Sediment-Water Interactions Affecting 

Dissolved-Mercury Distributions in 

Camp Far West Reservoir, California

Internet access at: http://pubs.water.usgs.gov/wri03-4140/

U.S. GEOLOGICAL SURVEY

Water Resources Investigations Report 03-4140

November 2002

U.S. Department of the Interior

Views from the Mid-Reservoir Site April 2002
Sediment-Water Interactions Affecting Dissolved-Mercury 

Distributions in Camp Far West Reservoir, California 

By James S. Kuwabara1

, Charles N. Alpers2, Mark Marvin-Dipasquale3, Brent R.Topping4, 

James L. Carter5

, A. Robin Stewart6, Steven V. Fend7, Francis Parchaso8, Gerald E. 

Moon9

 and David P. Krabbenhoft10 

U.S. GEOLOGICAL SURVEY 

Water-Resources Investigations Report 03-4140 

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

2 cnalpers@usgs.gov, U.S. Geological Survey, Sacramento, CA 

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

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

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

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

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

2
U.S. DEPARTMENT OF THE INTERIOR 

GALE A. NORTON, Secretary 

U.S. GEOLOGICAL SURVEY

CHARLES G. GROAT, Director 

For Additional Information 

Write to: 

Copies of this report may be 

obtained from the authors or 

James S. Kuwabara 

U.S. Geological Survey 

345 Middlefield Road, MS 439 

Menlo Park, CA 94025 

 

U.S. Geological Survey 

Information Center 

Box 25286, MS 517

Denver Federal Center 

Denver, CO 80225

3
TABLE OF CONTENTS

Executive Summary........................................................................................................................ 5 

Physical and Biological Characterizations.......................................................................... 6 

Chemical Characterizations................................................................................................ 7 

Remedial Implications........................................................................................................ 8 

Background..................................................................................................................................... 9 

Results and Discussion ................................................................................................................. 11 

Physical Data .................................................................................................................... 11 

Biological Data ................................................................................................................. 11 

Chemical Data................................................................................................................... 13 

Study Design and Methods........................................................................................................... 20 

Coring Operation .............................................................................................................. 20 

Core Incubations............................................................................................................... 20 

Physical Data .................................................................................................................... 21 

Biological Data ................................................................................................................. 21 

Chemical Parameters ........................................................................................................ 22 

References Cited ........................................................................................................................... 23 

Acknowledgments......................................................................................................................... 27 

Appendix 1: Comments on the Report Structure.......................................................................... 28 

Appendix 2: List of Figures.......................................................................................................... 29 

Appendix 3: List of Tables ........................................................................................................... 30

 

4
Sediment-Water Interactions Affecting Dissolved-Mercury 

Distributions in Camp Far West Reservoir, California 

by James S. Kuwabara, Charles N. Alpers, Mark Marvin-Dipasquale, Brent R. Topping, 

James L. Carter, A. Robin Stewart, Steven V. Fend, Francis Parchaso, 

Gerald E. Moon and David P. Krabbenhoft

Executive Summary 

Field and laboratory studies were conducted in April and November 2002 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 sampling locations within Camp Far West Reservoir, 

California: one near the Bear River inlet to the reservoir, a second at a mid-reservoir site of comparable depth to the 

inlet site, and the third at the deepest position in the reservoir near the dam (herein referred to as the inlet, midreservoir and near-dam sites, respectively; Background, Fig. 1). Because of interest in the effects of historic 

hydraulic mining and ore processing in the Sierra Nevada foothills just upstream of the reservoir, dissolved-mercury 

species and predominant ligands that often control the mercury speciation (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. 

Because of the affinity of mercury to adsorb onto particle surfaces and to form insoluble precipitates (particularly 

with sulfides), the mass transport of mercury in mining-affected watersheds is typically particle dominated. As 

these enriched particles accumulate at depositional sites such as reservoirs, benthic processes facilitate the 

repartitioning, transformation, and transport of mercury in dissolved, biologically reactive forms (dissolved methylmercury being the most bioavailable for trophic transfer). These are the forms of mercury examined in this study. 

In contrast to typical scientific manuscripts, this report is formatted in a pyramid-like structure to serve the 

needs of diverse groups who may be interested in reviewing or acquiring information at various levels of technical

detail (Appendix 1). The report enables quick transitions between the initial summary information (figuratively at 

the top of the pyramid) and the later details of methods or results (figuratively towards the base of the pyramid)

using hyperlinks to supporting figures and tables, and an electronically linked Table of Contents. 

During two sampling events, two replicate sediment cores (Coring methods; Fig. 2) from each of three 

reservoir locations (Fig. 1) were used in incubation experiments to provide flux estimates and benthic biological

characterizations. Incubation of these cores provided “snapshots” of solute flux across the sediment-water interface 

in the reservoir, under benthic, environmental conditions representative of the time and place of collection. 

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

to compare the results for mercury. The following major observations from interdependent physical, biological, and 

chemical data were made: 

5
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 rapid draw down of the reservoir during our work in 2002

(Fig. 3; U.S. Geological Survey, 2003a). Outflows are regulated to provide for agricultural needs 

during the summer months, but so are inflows (for example, through Rollins Dam up stream of the 

reservoir). Historical outflow from the reservoir (represented by Bear River discharge data at

Wheatland) before and after the dam construction in 1963 do not indicate marked differences 

(comparison of blue and green curves in Fig. 3) However, 2002 Wheatland discharge data

indicates two pulses of increased flow before our April sampling trip, and an early reduction in

discharge relative to historical averages following this trip. By the time of the second coring trip 

in November 2002, docking facilities for equipment and sample transfer were no longer available 

because of the 26-meter draw down (Fig. 4, see also the contrasting cover photographs). For

example, the water-column depths (in meters) at the coring sites on April 5, 2002 were 

approximately 45, 25, and 28 for the near-dam, mid-reservoir and inlet sites, respectively. By 

November, reservoir draw down caused the depth at the near-dam site to decrease to 19 meters, 

and restricted our coring efforts at the inlet and mid-reservoir sites to positions as close to the 

April 2002 locations that the research vessel could safely navigate (approximately 2-meter depth;

coordinates). 

2. Shifts in Redox Gradients: Dissolved-oxygen (DO) and temperature profiles during the coring trips 

indicate that an increase in bottom-water temperatures between the trips coincided with a decrease

in bottom-water dissolved-oxygen concentration at the deep near-dam site (Fig. 4). An oxycline 

was absent under more turbulent, high-flow spring conditions in April 2002. The water column

maintained a distinct oxycline in November 2002 at the near-dam site despite the impressive draw 

down described in item 1 above. 

 

3. Porosity: Porosity of the surficial sediments at the three coring sites was generally similar on both

sampling trips. Cores from the mid-reservoir site exhibited the most substantial change in porosity 

between April 2002 (0.88 and 0.94) and November 2002 (0.65 and 0.68). Like the mid-reservoir

site, elevated porosities were observed at the deep, near-dam site in April 2002 (0.96 and 0.98) 

relative to November (0.83 and 0.87). In contrast to the other two sites, the Bear River inlet site 

provided temporally consistent porosities (range of 0.84 to 0.90; Table 1). 

 

4. Benthic Biota: Macroinvertebrate populations in the Camp Far West samples were considered to be of

low importance to benthic flux (Table 2; Macrobenthos discussion). In other aquatic

environments, the presence of certain macroinvertebrate taxa may exponentially increase the 

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

and others, 1996; Kuwabara and others, 1999a). In an oligotrophic system like Camp Far West

Reservoir, however, densities of benthic macroinvertebrates are typically low relative to

mesotrophic and eutrophic environments that have elevated primary (food) production 

(Brinkhurst, 2002). Macroinvertebrate densities from the Camp Far West Reservoir sites were 

consistently highest at the near-dam site and appear to be similar to those from other oligotrophic 

sites where benthic-flux determinations have been previously reported (Kuwabara and others, 

2000; Kuwabara and others, 2001) (Fig. 5A). Benthic-chlorophyll concentrations mirrored the 

spatial pattern of macroinvertebrate densities during both coring trips with the highest 

concentrations at the near-dam site followed by the inlet site and the mid-reservoir site (Table 3; 

Fig. 5). 

6
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. Dissolved mercury in the water column: Elevated bottom-water concentrations relative to the surface 

at the near-dam site are consistent with a benthic source of solute. For methyl-mercury, elevated 

bottom-water concentrations relative to the surface were not observed, in part due to the overall 

tightness in the concentration range as mentioned below (Mercury results). 

Observed concentrations ranges for methyl-mercury (less than the detection of 0.20 picomolar to

0.50 picomolar, where picomolar units represent trillionths of a molar concentration) were tighter 

in magnitude than the range for total dissolved mercury concentrations (2.94 to 18.3 picomolar). 

The highest methyl-mercury concentrations in April 2002 were observed in the water column at 

the inlet site (0.30 + 0.01 picomolar), whereas in November, the highest concentrations were in the 

bottom waters of the deep, near-dam site (0.48 + 0.03 picomolar; Mercury results, Table 4). 

Dissolved total mercury concentrations in the bottom waters of Camp Far West were significantly 

lower under the low-flow conditions of November 2002 relative to high flow in April 2002, but

spatial trends were not temporally consistent. In terms of management implications, this contrast

suggests a potentially important shift between a predominance of upstream sources of methylmercury under high-flow conditions, and internal methyl-mercury sources under low-flow

conditions. 

2. Benthic flux of dissolved forms of mercury: Estimates of riverine and benthic flux indicate that

sediment sources of dissolved total mercury are seasonally variable, but consistently comparable 

to or greater in magnitude than riverine sources. Transport of dissolved, bioavailable mercury 

species between the reservoir bed and water column may therefore be a potentially critical process 

regulating the fate of mercury species in the reservoir water column (Mercury flux discussion). 

Benthic fluxes for dissolved methyl mercury were: (1) generally positive in April 2002, (2) 

negligible at all sites in November 2002, and (3) at least two orders of magnitude lower than totalmercury fluxes, roughly consistent with concentration differences between species. 

Similar to methyl-mercury fluxes, dissolved total-mercury fluxes, based on site averages from 2 

replicate cores, were generally higher in April than in November 2002 (Table 5; Fig. 6). Fluxes of 

36.3 and –28.4 picomoles per square meter per hour were measured in cores from the deep, neardam site which was suboxic in November, contrasting with positive fluxes of 306 and 272

picomoles per square meter per hour at that site in April 2002 when the bottom water was oxic 

(Fig. 4). All six measurements of total-mercury flux in April 2002 and five of six in November

2002 resulted in positive values (That is, out of the sediment into the overlying water column). 

3. Benthic flux of nutrients, and mercury-binding ligands: The reservoir was highly phosphate limited 

with molar N:P ratios in water column samples ranging from 136 to >5,000 (compared to the 

Redfield N:P molar ratio of 16; Table 6). Minor (that is, submicromolar) changes in water-column 

concentrations can significantly alter the balance of limiting nutrients in an oligotrophic system

like Camp Far West Reservoir. Because the flux of orthophosphate from bed sediments may be 

regulated by processes that control its adsorption and desorption onto metal-oxide surfaces 

(nutrient discussion; Kuwabara and others, 2003), sediment-water interactions may constrain 

orthophosphate flux relative to dissolved nitrogen species and thereby exacerbate phosphatelimiting conditions in the reservoir. 

7
Sulfide benthic fluxes were predictably highest (1,180 + 50 nanomoles per square meter per hour; 

Table 7) at the deep, near-dam site in November 2002 following at least 2 weeks of hypolimnetic

anoxia (based on prior water-column monitoring), and lowest at the same near-dam site in April 

(290 + 30 nanomoles per square meter per hour) under high-flow, oxic conditions. 

Consistent with sulfide fluxes, the highest benthic fluxes of dissolved organic carbon (DOC) were

also observed at the deep, near-dam site in November 2002 (520 + 50 micromoles per square 

meter per hour; Table 8), and were consistently negative (consumed by the sediment) at the midreservoir site. 

Remedial Implications 

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

important management implication. Remediation efforts and Total Maximum Daily Load (TMDL) allocations

along the Bear River have dual objectives of decreasing concentrations and loads to down-gradient systems and

reducing bioaccumulation of mercury in fish consumed by humans and wildlife. Using preliminary mercury-flux

estimates into the reservoir from the Bear River, our results indicate that with considerable seasonal variability a 

significant (and possibly predominant) percentage of dissolved mercury in the water column presently comes from

the reservoir bed (Mercury flux discussion). If upstream sources are controlled, which is desirable even apart from

reservoir effects, the present inflow loads are likely to be compensated in part by increases in benthic flux (Fig. 7). 

Concentrations of total mercury in the reservoir should be expected to decrease slowly after inflow loads are 

controlled. A previously developed management tool (Kuwabara and others, 2003) was applied to estimate a 

benthic response to regulating the riverine inputs of contaminants to the reservoir. 

8
Background 

Why may sediment-water interactions be important factors affecting mercury distributions in 

Camp Far West Reservoir?

Many fundamental processes affect the transport of dissolved chemical species (for example, nutrients, 

metals, 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 in a number of 

previous studies (for example, advective transport and point source inputs; Fisher and others, 1979). 

There are, however, terms in the conceptual model that have until recently 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 or other reservoirs in the Sierra Nevada. 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, 2002; Kuwabara and others, 2003). 

As a result of physical, chemical, and biological processes operating near the sediment-water interface, 

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

others, 2000; Fig. 9). Associated gradients in solute concentrations can thereby result in a benthic flux of that solute

that may be negative or consumed by the sediment, positive or released from the sediment, or insignificant when 

the gradient is indistinguishable. 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 sedimentwater interface only when another solute is depleted. Redox-sensitive solutes like dissolved iron often behaves in

this manner when suboxic conditions reduce it from ferric to ferrous forms, increasing its solubility and releasing

adsorbates. Macrofauna can also enhance benthic flux by irrigating, or perturbing surficial sediment layers 

(bioirrigation, bioturbation, or biologically enhanced advection). Certain productive benthic communities can 

enhance benthic flux by orders of magnitude beyond diffusive-control (Kuwabara and others, 1999a, Thibodeaux

and Bierman, 2003), while other communities are too sparse to generate this magnitude of enhancement (Kuwabara 

and others, 2003). Therefore, vertical chemical gradients generated by a variety of interdependent biogeochemical

processes can induce the transport of dissolved-mercury species across the sediment-water interface in lentic 

systems. 

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. Unfortunately, benthic-flux determinations are instrument and

manpower intensive, because each flux estimate requires a concentration time-series analysis or vertical-profile 

analysis. Benthic-flux studies have consequently lagged behind studies of other transport processes, but are spurring

interest in both the scientific and management communties (Kuwabara and others, 2003). 

Water-quality managers often assess and prioritize remediation strategies for aquatic systems that have

been adversely affected by anthropogenic activities. In the case of Camp Far West Reservoir along the Bear River, 

mercury associated with historic gold extraction has been fluvially transported and accumulated in the bottom

sediments. Frequent demands have been made by regional 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, 1999a). As part of a comprehensive examination of transport processes affecting mercury

dynamics in Camp Far West Reservoir, this study focuses on a poorly understood, yet potentially predominant, 

source of mercury to the reservoir water column: 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 (for example, advection and diffusion), chemical (that is, oxidation-reduction reactions,

9
complexation and repartitioning) and biological processes (for example, bioirrigation and bioturbation)( 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 California State Water Resources Control Board to provide initial determinations of dissolved total 

and methyl-mercury fluxes from the sediments into the water column of Camp Far West Reservoir. Recent mercury

distribution and bioaccumulation studies at Camp Far West Reservoir and other reservoirs in the Bear River and 

Yuba River watersheds suggested the potential importance of sediment-water interactions in describing mercury

speciation, transport and bioaccumulation (May and others, 2000). 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 appropriate water-quality models and remedial programs for this mining-affected system. 

With a wide a range of programmatic decisions to be made related to the improvement of water and 

fisheries resources in the Bear River, the question posed in this study was, “Are sources and sinks of dissolved total 

mercury and dissolved methyl-mercury associated with the bottom sediment within Camp Far West Reservoir 

significant relative to major surface-water inputs from the Bear River?” Various factors motivate this question. 

First, gold extraction upstream of the reservoir during the mid-nineteenth century provide a historic source of 

elemental mercury that continues to be fluvially transported in dissolved and particulate phases (Alpers and

Hunerlach, 2000). Therefore, determining whether some fraction of this sediment-associated mercury can 

remobilize for transport to the overlying water and subsequently to downstream portions of the Bear River is of

considerable environmental and ecological significance. 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, in pH, and in nutrient availability near the sediment-water interface (for example, 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; Tessier and others, 

1994). Finally, there is a growing body of evidence from other aquatic systems that benthic flux 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). Thus, the need clearly exists for more refined conceptual and 

numerical models describing mercury dynamics within the Bear River system. 

10
Results and Discussion 

Physical Data

Visually, there were general similarities in sediment texture among the three sampling sites. At all

sites, the cores were characterized by a bright orange (iron-oxide like), unconsolidated surficial 

layer overlying a dark gray to black mud. That surficial layer was most prominent in the April 

2002 sample and had a thickness greater than 2 centimeters. Given that core penetration depths 

were only approximately 10 centimeters to provide core volume for overlying water, sedimentcore retrieval was facilitated by the compacted silts and clays below the surficial layers 

(Unconsolidated or sandy material below surficial layers tend to cause a greater frequency of core 

material loss during retrieval). Porosity of the bottom sediment ranged from 0.65 at the midreservoir site in November 2002 to 0.98 at the near-dam site in April 2002 (Table 1). The lowest 

porosities (0.65 and 0.68) at the mid-reservoir site in November were coincident with the gradual 

filling of the reservoir after significant summer draw down. This generated rewetted, fractured

bottom sediment reminiscent of the summer-exposed, compacted bank sediment. 

Biological Data

1. Benthic macroinvertebrates: In certain aquatic environments, bioturbation and bioirrigation by benthic 

macroinvertebrates can significantly affect the benthic flux of solutes (Charbonneau and Hare, 

1997; Kuwabara and others, 1999a). At Camp Far West Reservoir, the benthic macroinvertebrates

were spatially and temporally dominated by immature tubificid worms (Table 2). Typically, 

immature tubificids are not quantitatively retained by 500-micrometer sieving. Reible and others 

(1996) reported that the flux of hydrophobic organics can be significantly increased when large, 

laboratory-cultured tubificids are present at densities between 6,000 to 27,000 per square meter. 

Not only were the densities of tubificids in Camp Far West Reservoir cores typically much lower, 

but they were also dominated by smaller, immature individuals (Table 2). Biologically enhanced

benthic flux through bioirrigation or bioturbation would not be expected from such low biomass of

the benthic assemblage at the time of coring. 

The average number of individuals in the benthic assemblage was similar between sampling dates 

in April 2002 and November 2002 (2,400+1,100 and 2,500+2,200 individuals per square meter, 

respectively). However, there were significant differences in the densities between sites (Fig. 5A). 

On both sampling dates, the lowest macroinvertebrate densities were found in the mid-reservoir 

sites, followed by Bear River inlet site, and highest densities at the deep, near-dam site. It is 

notable that the highest macroinvertebrate densities were observed at the near-dam site in

November 2002 when the highest benthic chlorophylls were also measured, but also when the

lowest dissolved-oxygen concentrations were present following at least 2 weeks of hypolimnetic 

anoxia. Immature tubificids, like those that dominated the benthic assemblage (Table 2), have

demonstrated tolerance to extended exposures (of the order of months) of suboxic and anoxic 

conditions (Jenderedjian, 1994; Rieradevall and Real, 1994; Smith, 2001). 

In summary, the low biomass of the benthic community at the coring sites suggest that there were 

insufficient biota to generate biologically enhanced benthic flux through bioirrigation or 

bioturbation. Macroinvertebrate communities were dominated by very small immature tubificid

worms and nematodes that typically pass through a conventional 500-micrometer mesh sieve. 

This is consistent with low benthic-chlorophyll concentrations particularly at the inlet and midreservoir sites (see section below and Fig. 5B), and with sulfide-flux determinations. However,

other lines of evidence such as bromide injections or radon emanation would be required to

confirm that bioturbation effects in the reservoir are negligible (Kuwabara and others, 2003). 

 

11
2. Benthic chlorophyll-a: Chlorophyll-a concentrations in surficial sediments ranged from 0.3 to 3.2

micrograms per square centimeter with highest concentrations observed at the near-dam site

during the fall sampling (November 2002), and the lowest concentrations at the mid-reservoir site 

during the spring sampling (April 2002) (Table 3). Benthic biological data from our study and 

previously published data from other oligotrophic lakes (Fig. 5B; Kuwabara and others, 2002;

Kuwabara and others, 2003) clearly indicated that the benthic algal biomass was sparse at all three 

coring sites and thus indicative of oligotrophic systems. Benthic chlorophylls in Camp Far West 

Reservoir (overall mean of 1.45 + 0.86 micrograms per square centimeter; Fig. 5B) are 

representative of low concentrations expected in an oligotrophic environment (Fig. 5B). The 

spatial distributions, coincident with those for macroinvertebrate densities, suggest a food

limitation in trophic pathways. The concentration ratio of chlorophyll-a to chlorophyll-a plus

phaeo-pigments (CCP ratio) is used as a crude indicator of the extent of senescence in an algal 

population (Thompson and others, 1981; Kuwabara and others, 1990). Ratios greater than 0.5 

indicate active growth, and values less than 0.5 suggest enhanced pigment degradation associated 

with cell senescence or predation pressures that process the chlorophyll-a to degradation products. 

The overall benthic CCP ratio for this study was 0.12 + 0.04 (n=36), indicating a predominance of 

degraded or biologically processed chlorophyll-a. 

Like Camp Far West Reservoir, sediment quality in San Francisco Bay is also affected by historic 

mining and ore processing in the Sierra Nevada (Jaffe and others, 1998). However unlike Camp

Far West Reservoir, the benthic CCP ratio is much more variable in San Francisco Bay sediments 

(0.20 + 0.19; n=414) because ratios greater than 0.3 and as high as 0.9 are associated with 

phytoplankton bloom periods in the northern and southern components of the estuary (Thompson

and others, 1981). Compared to benthic measurements, bottom waters of San Francisco Bay

display higher CCP ratios (0.50 + 0.14; n=330 from January 1999; water-quality database query at 

http://sfbay.wr.usgs.gov). The consistently low benthic CCP ratio at Camp Far West suggests a 

consistently senescent benthic community and therefore perhaps a primary planktonic source at

least during our sampling periods. 

3. Surface-water chlorophyll-a: Chlorophyll-a concentrations in the surface water (1-meter depth) ranged 

seasonally from low values of 1.6 to 5.0 micrograms per liter (μg-L-1) in the spring and summer 

(April and August 2002) to the higher values of 4.4 to 30.4 μg-L-1 in the fall and winter (October 

2001, February and November 2002, January 2003) (Table 9, Figure 10A). Marked differences

among stations within the reservoir were observed in the fall of 2001 and 2002 and in January of

2003 with chlorophyll-a concentrations significantly decreasing from the inlet to the near-dam

sites, except for January 2003 when concentrations were lowest at the inlet. These chlorophyll-a

concentrations are similar to those found in other reservoirs in California and Canada (Paterson 

and others, 1997; U.S. Geological Survey, 2003b) and surface waters of San Francisco Bay/Delta. 

Pheophytin pigments, which are markers of chlorophyll degradation, followed a similar seasonal 

cycle (Table 9, Figure 10B). Ingestion of algal cells by zooplankton and microherbivores results 

in the decomposition of chlorophyll pigments and the formation of pheophytins (Wetzel, 2001). 

Planktonic CCP ratios, which have been used as a coarse indicator of grazing by zooplankton, 

were consistently higher than observed in the benthic-chlorophyll samples (refer to previous 

section), and increased from 0.75 in the fall of 2001 to 0.90 in April of 2002, then declined to 0.65

in August of 2002 and remained relatively constant through January 2003 (Table 9, Figure 10C). 

These elevated water-column CCP ratios indicate that zooplankton grazing was not a significant 

removal mechanism for chlorophyll from the water column. Furthermore, it does not appear that 

the zooplankton community exerted significant grazing pressure on the phytoplankton stock.

12
4. Stable isotope and elemental composition of suspended particulate matter (SPM): Carbon isotopic 

ratios (δ13C) of the SPM were similar among stations and ranged from –34 ‰ in February 2002 to

–28 ‰ in April 2002 (Table 9, Figure 11A). Nitrogen isotopic ratios (δ15N) of the SPM were more 

variable among stations and ranged from a high of 10.64 in February 2002 to a low of 1.94 in

April 2002 (Table 9, Figure 11B). These patterns of enriched nitrogen isotopic values coincident

with depleted carbon isotopic values are typical of the water column in freshwater lakes (Vander 

Zanden and Rasmussen, 1999). The combination of surface-water chlorophyll-a concentrations 

and elemental and isotopic composition of the SPM suggests that the water column of Camp Far 

West Reservoir is a dominant source of production for the pelagic food web and possibly the 

benthic food web. 

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) benthic flux: Dissolved-oxygen fluxes were consistently negative at the deep, 

near-dam site, indicating benthic consumption, but were variable in direction at the shallower

upstream sites. DO flux ranged from –503 micromoles per square meter per hour at the inlet site

in April 2002 (average of two measurements, n=8) to 306 micromoles per square meter per hour at

the mid-reservoir site on the same date (Table 10). A comparison of average dissolved-oxygen 

flux measurements for the three sites indicates that the mid-reservoir site was the only site to

display positive fluxes in one of two cores in both April and November 2002. No temporal or

spatial variations in oxygen consumption at the inlet or the deep, near-dam sites were observed. 

With higher bottom-water temperatures (Fig. 4) and lower dissolved oxygen at the near-dam site 

in November 2002, significantly higher oxygen consumption at this site was expected relative to

April 2002, but this was not observed (Table 10). A partial mixing of the water column and 

transition from hypolimnetic anoxia to suboxia within 2 to 3 weeks of the November 2002

sampling may explain this result. A temporal comparison of water-column profiles for DO and

temperature at the inlet and mid-reservoir sites is not applicable here because of the reservoir draw

down. At these shallower depths (approximately 2-meter depth) the temperature and DO profiles 

were uniform. As with mercury-flux measurements, 95-percent confidence intervals were often

greater than the mean flux value despite acceptable coefficients of determination because only a 

single analytical replicate from each of four time intervals was used in the linear-regression for 

flux determinations and hence the confidence interval is based on only two degrees of freedom

(confidence interval discussion). More extensive follow up studies would warrant greater 

analytical replication, including at least three core replicates per site to better quantify small-scale 

(within site) variability. 

2. Dissolved mercury in reservoir bottom waters: At all three sampling sites, higher dissolved totalmercury concentrations (Table 4; Fig. 12) were observed in April (three-site mean: 16.7 + 2.0 

picomolar {n=11}) relative to November (4.6 + 2.2 {n=10}). These values include initial 

concentrations in the core overlying water. This difference may reflect the insolubility of mercury 

in the presence of elevated sulfide concentrations associated with the hypolimnetic anoxia and 

suboxia in November. This temporal trend was not observed for methyl-mercury in the bottom

waters with overall means for April and November of 0.24 + 0.09 (n=12) and 0.28 + 0.13

picomolar (n=11), respectively. In April, 2002, the inlet site displayed the highest methyl-mercury 

concentrations for bottom waters (0.33 + 0.06 picomolar; n=4) relative to the other two

downstream sites (0.19 + 0.06 picomolar, n=8). In contrast, November 2002 results indicate the 

13
highest methyl-mercury concentrations in the reservoir bottom waters at the deep, near-dam site 

(0.48 + 0.03 picomolar, n=3) relative to the two other upstream sites (all replicates <0.20 

picomolar, n=8). This contrast and the negligible benthic fluxes for methyl-mercury at all three 

sites suggest a possible switch between the dominance of allochthonous (external or upstream) 

methyl-mercury sources during high flow and autochthonous (internal including) sources during

low flow. By comparison, higher concentrations for dissolved mercury species were observed in

the bottom waters of Lahontan Reservoir in Nevada, another reservoir affected by historic gold 

mining (range of methyl-mercury from 0.3 to 1.1 picomolar and of total dissolved mercury from

613 to 2,068 picomolar; Kuwabara and others, 2002). 

Although available concentration data for dissolved total mercury near the surface of Camp Far 

West Reservoir are limited, near-surface concentrations were routinely lower than bottom-water 

concentrations. In April 2002, near-surface (3-meter depth) and bottom-water concentrations at

the near-dam site for dissolved total mercury were 4.5 picomolar (with an analytical error of + 0.5 

picomolar) and 17.5 + 1.2 picomolar (n=2, Table 4), respectively. Elevated bottom-water 

concentrations relative to the surface is consistent with a benthic source of solute. Vertical 

gradients for methyl-mercury in the water column were indistinguishable. 

3. Mercury benthic flux: Dissolved total-mercury fluxes were positive (that is, out of the bed sediment) in

all six incubated cores in April 2002 and five of six cores in November 2002. Dissolved totalmercury fluxes were greatest at the mid-reservoir site in April 2002. The lowest benthic fluxes

were at the deep, near-dam site in November, coincidently when the highest sulfide fluxes were 

observed (Table 5). Given lower solubility of mercury in the presence of sulfides (Hogfeldt, 

1983), a redox barrier to dissolved-mercury benthic flux may have been generated. 

In contrast to total-dissolved mercury, none of the three sites generated a significant benthic flux

for methyl-Hg in November, and in April 2002, the inlet site actually generated a negative flux

(Table 5). Dissolved methyl-mercury fluxes were highest at the mid-reservoir site in April 2002

(0.6 + 0.1 picomoles per square meter per hour). It should be noted that the imposed shift in the 

sampling locations for the inlet and mid-reservoir sites (that is, the effects of the draw down) to

some extent confound temporal with spatial comparisons at those two upstream sampling

locations. Variability in microbial activity and transformation rates for mercury species at these 

sites may also contribute to observed differences. Measurement of potential mercury

transformation (methylation and demethylation) rates were incorporated into the overall study of

this reservoir but are not yet available. For both sampling dates, the magnitude of the methylmercury fluxes were at least two orders of magnitude less than the flux estimates for total 

mercury, roughly consistent with concentration differences between species. 

A summary of mercury-flux data from Camp Far West Reservoir and Lahontan Reservoir along 

the Carson River (Fig. 6), illustrates that flux estimates not only vary over orders of magnitude, 

but also vary in direction where a positive flux represents 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 benthic-flux over the projected area of 

Camp Far West Reservoir for the two sampling dates (2.5 and 1.3 square kilometers in April and 

November 2002, respectively), and ranged between –1.0 to 0.5 and 3.9 to 300 picomoles mercury 

per square meter per hour for methyl and total mercury, respectively, using the mean of two

replicate cores per site. This corresponds to a mass range of –5 to 3 and 19 to 1,440 nanograms 

mercury per square meter per day for methyl and total mercury, respectively. Riverine fluxes into

the reservoir for the two sampling dates were based on stream-flow data available on the Internet

up to October 2002 (USGS, 2003) and provisional data from the Nevada Irrigation District for 

October through November 2002, the first two months of the 2003 Water Year. In addition, 

dissolved total and methyl-mercury data near the surface (3-meter depth) from the inlet site was

14
used as an estimate of the inflow concentration. For April 24 and November 4, 2002, the riverine 

fluxes were estimated at 0.9 and 0.4 nanograms methyl mercury per square meter per hour, and 52

and 36 nanograms total dissolved mercury per square meter per hour, respectively. Thus, based on 

our initial benthic and riverine flux estimates for dissolved mercury species, the reservoir bed

appears to represent a significant source of dissolved mercury to the water column, but that 

importance is at least seasonally and spatially variable. An analysis of total-mercury and methylmercury fluxes into and out of Camp Far West Reservoir, based on monthly water-quality 

sampling during water years 2000 to 2003 will be published separately. Additional work to

quantify flux variability over various spatial and temporal scales would be recommended in order 

for this benthic source to be quantitatively incorporated into mercury-transport, water-quality 

models for the reservoir. 

Gustin (cited in Kuwabara and others, 2002) estimated an elemental-mercury emission from

Lahontan Reservoir, Nevada, within the Carson River SuperFund site, at 4 nanograms mercury per 

square meter per hour. Corresponding mercury emissions have not been made at Camp Far West

Reservoir, but given that total dissolved-mercury concentrations in the water column of Camp Far

West Reservoir are two orders of magnitude lower than Lahontan Reservoir, and the surface area

is an order of magnitude lower than Lahontan Reservoir (20 to 40 square kilometers; Katzer, 

1971), it is reasonable to speculate that mercury release to the atmosphere from Camp Far West 

Reservoir is also lower than from Lahontan Reservoir in units of flux or load. On the basis of our 

current comparative information, it would seem prudent to consider benthic processes when trying

to quantitatively describe mercury-transport processes within and through this reservoir. 

Errors presented in the text and each of the tables represent 95-percent confidence intervals about

the mean value unless otherwise specified. In the case of the dissolved-mercury, nutrient and

oxygen fluxes, elevated confidence intervals relative to the flux estimate seem inconsistent with 

the goodness of fit to the linear regression (only one core with a coefficient of determination < 

0.79 for total-mercury flux). This inconsistency is due to the limitation of a single, qualitycontrolled concentration from each of four time intervals being used for the linear regression to 

determine benthic flux from each core. Therefore, only two degrees of freedom remain after the

slope and mean are determined. The confidence interval is therefore elevated above the standard 

deviation because a value from the t-distribution in the numerator increases nonlinearly as the 

degrees of freedom decrease. Because these results represent the first estimates of mercury 

benthic flux from Camp Far West Reservoir, cost efficiency in replication of mercury analytical

data relative to spatial and temporal coverage was a primary consideration. In future studies, 

greater replication of mercury-speciation data at each time interval during the core incubation 

would be required to decrease the confidence limits. However, the high coefficients of

determination and general consistency between cores from this study reflect positively on the 

reported fluxes. 

 

4. Dissolved trace metals in the water column (Application of a management tool): Increased

concentrations at depth relative to surface waters may indicate a positive flux of a solute from the 

sediment into the water column. Dissolved-metal concentrations for cadmium (Cd), cobalt (Co), 

copper (Cu), iron (Fe), lead (Pb), manganese (Mn), nickel (Ni), vanadium (V), and zinc (Zn) in the 

water column were determined as ancillary parameters to investigate the potential for significant 

benthic flux (Fig. 13, Table 11). As previously discussed, the shallow November 2002 depths

precluded any water-column profiling for the mid-reservoir and inlet stations, and thus any 

temporal comparisons for these two sites. At the near-dam site, concentrations for most metals 

increased throughout the water column between the sampling dates. This general increase may be

explained by: (1) evaporation leading to concentration of the solutes, or (2) seasonal changes in

inflow controlled upstream (for example, at Rollins Dam). However, the variability in the 

magnitude of the increases between individual metals suggests that repartitioning dynamics (for

15
example anoxia-induced reducing conditions) might be controlling some concentrations. For Zn

and Pb, bottom water concentrations in November were an order of magnitude greater than bottom

water concentrations in April (Table 11). Also, the ratio of bottom water to surface water 

concentrations (labeled “temporal ratio” in Table 11) rose significantly (>8 times) for Zn and Pb

between sampling dates. Considering the near-anoxic bottom water at the near-dam site in

November, it is likely that reducing conditions caused these trends for Zn and Pb. Some metals 

(Cd, V, Fe) showed smaller increases in this ratio (~1.5 times), while others (Cu and Ni) showed

temporal decreases in both this ratio and bottom-water concentrations. Also, the November data 

shows that the maximum concentration for most metals was observed at the depth just above the 

oxycline. These data indicate that further water-column profiling and benthic flux experiments 

would be needed to determine the potential significance of benthic flux in determining trace metal 

distributions in the reservoir. 

The consistent increase in dissolved-zinc concentrations near the sediment-water interface of

Camp Far West Reservoir (as mentioned above, and similar to those observed for dissolvedmercury species) suggests a benthic source of zinc and poses an interesting management

implication. Assuming that the observed water-column gradients for dissolved zinc resulted from

diffusion from the reservoir-bed, a benthic flux for zinc can be estimated for April 2002, in the 

absence of an oxycline, and for November 2002, under dissolved-oxygen stratification. The zinc

benthic flux (0.04 and 21 micromoles per square meter per hour, respectively; Table 12) is based 

on the eddy diffusivity of zinc near the reservoir bed, as reflected in the zinc concentrations of the

lentic water column (Kuwabara and others, 1999b). Applying those estimates for zinc benthic 

flux with an assumed diffusion coefficient (1x10-5 square centimeters per second; Li and Gregory,

1974) and a diffusion-layer thickness of 2 centimeters (Kuwabara and others, 2002), pore-water

zinc concentrations for April and November can be “back calculated” (0.25 and 120 micromolar, 

respectively; Table 12) using the standard expression for diffusion as described and applied by

Berelson and others (1982). Kuwabara and others (2003) described a management tool to

examine the effects of regulating upstream fluvial loads of contaminants. If the bottom-water and

estimated pore-water concentrations for dissolved zinc are plotted relative to other mining-affected

sites that have been previously examined, it would appear, at least for this zinc example, that April 

2002 results would not motivate upstream remedial action because both pore-water and bottomwater concentrations were low relative to the other lake and reservoir environments. The 

November 2002 results highlight the need to consider benthic flux of contaminants under these 

conditions in order to appropriately evaluate the effectiveness of upstream remediation, because 

benthic flux remains a dominant transport process with or without implemented remediation (Fig. 

14). Although desirable, data are not currently available to make analogous site comparisons for 

mercury to examine the indirect effects of upstream remediation. 

5. Nutrients in the water-column (The concept of a limiting nutrient): In an oligotrophic system like

Camp Far West Reservoir, ostensibly subtle shifts in solute concentrations can alter the balance of 

nutrients that regulate phytoplankton growth (that is, primary production in the water column). 

The molar nitrogen to phosphorus (N:P) ratio for assimilation by primary producers like 

phytoplankton is approximately 16 (often referred to as the Redfield N:P ratio; Wetzel, 2001). 

Dissolved-nutrient concentrations in the water column strongly indicate phosphate limiting

conditions in the reservoir with dissolved N:P molar ratios ranging from 136 to 5,191, with nine of

ten values of the ratio at least an order of magnitude greater than the Redfield ratio (Table 6). 

Molar carbon to nitrogen (C:N) ratios of the suspended particulate matter declined from

approximately 8 in October 2001 to 6.4 in February of 2002 and increased back to 8 in April of

2002 (Table 9, Figure 11C) following seasonal trends in phytoplankton abundance. These C:N

molar ratios are typical of phytoplankton-dominated suspended particulate matter (that is, similar

to the Redfield C:N molar ratio of 6.7) and consistent with the notion that neither carbon nor 

nitrogen was limiting phytoplankton growth in the reservoir. Furthermore, Wetzel (2001)

16
summarized that a molar C:N ratios less than 8.3 indicate no nitrogen limitation, an observation

that supports measured phosphate-limiting conditions in the reservoir water column (Table 6). 

Because orthophosphate has a high affinity to repartition to metal-oxide surfaces and algal cells 

(Goldberg, 1985; Kuwabara and others, 1986), one might hypothesize that sediment-water

interactions may play a significant role in regulating these nutrient limitations (see item 5 below). 

As reducing conditions facilitate the reduction of iron from ferric to ferrous form which has 

greater solubility, the dissolution of hydrous iron oxides releases adsorbed orthophosphate. 

During our two sampling dates, suboxic bottom-water conditions were only observed at the deep, 

near-dam site in November 2002. It is therefore notable that for that location and time, the oxic 

surface-water N:P molar ratio was at the highest observed value of 5,191, but the near-anoxic 

bottom water (approximately 1 milligram dissolved oxygen per liter), displayed an N:P molar ratio

that was an order of magnitude lower at 569. This order-of-magnitude difference above and below 

the oxycline was generated by increased bottom-water orthophosphate, not a water-column

gradient in dissolved nitrogen concentration (Table 6). 

6. Nutrient benthic flux: Forms of dissolved inorganic nitrogen (ammonia, nitrate, and nitrite) had a 

consistently positive flux in the reservoir (i.e., out of the sediment into the overlying water 

column) except at the mid-reservoir site in April 2002 (Table 6). The overall average for 

dissolved inorganic nitrogen flux for April and November 2002 was 74 and 143 μmoles-m-2-h-1, 

respectively. The corresponding averages for orthophosphate for April and November 2002 were

–0.16 and 0.29 μmoles-m-2-h-1. These average fluxes generate N:P ratios of –470 and 486 for 

April and November 2002, respectively (both conditions lead to absolute values that are an order 

of magnitude greater than the Redfield ratio described above). The flux of orthophosphate from

bed sediments may be regulated by processes that control its adsorption and desorption onto

metal-oxide surfaces (Kuwabara and others, 2003). For both April and November, sediment-water 

interactions may therefore constrain orthophosphate flux relative to dissolved nitrogen species and

thereby exacerbate phosphate-limiting conditions in the reservoir, as discussed in item 5 above. 

In aquatic systems where phytoplankton assemblages are dominated by diatoms, benthic fluxes of

silica are used to estimate the dissolution of biogenic bed material. Silica fluxes determined for 

Camp Far West Reservoir were anomalous because they exhibited fluxes that were not 

consistently positive (that is, sediment consumption of dissolved silica was measured in certain

cores during both incubation experiments; Table 6). The conventional assumption that diatom

frustules are the major silica source to the surficial bed sediment may be incorrect for Camp Far 

West Reservoir. Nutrient data also indicated anomalous results for Core 5 from the mid-reservoir

site in April 2002, reflected in the highly nonlinear time series for all solutes except silica. There 

were no visual indications of irregular characteristics in that core or in the core’s incubation that 

explained the disparate response. It would be strongly recommended that the incubation of at least 

three cores per site be used in subsequent studies for more quantitative understanding of smallscale (within site) variability (Kuwabara and others, 2003). As with mercury-flux measurements, 

95-percent confidence intervals about mean values were often greater than the mean despite 

acceptable coefficients of determination because only a single analytical replicate from each of 

four time intervals was used in the linear-regression for flux determinations (confidence interval 

discussion). 

7. Dissolved organic carbon (DOC): DOC concentrations serve as an indicator of the presence of organic 

ligands that can complex and enhance the solubility of trace metals including 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 this effect is

particularly important for mercury. Ligands represented by dissolved organic carbon can compete 

with sulfide to regulate mercury solubility (see section 8 below, Ravichandran and others, 1998) 

17
and can also decrease the bioavailability of mercury to certain aquatic biota (Sjoblom and others, 

2000). In April 2002, water-column DOC concentrations were highest at the near-dam site and 

were uniform with depth (184.3 to 185.8 micromolar; Table 8). At both upstream sites, a middepth minimum in DOC was observed with the lowest concentration at the inlet site (126.2 + 2.5 

micromolar). In contrast to the April 2002 results, the lowest DOC concentration in November 

2002 was observed in bottom waters of the near-dam site (187.0 + 2.6), yet that lowest 

concentration was higher than concentrations from April 2002. Higher DOC at all sites in

November relative to April 2002 is consistent with temporal trends observed at Lahontan

Reservoir in Nevada, another mining affected reservoir (Kuwabara and others 2001). This

temporal increase in the fall was probably caused by one or more of the following: snow-melt 

dilution (leading to lower DOC levels during April), a source of DOC available during the fall 

within the watershed, temporal changes in mass transport through the reservoir, or evaporation. In

contrast to our observed seasonal-scale trends, Yan and others (1996) hypothesized that long-term

(decadal) drought conditions and associated draw down at Swan Lake on the Canadian 

Precambrian shield caused reoxidation of exposed bank-sediment sulfur, resulting in acidification 

and decreased DOC concentrations in the lake water column. In April 2002, total dissolved

mercury concentration in the oxic bottom waters of Camp Far West Reservoir was highest at the 

deep near-dam site (17.5 + 1.2 picomolar) coincident with the highest bottom-water DOC

concentration (184.3 + 0.7 micromolar; Table 8). This solute interdependency is also consistent

with concentration trends observed at Lahontan Reservoir (Kuwabara and others 2001). DOC

serves as a ligand to enhance the dissolution of cinnabar (mercury sulfide, HgS), and hence 

increase mercury solubility and mobility (Ravichandran and others, 1998). It should be noted,

however, that the direct relationship between bottom-water mercury and DOC did not hold at the 

deep, near-dam site in November 2002, when the bottom water was suboxic (following weeks of

hypolimnetic anoxia). 

The benthic flux of DOC ranged from –169 micromoles per square meter per hour at the midreservoir site in April 2002 (average of two flux measurements, n=32 analytical measurements) to

518 micromoles per square meter per hour at the deep, near-dam site in November 2002 (Table 8). 

Negative DOC fluxes (that is DOC consumed by the sediment) has been reported, but is atypical

(Caffrey and others, 1996). For both sampling dates, the highest measured DOC fluxes out of the

sediment were at the near-dam site, followed by the Bear River inlet site. DOC benthic flux at the 

mid-reservoir site was consistently negative (consumed by the sediment). 

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

water column) and ranged from 290 to 1,180 nanomoles per square meter per hour (average of two 

replicates per site per date; Table 7). A spatial comparison of dissolved-sulfide benthic fluxes 

indicates the highest fluxes at the deep, near-dam site in November 2002 relative to the other two 

upstream sites. Given anoxic or suboxic conditions prior to and during our November 2002 

sampling, this observation is not surprising (Fig. 4). Conversely in April 2002, sulfide flux at the 

deep site was lowest among the three sites. As with porosities, the sulfide fluxes at the most

upstream Bear River inlet sites were temporally consistent. At the mid-reservoir sites, higher 

sulfide flux in November 2002 relative to April 2002 was coincident with bed sediment in 

November of lower surficial porosity (albeit fractured from rewetting). 

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). The calculated diffusive-flux estimates were consistently lower 

than directly measured sulfide-fluxes and 4 out of 6 diffusive-flux measurements were an order of

magnitude or more lower (Table 7). These differences between measured and calculated fluxes 

are curious and in need of further study because benthic biological measurements indicate that the 

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

18
bioirrigation effects. Benthic macroinvertebrate densities have been linked to relative 

enhancement of benthic flux in other mining-affect aquatic systems (Caffrey and others, 1996;

Kuwabara and others, 1999a; Kuwabara and others 2001; Kuwabara and others 2003; Thibodeaux

and Bierman, 2003). 

Dissolved sulfide represents a metastable ligand in most surficial environments which has a high

affinity to complex most divalent metals. This sulfide-complexing effect is particularly strong for 

mercury, which 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 therefore expected to be very low (of the

order of 10-50) due to the formation of mercuric sulfide. As the presence of DOC facilitates the 

dissolution or desorption of particle-associated mercury, and subsequent flux of dissolved mercury

species, the presence of sulfides can conversely facilitate precipitation and decreased solubility of

mercury, inhibiting benthic flux. 

19
Study Design and Methods 

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

experiments. Details (for example, quality control specifications) for each analysis have been previously 

documented (Woods and others, 1999; Praskins and others, 2001; Kuwabara and others, 2003). Sampling was

performed at three sites: a near inlet site (Fig. 1) (Map index 1, 39 degrees 2.039 minutes north latitude by 121 

degrees 16.409 minutes west longitude), a mid-reservoir site (Map index 2; 39 degrees 2.677 minutes N by 121

degrees 17.433 minutes W), and site near the dam at the deepest position in the reservoir (Map index 3;39 degrees 

3.140 minutes N by 121 degrees 18.687 minutes W). Because of significant draw down in the reservoir during the 

summer (Fig. 4), the inlet and mid-reservoir sites were sampled as close to the April 2002 sites as safely accessible 

by the sampling vessel (2 meter depth). These adjusted coordinates for the upstream sites in November 2002 were

39 degrees 2.124 minutes N by 121 degrees 16.939 minutes W, and 39 degrees 2.525 minutes N by 121 degrees

17.498 minutes W for the inlet and mid-reservoir sites, respectively. 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, and to facilitate the transport of cores to the incubation 

site. The site map (Fig. 1) also depicts reservoir outflow regulated at the Camp Far West Dam and regulated inflow

from the Bear River inlet to the east. 

As porosity measurements indicate (Table 1), the interface between the reservoir bed and the overlying

water column is characteristically a layer of extremely soft (unconsolidated) sediment 1 – 10 centimeters thick that

contains 65 to 95 percent water by volume. This intimate mixture of solid and liquid facilitates chemical and 

biological processes that move chemicals between solid and solution. Particularly under spring, high-flow

conditions, this unconsolidated surficial layer (overlying more compacted silts and clays) can be readily resuspended 

and transported to depositional areas near the dam. Similar materials are found in other lentic settings (Kuwabara 

and others, 2000). 

Coring Operation

On both sampling days (April 24 and November 4, 2002), three cores were collected at each of three sites 

using a 28-kilogram 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. Cores were collected in 10-centimeter diameter fluoroethylene polymer 

tubes 28-centimeters long. A tube, mounted on the coring device, was dropped onto the sediment by gravity, 

impacting the surface at about 0.2 meters per second and penetrating about 10 centimeters (depending on the 

porosity). Excessive penetration by the core was mitigated by four radially positioned, cylindrical weights that acted

as stabilizers (or feet) for the device. This sampling protocol was designed to provide minimal disturbance or

compression of soft sediments as visually verified by: (1) the presence oxic sediment (< 3 centimeters in thickness, 

often of brown or orange color) overlying gray to black anoxic sediment, (2) the presence of macroinvertebrates and

their tubes, and (3) minimal resuspension of surficial sediment in the overlying water when the core was retrieved.

If the sampler captured at least 2 centimeters of compacted silt or clay below the surficial layer, the sampler 

generally could be retrieved with the core and overlying water intact. Some samples were lost during retrieval, 

probably because the lowest sampled material was sandy or otherwise unconsolidated. 

Immediately after retrieval, the bottom and top of the sample tube were capped with an insert on the bottom

and a screw cap on the top (both made of fluoroethylene polymer). The tubes were stored upright in a padded plastic

container containing ice. 

Core Incubations

After cores were collected, samples were taken to Menlo Park, California for processing. Cores were left

overnight to equilibrate with wetted surfaces. Core incubations then began with the replacement of overlying water 

with bottom water that had been pumped from each site. Disturbance of the cores during handling was monitored 

by comparison of the chemical composition of the overlying water at the beginning of the incubation with bottom

20
waters. Significant deviations in the composition of these water samples would indicate significant disturbance of

the core during processing. 

Bottom water was aerobic on both sampling dates at the inlet and mid-reservoir sites. The twelve cores 

taken at those sites were aerated overnight to condition wetted surfaces before the incubation period (Topping and 

others, 2001). Although bottom water at the deep near-dam site was also aerobic in April, 2002, it was clearly 

suboxic (that is, less than 5 micromolar dissolved oxygen) in November, 2002. The April cores from this site were 

processed as described above for the other two sites, but the November samples from near-dam site were placed in

argon-purged and argon-filled bags during transport and incubation to avoid inadvertent aeration of the overlying

water. Two of the three cores per site were selected for incubation on the basis of visual inspection of the sedimentwater interface for any disturbances. Water overlying the sediment in the selected cores was sampled at four time 

intervals during a12-hour incubation at the bottom-water temperature. Overlying water sampled from the cores was 

replaced by pumped bottom water from the site where the core was collected. A linear regression of the time series 

of solute concentrations was used to determine the benthic flux (slope of the time series) from each of the incubated 

cores.

Physical Data 

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

collected from each core (top 3 centimeters) using a modified 10-cubic centimeter syringe. Wet 

weight and dry weight after lyophilization were measured to calculate porosity (Table 1). 

Biological Data

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

The sieved samples were fixed with 10-percent buffered formalin, later transferred to 70-percent 

ethanol, then sorted at 10× magnification and identified to the lowest practicable taxonomic level 

(Fig. 15). Samples were stained with rose bengal to facilitate sorting. No subsampling was used. 

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

Surficial sediment (top 0.5 centimeters) 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. 16). 

3. Stable isotope and elemental composition of suspended particulate matter (SPM): Water samples 

for chlorophyll-a, suspended particulate material (SPM), and elemental and stable isotopic carbon 

(C) and nitrogen (N) analyses were collected 1 meter below the surface using a van Dorn water 

sampler and filtered through a 210-micron plastic (Nitex) mesh to remove coarse particulates. 

Replicate water samples (n=2) were filtered within 2 hours of collection onto 25-millimeterdiameter and precombusted 13-millimeter-diameter glass-fiber filters and frozen on dry ice. SPM 

on 13-millimeter-diameter filters were fumed in a desiccator with HCl to remove carbonates, and 

filters were placed in tin capsules and held in a desiccator until analyzed for carbon and nitrogen

isotopes. Chlorophyll-a concentrations were determined on 25-millimeter-diameter filters 

fluorometrically (Eaton and others, 1995) within 6 months of collection. 

Isotope ratios and C and N concentrations of the SPM was determined at the Stable Isotope 

facility, Department of Agronomy and Range Science, University of California, Davis, using a 

Europa Scientific Hydra 20/20 continuous flow isotope ratio mass spectrometer and Europa

ANCA-SL elemental analyzer to convert organic C and N into CO2 and N2 gas. All samples were 

standardized against Pee Dee Belemnite (C) or N2 in air. The following example (for nitrogen) 

indicates how an isotopic ratio is calculated:

 δ15N‰ = [(Rsample/Rstandard)-1]×1000 

21
where R = 15N/14N. Instrument precision was 0.1 percent for carbon and 0.3 percent for nitrogen

on a per mil basis, based on replicate analyses of standard reference materials. 

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

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

using a micro-Winkler titration method (Fig. 17; Kuwabara and others, 2000). 

3. Dissolved Mercury and Methyl Mercury: Dissolved-mercury and methyl mercury samples were 

processed in a Class-100 laminar-flow hood. 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. 18, Krabbenhoft and others, 1998). Methodological details were reported by DeWild and

others (2002). 

4. Dissolved Trace Metals by ICP-MS: Water-column samples were also collected, filtered (0.2-

micrometer polycarbonate membrane) and acidified to provide dissolved trace-metal information 

for the reservoir by inductively coupled plasma mass spectrometry (Creed and others, 1994; Fig. 

19) 

5. Dissolved nutrients: Nutrient samples were filtered (0.2 micron polycarbonate membranes) and 

immediately refrigerated in darkness. Unlike trace-metal samples, nutrient samples were not 

acidified. Concentrations for dissolved (0.2-micron filtered) nitrate, ammonia, orthophosphate 

and silica were determined by automated spectrophotometry (Fig. 20). 

6. 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 carbon per liter) was generated from a doubledeionization unit with additional ultraviolet treatment (Milli-Q Gradient, Millipore Corporation)

(Fig. 21). 

7. Sulfide: Dissolved sulfides in overlying-water samples were analyzed by square-wave voltammetry 

(Fig. 22; Kuwabara and Luther, 1993). 

22
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26
Acknowledgments 

 

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

analyses, C. Lopez for benthic-chlorophyll analyses, and S. Hager for dissolved-nutrient analyses. Logistical and 

facilities support by M .Hunerlach and K. Rider is much appreciated. We are also grateful for discharge data and 

statistics by P. Shiffer (USGS, Sacramento, CA) and J. Erickson (Nevada Irrigation District, Colfax, CA). The 

California State Water Resources Control Board (Joint-funding Agreement 00-218-250-0) and in particular R.

Humphreys are gratefully acknowledged for 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. 

27
Appendix 1: Comments on the Report Structure 

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

range of interest groups. 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 this type of Web-based product may be improved to better serve readers are 

most welcome and may be directed to the major author (kuwabara@usgs.gov) so that they may be compiled for 

future revisions and reports. 

28
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 – Water-column chlorophyll and pheophytin pigment

Fig. 11 – Water-column carbon and nitrogen isotopes and elemental concentrations

Fig. 12 – Water-column total- and methyl-mercury

Fig. 13 – Water-column trace-metal concentrations 

Fig. 14 – Benthic response to upstream remediation (zinc) 

Fig. 15 – Photo of macroinvertebrate taxonomy methods 

Fig. 16 – Photo of benthic-chlorophyll analyses 

Fig. 17 – Photo of dissolved-oxygen analyses 

Fig. 18 – Photo of dissolved-mercury analyses 

Fig. 19 – Photo of dissolved-metal analyses 

Fig. 20 – Photo of dissolved-nutrient analyses 

Fig. 21 – Photo of dissolved organic-carbon (DOC) analyses 

Fig. 22 – Photo of dissolved-sulfide analyses 

29
 

Appendix 3: List of Tables 

Table 1 – Surficial-sediment porosities 

Table 2 – Benthic-invertebrate taxonomy

Table 3 – Benthic-chlorophyll analyses 

Table 4 – Water-column concentrations for mercury species in Camp Far West Reservoir. 

Table 5 – Dissolved-mercury fluxes 

Table 6 – Dissolved-nutrient fluxes and water-column concentrations 

Table 7 – Dissolved-sulfide fluxes 

Table 8 – Dissolved organic carbon (DOC) fluxes and water-column concentrations 

Table 9 – Photosynthetic pigments and stable isotopes in suspended particles 

Table 10 – Dissolved oxygen consumption

Table 11 – Dissolved-trace-metal water-column concentrations

Table 12 – Implications for upstream management

30
Fig. 1. Coring Sites at Camp Far West 

Fig. 1. Coring Sites at Camp Far West 

CA

1'

3

2'

2

39o 4.0’N39o 2.0’N

121o 18.0’W 

Bear River
Fig. 2. Coring Operatio

Fig. 2. Coring Operatio

Release 

Retrieval

Modifie
Bear River Discharge (m 3

 - s

-1) at Wheatland, CA

0.1

1

10

100

Storage

2002 Di

Pre-dam (1940 - 

Post-da(1962 - 2

March 1 June 1 Sept. 1 

Fig. 3. Hydrologic Condition

Fig. 3. Hydrologic Condition

2002 Date F

Sampling Trips
Fig. 4. Temporal Shifts in Water Fig. 4. Temporal Shifts in Water Fig. 4. Temporal Shifts in Water-column Conditions a

column Conditions a

lumn Conditions a

Draw down

DO (mg-L-1)

0 2 4 6 8 10

Vertical distance from lake bottom (m)

5040302010

0

Dissolved Oxygen

T

10 12 

Tem

4718
Fig. 5. Benthic Community Characte

Fig. 5. Benthic Community Characte

Location

A 

Benthic Chlorophyll-a

 ( μg-cm

-2)

024

In

Macroinvertebrate Density

 (individuals-m -2)

0

200040006000

Inlet Mid-reservoir Near Dam Other 

Studies

April 2002November 2002
Fig. 6. Hg-flux 

Comparisons between 

Mining-affected 

Reservoirs

Fig. 6. Hg Fig. 6. Hg

-flux 

Comparisons between Comparisons between 

Mining

-affected affected 

Reservoirs Reservoirs

Positive flux Flux into the water column
Fig. 7. Remediation efforts in the riv

concentration gradients in the re

Fig. 7. Remediation efforts in the riv

concentration gradients in the re

Incgraflu

Modifie
Fig. 8. Fig. 8.

Modifie
Fig. 9. Fig. 9.

Modifie
Figure 10. Seasonal dynamics of chlorophyll and pheophytin pigments in Camp Far West Reservoir.A. Chlorophylla concentration (μg-L-1). B. Pheophytin concentration (μg-L-1). C. Proportion of chlorophyll to 

pheophytin pigment. 

Values are means ± deviation (n=2). 

Figure 10. Seasonal dynamics of Figure 10. Seasonal dynamics of chlorophyll and chlorophyll and pheophytin pheophytin pigments in pigments in 

Camp Far West Reservoir. Camp Far West Reservoir.

A. Chlorophyll . Chlorophyll

-

a concentration (μg concentration (μg

-

L

-

1). 

B. Pheophytin Pheophytin concentration (μg concentration (μg

-

L

-

1). 

C. Proportion of chlorophyll to . Proportion of chlorophyll to 

pheophytin pheophytin pigment. pigment. 

Values are means ± deviation (n=2). Values are means ± deviation (n=2). 

Chl a (μg-L-1)

05

101520253035

Pheophytin (μg-L-1)

02468

101214161820

Jul-01 

Chl a/(Chl a + Pheophytin)

0.600.650.700.750.800.850.900.95
Figure 11. Seasonal trends in the elemental and isotopic composition of suspended particulate material (SPM) in Camp Far West Reservoir. A. δ13C of SPM. B. δ15N of SPM. C. Molar ratio of carbon to nitrogen 

in SPM relative to the molarRedfield ratio (6.7). 

Figure 11. Seasonal trends in the Figure 11. Seasonal trends in the elemental and isotopic composition elemental and isotopic composition of suspended particulate material of suspended particulate material (SPM) in Camp Far West Reservoir. (SPM) in Camp Far West Reservoir. 

A. δ13C of SPM. C of SPM. B. δ15N of SPM. N of SPM. C. Molar ratio of carbon to nitrogen . Molar ratio of carbon to nitrogen 

in SPM relative to the molar in SPM relative to the molarRedfield ratio (6.7). Redfield ratio (6.7). 

SPM δ13C

-36-34-32-30-28-26

SPM δ15N

02468

10121416

Jul-01 

N

Molar C:N ratio

456789

10
Fig. 12. Water Fig. 12. Water Fig. 12. Water-column concentrations of 

-column concentrations of 

column concentrations of 
Fig. 13. Water-column trace-metal co

at Deep, near-dam site

Fig. 13. Water Fig. 13. Water

-column trace column trace

-metal co

metal co

at Deep, near at Deep, near

-dam site dam site

Legend

MnZnCdNiCuPbCoVFe

0.1 

Depth (m)

05

1015

No

Concentrations normalized relative to 

0.1 1 10 100

Depth (m)

0

10203040

April 2002
Fig. 14. Benthic response to upstream remediat

(An extended example for zin

Fig. 14. Benthic response to upstream remediat

Fig. 14. Benthic response to upstream remediat

(An extended example for zin

(An extended example for zin

* Data points indicate calculated changes in diffusion-controlled benthic flux resulting from a hypothetical 50% decrease in dissolved-zinc concentration in the reservoir. Curves represent changes in benthic flux for lake environments that reflect contrasting pore-water quality (modified from Kuwabara and others, 2003).

* Data points indicate calculated changes in diffusion-controlled benthic flux resulting from a hypothetical 50% decrease in dissolved-zinc concentration in the reservoir. Curves represent changes in benthic flux for lake environments that reflect contrasting pore-water quality (modified from Kuwabara and others, 2003).
Fig. 15. Benthic Macroinvertebrate 

Fig. 15. Benthic Macroinvertebrate 

Modifie
Fig. 16. Benthic Chlorophyll Ana

Fig. 16. Benthic Chlorophyll Ana

Modifie
Fig. 17. Dissolved Fig. 17. Dissolved Fig. 17. Dissolved-oxygen Analyses by Micro-

-oxygen Analyses by Micro oxygen Analyses by Micro

-

Figures of merit for dissolved oxygen

Detection limit

0.5 mg-L-1
Fig. 18. Cold Fig. 18. Cold Fig. 18. Cold-vapor Atomic Fl -vapor Atomic Fluorescence Spe

vapor Atomic Fluorescence Spe

uorescence Spe

Modifie

Figures of merit for mercury analysis

Detection limit

0.2 pM
Fig. 19. Dissolved Trace-Metal Analyses b

Plasma Mass Spectrometry (I

Fig. 19. Dissolved Trace Fig. 19. Dissolved Trace

-Metal Analyses by 

Metal Analyses b

Plasma Mass Spectrometry (I

Plasma Mass Spectrometry (I

Modifie
Fig. 20. Dissolved Nutrient Anal

Automated Spectrophotome

Fig. 20. Dissolved Nutrient Anal

Automated Automated Spectrophotome

Spectrophotome

Figures of merit for nutrient analysis

Detection limit

Ortho-P 2.5 μg/L
Fig. 21. Dissolved Organic Carbon (D

High-Temperature Non-Catalyti

Fig. 21. Dissolved Organic Carbon (D

High

-Temperature Non Temperature Non

-Catalyti

Catalyti

Modifie

Figures of merit for DOC analysis

Detection limit

0.1 mg/L
Fig. 22. Dissolved Fig. 22. Dissolved Fig. 22. Dissolved-sulfide Analy -sulfide Analyses by Square sulfide Analyses by Square ses by Square

Modifie

Figures of merit for sulfide analysis

Detection limit

0.1 μg/L (4 nM)
1

 Definitions for abbreviations used in this table:

Inlet = Bear River inlet site

Mid-reservoir = mid-reservoir site

Near dam = deep, near-dam site
Processed with 500-micrometer sieve1 April 24, 2002 November 4, 2002

Site2 Inlet Inlet Mid Mid Dam Dam Inlet Inlet Mid Mid Dam Dam

Core Number 2 4 5 6 10 11 2 3 6 7 9 10

Phylum Nematoda

Family Mermithidae

undet. mermithid 1

Phylum Annelida

Class Oligochaeta

Order Tubificida

Family Naididae

Dero digitata 1 2

Uncinais uncinata 1

Family Tubificidae

Bothrioneurum vejdovskyanum 1

Limnodrilus hoffmeisteri 1 1 3 1 3 1 1

immature tubificids cf. Limnodrilus 10 18 7 5 19 27 3 13 4 4 37 39

Ilyodrilus templetoni 1 1 1

immature tubificids cf. I. templetoni 1

Phylum Arthropoda

Class Branchiopoda

Order Anomopoda

Family Daphnidae

cf Daphnia 1

Class Copepoda

Order Cyclopoida

undet. Cyclopoid 1 2

Class Ostracoda

undet. ostracod 1

Class Insecta

Order Diptera

Family Chironomidae

Procladius sp. 1 1

Chironomus cf. decorus group 1 2 1 1

Chironomus cf. plumosus 1 2 1 1 1

Phaenopsectra sp 1

Micropsectra sp. 1

Totals 18 22 10 7 24 30 4 17 8 6 39 42

Density (ind-m-2) 2338 2857 1299 909 3117 3896 519 2208 1039 779 5065 5455

Site Mean Density (ind-m-2) 2597 1104 3506 1364 909 5260

Site STD (ind-m-2) 367 275 551 1194 184 275

Overall Mean Density (ind-m-2) 2403 2511

S.D. (ind-m-2) 1131 2210

1

 Additional taxa present as inactive stages in all samples, but not counted

Phylum Ectoprocta Pectinatella magnifica (statoblasts)

Phylum Arthropoda Daphnidea spp. (epihippia)

2

 Definitions for abbreviations used in this table:

Inlet = Bear River inlet site

Mid = mid-reservoir site

Dam = deep, near-dam site

ind = individual macroinvertebrates 

m-2 = per square meter

S.D. = standard deviation
1

 Definitions for abbreviations used in this table:

Inlet = Bear River inlet site

Mid-reservoir = mid-reservoir site

Near dam = deep, near-dam site

S.D. = standard deviation

μg-cm-2 = micrograms per square centimeter
Table 4. Filtered (0.7

micrometer quartz membrane) bottomwater concentrations of 

mercury species at Camp Far West Reservoir

April 24, 2002

Methyl-Hg (pM)1 Total-Hg (pM)

Site Ind. Rep. Site ave. S.D. Ind. Rep. Site ave. S.D. 

Inlet 0.30 0.33 0.06 14.3 15.3 1.0

0.30 15.5

0.31 16.3

0.43

Mid-reservoir < 0.20 < 0.20 n/a 14.8 16.7 3.3

< 0.20 15.1

0.21 15.3

0.23 21.6

Near-dam 0.21 0.23 0.01 16.6 17.4 0.7

0.22 17.1

0.23 17.7

0.24 18.3

Near-dam shallow < 0.202 0.20 n/a 4.5

< 0.20 n/a

November 4, 2002

Methyl-Hg (pM) Total-Hg (pM)

Site Ind. Rep. Site ave. S.D. Ind. Rep. Site ave. S.D. 

Inlet < 0.20 < 0.20 n/a 3.1 3.7 0.7

< 0.20 3.2

< 0.20 3.8

< 0.20 4.6

Mid-reservoir < 0.20 < 0.20 n/a 2.9 2.9 0.3

< 0.20 3.1

< 0.20 2.6

< 0.20 n/a

Near dam 0.50 0.48 0.03 7.2 7.5 2.0

0.50 9.6

0.45 5.7

1

 Definitions for abbreviations used in this table:

Inlet = Bear River inlet site

Mid-reservoir = mid-reservoir site

Near dam = deep, near-dam site

pM = picomolar

Ind. Rep. = individual sample replicates

ave. = average

S.D. = standard deviation

n/a = not applicable or not available

2

 The detection limit for the dissolved total-mercury and dissolved 

methyl-mercury analysis was 0.20 pM.
10 Near dam 0.5 2.6

11 Near Dam 0.2 1.0 1.8

Total Mercury

Core Mercury Flux (pmoles-m-2-h-1) Mercury Flux Site ave.

Number Station mean ci r-squared n (ng-m-2-d-1) (ng-m-2-d-1)

2 Inlet 260 +/- 273 0.85 4 1254

4 Inlet 337 +/- 394 0.82 4 1623 1439

5 Mid-Res 308 +/- 149 0.96 4 1481

6 Mid-Res 210 2 +/- 368 0.99 3 1010 1245

10 Near dam 306 +/- 372 0.81 4 1475

11 Near Dam 272 +/- 316 0.83 4 1311 1393

November 4, 2002 *Note: All methyl-Hg concentrations for all cores, for all times < detection limit of 0.2 

Methyl-Hg

Core Mercury Flux Site ave.

Number Station (pmoles-m-2-h-1) (ng-m-2-d-1) (ng-m-2-d-1)

2 Inlet < 0.2 < 0.9

3 Inlet < 0.2 < 0.9 < 0.9

6 Mid-Res < 0.2 < 0.9

7 Mid-Res < 0.2 < 0.9 < 0.9

9 Near dam < 0.2 < 0.9

10 Near Dam < 0.2 < 0.9 < 0.9

Total Mercury

Core Mercury Flux (pmoles-m-2-h-1) Mercury Flux Site ave.

Number Site mean ci r-squared n (ng-m-2-d-1) (ng-m-2-d-1)

2 Inlet 43 +/- 19 0.97 4 209

3 Inlet 106 +/- 52 0.96 4 508 359

6 Mid-Res 124 +/- 161 0.79 4 597

7 Mid-Res 195 +/- 158 0.91 4 938 767

9 Near dam 36 +/- 163 0.24 4 175

10 Near Dam -28 +/- 13 0.97 4 -137 19

1

 Definitions for abbreviations used in this table:

Inlet = Bear River inlet site

Mid-Res = mid-reservoir site

Near dam = deep, near-dam site

Site ave. = site average

ci = 95-percent confidence interval

r-squared = coefficient of determination for flux regression

pmoles-m-2-h-1 = picomoles per square meter per hour

ng-m-2-d-1 = nanograms per square meter per day

2

 The flux calculation did not include the concentration at the final increment because it was approximately 

10 standard deviations greater than the mean for the final time increment 

from the other April cores. With this outlier included, the flux was 1,217 +/- 16,940 pmoles-m-2-h-1.
Number Site mean ci

r

2

n Number Site mean ci

r

2

n

2 Inlet 799 +/- 782 0.7552 Inlet 52 +/- 166 0.224

4 Inlet 755 +/- 769 0.7453 Inlet -52 +/- 29 0.904

5

2 Mid-Res -100 +/- 95 0.61

56 Mid-Res 86 +/- 38 0.944

6 Mid-Res -46 +/- 77 0.5147 Mid-Res 160 +/- 121 0.844

10 Near dam 100 +/- 83 0.8149 Near dam -213 +/- 178 0.814

11 Near dam 157 +/- 73 0.934 10 Near dam 39 +/- 475 0.044

r-squared = coefficient of determination for flux regression

m = meter

n = analytical replicates used in flux determination

μmoles-m-2-h-1 micromoles of DOC (in terms of carbon) per square meter per hour

μΜ = micromoles of DOC (in terms of carbon) per liter

2

 All 12 hours used in DOC flux calculations

3

 Not available because the site depth was only 2 meters to accommodate the draft of the sampling vessel

4

 For its potential chemical significance, mid-depth was taken just above the oxycline.

3 Inlet -1205 +/- 1165 0.87 4

6 Mid-Res -310 +/- 327 0.85 4 -34

7 Mid-Res 242 +/- 664 0.46 4

9 Near dam -426 +/- 459 0.85 4 -411 purged

10 Near dam -396 +/- 600 0.74 4 purged

1

 Definitions for abbreviations used in this table:

Inlet = Bear River inlet site

Mid-Res = mid-reservoir site

Near dam = deep, near-dam site

μmoles-m-2-h-1 = micromoles of dissolved oxygen per square meter per hour

ci = 95-percent confidence interval

r-squared = coefficient of determination for flux regression

n = analytical replicates used in flux determination

2

 Negative fluxes indicate consumption by the sediment
Deep/Surface April 2.9 6.2 2.9 1.2 1.5 1.1 1.0 0.9 1.1

ratio November 1.0 54.6 4.4 0.1 1.2 9.8 0.8 1.1 2.0

Temporal ratio3 0.3 8.8 1.5 0.1 0.8 9.2 0.8 1.2 1.8

1

 Definitions for abbreviations used in this table:

Inlet = Bear River inlet site

Mid-Res = mid-reservoir site

Near dam = deep, near-dam site

nM nanomoles per liter

2 In the interest of simplicity, confidence intervals are not given for these data. For general reference, tabulated concentrations reflect 

the mean of at least 4 analytical replicates with a 95-percent confidence interval 10% or less of the mean. 

3

 Temporal ratio = the deep/surface ratio in November divided by the deep/surface ratio in April;

this is used to demonstrate the temporal increase in concentration at depth relative to the surface.