from diel-flux study in July, 2001
Fig. 7. Fig. 7. Illustration sh Illustration showing how rem owing how remediation efforts in the forts in the river 

 river 

affect gradients in th affect gradients in the lake. A e lake. A regulated regulated decrease i crease in coin the water-column ( column (green arrow) ge arrow) generates an increas

an increas

across the sediment across the sediment-water interface resulting in water interface resulting in condit

enhanced benthic flux (yellow arrow) enhanced benthic flux (yellow arrow).

Incrgradflux
Modified 

Fig. 8. Fig. 8. Conceptual mo tual model of solute transport through a 

 transport through a 
Fig. 9. Fig. 9.

Modified 

Solute Transformation

and Repartitioning

Fig. 9. Fig. 9. Processes regulating the benthic flux of solutes

Processes regulating the benthic flux of solutes
Fig. 10. Fig. 10. Abundance o Abundance of benthic macroinverte macroinvertebrates at each of the th

at each of the th

sampling periods. The predominance of nematodes and tubif

tubif

Individuals/m

2
Fig. 11. Fig. 11. Photomicrographs showing size va icrographs showing size variation of individuals ret

individuals ret

Mature tubificidImmature tubificids Immature tu

Naidids

500 

500 μm sieved m sieved 100 μm sie

m sie

Ph

Ph
Fig. 12. Fig. 12. Photograph showing the process of Photograph showing the process of sorting and ide sorting and identifi

tifi

practicable taxono practicable taxonomic level.

Modified 

Modified 
Fig. 13. Fig. 13. Photograph showing spectroph spectrophotometric equipment 

benthic benthic-chlorophyll analyses. chlorophyll analyses.

Modified 

Modified 
Fig. 14. Fig. 14. Photograph showing Micro-Winkler Titration eq

Winkler Titration eq

dissolve dissolved-oxygen analyses during benthic oxygen analyses during benthic-flux e
Fig. 15. Fig. 15. Photograph of Cold-vapor Atomic Fluorescence Spectr

vapor Atomic Fluorescence Spectr

equipme equipment used for an nt used for analysis o alysis of dissolve dissolved-mercury spec

y spec

Modified 

Modified 
Fig. 16. Fig. 16. Photograph showing equipme raph showing equipment used for Dissolve ed for Dissolved Organic C

 Organic C

high-temperatu temperature non-catalytic oxidation. catalytic oxidation.

Modified 

Modified 
Fig. 17. Fig. 17. Photograph showing polarogr polarographic equipment use quipment used 

analyses analyses by square by square-wave voltammetry.

Modified 

Modified 
Table 1. Surficial-sediment Porosities for Lahontan Re

5/1/2001 

Site# Core# Porosity 

1 1 0.65 1 3 0.65 2 4 0.81 2 9 0.80 3 12 0.89 3 13 0.91 

7/19/2001 

Site# Core# Porosity 

1 1 0.73 1 3 0.70 2 6 0.80 2 7 0.85 3 10 0.93 3 11 0.92 
Table 2. Macro-invertebrate taxonomy for Lahontan core samples 5/1/01, 500 uM sieve 

Mean number of individuals per square meter based on core surface area of 77 square centimeters. 

Note:Species are listed below in the individuals/m2 

same general order as on the plots Site 1 Site 2 Site 3 

Core Number: 1 3 4 9 12 13 

Phylum Platyhelminthes 

Class Turbellaria 

Order Rhabdocoela 

Gyratrix sp. 

Phylum Nematoda 

Nematodes 

Phylum Annelida 

Class Oligochaeta 

Order Tubificida 

Family Naididae 

Amphichaeta leydigi Chaetogaster diastrophus Dero spp. (digitata and nivea ) 

cf. Pristinells acuminata 

Vejdovskyella comata Vejdovskyella intermedia 

Family Tubificidae 

390 260 260 130 130 390 260 260 

Limnodrilus hoffmeisteri 390 260immature tubificids 390 1039 779 519 130 260Tubifex tubifex 

Phylum Arthropoda 

Class Ostracoda 

Order Podocopida 

Ostracod 

Class Branchiopoda 

130 
Table 3. Macro-invertebrate taxonomy for Lahontan core samples 07/19/01, 500 uM sieve 

Mean number of individuals per square meter based on core surface area of 77 square centimeters. 

Note:Species are listed below in the individuals/m2 

same general order as on the plots Site 1 Site 2 Site 3 

Core Number: 1 3 6 7 10 11 

Phylum Platyhelminthes 

Class Turbellaria 

Order Rhabdocoela 

Gyratrix sp. 

Phylum Nematoda 

Nematodes 779 2208 260 1169 1039 260 

Phylum Annelida 

Class Oligochaeta 

Order Tubificida 

Family Naididae 

Amphichaeta leydigiChaetogaster diastrophusDero spp. (digitata and nivea ) 130 779 649 260

cf. Pristinells acuminataVejdovskyella comataVejdovskyella intermedia

Family Tubificidae 

Limnodrilus hoffmeisteri 260 260 

immature tubificids 649 1039 909 4675 390 260 Tubifex tubifex 

Phylum Arthropoda 

Class Ostracoda 

Order Podocopida 

Ostracod 

Class Branchiopoda 
Table 4. Macro-invertebrate taxonomy for Lahontan core samples 07/19/01, 106 uM sieve 

Mean number of individuals per square meter based on core surface area of 77 square centimeters. 

Note:Species are listed below in the individuals/m2 

same general order as on the plots Site 1 Site 2 Site 3 

Core Number: 1 3 6 7 10 11 

Phylum Platyhelminthes 

Class Turbellaria 

Order Rhabdocoela 

Gyratrix sp. 260 

Phylum Nematoda 

Nematodes 519 260 1429 130 6364 909 

Phylum Annelida 

Class Oligochaeta 

Order Tubificida 

Family Naididae 

Amphichaeta leydigi 

Chaetogaster diastrophus Dero spp. (digitata and nivea ) 

cf. Pristinells acuminata Vejdovskyella comata 

2208 1429 

130 

130 

130 

260 130 1169 130 

Vejdovskyella intermedia 260 130 

Family Tubificidae 

Limnodrilus hoffmeisteri 

immature tubificids 260 519 909 2338 649 130 

Tubifex tubifex 

Phylum Arthropoda 

Class Ostracoda 

Order Podocopida 

Ostracod 130 

Table 6. Dissolved-oxygen flux for Lahontan Reservoir, 2001 

May 1, 2001 

DO Flux (umoles/m2-h) 

Core# Site# Mean CI r2 n 

1 -843 +/- 2061 0.52 4 3 -1088 +/- 1995 0.66 4 4 -1072 +/- 416 0.98 4 

9 -585 +/- 1676 0.44 4 

12 3 -756 +/- 1470 0.63 4 

13 3 -1034 +/- 2308 0.56 4 

July 19, 2001 

DO Flux (umoles/m2-h) 

Core# Site# Mean CI r2 n 

1 -860 +/- 1928 0.56 4 3 -1268 +/- 2307 0.66 4 6 -1331 +/- 1942 0.75 4 7 -1149 +/- 1836 0.72 4 

10 3 -721 +/- 491 0.93 4 11 3 -245 +/- 1185 0.21 4 

1 1 2 2 1 1 2 2 

CI=95% confidence interval 
Table 7. Bottom-water concentrations for mercury species in Lahontan 

April 30, 2001 

Methy-Hg (pM) Total-Hg (pM) 

Site# Ind. Rep Site avg. S.D. Ind. Rep Site avg. S.D

1 0.7 0.9 0.3 647 671 3

1.1 695 

2 0.8 0.8 0.0 613 632 2

0.8 651 

3 0.4 0.4 0.0 373 375 

0.5 377 

July 18, 2001 

Methy-Hg (pM) Total-Hg (pM) 

Site# Ind. Rep Site avg. S.D. Ind. Rep Site avg. S.D

1 0.7 0.8 0.1 1907 1930 3

0.9 1952 

2 0.6 0.7 0.1 1622 1845 31

0.7 2068 

3 0.3 0.4 0.2 676 732 7

0.5 787 

pM=picomolarS.D.=Standard deviationInd. Rep. = Individual sample replicates
Table 8. Dissolved-mercury Fluxes for Lahontan Reservoir, 2001 

May 1, 2001 July 19, 200

Methyl-mercury flux Methyl-mercur

in pmoles/m2-h in ng/m2-day 

Core# Site# Site avg. Site avg. Core# 

1 1 -1.9 -9.3 1 3 1 1.7 -0.1 8.3 -0.5 3 

4 2 3.3 15.8 6 9 2 1.6 2.4 7.5 11.7 7 

12 3 1.8 8.5 10 13 3 2.9 2.3 13.9 11.2 11 

Total Mercury flux (End points only) Total Mercury 

in pmoles/m2-h in ng/m2-day 

Core# Site# Site avg. Site avg. Core# 

1 1 2113 10174 1 3 1 1362 1738 6554 8364 3 4 2 -1401 -6745 6 9 2 2030 314 9771 1513 7 

12 3 1358 6539 10 13 3 1266 1312 6094 6316 11 

Total Mercury (Entire time series) Total Mercury 

in pmoles/m2-h in ng/m2-day 

Core# Site# Mean CI Mean CI r2 n Core# 

1 1770 +/- 18345 8521 +/- 88312 0.06 4 1 

3 918 +/- 20810 4420 +/- 100177 0.01 4 3 4 -331 +/- 8091 -1594 +/- 38947 0.01 4 6 9 1560 +/- 16940 7512 +/- 81548 0.05 4 7 

12 3 1310 +/- 10515 6304 +/- 50620 0.09 4 10 13 3 1633 +/- 3740 7863 +/- 18005 0.06 4 11 

1 1 2 2 

CI=95% confidence interval 
Table 9. Bottom-water DOC Concentrations for L

Rep 

Dissolved Organic Carb

30-Apr-01 

DOC (uM) CI n 

Site 1 A 

B 

Site 2 A 

B 

Site 3 A 

B 

275.5 +/- 2.0 5 

275.6 +/- 0.7 3 

276.8 +/- 0.8 3 

282.0 +/- 0.7 4 

211.9 +/- 1.8 5 

207.0 +/- 0.3 3 

Rep= Individual sample replicate CI= 95% Confidence Interval 
Table 10. Dissolved Organic Carbon (DOC) Fluxes for Lahontan Res

May 1, 2001 

DOC Flux (umoles/m2-h) 

Core# Site# Mean CI r2 n Site avg. 

1 -19 +/- 17 0.05 14 23 3 64 +/- 11 0.49 16 

4 96 +/- 25 0.35 15 116 9 136 +/- 13 0.79 15 

12 3 89 +/- 15 0.47 17 66 

13 3 42 +/- 13 0.24 16 

July 19, 2001 

DOC Flux (umoles/m2-h) 

Core# Site# Mean CI r2 n Site avg. 

1 19 +/- 2 0.79 14 26 

3 32 +/- 2 0.90 17 

6 25 +/- 12 0.11 16 49 

7 72 +/- 14 0.53 14 

10 3 -9 +/- 10 0.04 14 -45 11 3 -80 +/- 18 0.45 14 

1 1 

2 2 

1 

1 

2 

2 

CI= 95% Confidence Interval 
Table 11. Dissolved Sulfide Flux and Diffusive Estimates for Lahontan Reserv

May 1, 2001 Bottom- 

Sulfide Flux (nmoles/m2-h) Porewater water 

Core# Site# Mean CI r2 n S-2 (uM) 

S-2 (nM) 

1 341 +/- 33 0.96 8 

3 497 +/- 40 0.96 9 0.70 1.1

4 500 +/- 29 0.98 8 

9 349 +/- 63 0.87 8 2.46 5.1

12 3 705 +/- 128 0.81 9 

13 3 527 +/- 69 0.89 9 2.55 5.2

July 19, 2001 Bottom- 

Sulfide Flux (nmoles/m2-h) Porewater water 

Core# Site# Mean CI r2 n S-2 (uM) 

S-2 (nM) 

1 778 +/- 133 0.88 8 

3 754 +/- 48 0.98 8 1.16 2.6

6 838 +/- 72 0.97 8 

7 630 +/- 36 0.98 8 1.53 3.2

10 3 594 +/- 41 0.97 9 11 3 870 +/- 32 0.99 9 0.51 1.7

1 1 2 2 1 1 2 2 

* Diffusive flux was based on 2cm diffusion layer thickness and 

a diffusion coefficient='0.00001 cm2 s-1 (Li and Gregory, 1974) 

CI= 95% Confidence Interval