POSTER SESSION ABSTRACTS
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

69

Transport Model of Contaminants

in Stream-Shallow Groundwater 

System

Multi-Parametric

Sensitivity Analysis

Relative Importance of 

involved Physical and 

Biogeochemical processes

Physical Processes

•Advection 

•Dispersion

•Hyporheic exchange

•G.W. interaction

Biogeochemical Processes

•Aquatic chemistry

 (for example, O2 and pH)

•Precipitation/Dissolution

•Redox processes

•Sorption/Desorption

•Microbial activities

Stream-tracer

experiments

Detailed field 

and laboratory 

measurements

Figure 1. Coupling MPSA with transport model to identify the relative importance of

physical and biogeochemical processes. 

Use of Multi-Parameter Sensitivity Analysis to

Determine Relative Importance of Processes

Involved in Transport of Mining Contaminants

By Jungyill Choi, Judson W. Harvey, and Martha H. Conklin

ABSTRACT 

Combining multi-parametric sensitivity analysis (MPSA) with stream transport modeling is

proposed to determine the relative importance of physical and biogeochemical processes controlling

transport of mining contaminants in natural stream systems. The MPSA is based on a large number of

Monte-Carlo simulations to identify the sensitive parameters over a broad range of each parameter.

This combined approach can provide an integrated view of transport processes of contaminants in

natural stream system.

INTRODUCTION

The fate and transport of contaminants in streams and rivers are controlled by a variety of physical

and biogeochemical processes. The physical processes play an important role in determining the fate of

solutes in surface-water environments. These physical processes include advection, dispersion,

hyporheic exchange, and ground-water interaction. In many situations, however, the transport of

contaminants are also greatly affected by biogeochemical processes, such as sorption/desorption,

oxidation/reduction, volatilization, hydrolysis, biodegradation, and other biochemical reactions.

Therefore, transport of contaminants in natural streams and rivers is best described by considering all

of the relevant physical and biogeochemical processes simultaneously (fig. 1).

To answer the question about relative importance of factors, the sensitivity of a numerical transport

model needs to be tested for the physical and biogeochemical parameters (processes) that are involved

in the forward transport model. However, traditional parameter-sensitivity analysis pertains to a
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

70

particular point (localized) in the parameter space, which is defined by all possible combinations of

parameter values. Also, in the localized sensitivity analysis, the importance or sensitivity of a selected

parameter can be affected greatly by the values of other parameters, because the significance of one

selected process is usually dependent on other processes. Typically, the importance of biogeochemical

processes are highly dependent on the physical processes, whereas the physical processes are not

affected by the biogeochemical processes. For example, the biogeochemical reactions of solutes in the

hyporheic sediments are enhanced by the prolonged retention time of solutes in these sediments.

Therefore, to account for parameter interactions, the relative importance of the physical and

biogeochemical processes of the transport model can be evaluated more accurately by a generalized

(multi)-parameter sensitivity analysis, which encompasses the entire parameter space (fig. 1).

This paper presents the concepts and procedures of multi-parameter sensitivity analysis (MPSA)

that is used to determine the relative importance of transport processes

METHODOLOGY

A numerical transport model may include detailed field measurements as well as ill-defined

parameters that cannot be measured with a high degree of accuracy in the field or in the laboratory.

These ill-defined parameters will severely limit the accuracy of any single simulation and increase the

difficulty of assessing the relative importance. In an attempt to overcome this difficulty and to

recognize the relative significance of parameters involved in the model, the sensitivities of simulations

results to input parameters need to be evaluated by assigning either a range of variation or a degree of

uncertainty to each parameter and implementing a generalized sensitivity analysis (Hornberger and

Spear, 1980; Chang and Delleur, 1992; Choi, et al., 1998; Choi, 1998). This multi-parametric

sensitivity analysis (MPSA) followed the procedure proposed by Chang and Delleur (1992) and Choi,

et al. (1998). The procedure includes the following steps:

 • Select the parameters to be tested.

 • Set the range of each parameter to include the variations experienced in the field and laboratory

measurement.

 • For each selected parameter, generate a series of, for example, 500 independent random numbers

with a uniform distribution within the design range.

 • Run the model using selected 500 parameter sets and calculate the objective function values.

 • Determine whether the 500 parameter sets are ‘acceptable’ or ‘unacceptable’ by comparing the

objective function values to a given criterion (R).

 • Statistically evaluate parametric sensitivity. For each parameter, compare the distributions of the

parameter values associated with the acceptable and unacceptable results. If the two distributions

are not statistically different, the parameter is classified as insensitive; otherwise, the parameter is

classified as sensitive. Relative importance can be evaluated statistically if desired.

The objective function values of the sensitivity analysis usually are calculated from the sum of

squared errors between observed and modeled values:

f = [x (i) - x (i)] o c 2

i=1

n

∑
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

71

Select parameters

to be analyzed

Make parameter

set by random 

number generation

OTIS

implementation

Calculation of

objective function

value

Comparison of the

distribution between

acceptable and

unacceptable cases

Evaluate the

sensitivity of each 

selected parameter

ic=500

(ic:repeating times)

No

Yes

Monte-CarloSimulation

Figure 3. Flow chart illustrating the

procedure of multiparametric sensitivity

analysis (MPSA).

“Insensitive”

p2

Cumulativefrequency

p1

Frequency

p2

Frequency

“Sensitive”

p1

Cumulativefrequency

Acceptable

Unacceptable

p1

R

p2

R

Acceptable

Unacceptable

Figure 2. Basic concept of multi-parametric sensitivity analysis (MPSA)

using a hypothetical model with only two parameters.

where f is the objective function value and xC(I) and x0(I) are calculated and observed values,

respectively. Observed values often are obtained from simulations that used the mid-points of the

characteristic range for each parameters. The ranges for each parameter are determined from minimum

to maximum values that are obtained from parameter estimations and field measurements through the

study reaches. If the objective function value obtained from the simulation is less than a subjective

criterion then the result is classified as acceptable, otherwise the result is classified as unacceptable.

Three different objective function values often are tested for a subjective criterion. Those values

typically define the 33, 50 and 66% divisions of 500 sorted objective functions.

The basic concept of MPSA is illustrated by using a hypothetical model with only two parameters

(Figure 2). In addition, the modeling procedure of MPSA described above is summarized using a

flowchart (Figure 3).

CONCLUSIONS

The combined efforts of forward modeling approach and generalized sensitivity analysis can

provide an integrated view and better understanding of contaminant transport processes in natural

stream systems. The multi- parametric sensitivity analysis especially helps identify the relative

importance of physical and biogeochemical processes controlling the transport of contaminants.

Furthermore, this methodology can provide a guide for future data-collection efforts and to order

research priorities.
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

72

REFERENCES 

Chang, F., and J.W. Delleur, 1992. Systematic parameter estimation of watershed acidification model,

Hydrological Processes, v. 6, pp. 29-44.

Choi, J, 1998. Transport modeling of metal contaminants in a stream-aquifer system; University of

Arizona, Department of Hydrology and Water Resources, unpublished PhD thesis, 225 pp.

Choi, J.Y., Hulseapple, S.M., Conklin, M.H., J.W. Harvey, 1998. Modeling CO2 degassing and pH in a

stream-aquifer system, Journal of Hydrology, v. 209, pp. 297-310.

Hornberger, G.M. and R.C. Spear, 1980. An approach to the preliminary analysis of environmental

system, Journal of Environmental Management, v. 12, pp. 7-18.
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

73

Figure 1. Study reach and sampling sites at Pinal

Creek. Median pH values are shown for the study

period.

Measurements of Plant and Algal

Bioaccumulation of Metals in Pinal and Pinto

Creeks, Arizona

By Justin C. Marble, Timothy L. Corley, and Martha H. Conklin

Dissolved Mn is an essential element for higher plant systems and is involved in photosynthesis

(the Hill reaction) and activation of different enzyme systems (e.g., superoxide dismutase production)

(Mukhopadhyay and Sharma, 1991). Critical deficiency levels of Mn(II) range between 0.01 to 0.02

microgram Mn per gram (mg Mn(II) g-1) dry weight in dry mature leaves but vary tremendously

between plants (Mukhopadhyay and Sharma, 1991). Vascular plants and algae also require certain

amounts of other trace metals for normal plant growth (Zn, Ni, Cu, Fe, Co, Ca, and Mg). 

Although Mn(II) supplements can increase growth yields of plants, large amounts of Mn(II) can

interfere with the uptake of other trace metals (Mukhopadhyay and Sharma, 1991). In addition, excess

concentrations of Zn, Ni, Cu, Fe, and Co can trigger an inherent defense mechanism that plants have

developed that involves production of phytochelatins—polypeptides that bind metals (Ahner, et al.,

1995). Phytochelatin production in response to high metal levels has been identified in land plants,

vascular aquatic plants, fungi, and marine and freshwater algae. This mechanism results in an

accumulation of the excess metals within the plants with the final metal concentration often being

significantly higher than found in water supplied to the plants.

The work reported in this paper focuses on bioaccumulation of metals by aquatic plants, algae, and

moss in Pinal Creek, an Arizona State Superfund site, near Globe, Arizona, that has been contaminated

by acid-mining activities in the area. The primary purposes of this study were to determine the extent

to which metals were taken up by the diverse plant community at Pinal Creek and to determine which

plants were particularly effective at bioaccumulation of metals. To further aid in our assessment of the

potential role of plants as a sink for metal contaminants in Pinal Creek (Figure 1), comparisons of

metals uptake were made with other measurements reported for similar plants in Pinto Creek, also near

Globe, Arizona. A comparison of typical surface

water data for Pinal and Pinto Creeks is given in

Table 1.

Plant grab samples were collected from several

locations and rinsed with creek water to remove 

insects and loosely attached sediment material. At

Pinal Creek, plant samples were collected from

sites Z0, J2

-1, J2-5, J2-15, and Z11 (Figure 1). At

Pinto Creek, grab samples were collected from two

USGS stream gaging sites 09498501 (below

Haunted Canyon near Miami, Arizona) and

09498502 (Pinto Creek near Miami, Arizona).
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

74

After rinsing with creek water, the plant samples were placed in plastic bags and put into a cooler.

Upon arrival at the laboratory, samples were dried at 60°C for 24 hours. Dried samples were ground

and sieved, then digested with nitric acid. Digested plant samples were analyzed by flame or graphite

atomic absorption spectroscopy for different metal concentrations. Results are reported as

bioaccumulation, i.e., mg of metal per kg of dried plant material (mg kg-1). The values reported

represent the average of 2 subsamples with the maximum and minimum values measured being within

±2 percent of the average value.

The aquatic plant species found at Pinal Creek varied in type and density depending upon the time

of year and the location. Before plant sampling started in 1996, water speedwell (Veronica anagallisaquatica) and rabbitfoot grass (Polypogon monspeliensis [L.] [Desf.]) dominated the upstream portion

of Pinal Creek (J2

-1) and algae (e.g., Microcystis, Vaucheria, and Oocystis) dominated in the

downstream section (J2

-15). However, over the study period (November, 1996 through June, 1997),

water speedwell, rabbitfoot grass, and algae were found along the entire study reach. 

Water speedwell from Pinal Creek was collected from several field locations (Z0, J2

-1, and J2-15)

over a period of 8 months and analyzed for Mn(II). There was no obvious correlation between

sampling date and bioaccumulation of Mn at J2

-1 and J2-15 . A subset of the water speedwell samples

from sites J2

-1 and J2-15 were analyzed for other trace metals (Table 2). No trend with location was

observed for concentrations of Fe, but Zn and Ni were higher at J2

-15 than at J2-1 and Cu was higher at

J2

-1 than at J2-15. Bioaccumulation of Mn and Co exhibited consistently higher bioaccumulation at J2-

15 compared to J2

-1, about a factor of 2 difference.

Pinal Pinto

Parameter Value Value

pH 6.4 7.8

Oxygen 6.9 9.0

Alkalinity1 51 180

TDS 2,640 531

Co(II) 0.410 0.003

Cu(II) 0.050 0.010

Fe(II) <0.130 0.0053

Mn(II) 72.0 0.0038

Ni(II) 0.790 0.010

Zn(II) 0.500 0.0060 1 As CaCO3.

Table 1. Physical and Chemical Values for Pinal Creek

(Z0 on January 25, 1995) and Pinto Creek (near Miami,

Arizona, on June 18, 1997, USGS). (mg L-1 except for pH

which is in standard pH units).

Metal Date J2-1 J2-15

Mn 12/13/96 6450 18600

Mn 1/31/97 7990 16400

Fe 12/13/96 4400 1880

Fe 1/31/97 2520 2670

Ni 12/13/96 109 151

Ni 1/31/97 148 182

Cu 12/13/96 901 824

Cu 1/31/97 1750 1130

Co 12/13/96 80.5 158

Co 1/31/97 134 279

Zn 12/13/96 516 665

Zn 1/31/97 772 801

Table 2. Water speedwell bioaccumulation from Pinal

Creek collected on December 1996 and January 1997 for

Mn, Zn, Ni, Cu, Co, and Fe (units are mg kg-1).
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

75

Rabbitfoot grass samples from J2-1 and J2-15 were also analyzed for Mn, Zn, Ni, Co, and Fe (Table

3). Both upstream and downstream sampling sites had similar bioaccumulation values for Zn and Ni,

but Mn, Cu, Co, and Fe values were larger at site J2

-1 than J2-15. A factor of about 2 between values at

J2

-1 and J2-15was observed for Mn, Co, and Cu, and a factor of about 10 for Fe. Bioaccumulation of

Mn at both sites was also greater than the other metals. Duckweed (Lemna minor) was less widely

distributed than either water speedwell or rabbitfoot grass, and was typically only found in slow

moving or stagnant water near the banks of the creek. However, a sample collected from J2

-5 on June

25, 1997 (pH 7.1, Mn(II) concentration 47.0 mg L-1) had an Mn concentration of 10760 mg L-1.

Metal J2

-1 J2-15 Table 3. Rabbitfoot grass bioaccumulation in samples from Pinal

Creek collected on January 31, 1997 (units are mg kg-1

Mn 13600 5240 ).

Fe 6890 691

Cu 1640 828

Ni 163 161

Co 237 130

Zn 581 534

Algae is prolific at both Pinal Creek and Pinto Creek and grab samples at both sites included the species

Microcystis, Vaucheria, and Oocystis. Samples were collected from both creeks to compare

bioaccumulation of Mn (Table 4). Although Pinal Creek samples had more bioaccumulation, the ratios of

plant concentration to surface water concentration were greater in the Pinto Creek samples.

Site Date Mn Table 4. Algae samples from Pinal Creek and Pinto Creek:

bioaccumulation of Mn (mg kg-1). Pinal, Z11 7/17/96 49700

Pinal, Z11 11/15/96 90200

Pinal, J2-15 12/12/96 5550

Pinal, J2

-15 1/31/97 79300

Pinto, Miami

Pinto, Miami

6/18/97

6/18/97

240

1460

Water speedwell collected from Pinto Creek had lower bioaccumulation of Mn than samples collected

from Pinal Creek (Table 5). However, the ratios of Mn concentrations in water speedwell to the surface

water concentrations were significantly higher in Pinto Creek than found for Pinal Creek. Similar

differences were apparent from comparisons of the algae samples collected at Pinto and Pinal Creeks

Site Date Mn Table 5. Water speedwell bioaccumulation

(mg kg-1) in Pinal and Pinto Creeks. Pinal, J2

-1 6/25/97 3870

Pinto, Haunted

Canyon

6/18/97

6/18/97

505

97

Pinto, Miami 6/18/97 47

These studies indicate that water speedwell, rabbitfoot grass, and algae bioaccumulate Mn.

Bioaccumulation of Zn, Ni, Co, Cu, and Fe was also observed in water speedwell and rabbitfoot grass.

Comparisons between water speedwell and algae samples collected from Pinal Creek and Pinto Creek

suggest that at Pinal Creek the plant capacity for metal uptake may have been reached and/or that metal

toxicity effects must be considered. Water speedwell and other aquatic plants are prolific in Pinal Creek

and could play a significant role in determining the fate of metal contaminants entering the stream.

Additional data concerning the total biomass in the system, and the potential release of metals as plants
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

76

die and decay, are required to assess the potential and actual contribution of plants to total metals removal

in this system. 

ACKNOWLEDGMENTS

This publication was made possible by grant number P42 ESO4940 from the National Institute of

Environmental Health Science with funding provided by EPA and by grant number EAR-95-23881 from

the NSF. Its contents are solely the responsibility of the authors and do not necessarily represent the

official views of the NIEHS, NIH, or EPA, or NSF.

REFERENCES

Ahner, B.A., Kong, S., and F.M.M. Morel, 1995. Phytochelatin production in marine algae—An

interspecies comparison, Limnology and Oceanography, v. 40, pp. 649-657.

Mukhopadhyay, M.J. and A. Sharma, 1991. Manganese in Cell Metabolism of Higher Plants, The

Botanical Review, v. 57, pp. 17-149.

U.S. Geological Survey, 1997. unpublished data of chemical and physical parameters from Pinto Creek

near Miami, AZ, Station Number 09498502, Lab ID 1830173, June 18.

AUTHOR INFORMATION

Justin C. Marble, Timothy L. Corley, and Martha H. Conklin, The University of Arizona, Department

of Hydrology & Water Resources, Tucson, Arizona.
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

77

Shark River 

Slough

Taylor Slough

C-111

Miami

Bicayne 

Bay

Long Key

Big Pine

Key Largo

Florida Bay

Offshore wells

Flamingo

A

B

C

D

E

F

G

H

Solution Hole

Figure 1. Florida Bay separates the Florida Keys, located off the southern tip of

Florida, from the mainland. Water samples were collected primarily from north of

Long Key and East of Flamingo. Groundwater samples from offshore wells were

collected where indicated by the circles. Letters refer to locations mentioned in

the text: A. Carysfort Reef; B. Algae Reef; C. French Reef; D. Molasses Reef; E.

Rock Harbor; F. Porjoe Key; G. Black Betsy Keys; and H. Tavanier Basin.

(Figure from Corbett et al., 1999)

Tracing Groundwater Flow into Surface Waters

by Application of Natural and Artificial Tracers

By D. Reide Corbett, William Burnett, Jeffrey Chanton, and Kevin Dillon

Submarine groundwater discharge (SGD) is an often overlooked yet possibly significant process in

the geochemical and nutrient budgets of marine nearshore waters. According to Johannes (1980) “SGD

should occur anywhere that an aquifer is hydraulically connected with the sea through permeable rocks

or bottom sediments and where the head is above sea level.” Such conditions are met in most coastal

areas. This process may be significant for transport of limiting nutrients in pristine coastal areas or, in

the case of polluted aquifers, could be an important source of contamination to the marine

environments. The problem is how to assess the extent of the groundwater flow and how to link

environmental problems with specific sites of contamination. Due to the extreme temporal and spatial

variability of many of these variables, the exact location of problematic discharges into coastal regions

may be difficult to determine by monitoring standard water quality constituents (e.g., NOx, turbidity).

In this research, subsurface water movement was evaluated with natural and artificial tracers in the

karst limestone of the Florida Keys 

(Figure 1).

In the Florida Keys, natural

tracers (222Rn and CH4) were

used to locate areas of

increased groundwater/surface

water interactions by

reconnaissance surveys of the

concentrations of radon and

methane in the bay waters

(Corbett, et al., 1999). These

trace gases function as natural

indicators of submarine

groundwater discharge into

standing bodies of water due to

their significantly higher

concentrations in groundwaters

(Cable, et al., 1996a, b; Bugna,

et al., 1996). General trends in

surface water concentration

were established by contouring

data from each tracer survey

with a kriging method by use

of the software package

Surfer® (Golden Software).

Although kriging interpolates

between data points, creating

some artifacts, the general

trends described are

independent of the contouring
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

78

Figure 2. Contours of radon (A) in dpm L-1 and methane (B) in nM for

samples collected in June/July 1997. Solid crosses indicate sampling

locations. Note the darker contours, indicating higher concentrations of

both parameters, near the upper Keys. (Figure from Corbett et al., 1999)

method or a reasonable change

in contouring concentration.

Examination of these contour

plots showed very little

apparent seasonal variation

throughout the study period.

During each period we

sampled, high concentrations

of both tracers were observed

near the Keys. Plots for 222Rn

and CH4 in summer 1997 show

the typical trends observed

(Figure 2). Direct

measurements of groundwater

flux via seepage meters were

also made in several different

areas of Florida Bay. Radon

and methane concentrations in

water samples collected from

wells, springs, canals, and

Florida Bay showed a

significant correlation, despite

the fact that the two trace

gases have independent source

terms (Figure 3). Natural

abundance of nitrogen

isotopes measured on attached

algae and seagrass also show

greatest 15N enrichment in

areas near the keys. We

observe a strong spatial

gradient in 15N of macrophytes

(seagrasses and macroalgae)

in Florida Bay, with relatively

light (-1 to 4 ‰) macrophytes

in western Florida Bay and

relatively heavy (6 to 13 ‰)

macrophytes in northeastern

Florida Bay (Figure 4). This

gradient is likely a function of

two processes: (1) progressive

denitrification of N brought

into Florida Bay via tidal exchange with the Gulf of Mexico; and (2) entry of 15N-enriched water from

the subsurface adjacent to the Keys in northeastern Florida Bay. Collectively, these results indicate a

greater flow of groundwater along the inside of the keys. Nutrient flux estimates, based on interstitial

nutrient concentrations and groundwater flux measurements, suggest that groundwater in the eastern

area of Florida Bay may provide as much nitrogen (110 ± 60 mmol N m-2 y-1) and phosphate (0.21 ±

0.11 mmol PO4

3- m-2 y-1) as surface freshwater sources from the Everglades (i.e., Taylor Slough and C111). However, the inputs are clearly not uniform and areas near solution holes/tidal springs may have

a substantially greater nutrient flux into surface waters then these estimates (Corbett, et al., 1999).
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

79

0

500

1000

1500

2000

Methane (nM)

0 100 200 300 400 500

Excess Radon (dpm L-1)

Groundwater Average

Porjoe Surface

Porjoe Interstitial

Garden Cove Spring

Garden Cove Surface

Lois Key Spring

Figure 3. Radon and methane concentrations in waters

sampled throughout the Keys. The groundwater tracer

concentrations are based on the overall average of all

samples collected. (Figure from Corbett et al., 1999).

Figure 4. Contours of 15N in macroalgae collected

throughout the study period. Solid crosses indicate

sampling locations. Note the darker contours, indicating

higher enrichment, near the upper Keys. (Figure from

Corbett et al., 1999).

Figure 5. I-131 in surface waters sampled on both sides

of Long Key, Florida Bay (closed square) and the

Atlantic via a canal (closed diamonds). Radioactive

iodine was added to a sewage injection well as a

conservative tracer to track wastewater movement in the

subsurface. Due to the rapid vertical and horizontal

movement of wastewater in the subsurface, injected

tracers appear in surface waters within 3 days after

injection.

Artificial tracers (SF6, 131I, 32P) were used to

establish a direct link of contaminated groundwaters to surface waters. Tracers injected

directly into sewage injection wells indicate

rapid flow of groundwater beneath the keys.

Experiments conducted on Long Key indicate

two different types of transport: (1) rapid flow

(0.20–2.20 m/hr), presumably through cracks

and conduits present in the limestone; and (2)

slow diffusive flow (<0.003-0.14 m/hr),

associated with the limestone’s primary porosity

(Dillon, et al., 1999). Vertical flow of the wastewater effluent was comparable to horizontal flow due

to the buoyancy of the relatively fresh wastewater compared to the surrounding saline groundwater.

These experiments showed that solutes injected into the Key’s subsurface have the potential to reach

surface waters within a few days (Figure 5). Tracer experiments conducted using both a conservative

tracer (SF6, 131I) and nutrients of interest (nitrogen and phosphorous) showed that both nitrate and

phosphate have some non-conservative behavior. Either through microbial alteration or interaction

with the limestone matrix, water from the wastewater injection appears to be polished as it flows

through the subsurface.

In a review of the general subject of SGD, Johannes (1980) stated that “It is...clear that submarine

groundwater discharge is widespread and, in some areas, of greater ecological significance than

surface runoff.” I agree with this appraisal and add that from my review of the available literature, I
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

80

find that it has been largely the biological community which has recognized the potential importance of

SGD. I maintain that the process of SGD may also be an important and overlooked part of the

geochemical cycles of many elements. In the case of Florida Bay and the Florida Keys, SGD has been

completely ignored in previous nutrient budgets in the area and has therefore not been considered as a

potential threat of contaminants (e.g., sewage effluent) to the ecosystem. It is hoped that one of the

main outcomes of this research will be the development of an approach which integrates geochemical

and hydrogeological techniques for assessing directions and rates of subsurface flow and, specifically,

how to quantify the flow into surface waters.

REFERENCES

Corbett, D.R., J. Chanton, W.C. Burnett, K. Dillon, C.M. Rutkowski, and J. Fourqurean, 1999.

Patterns Of Groundwater Discharge Into Florida Bay, Limnol. Oceanog.

Dillon, K., D.R. Corbett, J.P. Chanton, W.C. Burnett, and L. Kump, 1999. Rapid transport of a

wastewater plume injected into saline groundwaters of the Florida Keys, USA. Submitted to

Groundwater.

Johannes, R.E., 1980. The ecological significance of the submarine discharge of groundwater, Mar.

Ecol. Prog., Ser. 3, pp. 365-373.

AUTHOR INFORMATION

D. Reide Corbett, William Burnett, Jeffrey Chanton, and Kevin Dillon, Department of Oceanography,

Florida State University, Tallahassee, Florida 32306-4320. email: rcorbett@ocean.fsu.edu; Tel:

850-644-9914, Fax: 850-644-2581
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

81

Considerations for Calculating the Mass Loading

of Metal Contaminants to a Marine Embayment:

ASARCO Superfund Site, Tacoma, WA

By Gayle Garman and ASARCO Sediments/Groundwater Task Force

INTRODUCTION

The Asarco Superfund Site is located along the southern shore of Commencement Bay, an

industrialized marine embayment in southeastern Puget Sound, approximately 30 miles south of

Seattle, WA. The first industries on the site were sawmills that deposited woodwaste along the

shoreline. A lead smelter that began operations in 1890, was purchased by Asarco in 1905 to process

copper ore from other locations. By-products of copper smelting were further refined to produce

additional products, including arsenic, sulfuric acid, liquid sulfur dioxide, and slag. Smelter operations

ended in 1985 (Hydrometrics, 1996). 

Arsenic, cadmium, copper, lead, zinc, and other trace elements were released into soil, air, and

surface water as a result of the smelting and refining operations. There are six upland source areas

where the highest measured concentrations of contaminants in soils are found: The Stack Hill area,

Cooling Pond area, Arsenic Kitchen area, Copper Refinery area, the Fine Ore Bins building, and the

Southeast Plant/DMA area where sulfuric acid spills were frequent. Metals from soil releases and from

slag have migrated to groundwater at the Site (Hydrometrics, 1996).

Many of the smelter buildings and structures are on slag fill. In addition, Asarco extended the

existing shoreline by pouring molten slag into Commencement Bay. The upland area consists of both

gradual and steep slopes extending down to the slag filled shoreline, where slag bluffs extend as much

as 30 ft above the natural sandy substrate. These slag bluffs are very porous, and are subject to twice

daily tides that fluctuate up to 12 ft. vertical (Cross-Section D-D’).

The adjacent 23-acre Breakwater Peninsula is composed of massive and granulated slag that were

placed into Commencement Bay between 1917 and 1970. An estimated 15 million tons of slag exist at

the smelter property and slag peninsula (Hydrometrics, 1996).

SEDIMENTS/GROUNDWATER EVALUATION

Sediments as far as 1,000 ft from shore exhibit toxicity in bioassays, and are being evaluated for

remedial action. The preferred alternative is to place a clean sand cap over the contaminated sediment

to prohibit slag contact with marine organisms (Parametrix 1996). Reviewers of the proposed remedial

action asked whether metal contaminants in site groundwater would recontaminate the clean cap. The

Asarco Sediments Groundwater Task Force (ASGTF) was organized to evaluate this question.

Upland geological cross-sections were extended to the shoreline based on boring logs for nearshore

monitoring wells and offshore cores. These cross-sections indicated that the slag formation did not

discharge through sediment to the Bay, but rather, discharged directly to the Bay surface water. In

order to assess the influence of the tides on the hydrologic parameters of the site, 15 slag wells were

monitored over four complete tidal cycles in January 1998 and a multi-well pump-test was conducted
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

1

 Editor’s note: Figures follow the text. Figures are not consecutively numbered.

82

at new slag well MW-206. This new data was combined with slug-test data from the upland Remedial

Investigation. The distribution of hydraulic conductivities suggested the slag was characterized by four

corridors, indicated as A, B, C and D on the Figures. Later, corridor A was subdivided into corridors

A1 and A2 (Figure 1). Discharge rates were calculated for the slag aquifer and the underlying marine

sand aquifer in each corridor by using site data and Darcy’s law (ASGTF Group 1, 1998):

Q = Kibw

Where:

 Q = groundwater discharge rate (ft3/day)

K = hydraulic conductivity (ft/day)

i = hydraulic gradient (ft/ft)

b = aquifer thickness at the shoreline (ft)

w = width of groundwater flow path (ft) 

APPROACH TO CONTAMINANT FLUX ESTIMATION

The ASGTF recognized that Darcy’s law “provides an estimate of the net groundwater flow

discharging from the site to Commencement Bay.” This net groundwater flow originates as recharge in

upgradient water-bearing zones, as infiltration of surface water run-on, and as precipitation onto the

slag. While the twice daily tides, with ranges to 12 vertical feet, are known to cause recurrent inflow

and outflow of seawater in nearshore areas of both the slag and marine sand aquifers, it is assumed that

the net tidal flow is zero (ASGTF, Group 1 Memo). 

The mass flux of a contaminant is calculated by multiplying the groundwater discharge rate by the

contaminant concentration. However, when (fresh) groundwater mixes with seawater, there are

changes in geochemistry that alter the solubility, and consequently the mobility, of the metal

contaminants of concern at this site. 

Arsenic (As) is the primary contaminant of concern in upland areas of the site. Dissolved arsenic

concentrations to 30 mg/5 have been measured in groundwater near the Fine Ore Bins. The chronic

marine AWQC (Ambient Water Quality Criterion) for arsenic is 0.036 mg/5.

DATA AND FIGURES 1

Data are collected at site monitoring wells each Spring and Fall. The ASGTF used data from

March 1994 through September 1998, thus, the Figures show the mean of ten measurements for each

parameter. The Figures are taken directly from the ASGTF Group 4 Technical Memorandum,

(December 1998) and consequently, are not numbered sequentially in this presentation. The

contaminant isopleths were drawn by hand.

ARSENIC ATTENUATION

Figure 10 shows the intrusion of seawater (chloride) for hundreds of feet into the slag formation

along the shoreline. Chloride in upland groundwater is negligible. The landward intrusion of seawater

into the slag is least in corridor D and greatest in corridor A1. The chloride concentrations in both deep

and shallow wells on the breakwater peninsula approximate the chloride concentrations in

Commencement Bay surface water.
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

83

Figure 2 shows that the arsenic concentration is an order of magnitude greater in the deeper

breakwater wells (B) than in the shallow breakwater wells (A). The breakwater peninsula is composed

entirely of smelter slag. The ASGTF concluded that the lower oxygen exchange capacity for water

deep within the breakwater peninsula, and the associated lowered redox condition, increase the

solubility of slag arsenic.

Figure 11 shows that dissolved oxygen, in general, diminishes in proportion to distance from the

shoreline and more rapidly in the less permeable corridors, e.g., Corridor D. However, it is difficult to

get accurate field measurements of dissolved oxygen, so the distribution of manganese, which rapidly

precipitates in the presence of dissolved oxygen, also was evaluated.

Figure 6 shows the distribution of dissolved manganese which corroborates the mechanism of

arsenic precipitation, described below. Manganese concentrations decline by an order of magnitude as

site groundwater approaches the shoreline and mixes with oxygen-rich seawater that has intruded into

shoreline slag.

In upland areas of the site (not the slag peninsula, which is entirely slag), the greatest groundwater

flow occurs in the slag formation, which is above the natural geologic formations. The slag, in turn, is

topped by a thin layer of filled soil. Upland groundwater has low oxygen content relative to seawater,

so the geochemistry of the upland groundwater is reducing in comparison to the water of

Commencement Bay. At the shoreline, the tides of Commencement Bay enter the porous seaward face

of the slag formation, forcing seawater into the slag. Thus, as the upland reduced groundwater migrates

toward the shoreline, it gains oxygen by mixing with tidal seawater within the slag. The solubility of

the arsenic then decreases, and most of the dissolved arsenic is precipitated as secondary minerals in

the slag and does not discharge to Commencement Bay. The presence of secondary arsenic minerals

has been confirmed by a mineralogic study of material recovered when MW-206 was installed (US

EPA, 1998). Thus, the changing redox condition of the groundwater explains the attenuation of the

primary contaminant of concern, arsenic. However, understanding the mechanism that controls arsenic

solubility does not answer the question of how to calculate the mass flux of arsenic to the Bay (ASTGF

1998, Group 4). 

RELEASE OF COPPER

Figure 3 shows an area in Corridor D where the average dissolved copper concentration in

groundwater is greater than in any other nearshore area. Unlike arsenic, copper is generally more

soluble when there is more dissolved oxygen. However, the shallow (A) wells on the Breakwater

Peninsula, where oxygen is available from seawater and atmospheric exchange, do not have copper

concentrations as great as the wells in Corridor D. The ASGTF concluded there must be another

geochemical parameter causing copper to dissolve from slag in corridor D (ASGTF 1998, Group 4). 

Figure 14 shows that acidity may be controlling copper concentrations in Corridor D. Acidity is

measured in logarithmic pH units. The pH of Commencement Bay water is about 8.0. Wells in the

southeast plant/DMA area have average pH values less than 6.0, indicating acid concentrations two

orders of magnitude greater than Commencement Bay. The southeast plant/DMA area is the location

of previous liquid sulfur dioxide and sulfuric acid manufacture. Materials remaining at this location

apparently continue to acidify the groundwater, releasing copper from the slag matrix. Even the

intrusion of seawater does not overcome this effect, as the Corridor D well nearest the shoreline has an

average dissolved copper concentration of 3.6 mg/5, more than an order of magnitude greater than any

other shoreline well. By comparison, the copper acute marine AWQC is 0.0029 mg/5 (a chronic copper

marine AWQC has not been adopted).
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

84

Figure 13 shows the distribution of dissolved iron in shoreline wells, which helps corroborate the

mechanism of copper solubility. Like copper, iron is more soluble in a low pH (acid) environment

(e.g., landfill leachate). Higher concentrations of iron are found in the Southeast Plant/DMA source

area where pH was low and copper was high. Thus, both copper and iron exhibit increased solubility

here because of the lower pH (higher acidity). However, unlike copper, the average dissolved iron

concentrations are quite similar in all the wells closest to the shoreline (ASGTF 1998, Group 4). This

suggests that when the dissolved copper and iron in acidified groundwater in corridor D encounter

intruding seawater within the shoreline slag, that the reaction of iron with the oxygen and alkalinity of

the seawater forming an iron precipitate is more rapid than the similar reaction of copper.

CONCLUSIONS

The wells closest to the shoreline have mean dissolved copper concentrations at least three orders

of magnitude greater than the marine acute/chronic AWQC of 0.0029 mg/5. The greatest volume of

groundwater discharges through the fractured slag into the marine water column rather than through

contaminated subtidal sediment. Thus, the groundwater process that is having the greatest effect on

marine biota is probably copper discharge to water, not arsenic discharge to sediment. 

Mass loading of contaminants in general is a simple calculation that multiplies the average

contaminant concentration by the corresponding average water (volume) discharge rate. This calculation is valid as long as the contaminant concentration is the concentration in the volume of water that is

discharged. The calculation of contaminant mass loading rates to Commencement Bay is complicated

by the geochemical changes that occur in the shoreline slag as the fresh groundwater mixes with

marine surface water, altering the solubility of the metal contaminants; and by the difficulty in

determining the corresponding volume (discharge rate) of water, which is influenced at the shoreline

by the twice daily tidal flux. Wells nearest the shoreline, where contaminant concentrations are most

representative of discharges to the Bay, are affected by the influx and efflux of tidal water, i.e., the

volume of discharging water characterized by the contaminant concentration measured in the shoreline

wells, is likely greater than the net groundwater flux from the site. 

The ASGTF has not yet found a satisfactory method for resolving this problem.

REFERENCES

ASGTF Group 1, 1998. Draft Technical Memorandum on Groundwater Discharge Estimates,

November 1998. 25 pp. plus attachments A and B.

ASGTF Group 4, 1998. Draft Technical Memorandum for the Asarco Sediments/Groundwater Task

Force, December 1998. 30 pp. plus Figures.

US EPA, 1998. Mineralogical Study of Borehole MW-206, Asarco Smelter Site, Tacoma, Washington.

US Environmental Protection Agency Region 10, Office of Environmental Assessment, Seattle,

WA, 31 pp. plus Appendices A-D.

Hydrometrics, 1996. Tacoma Smelter Post-Remediation Surface Water Evaluation and Technical

Impracticability Demonstration. Draft Revision 1 for ASARCO, Inc., Tacoma, WA, June 1996.

Parametrix, Inc. 1996. Asarco Sediments Superfund Site Expanded Remedial Investigation and

Feasibility Study, Phase 2 Refinement of Options Report, Volume I. For ASARCO, Inc., Tacoma,

WA, December 1996.
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

85

AUTHOR INFORMATION

Gayle Garman, NOAA Office of Response and Restoration, Coastal Protection and Restoration

Division, 7600 Sand Point Way, NE, Seattle, WA 98115-0070. gayle.garman@noaa.gov

MEMBERS OF THE ASARCO SEDIMENT/GROUNDWATER TASK FORCE (ASGTF)

Marian Abbett, State of Washington, Department of Ecology, Olympia, WA

Thomas Aldrich, Asarco, Inc., Tacoma, WA

Bruce Cochran, State of Washington, Department of Ecology, Olympia, WA, David Frank, US EPA,

Seattle, WA

Gayle Garman, NOAA, CPRD, Seattle, WA

James Good, Parametrix, Inc., Kirkland, WA

Douglas Holsten, CH2M-Hill, Inc., Bellevue, WA

Lee Marshall, US EPA, Seattle, WA

Scott Mason, Hydrometrics, Inc., Kalispell, MT

Roger McGinnis, Roy F. Weston, Inc., Seattle, WA

Robert Miller, Hydrometrics, Inc., Tacoma, WA 

David Nation, Hydrometrics, Inc., Tacoma, WA

Karen Stash, Roy F. Weston, Inc., Seattle, WA

Carl Stivers, Parametrix, Inc., Kirkland, WA

Donald Weitkamp, Parametrix, Inc., Kirkland, WA

Bernie Zavala, U.S. EPA, Seattle, WA
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

86

Conceptual Cross-Section D1 – D1'
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87

Figure 1
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Figure 2
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Figure 3
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Figure 6
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91

Figure 10
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Figure 11
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Figure 13
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Figure 14
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

95

The Interaction of Ground Water and Surface

Water within Fall Chinook Salmon Spawning

Areas in the Hanford Reach of the Columbia

River

By David R. Geist

INTRODUCTION

The Hanford Reach is the last unimpounded section of the mainstem Columbia River in the United

States and supports a large run of fall chinook salmon (Oncorhynchus tshawytscha) that returns there

annually to spawn (Dauble and Watson 1997). Previous studies have shown that adult salmon

repeatedly spawn in definite locations within the Reach (Geist 1999; Geist and Dauble 1998; Dauble

and Watson 1997), but the physical characteristics associated with these areas are variable and poorly

understood. More information on the spawning habitat characteristics of fall chinook salmon that

utilize large rivers is needed to recover stocks listed on the Endangered Species Act.

The association between fall chinook salmon spawning and physical habitat characteristics was

previously examined in the Hanford Reach at Locke Island and Wooded Island (Geist 1999). Although

the physical habitat characteristics, e.g., depth, substrate, and water velocity, at the two sites were

similar, only the Locke Island site had extensive salmon spawning. Additional measurements were

taken to determine if the interaction of ground water and surface water within the hyporheic zone could

explain this discrepancy in habitat use between the two sites. Hyporheic discharge was assumed to

affect spawning site selection by providing cues (chemical, temperature, and physical) for prespawning adults to locate spawning reaches (usually 2 to 5 km in length). Once these reaches were

“discovered,” hyporheic discharge was assumed to correlate with the distribution of redd clusters (500

to 800 m in length, 120 m in width; Geist 1999) within these river reaches

METHODS

During the fall chinook salmon spawning seasons (October and November; Dauble and Watson

1997) from 1995 to 1997, mini-piezometers (Lee and Cherry 1979) and internal-drive-rod piezometers

(Geist, et al. 1998) were installed within the two sites. Piezometers were installed within the river

channel in groups of three or four, and hyporheic water within the piezometers was sampled 2 to 7

times each year for specific conductance (μS/cm at 25(C), water temperature (T, (C), dissolved

oxygen (DO, mg/5), and hydraulic head (h, cm). These same parameters were also measured on a

contiguous river sample. 

It was assumed that water discharging from the hyporheic zone into the river was a combination of

ground water and surface water. Specific conductance was the primary measure used to differentiate

undiluted ground water from surface water; specific conductance of undiluted ground water adjacent to

the Hanford Reach averages 300-400 μS/cm while the river water averages ~150 μS/cm. The

differences in temperature (T) and hydraulic head (h) were based on the piezometer reading (hyporheic

water) minus the reading from its paired river sample. Differences in hydraulic head between the river

and hyporheic waters were used to calculate a vertical hydraulic gradient (VHG) between the two
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

96

125.0

150.0

175.0

200.0

225.0

Specific conduc. (

Locke Wooded

S/cm)

B.

River

-0.3

-0.2

-0.1

0.0

0.1

0.2

0.3

Locke Wooded

 temp. (hypor. - river; C)

C.

-0.02

0.00

0.02

0.04

0.06

0.08

Locke Wooded

VHG (cm/cm)

A.

Figure 1. Physiochemical data collected from

piezometers installed and monitored at the Locke Island

and Wooded Island study sites during October and

November, 1995. Bars above and below the points

represent the 95% confidence interval of the mean. (A)

Vertical hydraulic gradient (VHG) between hyporheic and

surface waters where positive values indicated potential

upwelling and negative values downwelling, (B) specific

conductance of hyporheic and surface waters, and (C)

differences in water temperatures of hyporheic and

surface waters.

(VHG = h/depth of piezometer). Slug tests were used to estimate the volume of hyporheic discharge

from the sediments into the river channel.

RESULTS AND DISCUSSION

The results showed that fall chinook salmon

spawning locations were highly correlated with

hyporheic discharge that was composed of mostly

river water and not undiluted ground water.

Hyporheic water that discharged into fall chinook

salmon spawning locations was consistently

greater in magnitude, and had higher dissolved

oxygen and lower specific conductance than

discharge into non-spawning locations. However,

there was no significant difference in temperature

between hyporheic and river water. These results

were true when comparisons were performed

between Locke Island (spawning site) and Wooded

Island (non-spawning site) (Figure 1), and also true

when spawning and non-spawning clusters within

the Locke Island site were evaluated (Figure 2). 

Slug tests showed that substrate permeability

decreased with increasing distance below the river

bed at Wooded Island but did not change over the

depths monitored at Locke Island (Figure 3). This

suggested the mixing zone where river water

penetrated into the river bed was greater within the

spawning site than within the non-spawning site.

Specific discharge calculations gave an average

flux out of the sediments on the order of 9.0 x 10-4

cm/s at Locke Island and 3.0 x 10-4 cm/s at

Wooded Island. Thus, specific discharge of

hyporheic waters was approximately 3 times larger

at Locke Island than Wooded Island.

River water was presumed to have entered

highly permeable riverbed substrate at locations

upstream of spawning areas. Geomorphic bed

features (i.e., islands, gravel bars, riffles) of

alluvial rivers are able to create hydraulic gradients

sufficient to direct surface water into the bed

(Stanford, et al. 1996; Brunke and Gonser 1997).

River water is able to penetrate deeper into

hyporheic habitats if the riverbed is composed of

alluvium that is highly permeable (Vaux 1962,

1968; White 1993). The more permeable the

alluvium, the more that the physiochemical characteristics of the hyporheic waters will resemble

surface water rather than ground water. In contrast, the relative proportion of phreatic ground water in

hyporheic waters will be greater if the riverbed sediments are of low hydraulic permeability because
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

97

100.0

125.0

150.0

175.0

200.0

225.0

Specific conduc. (

96 S 96 NS 97 S 97 NS

S/cm)

Riv er

B.

-0.5

0.0

0.5

1.0

1.5

2.0

96 S 96 NS 97 S 97 NS

 temp. (hypor.-river;C)

C.

0.00

0.02

0.04

0.06

0.08

0.10

96 S 96 NS 97 S 97 NS

VHG (cm/cm)

A.

nd nd

Figure 2. Physiochemical data collected from

piezometers installed and monitored within the Locke

Island site at spawning (S) and non-spawning (NS) sites

during October and November, 1996 and 1997. Bars

above and below the points represent the 95% confidence

interval of the mean. ND = no data.

0

5

1 0

1 5

2 0

2 5

3 0

3 5

0 50 100 150 200

Pie z o me te r d e p th (c m)

Recovery time (secs.)

L oc k e Is la nd

W oo d ed Is la nd

Figure 3. Recovery time to 37% of the initial hydraulic head

following a slug-test within piezometers installed at Locke

Island and Wooded Island. The trend line at the Wooded

Island site (dashed line) was significant (P = 0.02, r2

=0.78) but

not significant at Locke Island (P = 0.42, r2

= 0.08).

river water will not be able to readily enter the

substrate and dilute the ground water (White 1993;

Brunke and Gonser 1997). I concluded river water

that became entrained into the “hyporheic

corridor” had a strong influence on vertical

hydraulic gradients and influenced the use of

salmon spawning habitat. Knowledge of the threedimensional connectivity between rivers and

ground water within the hyporheic zone can be

used to improve the definition of fall chinook

salmon spawning habitat. 

REFERENCES

Brunke, M. and T. Gonser, 1997. The ecological

significance of exchange processes between

rivers and groundwater. Freshwater Biology v.

37, pp. 1-33.

Dauble, D.D., and D.G. Watson, 1997. Status of

fall chinook salmon populations in the midColumbia River, 1948-1992, North American

Journal of Fisheries Management, v. 17, pp.

283-300.

Geist, D.R., 1999. Redd site selection and

spawning habitat use by fall chinook salmon.

Ph.D. Dissertation, Oregon State University,

Corvallis, Oregon.

Geist, D.R., and D.D. Dauble, 1998. Redd site

selection and spawning habitat use by fall

chinook salmon: the importance of geomorphic

features in large rivers, Environmental

Management, v. 22, pp. 655-669.

Geist, D.R., M.C. Joy, D.R. Lee, and T. Gonser,

1998. A method for installing piezometers in

large cobble-bed rivers, Ground Water

Monitoring and Remediation, v. 18, pp. 78-

82.

Lee, D.R., and J.A. Cherry. 1978. A field

exercise on groundwater flow using

seepage meters and mini-piezometers,

Journal of Geological Education, v.27,

pp. 6-10.

Stanford, J.A., J.V. Ward, W.J. Liss, C.A.

Frissell, R.N. Williams, J.A.

Lichatowich, and C.C. Coutant, 1996. A

general protocol for restoration of

regulated rivers, Regulated Rivers:

Research and Management, v. 12, pp.

391-413.
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

98

Vaux, W.G., 1962. Interchange of stream and intragravel water in a salmon spawning riffle. Special

Scientific Report—Fisheries, No. 405. U.S. Fish and Wildlife Service, Bureau of Commercial

Fisheries, Washington, DC

Vaux, W.G., 1968. Intragravel flow and interchange of water in a streambed. Fishery Bulletin 66:479-

489.White, D.S. 1993. Perspectives on defining and delineating hyporheic zones, Journal of the

North American Benthological Society, v. 12, pp. 61-69.

AUTHOR INFORMATION

David R. Geist, Ecology Group, Pacific Northwest National Laboratory, MS K6-85, Post Office Box

999, Richland, Washington, 99352; 509-372-0590; fax: 509/372-3515; david.geist@PNL.gov.
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

99

Integrated Acoustic Mapping of Surface Waters:

Implications for Ground-Water/Surface-Water

Linkages

By Chad P. Gubala, Ullrich Krull, Joseph M. Eilers, Mike Montoya, and Jeff

Condiotty

The study of aquatic systems has historically been approached in a traditional scientific manner.

“Representative” sections or components of lakes and rivers have been examined intensively through a

combination of laborious sampling methods. Broader assessments of specific aquatic ecosystems have

then been statistically constructed through the assembly of discrete study elements. Changes in aquatic

ecosystems have then been documented by repeating a similar regimen of sampling at varying time

intervals. Aquatic ecosystem analyses and risk-based management plans have been developed on the

basis of discreet and/or empirical numeric models of aquatic ecosystems, deriving from the original

field investigations. 

The efficacy of aquatic assessments and/or risk-based management plans depends upon the

completeness and accuracy of the original data collection and analysis scheme. In order to assemble an

accurate model of an entire aquatic ecosystem, data must be collected in a manner that minimizes the

major components of uncertainty: measurement, spatial and temporal. Most researchers have been able

to adequately minimize measurement error throughout intensive, small-scale research studies or

monitoring exercises. However, precise and accurate measurements distributed over a small section of

a large domain frequently lead to inaccurate conclusions. This phenomenon derives from the

uncertainty of interpolating the conditions of an unknown domain, such as a river reach or lake region,

through interpolation or extrapolation from a limited data-base. 

A need exists to develop better monitoring techniques for the dynamic management of aquatic

ecosystems. Combinations of current and emerging technologies, drawn from a variety of application

areas may provide for faster, more cost-effective means of acquiring aquatic systems data and

information. Linking mobile sensors such as hydroacoustic arrays with Global Positioning System

(GPS) navigation have already yielded effective methods for rapidly delineating the bathymetric,

morphometric and hydrologic features of lakes and rivers. Expansion of the role of acoustics has also

permitted the spatial analysis of aquatic biological communities within complex spatial domains. The

coupling of standard aquatic sensor arrays, such as temperature, conductivity or velocity probes, with

GPS will provide multiple parameters for a system in a cost-effective manner. Integration of advanced

sensor probes, such as real-time DNA detectors for identification of aquatic microorganisms will also

greatly enhance the ability to detect and manage change in aquatic ecosystems.

AUTHOR INFORMATION

Chad P. Gubala, SATL (The Scientific Assessment Technologies Laboratory),The University of

Toronto at Mississauga, 3359 Mississauga Road North, Mississauga, ON L5L 1C6; 905-828-3863

(voice); cgubala@credit.erin.utoronto.ca.

Ullrich Krull, University of Toronto, Joseph M. Eilers, JC Headwaters, Inc.

Mike Montoya, Ute Tribal Nation, Jeff Condiotty, Simrad, Inc.
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

100

Delineation of VOC-Contaminated Groundwater

Discharge Zone, St. Joseph River, Elkhart,

Indiana

By John H. Guswa, Jonathan R. Bridge, and Michael J. Jordan

A hydrogeologic study was conducted to locate and delineate the portion of the St. Joseph river

within which VOC contaminated groundwater observed in monitoring wells in the study area was

discharging. The principal groundwater contaminants are trichloroethene (TCE) and carbon

tetrachloride (CCl4). Water samples were collected at a depth of approximately two to five feet below

the river bed using a GeoProbe® from a pontoon boat. The samples were analyzed for the purpose of

delineating VOC concentrations in groundwater directly beneath the river bed. The results of this

hydrogeologic study were used to select sampling locations for a benthic macroinvertebrate

investigation to determine if there were any ecological effects resulting from the discharge of VOCcontaminated groundwater to the river.

Groundwater and surface water samples were initially collected from selected wells and sampling

stations on the river and analyzed for the inorganic analytes listed in Table 1. The purpose of this

sampling was to identify whether there were inorganic analytes that could be used as “tracers” to

ensure that the samples collected from beneath the river bed were groundwater samples and not

induced river water. The concentrations of none of the inorganic analytes proved to be consistently

different between the surface water samples and the groundwater samples. Therefore this group of

inorganic analytes could not be used as “tracers.” Other field measured parameters, in particular pH,

temperature and specific conductance were more useful in this regard, and were used for that purpose.

The pH of the river water was approximately one pH unit higher than the groundwater. The river water

temperature was approximately 8(C higher than the groundwater temperature and the specific

conductance of the groundwater was generally higher than the river water. 

Table 1. Results of preliminary inorganic analyses, in mg/5.

Sampling

Location

CATIONS ANIONS

calcium magnesium sodium potassium bicarbonate carbonate sulfate

Groundwater

MW-7S

MW-7D

MW-8S

MW-8D

MW-9

MW-10S

94.3

89.9

76.8

81.0

75.9

82.8

25.6

24.3

18.6

19.9

19.5

20.1

7.7

31.7

22.2

31.4

ND 

14.5

 ND*

ND

ND

ND

ND

ND

280

230

230

230

210

230

ND

ND

ND

ND

ND

ND

28.7

41.2

20.3

26.3

10.9

28.6

Surface Water

SW-01

SW-02

SW-03

83.0

81.1

66.5

23.4

22.8

22.7

15.8

13.7

21.1

ND

ND

ND

180

220

210

ND

ND

ND

2.7

43.5

27.0

*ND= Not Detected
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

101

FLOW

THROUGH

CELL

PERISTALTIC

PUMP

TU RBID ITY

METER

GEOPROBE  SCREEN

PO INT G RO U ND W ATER

SAMPLER

Figure 1. Schematic of sub-riverbed sampling equipment

Groundwater samples were collected from beneath the river bed at 73 locations located along 19

transects. The initial sampling locations were based on evaluation of water quality data from

upgradient monitoring wells and evaluation of groundwater flow direction. Subsequent transect and

sampling locations were selected by reviewing daily analytical results received from an on-site field

laboratory. The water samples were collected by driving a GeoProbe®, from a pontoon boat,

approximately two to five feet below the bed of the river. The one foot long GeoProbe® screen was

then exposed, and water was pumped to the surface using a peristaltic pump. The GeoProbe® was fitted

with a thermocouple to permit in-situ measurement of groundwater temperature. A schematic of the

GeoProbe® sampling device is shown on Figure 1. At some locations the GeoProbe® screen became

clogged with fine sediment from the river bed. When this happened the screen was flushed with

deionized water to clear the screen. The sampling then proceeded using a low-flow sampling protocol.

During pumping, the water quality parameters pH, Eh, specific conductance, dissolved oxygen,

temperature, and turbidity were measured in the field. The pH, Eh, specific conductance dissolved

oxygen and temperature were measured utilizing a flow through cell with a YSI model 6820 multi

parameters instrument. Turbidity was measured using an HF model DRT-15CE turbidity meter. A

surface water sample at each sampling location was also analyzed for the field water quality

parameters. The sampling point was purged until the field water quality parameters stabilized. After

the field parameters stabilized the surface water results and the stabilized results from the sampling

point were compared to be certain that groundwater, and not surface water, was being sampled. The

field analyses indicated that pH, temperature and specific conductance were good indicator parameters

for this comparison. The collected sub-river bed samples were then analyzed for the nine VOCs listed

in Table 2. VOC analyses were performed in the field laboratory using a modification of EPA Method

SW-846 8021.
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

102

Figure 2. Sampling grid and contaminant discharge area

Table 2. VOC Analytes

Trichloroethene

Carbon Tetrachloride

1,1,1-Trichloroethane

Chloroform

Vinyl Chloride

Tetrachloroethene

 1,1-Dichloroethene

Chloromethane

1,2-Dichloroethene

The results of the sampling and analyses indicated that TCE contaminated groundwater is

discharging into the river along a 5'500 foot length of the river. The maximum TCE concentration

detected was 1'600 micrograms per liter (μg/5). Groundwater containing CCL4 is discharging to the

river along a 2'500 foot length of the river, and the CCl4 discharge zone is contained within the TCE

discharge zone. The maximum CCL4 concentration was 940 μg/5. The sampling locations, and the TCE

and CCl4 distribution in the sub-riverbed groundwater, are shown on Figure 2. Based on the results of

this investigation, sampling locations for a benthic macroinvertebrate investigation were selected.

AUTHOR INFORMATION

John H. Guswa, Jonathan R. Bridge, and Michael J. Jordan, HSI GeoTrans, Inc. 6 Lancaster 

County Road, Harvard, MA 01451
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

103

Measuring Enhanced Removal of Dissolved

Contaminants in Hyporheic Zones and

Characterizing Causes and Consequences for

Water Quality 

By Judson W. Harvey, Christopher C. Fuller, and Martha H. Conklin

ABSTRACT

Characterizing both the causes and consequences of enhanced oxidation of dissolved manganese

(Mn) in the hyporheic zone at Pinal Creek basin, AZ required measurements with spatial resolution

varying across five orders of magnitude. Our measurements ranged in scale from that of the

fundamental interactions between surface and ground water (centimeters) to the scale of the perennial

stream that receives ground-water discharge from the entire drainage basin (kilometers). Because of

the lower uncertainty of the stream-tracer approach for estimating the average reaction rate, that

method provided the most reliable basin-scale simulation of the effects of enhanced Mn-removal in

hyporheic zones. The stream-tracer characterization alone, however, could not determine that the

removal of manganese was pH-dependent, or even that the reaction occurred in hyporheic zones (as

opposed to slow-moving zones in surface water). Laboratory and in situ measurements within

hyporheic zones provided the crucial evidence to support interpretations about the causal processes.

INTRODUCTION

Hydrologic exchange of streamwater and ground water back and forth across channel beds of

rivers and streams enhances chemical transformations in shallow groundwater beneath the streambed

(hyporheic zone). The hyporheic zone is defined hydrologically by flow paths that route streamwater

temporarily through the subsurface and chemically by subsurface water that can be shown to receive

greater than 10% of its water from the surface (Triska and others, 1993). Steep chemical gradients in

dissolved oxygen, dissolved organic carbon, and pH in hyporheic zones enhance biogeochemically

mediated transformations of solutes, such as nitrification and denitrification (Grimm and Fisher, 1984;

Triska and others, 1993), oxidation of metals (Benner and others, 1995), and biodegradation of volatile

organic compounds (Heekyung and others, 1995). Hyporheic flow paths are typically small in their

spatial dimensions, but if chemical reaction rates are fast enough, and if enough exchange occurs

between flowing water and sediment, then the effects can accumulate downstream and affect water

quality (Harvey and Fuller, 1998).

This short paper considers three types of measurements at different spatial scales of resolution. The

three measurement types are: (1) laboratory-batch experiments that quantify solute-sediment

interactions at the millimeter-scale, i.e. the scale of individual sediment grains, (2) in situ

measurements in hyporheic flow paths at the scale of centimeters beneath the streambed, and (3)

stream-tracer experiments that quantify removal rates at the scale of experimental subreaches in the

perennial stream (approximately 500 meters) or at the scale of the perennial stream that receives

ground-water discharge from the entire drainage basin (3 kilometers). 

A number of physical and chemical measurements of the hyporheic zone have been made as part

of our investigations, including the hyporheic-zone depth, hydrologic residence time in the hyporheic
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

104

zone, net removal-rate constant for dissolved manganese (Mn), and percent removal of Mn in

hyporheic flow paths. Previously, we found good agreement across scales of measurement based on a

relatively limited data set (Harvey and Fuller, 1998). In this paper, we update with new data the means

and standard deviations for manganese removal-rate constants and compare them among the three

measurement types. Field methods, analyses, and modeling calculations for reach-scale and in-situ

measurements are presented in Harvey and Fuller (1998), Fuller and Harvey (1999), and Duff, et al.,

(1998). Laboratory methods and analyses are given by Marble and others (1999) and Harvey and Fuller

(1998). 

COMPARISON OF REMOVAL-RATE CONSTANTS ACROSS SCALES 

In situ rate constants (cm-scale) in the hyporheic zone were determined at a total of eleven sites in

1994, 1995, and 1997. Rate constants were determined for sub-reaches of the perennial stream (500-m

scale) by averaging results from stream-tracer injections in 1994 and 1995. The basin-scale estimates

(3-km scale) were computed by averaging the mean rate constant from the four subreaches in 1994

with the mean for the five subreaches in 1995. Laboratory rate constants estimates were computed

using data from the subset of unpoisoned experiments conducted between pH 6 and 6.9, which

matches the range of pH’s that were measured in situ. We chose the coefficient of variation (standard

deviation divided by the mean) as a measure of uncertainty.

The mean rate constant for the three field estimates (e.g. in situ, sub-reach, and basin-scale) was

approximately 2.3 x 10-4 per second. The mean rate constant determined in the laboratory was

approximately 30% lower. An average rate constant of 2.3 x 10-4 per second for removal of manganese

corresponds to a time constant (inverse of rate constant) of approximately 1.3 hours, which is

comparatively fast in a drainage basin where the hydrologic residence time in surface water of Pinal

Creek is approximately 1 day. Although the mean estimates for each field technique varied little (4%

coefficient of variation), standard deviations varied by approximately a factor of four. The in situ

estimate of the removal-rate constant was most uncertain with a coefficient of variation of 107%.

Estimates made at the kilometer-scale based using the stream-tracer approach were least uncertain,

with a coefficient of variation equal to 26%. The coefficient of variation for laboratory and sub-reaches

had intermediate values of 84% and 56%, respectively.

DISCUSSION

Rate constants for removal of manganese differed little between laboratory experiments, in-situ

field measurements, and measurements based on stream-tracer experimentation. The advantage of

laboratory experiments was the isolation of the effects of microbial colonies and pH. Marble and

others (1999) discuss pH and other factors affecting Mn-oxidation reactions. One problem of the

batch-laboratory experiments is extrapolating results to sediment-water ratios that more accurately

approximate field conditions. Following Harvey and Fuller (1998), we scaled laboratory-rate constants

by multiplying them times the ratio between the average sediment concentration (grams/liter) in the

streambed at Pinal Creek and the sediment concentration used in laboratory experiments. That

adjustment assumes that grain-size variations, which are likely to affect sediment-surface area

available for oxidation of manganese, are the same in laboratory experiments and in the streambed.

Another possible problem of the laboratory experiments is controlling for variation in activity levels of

microbial colonies. For example, Marble and others (this volume) report a significant time lag before

removal in Mn begins in sediment samples that were stored before usage in experiments. Either of

those possible problems might explain the lower Mn removal-rate constant compared with in-situ and

stream-tracer estimates
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

105

In situ sampling within hyporheic flow paths addresses the problem of realistic field conditions by

quantifying rates of removal without disturbing the sediments or natural hydrologic fluxes. But this

method has practical limitations, however. In situ sampling has the disadvantage that the

measurements are difficult and time consuming to make in the field, which limits sample sizes. In

addition there is also the problem that ancillary physical and chemical factors cannot be varied except

though careful site selection. The principal advantage of in-situ field measurements is that interactions

between flow and biogeochemical processes are preserved, which potentially could reveal findings that

would be difficult to detect in a laboratory setting. 

Stream-tracer experiments provided the most reliable reach-averaged rate constants for modeling

the basin-scale consequences of enhanced chemical reactions in hyporheic zones. Nevertheless, there

remains a major disadvantage of the stream-tracer approach for quantifying hyporheic-zone processes.

On the basis of stream-tracer experiments alone, we cannot be sure that the removal of reactive solutes

actually occurs in hyporheic-zones, or on the leaves of aquatic vegetation in slowly-moving surface

water at channel margins or behind channel obstructions. Another problem with stream-tracer methods

is that the detection sensitivity for hyporheic zones is not equal across the multiple types of hyporheic

zones that may be present in a given system (Harvey and others, 1996). Only direct sampling of

hyporheic zones using in-situ methods can provide the independent confirmation needed to support

physical interpretations at larger spatial scales.

SUMMARY AND CONCLUSION

Mean rate constants for the removal of dissolved manganese agreed closely between three scales of

resolution in the field, ranging from centimeter-scale field measurements acquired in situ in hyporheic

zones to kilometer-scale estimates determined using stream tracers. The laboratory estimate of the Mn

removal-rate constant was approximately 30% lower than field estimates. In situ and laboratory rate

constants had relatively large coefficients of variation (107% and 84%, respectively), which may be

too large to be used reliably in transport simulations. Stream-tracer experiments provided estimates of

the rate constant with lower uncertainties; 56% when averaged at the reach-scale (approximately 500

meters) and 26% when averaged at the basin-scale (3 kilometers). Our experience at Pinal Creek basin

leads us to conclude that a multi-scale approach is a necessity for characterizing enhanced

biogeochemical reactions in hyporheic zones. 

REFERENCES

Benner, S.G., Smart, E.W., and J. N. Moore, 1995. Metal behavior during surface-groundwater

interaction, Silver Bow Creek, Montana, Environmental Science and Technology, v. 29, pp. 1789-

1795.

Duff, J.H., Murphy, F., Fuller, C.C., Triska, F.J., Harvey, J.W., and A.P. Jackman, 1998. A mini

drivepoint sampler for measuring pore water solute concentrations in the hyporheic zone of sandbottom streams, Limnology and Oceanography, v. 43, no. 6, pp. 1378-1383.

Fuller, C.C. and J.W. Harvey, 1999. The effect of trace-metal reactive uptake in the hyporheic zone on

reach-scale metal transport in Pinal Creek, Arizona, 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 1—Contamination from Hard

Rock Mining: U.S. Geological Survey Water-Resources Investigations Report 99-4018A, this

volume.

Grimm, N.B. and S.G. Fisher, 1984. Exchange between interstitial and surface water: implications for

stream metabolism and nutrient cycling, Hydrobiologia, v. 111, pp. 219-228. 
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

106

Harvey, J.W. and C.C. Fuller, 1998. Effect of enhanced manganese oxidation in the hyporheic zone on

basin-scale geochemical mass balance, Water Resources Research, v. 34, pp. 623-636.

Harvey, J.W., Wagner, B.J., and Bencala, K.E., 1996. Evaluating the reliability of the stream-tracer

approach to characterize stream-subsurface water exchange, Water Resources Research, v. 32, no.

8, pp. 2441-2451.

Heekyung, K., Hemond, H.F., Krumholz, L.R., and B.A. Cohen, 1995. In-situ biodegradation of

toluene in a contaminated stream, 1, Field Studies, Environmental Sciences and Technology, v. 29,

pp. 108-116.

Marble, J.C., Corley, T.L., Conklin, M.H., and C.C. Fuller, 1999. Environmental factors affecting

oxidation of manganese in Pinal Creek, Arizona., 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 1—Contamination from Hard

Rock Mining: U.S. Geological Survey Water-Resources Investigations Report 99-4018A, this

volume.

Marble, J.C., 1998. Biotic contribution of Mn(II) removal at Pinal Creek, Globe, Arizona, unpublished

M.S. thesis, University of Arizona, Department of Hydrology and Water Resources, Tucson. 91

pp.

Triska, F.J., Duff, J.H., and R.J. Avanzino, 1993. The role of water exchange between a stream

channel and its hyporheic zone on nitrogen cycling at the terrestrial-aquatic interface,

Hydrobiologia, v. 251, pp. 167-184.
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

107

Bioassessment of Hyporheic Microbial

Communities Using a Specially-designed

Sediment Colonization Chamber

By Susan P. Hendricks

Streambed sediments are often very heterogeneous in particle size distribution and permeability,

creating mosaics of hyporheic habitats and biotic communities. Biotic patchiness and difficulties

sampling the physical environment present challenges for researchers investigating hyporheic

microbial transformation processes. Heterogeneous sediments may not be conducive to conventional

sampling techniques (Fraser and Williams 1997, Mauclaire, et al. 1998). For example, streambeds

composed of large gravel and cobble are prohibitive to mechanical or hand-coring of sediments.

Freeze-core sampling techniques often used in heterogeneous sediments for invertebrates are not

appropriate because freezing alters microbial activity (Humpesch and Niederreiter 1993, Claret 1998a,

1998b). 

Some sampling difficulties may be overcome using artificial chambers. Many investigators have

designed chambers for various specific monitoring purposes, including water chemistry,

macroinvertebrates, and in situ microbial metabolism measurements (Danielopol and Niederreiter

1987, Dodds, et al. 1996, Shati, et al. 1997). There also have been several site- and/or question-specific

chamber designs, particularly for sediment microbial studies (e.g., Fischer, et al. 1996, Frazer, et al.

1996, Eisenmann, et al. 1997, Claret 1998a, 1998b). The purpose of this paper is to describe a multipurpose sediment microbial colonization chamber that combines attributes of several previous designs.

The chamber has been used successfully in heterogeneous cherty western Kentucky and Tennessee

streams and can be used not only for chemical and microbial monitoring, but also for experimental

manipulations in situ. 

DESIGN CONCEPT

Gravel bars are conspicuous geomorphic features of many mid-reach streams (Figure 1).

Conservative tracer experiments in the study streams have shown that both gravel bars and the

hyporheic zone are important transient storage zones where dissolved organic matter, nutrients and

contaminants may be retained for periods of time and transformed before re-entering the surface

stream environment. Microbial activity within these subsurface regions is important in mediating

nutrient and carbon cycling (Hendricks and White 1991, 1995; Hendricks 1993, 1996) and potentially

important in transforming contaminants. 

The chamber was designed for maximum flexibility in placement and function within both the

hyporheic zone and within gravel bars lateral to the stream (Figure 1). The design (Figure 2) has

allowed us to 1) collect interstitial water samples for nutrient chemistry and dissolved oxygen, 2) subsample sediments for various microbial assays (e.g., bacterial productivity, phosphatase activity), and

3) carry out time-course in situ experiments for determining transformation rates of various substances

(e.g., respiration rates, nitrification rates). Additionally, the sediment volume within the chamber

enables 1) collection of small test-tube sized cores from each of the depth intervals for further

laboratory microbial analyses (e.g., microbial diversity studies using amplified rDNA restriction

analysis, fatty acid analysis, perfusion experiments) and 2) transplant experiments between sites and

streams and monitor subsequent changes in activities and other effects. Our initial experiments have
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

108

Figure 1. Study sites at Ledbetter and Panther Creeks. Black

arrows indicate groundwater inflow: Gray arrows indicate

flowpaths of surface water into gravel bars and hyporheic zone

beneath stream.

been comparisons between an

agriculturally impacted third-order

stream (Ledbetter Creek) and a pristine

third-order stream (Panther Creek)

with similar sediment heterogeneities

and watershed characteristics. 

METHODS

Sediments collected from each

stream were sieved (£ 3 mm size

fraction), autoclaved, and placed into

chamber baskets (Figure 2). The

baskets were stacked such that each

represented a hyporheic depth interval

(0-10 cm, 15-25 cm, and 30-40 cm).

Chambers were placed just below the

top of the water table along previously

determined subsurface flow-paths

within the gravel bars and below the

sediment-water interface within

hyporheic regions (Figure 1). Ports

between inner and outer cylinders were

aligned in the open position to allow

interstitial flow and colonization with

microflora for approximately 8-10

weeks.

Following colonization, the

chamber caps were removed and

interstitial water was withdrawn for

dissolved O2, nutrients, and DOC from

each of the stacked sediment baskets by attaching a syringe to the tubing embedded in the basket

(Figure 2). Baskets were then retrieved from the inner cylinder by pulling up on the central stem.

Sediment sub-samples were collected from each basket, placed into sterile containers, and transported

to the laboratory for phosphatase activity, bacterial productivity and microbial diversity assays.

In situ experiments also were conducted. For example, interstitial dissolved O2 and NO3-N samples

were collected from each depth interval within the chamber as pre-incubation references (i.e., ports on

both cylinders aligned in the open position). Twisting the inner cylinder in the opposite direction

resulted in closing of the ports and isolation of sediments from interstitial flow. Following incubation

in the closed position for a chosen period of time (e.g., 24-72 hours), interstitial water was re-sampled.

Oxygen consumption (respiration rate) and NO3-N accumulation (nitrification rate) were estimated as

the difference between dissolved O2 or NO3-N concentrations before and after incubation over time

and expressed as change in concentration g-1 wet weight or L-1 sediment h-1. 
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

109

Hyporheic/Gravel Bar Sediment Chamber

CAP

Chambers are buried in 

streambed to just below the 

sediment surface and in 

gravel bars to just below the 

top of the water table

Chamber constructed of PVC 

pipes - Outer chamber is 

casing for inner chamber, 

inner chamber contains 

sediment baskets made of 

screening.

The inner cylinder holding the 

sediment baskets can be 

turned to align the openings 

with outer cylinder exposing 

sediments to interstitial flow, 

or closed off to isolate 

sediments from flow.

Tubing from which interstitial water samples can 

be collected from each depth interval. Syringe

Valve

Connector tubing

Stem to facilitate basket removal

0-10 cm depth

15-25 cm depth

30-40 cm depth

spacers

Spacers for isolating depth interv

OFF

60c

5040302010

20oz

1 1⁄2

1

1⁄2

B-D

Figure 2. Schematic of gravel bar/hyporheic chamber design.

RESULTS

Some examples of subsurface

microbial activities important in P,

N and C cycles are summarized in

Tables 1-3. The tabulated data are

means calculated from all depths (3)

within replicate chambers (2) within

a site (gravel bar=3, hyporheic=2)

for each stream. Alkaline

phosphatase activity (Sayler, et al.

1979) was higher in the gravel bar

than in the hyporheic zone at

Ledbetter Creek during both summer

and spring sampling periods, and

higher in general than in Panther

Creek (Table 1), indicating

differences in phosphorus demand

by the microbial community

between the two streams and among

sites within the streams. 

The Ledbetter Creek gravel bar

generally showed higher bacterial

productivity (methods modified

from Findlay 1993) than the Panther

Creek gravel bar (Table 2). Bacterial

productivity was highest where

interaction between the subsurface

gravel bar and stream surface water

was greatest as indicated by

dissolved O2 gradients in both gravel bars (data not shown) along subsurface flow-paths (Figure 1). 

Nitrification rate (methods of Jones, et al. 1995), reported as the increase in μg NO3+NO2 L-1

sediment h-1, was higher in Ledbetter Creek than in Panther Creek (Table 3). The Ledbetter Creek

sediment bacterial community is composted of taxa, which appear adapted to high levels of NH4NO3

fertilizer applications.

Table 1. Alkaline phosphatase activity (APA = μm nitro-phenylphosphate

reduced g-1 sediment dry wt.) in hyporheic and gravel bar chambers at

Ledbetter and Panther Creeks. AG = agriculturally impacted, P = pristine.

Stream Site Spring Summer

Ledbetter (AG) Gravel Bar 1020.5 1942.9

Hyporheic 700.9 205.4

Panther (P) Gravel Bar 687.7 370.9

Hyporheic no data no data
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

110

Table 2. Subsurface gravel bar bacterial productivity (μg C m-2 h-1) at Ledbetter and

Panther Creek as estimated from incorporation of 3

H-thymidine into bacterial DNA. AG =

agriculturally impacted, P = pristine.

Gravel Bar Position

Stream Month Upstream Mid-Bar Downstream

Ledbetter (AG) Jan 3.13 11.7 32.3

Jun 64.5 73.6 225.9

Sep 37.7 67.1 67.1

Panther (P) May 72.1 58.6 44.9

Aug 147.0 51.6 48.0

Table 3. Nitrification rates measured as the increase in NO3+NO2 in sediments (μg NO3+NO2

produced L-1 sediment h-1) in Ledbetter and Panther Creek chambers. AG = agriculturally

impacted, P = pristine.

Stream Site ûNO3 (μg L-1) % Increase

Rate

(μg NO3 L-1

sediment h-1)

Ledbetter(AG) Gravel Bar 801.2 +340 11.13

Hyporheic 238 +260 3.30

Panther (P) Gravel Bar 4.5 +5 0.06

Hyporheic 8.0 +8 0.11

DISCUSSION

It is well known that agricultural practices increase sedimentation and greatly alter the chemistry of

surface waters. However, land-use effects on microbial and biogeochemical processes at the

groundwater-surface water interface (hyporheic zone) mediated by increased nutrient, carbon, and

sediment loads are largely unknown. Contaminants reaching streams from subsurface sources such as

groundwater are expected to be processed/transformed at the groundwater-surface water interface

depending on heterogeneity and permeability of sediments and subsurface flow-path complexity.

Methods and data presented here have focused primarily on delineating differences in hyporheic zone

function that mediate agricultural and suburban runoff between and within streams. Results presented

above are limited examples of data which might be obtained from colonization chambers. It is feasible

to examine other processes which indicate disturbance or alteration of function by other contaminants

entering streams from either point or non-point sources (e.g., contaminated groundwaters). 

CONCLUSIONS

The sediment microbial colonization chamber described in this paper appears to be a reasonable

device for examining microbial activities and biogeochemical transformations within the hyporheic

zone or at the groundwater-surface water interface within streambeds. 

Advantages of using the sediment colonization chamber described here are 1) construction

materials are inexpensive and the design is flexible for a variety of streambed types, 2) chambers are

multi-purpose in that both interstitial water chemistry and sediment sampling can be done, 3) time

course incubations can be carried out, 4) transplanting of chambers for inter- and intra-site
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

111

comparisons may be carried out, and 5) either natural sediments or more homogeneous artificial

particles (e.g., glass or ceramic beads) may be used in the chambers, 6) replication is quite good and

data are consistent between replicate chambers. 

Some disadvantages may include 1) installation that may require 2 or more people, 2) colonization

periods may be long (6 weeks minimum, 8-10 weeks preferred), 3) chamber sediments may not reflect

actual particle size distributions found in streambeds, 4) chambers may prohibit infiltration of natural

CPOM (microbial fuel), and 5) chambers may alter local subsurface hydraulics. Continued monitoring

of the chambers over time and space will help evaluate their ultimate usefulness in stream ecosystems.

REFERENCES

Claret, C., 1998a. A method based on artificial substrates to monitor hyporheic biofilm development,

Internat. Rev. Hydrobiol., v. 83, pp. 135-143.

Claret, C., 1998b. Hyporheic biofilm development on artificial substrata, as a tool for assessing trophic

status of aquatic systems: first results. Annls. Limnol., v. 34, pp. 119-128.

Danielopol, D.L. and R. Niederreiter, 1987. A sampling device for groundwater organisms and oxygen

measurement in multi-level monitoring wells, Stygologia, v. 3, pp. 252-263.

Dodds, W.K., C.A. Randel, and C.C. Edler, 1996. Microcosms for aquifer research: application to

colonization of various sized particles by ground-water microorganisms, Ground Water, v. 34,

pp.756-759.

Eisenmann, H., W. Traunspurger, and E.I. Meyer, 1997. A new device to extract sediment cages

colonized by microfauna from coarse gravel river sediments, Arch. Hydrobiol., v.139, pp. 547-561.

Findlay, S., 1993. Thymidine incorporation into DNA as an estimate of sediment bacterial production,

in: Kemp, P., B. Sherr, E. Sherr, and J. Cole (eds), Handbook of methods in aquatic microbial

ecology, Lewis Publishers, Boca Raton, FL.

Fischer, H., M. Pusch, and J. Schwoerbel, 1997. Spatial distribution and respiration of bacteria in

streambed sediments, Arch. Hydrobiol., v.137, pp. 281-300.

Fraser, B.G. and D.D. Williams, 1997. Accuracy and precision in sampling hyporheic fauna, Can. J.

Fish. Aquat. Sci., v. 54, pp. 1135-1141.

Fraser, B.G., D.D. Williams, and K.W.F. Howard. 1996. Monitoring biotic and abiotic processes

across the hyporheic/groundwater interface. Hydrogeology Journal, v. 4, pp. 36-50.

Hendricks, S.P. 1996. Bacterial biomass, activity, and production within the hyporheic zone of a northtemperate stream, Archiv für Hydrobiologie, v.135, pp. 467-487.

Hendricks, S.P. and D.S. White, 1995. Seasonal biogeochemical patterns in surface water, subsurface

hyporheic, and riparian groundwater in a temperate stream ecosystem, Archiv für Hydrobiologie, v.

134, pp. 459-490.

Hendricks, S.P. and D.S. White, 1991. Physicochemical patterns within a hyporheic zone of a northern

Michigan river, with comments on surface water patterns, Can. J. Fish. Aquat. Sci., v.48, pp. 1645-

1654.

Hendricks, S.P., 1993. Microbial ecology of the hyporheic zone: a perspective on the integration of

hydrology and biology, J. No. Amer. Benthol. Soc., v. 12, pp. 70-78.

Humpesch, U.H., and R. Niederreiter, 1993. Freeze-core method for sampling the vertical distribution

of the macrozoobenthos in the main channel of a large deep river, the River Danube at river

kilometre 1889, Arch. Hydrobiol. Suppl., v. 101, pp. 87-90.

Jones, J.J., S.G. Fisher, and N.B. Grimm, 1995. Nitrification in the hyporheic zone of a desert stream

ecosystem, J. N. Am. Benthol. Soc., v. 14, pp. 249-258.

Mauclaire, L., P. Marmonier, and J. Gibert, 1998. Sampling water and sediment in interstitial habitats:

a comparison of coring and pumping techniques, Arch. Hydrobiol., v.142, pp. 111-123.
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

112

Sayler, G.S., M. Puziss, and M. Silver, 1979. Alkaline phosphatase assay for freshwater sediments:

application to perturbed sediment systems, Appl. Environm. Microbiol., v. 38, pp. 922-927.

Shati, M.R., D. Rönen, and R. Mandelbaum, 1996. Method for in situ study of bacterial activity in

aquifers, Environm. Sci. Tech., v. 30, pp. 2646-2653.

AUTHOR INFORMATION

Susan Hendricks, Hancock Biological Station, Murray State University, Murray, KY 42071;

susan.hendricks@murraystate.edu.
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

113

Figure 1. A standard lipid-containing SPMD with three molecular

welds near each end. Note that low interfacial tension causes

intimate contact (i.e., the present of a lipid film on the membrane

interior surface) between the triolein and the membrane even

where air bubbles exists.

Fundamentals of SPMD Sampling, Performance,

and Comparability to Biomonitoring Organisms

By J.N. Huckins, J.D. Petty, H.F. Prest, J.A. Lebo, C.E. Orazio, J. Eidelberg, W.L.

Cranor, R.W. Gale, and R.C. Clark

INTRODUCTION

Passive monitoring devices have long been used by the chemical industry and governmental 

agencies to ensure compliance to OSHA standards for time-weighted-average (TWA) concentrations

of organic vapors in the work environment. With the recent development of lipid-containing

semipermeable membrane devices (SPMDs) and samplers based on diffusive gradients across thin

polymeric films (DGTs), the passive in situ monitoring approach can now be applied to determining

TWA concentrations of both hydrophobic organics (SPMDs) and heavy metals (DGTs) in aquatic

environments (1-3). In this work, we focus on the fundamentals of SPMD technology, and the potential

utility of the approach for monitoring organic contaminants in groundwater (includes the hyporheic

zone).

SPMD SOURCE, DESIGN AND THEORY

The SPMD technology is the subject of two government patents and the devices are 

commercially available from Environmental Sampling Technologies, 1717 Commercial Drive, St.

Joseph, MO 64503. A standard SPMD consists of a thin walled (75-95 m) layflat tube of low density

polyethylene (LDPE) containing a thin film of 95% pure triolein (Figure 1). The ends of the LDPE, are

welded by heat-sealing. The length or size of the standard device can be customized to fit an

investigator’s sampling needs. The membrane surface area-to-lipid-volume ratio of a standard SPMD

is 450 cm2

/m5 triolein. Because the LDPE membrane is nonporous (i.e., fixed pores do not constitute a

significant portion of the polymer free volume) and transient cavities in the membranes are generally <

10 in diameter, only dissolved or readily bioavailable compounds are sampled.

A fundamental characteristic of passive monitors used to estimate TWAs of contaminants is the

proportionality of sampler concentrations to ambient environmental concentrations. Also, such

monitors are typically designed to be an infinite sink for target analytes. This characteristic results in a
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

114

constant volume of air or water extracted per unit time (i.e., linear uptake of analytes at constant

ambient concentrations), and the following equation applies

Cw = CSPMD VSPMD/Rs t (1)

where Cw is the TWA analyte concentration in water (ng/5), CSPMD is the concentration in the whole

device (membrane + lipid) in ng/5, VSPMD is the volume of the SPMD (L), Rs is the sampling rate for a

chemical in L/d, and t is exposure time in days (d).

Huckins et. al. (4) have shown that Rs values for PAHs and OCs ranged from 1 to 8 L/d (exposure

conditions: 10-26 oC, water velocity < 1cm/sec), when using a standard SPMD with VSPMD = 5mL.

Thus, in the linear region of analyte uptake, an SPMD with VSPMD = 1 m5 will daily extract dissolved

contaminants from 200 to 1,600 m5 of water. If the target compounds have relatively low octanol-water

partition coefficients (i.e., log Kow < 4), such as VOCs, equilibrium between the device and the

surrounding water is often achieved in < 1 week. In that case, the following simple model can be used

for water concentration estimates:

Cw = CSPMD/ KSPMD (2)

where KSPMD is the equilibrium SPMD-water partition coefficient. For compounds with log Kow of 4.0

and > 4.0, KSPMD is 0.75 Kow and 0.3 Kow, respectively.

APPLICABILITY OF THE APPROACH

Standard SPMDs are designed to sample nonionic hydrophobic compounds. The total volume of

water extracted by an SPMD at equilibrium is estimated by 0.75 Kow VSPMD (log Kows 4.0) and 0.3 Kow

VSPMD (log Kows > 4.0). For example, if the target analyte has a Kow of 300 the maximum (equilibrium)

volume of water extracted by a standard 1 m5 triolein SPMD (VSPMD 5 mL) is only a little more (i.e.,

1.1L) than a 1L grab sample. Thus, standard SPMDs are useful only for sampling compounds with

Kows > 300, unless an adsorbent is dispersed in the triolein (4). 

Trace levels of a variety of classes of organic contaminants have been successfully determined in

aquatic systems using SPMDs. These include but are not limited to the following: polycyclic aromatic

hydrocarbons, polychlorinated -biphenyls and -terphenyls, organochlorine pesticides, polychlorinateddibenzodioxins and -dibenzofurans, chlorinated and brominated diphenyl ethers, chlorinated -

benzenes, -anisoles and -veratroles, certain alkylated and chlorinated phenols, heterocyclic aromatics,

pyrethroid and nonpolar to moderately polar organophosphate pesticides, and nonionic organometals.

DEPLOYMENT CONSIDERATIONS

Because SPMDs readily sample a broad spectrum of chemicals from air, exposure to organic 

vapors is minimized by transport to and from the sampling site in clean gas-tight metal cans. Other

precautions are similar to those used for standard grab sampling methods. 

The appropriate exposure duration is dependent on the physicochemical properties (e.g., Kows) of

the target analytes, analytical sensitivity needed, choice of sampling approach (i.e., integrative or

equilibrium), environmental conditions, and the potential for vandalism. When calibration data are
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

115

available (4,6) or can be estimated for contaminants of concern, the following model can be used to

determine the time required to reach half of the SPMD equilibrium concentration (t1/2):

t1/2 = ln 0.5 KSPMD VSPMD/Rs (3)

Knowledge of these half-times or half-lives is useful because the uptakes of analytes are linear during

1 t1/2.

Thus, within 1 t1/2 equation 1 can be used for water concentration estimates. If an equilibrium

sampling approach is used, exposure time should be > 4 t1/2s and equation 2 is applicable. SPMD

sampling rates are affected by temperature, water velocity, and biofouling. Fortunately, temperature,

and to a lesser extent hydraulic conductivity (velocity), are generally more constant in ground water

systems than in surface waters. In surface waters, biofouling impedes analyte uptake, often limiting the

utility of long exposure times. However, biofouling is much reduced in groundwater, thereby

permitting significantly longer exposures (months instead of weeks) for compounds with high Kows

(i.e., 6.0).

Even with the effects of the aforementioned environmental variables on sampling rates, Ellis et. al.

(5) have shown that river water concentrations (dissolved phase) of trace organic contaminants can be

estimated from SPMD concentrations within two-fold accuracy. Huckins, et al. (1,4) have suggested

that the use of permeability reference compounds (PRCs) may further reduce errors in water

concentration estimates. PRCs are analytically noninterfering compounds, such as deuterated PAHs

with log Kows < 5.0, that are added to SPMD lipid before deployment. By determining PRC loss rates

(Rs/KSPMD VSPMD or k2 ) from SPMDs during an environmental exposure and comparing them to PRC

k2s measured during the experimental determination of sampling rates (Rss), the laboratory derived Rss

of analytes can be adjusted to account for the effects of field exposure conditions.

COMPARABILITY TO BIOMONITORING ORGANISMS

The accumulations of organic contaminants by SPMDs and aquatic organisms have been compared

in a number of studies (2,4-6). Some studies have shown that the concentration patterns and uptake

rates of several classes of chemicals by SPMDs and fishes are similar (6,7). However, it is unrealistic

to expect SPMDs to mimic the uptake of all organic contaminants by all aquatic species because large

differences exist among species in regard to diet/source of energy, metabolic activities (xenobiotics),

and lipid composition and percentages. For example, few aquatic organisms contain 20% lipid by

weight as SPMDs do. Thus, the capacities of most aquatic organisms to retain accumulated residues

are nearly always less than SPMDs (i.e., organism k2>>SPMD k2), which suggests that organisms used

as biomonitors may not retain detectable levels of some residues several days after an episodic

contaminant release. 

GROUNDWATER APPLICATION

Although SPMDs are used extensively in surface waters (2,4) and in the atmosphere (8), their

application to ground water systems has been limited to a few studies. This is surprising because

laboratory studies performed to determine SPMD sampling rates (4) more closely simulate the

relatively constant conditions existing in some ground water systems.

Herein, we highlight a pilot study on the use of SPMDs by EPA Region 9 personnel for sampling

dieldrin in groundwater at the George Air Force Base, California. The devices were deployed in two
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

116

ground water monitoring wells where dieldrin residues were previously determined. The exposure

period was 29 days. Based on laboratory calibration studies (4), each standard SPMD sampled a total

volume of 58 5 of groundwater during the 29-day exposure. Figure 2 illustrates the results of the

analysis (high resolution gas chromatography-electron capture detection [ECD]) of SPMD extracts

from well water at one of the sites, and of an associated SPMD control. Microgram quantities of

dieldrin were concentrated in each of two replicate SPMDs and the SPMD-derived TWA water

concentration was estimated at 69 ng/5 (recovery-corrected). Earlier, a grab sample at the same site was

found to contain 110 ng/5 of dieldrin (includes sorbed fraction). Much lower levels (140 to 2,800-fold

less) of fifteen other chlorinated pesticides (e.g., chlordane components, endrin, endosulfan II and

sulfate, etc.) were also detected in the SPMDs. These trace contaminants were not detected using

EPA’s (CLP) low-level pesticide method. In summary, SPMDs appear to be well suited for monitoring

trace hydrophobic organics in ground water systems.

Figure 2. ECD Chromatograms of an SPMD sample extract and a control SPMD. SPMDs were deployed for

29 days in a ground water monitoring well at George Air Force Base, CA. Dieldrin and 15 other organochlorine pesticides were detected in the sample. The bottom two chromatograms were diluted 100-fold to

keep dieldrin on scale, and octachloronaphthalene (OCN) was used as an instrumental internal standard.

REFERENCES

(1) Huckins, J.N., Manuweera, G.K., Petty, J.D., Mackay, D., and J.A. Lebo, 1993. Environ. Sci.

Technology, v. 27, pp. 2489-2496.

(2) Huckins, J.N., Petty, J.D., Lebo, J.A., Orazio, C.E., Prest, H.F., Tillitt, D.E., Ellis, G.S., Johnson,

B.T., and G.K. Manuweera, 1996. In Techniques In Aquatic Toxicology; G.K. Ostrander (Ed.),

CRC-Lewis Publishers, Boca Raton, FL, pp. 625-655.

(3) Davison, W. and H. Zhang, 1994. Nature, pp. 367,545.

(4) Huckins, J.N., Petty, J.D., Prest, H.F., Orazio, C.E., and R.C. Clark, 1999. Guide for the Use of

Semipermeable Membrane Devices (SPMDs) as Samplers of Waterborne Hydrophobic Organic

Contaminants, Reports for the American Petroleum Institute (API), API: 1220

(5) Ellis, G.S., Huckins, J.N., Rostad, C.E., Schmitt, C.J., Petty, J.D., and P. MacCarthy, 1995.

Environ. Toxicol. Chem., v.14, pp. 1875-1884.
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

117

(6) Meadows, J.C., Echols, K.R., Huckins, J.N., Borsuk, F.A., Carline, R.F., and D.E. Tillitt, 1998.

Environ. Sci. Technol., v. 32, pp. 1847-1852.

(7) Peven, C.S., Uhler, A.D., and F.J. Querzoli, 1996. Environ. Toxicol. Chem., v. 15, pp. 144-149.

(8) Ockenden, W.A., Prest, H.F., Thomas, G.O., Sweetman, A., and K.C. Jones, 1998. Environ. Sci.

Technol., v. 32, pp. 1538-1543.

AUTHOR INFORMATION

J.N. Huckins and J.D. Petty, Columbia Environmental Research Center, BRD, USGS, 4200 New

Haven Road, Columbia, MO. 65201.

H.F. Prest, J.A. Lebo, and C.E. Orazio, Institute of Marine Science, Long Marine Laboratory,

University of California Santa Cruz, Santa Cruz, CA 95060.

J. Eidelberg, W.L. Cranor, R.W. Gale, and R.C. Clark, U.S. EPA, 75 Hawthorne Street, San Francisco,

CA 94105.
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

118

Acid Mine Drainage—The Role of Science

By Briant Kimball

Thousands of abandoned and inactive mines are located in environmentally sensitive mountain

watersheds. Cost-effective remediation of the effects of metals from mining in these watersheds

requires knowledge of the most significant sources of metals. The significance of a given source not

only depends on the concentration of a toxic metal, but also on the total mass of metal added to the

stream. This discussion deals with accounting for the mass of metal that enters the stream, which is

called the mass loading. It is calculated as the product of metal concentration and stream discharge.

Without discharge measurements, the overall effect of high metal concentrations on streams and

aquatic organisms is unclear. A traditional discharge measurement is obtained by dividing a stream

into small sections and measuring cross-sectional area and average water velocity in each section.

Because the channel bottom in mountain streams is not smooth and much of the flow is among the

streambed cobbles, accurate discharge measurements are difficult to obtain, even under the best

conditions.

AN APPROACH FOR MOUNTAIN STREAMS

A recent study by the U.S. Geological Survey Toxic Substances Hydrology Program illustrates a

practical approach to obtaining and using discharge measurements in mountain streams. Chalk Creek,

a tributary of the Arkansas River in Colorado, receives mine drainage from the Golf Tunnel adit.

Metal-rich mine drainage from the Golf Tunnel is routed around waste rock and a capped tailings pile

into a constructed wetland. From the wetland, the mine drainage enters Chalk Creek from small

springs and seeps along the stream. Regulatory and land management agencies have asked three basic

questions about Chalk Creek. First, is there more than one source of mine drainage that affects the

stream? Second, does a remediation plan need to account for drainage from more than one source?

Finally, have past remediation efforts been successful? To address these questions, we employed a

tracer-dilution study to determine discharge and synoptic sampling to obtain detailed chemical

composition from many locations. The synoptic samples are collected during a short period of time,

typically a few hours, providing a “snapshot” of the changes along a stream at a given point in time.

ADDING A TRACER: DISCHARGE BY DILUTION

Discharge in mountain streams can be measured precisely by adding a dye or salt tracer to a

stream, measuring the dilution of the tracer as it moves downstream, and calculating discharge from

the amount of dilution. Because we know the concentration of the injected tracer and the rate at which

it is added to the stream, we know the mass added to the stream. By measuring the concentration of the

tracer upstream and downstream from the injection point, we can calculate the discharge by dilution of

the tracer in the stream. To define discharge in Chalk Creek, a sodium chloride tracer was added at a

constant rate for 24 hours at a point upstream from the mine drainage. The chloride injection was

monitored at several sites downstream from the injection point, documenting the incremental increase

of discharge due to water entering the stream. The difference in discharge between two stream sites

gives the total amount of inflow from surface- and ground-water sources in that small reach.
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

119

SYNOPTIC SAMPLING: A “SNAPSHOT” IN TIME TO COMPARE METAL LOADING OF SOURCES

Synoptic samples provided metal concentrations, giving a detailed profile of zinc concentrations in

both the stream and inflows along the stream reach. To evaluate these concentrations, a mass-loading

profile was calculated from the concentrations and the discharge values. The concentrations and the

mass-loading profile help answer the basic questions about the sources of metals and the effectiveness

of remediation. First, there appears to be more than one source of mine drainage, because the high

concentration at 252 meters could not be from the Golf Tunnel. Second, despite the higher

concentration of zinc in water from the second source, the loading profile shows that the highconcentration water only contributes about 8 percent of the zinc load and would not require a separate

remediation plan. Finally, there are still effects on metals in the stream where old tailings were

removed downstream from 300 meters; these effects will likely decrease with time. The example of

Chalk Creek shows that the highest inflow concentrations do not always result in the most significant

sources of metal loading. Our results show that a site can be investigated in great detail to help make

decisions by using tracer injections and synoptic sampling.

AUTHOR INFORMATION

Briant Kimball, U.S. Geological Survey; bkimball@usgs.gov. 

(Mr. Kimball’s article is adapted, by permission, from an article published by the USGS. in January

1997. More information on the Toxic Substances Hydrology Program can be obtained at

http://toxics.usgs.gov/toxics.)
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

120

Temporal and Spatial Trends in Biogeochemical

Conditions at a Groundwater-Surfacewater

Interface

By John M. Lendvay and Peter Adriaens

BACKGROUND

The Bendix, Allied Signal National Priority List (NPL) site at St. Joseph, Michigan, has been

extensively characterized for contaminant distribution and biogeochemical conditions between the

contaminant source and zone of emergence in Lake Michigan [Tiedeman and Gorelick, 1993;

Semprini, et al., 1995; Lendvay, et al., 1998a; Lendvay, et al., 1998b]. The source of the contaminant

plume at the site consists of trichloroethene (TCE) and minor contamination with hydrocarbons.

Contaminant hydrocarbons and natural organic matter have stimulated sufficient indigenous microbial

activity in the groundwater to result in anaerobic conditions, predominantly sulfate-reducing and

methanogenic. Under these terminal electron accepting processes (TEAPs), TCE has been reductively

dechlorinated to predominantly cis-1,2-DCE, chloroethene, and ethene with minor production of 1,1-

DCE, trans-1,2-DCE, and ethane. 

Since the source of contamination is located approximately 750m up-gradient of Lake Michigan,

the flow of groundwater toward the lake raised concern about the potential contamination of Lake

Michigan with reductive dechlorination products, particularly chloroethene. Between 1994 and 1996,

three transects of temporary bore-holes were established on the beach and approximately 100 meters

from shore under the lake bottom to monitor the contaminant distribution and oxidation-reduction

conditions at the GSI [Lendvay, et al., 1998a]. This study of the GSI suggested that the predominant

TEAPS under Lake Michigan and in the zone along the beach was sulfate-reduction interspersed with

methanogenic conditions. Furthermore, reductive dechlorination was the predominant contaminant

transformation where these TEAPs predominated. However, in shallow regions of the contaminant

plume near the lake shoreline, iron-reducing conditions predominated, which are conducive to either

dechlorination [McCormick and Adriaens, 1998] or oxidation [Bradley and Chapelle, 1996] reactions.

The more oxidized iron-reducing zone along the top of the GSI was hypothesized to result from reoxygenation of the plume by surface water run-up, infiltration, and wave activity. Furthermore, it was

hypothesized that re-oxygenation of the aquifer to hypoxic conditions resulting from wave activity

might provide a suitable environment for aerobic commensalic or cometabolic biodegradation

processes in the shallow zone along the beach.

The goals of this current study were to: i) evaluate the temporal effects of increased wave 

activity on the TEAPs and contaminant distribution; ii) determine the most likely microbial processes

affecting intrinsic remediation of the contaminants at the GSI; iii) present laboratory data to

corroborate field observations; and iv) determine field oxidative flux of chloroethene at the GSI.

METHODS

Multi-level arrays were placed to capture spatial variations in contaminant distribution and

predominant TEAPs, seasonal changes and effect of wave activity, and spatial infiltration of lake water

into the GSI [Lendvay, et al., 1998b; Lendvay, et al., 1999b]. The arrays were semi-permanent to

evaluate temporal effects of lake-activity on contaminant distribution and TEAPs during a six month
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

121

Lake Michigan Wave Data

(NOAA Buoy 45007)

0.0

0.5

1.0

1.5

2.0

Jun Jul Aug Sep Oct Nov Dec

Month 1996

Avg. Wave Height (m)

Figure 1: Average wave height for Lake Michigan as measured by

NOAA buoy 45007 during the sample time study (left). Error bars

represent one standard deviation. Picture of GSI after a major storm

event (right). Notice the scatering of large debris, not present prior to

the storm.

period (July – December)

encompassing seasonal changes in

weather and lake activity. Seasonal

changes were evidenced by changes in

Lake Michigan wave height, as

measured by National Oceanographic

and Atmospheric Administration

(NOAA) buoy 45007, averaged 0.53m

for August and September and 1.17 m

for November and December with

height increasing steadily between

June and December (Figure 1)

[Lendvay, et al., 1998b]. 

Groundwater was sampled five

times, in four to six week intervals,

from each sample point between July

and December [Lendvay, et al.,

1998b]. Specifically, temperature, pH, reduction potential (redox), specific conductance, and dissolved

oxygen were measured using a QED flow cell [Ann Arbor, MI]. Dissolved oxygen, aqueous ferrous

iron, and aqueous sulfide were determined colorimetrically using a Chemetrics [Calverton, VA] field

sampling kit. Dissolved hydrogen gas in the groundwater was determined as previously described

[Lovley, et al., 1994]. Samples for contaminants, methane, sulfate, and short chain organic acids were

collected, preserved, and analyzed using headspace gas chromatography, ion chromatography, or

HPLC as previously described [Lendvay, et al., 1998b].

Biomass was separated from aquifer solids to evaluate transformation potential by indigenous

methane-oxidizing microorganisms as previously described [Lendvay and Adriaens, 1999a]. The

resulting liquid suspension was plated and grown in an atmosphere of 50% methane and 50% air at

ambient temperature (22( ± 1(C). Methane-oxidizing colonies were washed off the plates and grown

in liquid culture. Finally, cells were harvested by centrifugation and washing then resuspended in

liquid culture to obtain a cell density of 2.89 mg of cells/m5 of suspension. Aqueous batch

transformation experiments were conducted at ambient temperature to evaluate transformation kinetics

of cis-DCE, chloroethene and ethene by resting cells.

To evaluate the impact of oxidation on the flux of chloroethene into Lake Michigan, laboratory and

field results were transformed to an aquifer oxidative flux (Equation 1) as previously described

[Lendvay and Adriaens, 1999a].

0 (1) Field C

day

[VC]

dt

dC = λ ∆ ≈

The temporal change in field concentration of chloroethene (dC/dt) was assumed to equal the

observed rate of change for chloroethene concentration in the field, and the initial concentration of

chloroethene (C0) was assumed to be the average chloroethene concentration upgradient of the GSI

[Weaver, et al., 1995]. The first order decay constant () was then determined. For this calculation,

only the sample points that provided evidence of chloroethene oxidation were considered to contribute

to the field oxidative flux.
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

122

Oxygen

166

167

168

169

170

171

172

173

0.00 0.02 0.04 0.06 0.08 0.10

[Oxygen] (m M)

Elev ation (MAMSL)

July S eptember December

S hallow Zone

Deep Zone

B Methane

166

167

168

169

170

171

172

173

0.0 0.2 0.4 0.6

[Methane] (m M)

Elev ation (MAMSL)

July S eptember December

S hallow Zone

Deep Zone

Chloroethene C

166

167

168

169

170

171

172

173

0.0 2.0 4.0 6.0

[VC] (μ M )

Elev ation (MAMSL)

July S eptember December

S hallow Zone

Deep Zone

A

Figure 2: Vertical concentration profiles for chloroethene (A), oxygen (B), and

methane (C) at ML-3 for three different time periods. 

To evaluate the impact of oxidation on the flux of chloroethene into Lake Michigan, laboratory and

field first order decay constants were transformed to an aquifer oxidative flux at the GSI using

Equation 2 [Lendvay and Adriaens, 1999a]. The porosity (n) was assumed to be 0.32, a nominal value

for sandy aquifers. The horizontal distance of the zone where chloroethene is oxidized (x) is the

horizontal coverage of ML-2 and ML-3. The vertical coverage of the oxidative zone (zshallow) is ratioed

to the total depth (ztotal) of 6.3 meters to only consider the oxidative zone. Finally, the storm activity of

the lake is assumed to be sufficient to oxidize chloroethene for only 180 days/year. 

0 (2) Field C

day

[VC]

dt

dC = λ ∆ ≈

RESULTS AND DISCUSSION

Vertical profiles of contaminants, geochemically relevant species, specific conductance, reduction

potential, and dissolved hydrogen gas concentrations were previously reported [Lendvay, et al., 1998b]

with vertical profiles for chloroethene, oxygen and methane presented here (Figure 2). In addition to

these profiles, a lack of quantifiable contaminant concentration, high dissolved oxygen, and specific

conductance measurements corresponding to typical lake water values provided direct evidence of lake

water infiltration at the shallowest sample point as wave activity increased suggesting that lake water

penetrated the top 4 meters of the aquifer. Products of electron acceptor reduction, ferrous iron and

sulfide, provided evidence for iron-reducing conditions in the shallow zone and iron- and sulfatereducing conditions in the deep zone of the plume. The methane concentration profile decreased with

time in the shallow zone as was noticeable for the December sampling (Figure 2C). 

Reduction potential measurements increased in value with time at all elevations suggesting reoxidation occurred across the entire depth profile between August and December [Lendvay, et al.,

1998b]. With the caveat that reduction potential measurements are biased towards the iron couple

[Barcelona, et al., 1989; Barcelona and Holm, 1991], observed values were indicative of denitrifying to

iron-reducing conditions. Dissolved hydrogen values were indicative for iron-reducing to sulfatereducing conditions in the shallow zone, and iron-reducing to methanogenic in the deep zone [Lendvay

et al., 1998b]. Furthermore, a temporal decrease in dissolved hydrogen concentrations (less reducing

TEAPs) corroborated temporal trends in redox measurements. 

To be able to discern temporal effects in the contaminant plume, selected contaminant and

methane concentrations were compared at the same location in the plume over the time period of

interest, using quantile-quantile plots (Figure 3) [Lendvay, et al., 1998b]. These plots show that the

concentration of methane was

lower for the November and

December data compared to

the August and September

data (Figure 3A). Additionally,

these plots showed that

chloroethene concentrations

decrease with time only in the

shallow zone, and slightly

increased in the deep zones of

the plume (Figure 3C). In

contrast, the concentration of

cis-DCE increased in both the

shallow and deep zones of the

contaminant plume with time
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

123

cis-DCE Oxidation

0

2

4

6

8

10

12

0 20 40 60 80

Time (hours)

[cis-D C E] ( M)

Active cis-DCE Ethyne Control

Azide Control

A Ethene Oxidation

y = -2.0E-02x + 9.9E+00

R2

 = 9.6E-01

0

2

4

6

8

10

12

0 20 40 60 80

Time (hours)

[Ethene] ( M)

Active Ethene Ethyne Control

Azide Control

Chloroethene Oxidation C

y = -6.2E-02x + 1.1E+01

R2

 = 9.9E-01

0

2

4

6

8

10

12

0 20 40 60 80

Time (hours)

[Chloroethene] ( M) Active Chloroethene Ethyne Control

Azide Control

B

Figure 4: Resting Cell Oxidation of cis-DCE (A), chloroethene (B), and ethene (C) by methane oxidizing

microorganisms.

Te mporal Variation in M e thane

0.0

0.2

0.4

0.6

0.8

1.0

0.0 0.2 0.4 0.6 0.8 1.0

[M ethane] Aug./S ep. (m M )

[M ethane] N ov./D ec. (m M ) 

A

Te mporal Variation in Chloroe the ne

0.0

1.0

2.0

3.0

4.0

5.0

0.0 1.0 2.0 3.0 4.0 5.0

[V C ] Aug./S ep. (μ M )

[V C ] N ov./D ec. ( M )

Shallow Deep

Te mporal Variation in cis -DCE C

0.0

1.0

2.0

3.0

4.0

5.0

0.0 1.0 2.0 3.0 4.0 5.0

[cis-D C E ] Aug./S ep. (μ M )

[cis-D C E ] N ov./D ec. ( M )

Shallow Deep

B

Figure 3: Quantile-quantile plots for the data indicated. For

each case, all data (ML-2 and ML-3) are considered for

November and December samples on the ordinate, and

August and September samples on the abscissa.

(Figure 3B). Combining the results for chloroethene and cis-DCE suggests the decrease in

chloroethene concentration in the shallow zone was not a result of dilution by infiltration of lake water.

To elucidate specific field microbial processes responsible for chloroethene transformation,

scattergrams comparing chloroethene or methane with oxygen in the shallow zone were evaluated for

statistical correlations [Lendvay, et al., 1998a]. In the case of both chloroethene and methane, peak

concentrations occurred at low oxygen

concentrations and low chloroethene and

methane concentrations occurred at high

oxygen concentrations. Combined, these trends

suggest that chloroethene may be co-oxidized

by methane-oxidizing microorganisms in the

shallow zone of the GSI as lake activity

increases.

To test this field-derived hypothesis and

corroborate the field measurements, laboratory

studies were conducted using groundwater and

aquifer solids collected from the GSI. The

mixed cultures were grown on methane as a

sole source of carbon and energy under aerobic

conditions. A resting cell oxidation experiment

was conducted for cis-DCE, chloroethene, and

ethene in triplicate (Figure 4). Separate controls

of either 220 mg/5 sodium azide or 0.3%

(vol./vol. gas phase) ethyne were effective at

suppressing all transformation of contaminants. Considering Figure 4, transformation of both

chloroethene and ethene is evident relative to controls, with the rate of chloroethene oxidation being

2.9 nmoles/(day-mg protein) and ethene oxidation being 0.9 nmoles/(day-mg protein). No

transformation of cis-DCE was observed relative to controls. These results support the field findings,

that chloroethene and possibly ethene are co-oxidized by methane-oxidizing microorganisms while cisDCE is not.

To evaluate the impact of chloroethene oxidation on the flux of chloroethene into Lake Michigan,
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

124

Table 1: Reported value for chloroethene flux at transect-5 (upgradient transect) using a MODFLOW model [Weaver, et al., 1995], and values for chloroethene flux at the GSI using

calculated field and laboratory rates. Percent values represent the percent of chloroethene flux

at transect-5 that is oxidized by the reported value.

Defined Value Flux (g/(year-m2))

Chloroethene Flux at

Transect-5

0.86

Oxidative Flux of

Chloroethene by Field

Measurement

0.063 (~7%)

Oxidative Flux of

Chloroethene by Laboratory

Measurement

0.0007 (~0.1%

of 0.86 or ~1%

of Field Flux)

laboratory and field results were transformed to a field oxidative flux [Lendvay and Adriaens, 1999a].

Using the calculated field and laboratory chloroethene oxidation rates, field flux calculations were

possible (Equation 2). The calculated oxidation fluxes were compared to the chloroethene flux

upgradient as calculated by a MOD-FLOW model (Table 1) [Wilson, et al., 1994].

Field calculations of the oxidative flux for chloroethene suggest that only about 7% of the annual

chloroethene flux into Lake Michigan is mitigated by measurable biogeochemical processes at the GSI

as a result of increased storm activity. Of this observed field flux, only 1% could be associated with a

particular microbial culture namely methane-oxidizing microorganisms. The remaining catalytic

activity affecting the contaminant plume may be due to other physiological types in the aquifer solids

such as non-culturable methane-oxidizing, heterotrophic [Wackett, et al., 1989; Davis and Carpenter,

1990], autotrophic [Vannelli et al, 1990], ethene oxidizing [Freedman and Herz, 1996], iron-reducing

[Bradley and Chapelle, 1996], or fermentative microorganisms [Bradley, et al., 1998]. This study has

contributed to our understanding of the dynamics of GSIs with respect to microbial activity and

geochemistry, and points towards a possible role of methane-oxidizers in mitigating chloroethene

imparted toxicity.

REFERENCES

Barcelona, M.J. and T.R. Holm, 1991. Oxidation-Reduction Capacities of Aquifer Solids,

Environmental Science and Technology, v. 25, pp.1565-1572.

Barcelona, M.J., Holm, T.R., Schock, M.R., and G.K. George, 1989. Spatial and Temporal Gradients

in Aquifer Oxidation-Reduction Conditions, Water Resources Research, v. 25, pp. 991-1003.

Bradley, P.M. and F.H. Chapelle, 1996. Anaerobic Mineralization of Vinyl Chloride in Fe (III)-

Reducing, Aquifer Sediments, Environmental Science and Technology, v. 30, pp. 2084-2086.

Bradley, P.M. and F.H. Chapelle, 1998. Effect of Contaminant Concentration on Aerobic Microbial

Mineralization of DCE and VC in Stream-Bed Sediments, Environmental Science and Technology,

v. 32, pp.553-557.

Davis, J.W. and C.L. Carpenter, 1990. Aerobic Biodegradation of Vinyl Chloride in Groundwater

Samples, Applied and Environmental Microbiology, v. 56, pp. 3878-3880.
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

125

Freedman, D.L. and S.D. Herz, 1996. Use of Ethylene and Ethane as Primary Substrates for Aerobic

Cometabolism of Vinyl Chloride, Water Environment Research, v. 68, pp. 320-328.

Lendvay, J.M., Sauck, W.A., McCormick, M.L., Barcelona, M.J., Kampbell, D.H., Wilson, J.T., and

P. Adriaens, 1998a. Geophysical Characterization, Redox Zonation, and Contaminant Distribution

at a Groundwater/Surface Water Interface, Water Resources Research, v. 34, pp. 3545-3559.

Lendvay, J.M., Dean, S.M., and P. Adriaens, 1998b. Temporal and Spatial Trends in Biogeochemical

Conditions at a Groundwater-Surface Water Interface: Implications for Natural Bioattenuation,

Environmental Science and Technology, v. 32, pp. 3472-3478.

Lendvay, J.M. and P. Adriaens, 1999a. Laboratory evaluation of temporal trends in biogeochemical

conditions at a groundwater–surface water interface, Physics and Chemistry of the Earth.

Lendvay, J.M., Dean, S.M., Barcelona, M., Adriaens, P., and N.D. Katopodes, 1996b. Installing MultiLevel Sampling Arrays to Monitor Groundwater and Contaminant Discharge to a Surface Water

Body, Groundwater Monitoring and Remediation.

Lovley, D.R., Chapelle, F.H., and J.C. Woodward, 1994. Use of Dissolved H2 Concentration to

Determine Distribution of Microbially Catalyzed Redox Reactions in Anoxic Groundwater,

Environmental Science and Technology, v. 28, pp. 1205-1210. 

McCormick, M.L. and P. Adriaens, 1998. Tetrachloroethylene Transformation in an Iron Reducing

Enrichment Culture, Abstract, 98th General Meeting of the American Society for Microbiology,

Atlanta, GA, p.453, American Society for Microbiology, Washington, DC, 1998.

Semprini, L., P.K. Kitanidis, D.H. Kampbell, and J.T. Wilson, 1995. Anaerobic Transformation of

Chlorinated Aliphatic Hydrocarbons in a Sand Aquifer Based on Spatial Chemical Distributions,

Water Resources Research, v. 31, pp.1051-1062.

Tiedeman, C. and S.M. Gorelick, 1993. Analysis of Uncertainty in Optimal Groundwater Contaminant

Capture Design, Water Resources Research, v. 29, pp. 2139-2153.

Vannelli, T., Logan, M., Arciero, D.M., and A.B. Hooper, 1990. Degradation of Halogenated Aliphatic

Compounds by the Ammonia-Oxidizing Bacterium Nirosomonas europaea, Applied and

Environmental Microbiology, v. 56, pp.1169-1171.

Wackett, L.P., Brusseau, G.A., Householder, S.R., and R.S. Hanson, 1989. Survey of Microbial

Oxygenases: Trichloroethylene Degradation by Propane-Oxidizing Bacteria, Applied and

Environmental Microbiology, v. 55, pp. 2960-2964.

Weaver, J.W., J.T. Wilson, D.H. Kampbell, and M.E. Randolph, 1995. Field Derived Transformation

Rates for Modeling Natural Bioattenuation or Trichloroethylene and its Degradation Products, in

Proceedings of the Next Generation Environmental Models and Computational Methods, Bay City,

MI.

Wilson, J.T., Weaver, J.W., and D.H. Kampbell, 1994. Intrinsic Bioremediation of TCE in Ground

Water at an NPL Site in St. Joseph, Michigan, EPA/540/R-94/515, pp.154-160.
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

126

Natural Attenuation of Chlorinated Solvents in a

Freshwater Tidal Wetland, Aberdeen Proving

Ground, Maryland

By Michelle M. Lorah and Lisa D. Olsen

Ground-water contaminant plumes that are flowing toward or currently discharging into wetland

areas present unique remediation problems because of the hydrologic connections between ground

water and surface water and the sensitive habitats in wetlands. Because wetland sediments typically

have a large diversity of microorganisms and redox conditions that could enhance biodegradation, they

are ideal environments for natural attenuation of organic contaminants. “Natural attenuation” is a

general term that includes all naturally occurring physical, chemical, and biological processes that can

reduce contaminant concentrations or toxicity without human intervention. Biodegradation generally is

considered the most important of these processes for ground-water remediation purposes because it is a

destructive process, unlike sorption, dilution, or volatilization. Natural attenuation is a treatment

method that would leave the ecosystem largely undisturbed and be cost-effective. In this research, the

natural attenuation of chlorinated volatile organic compounds (VOCs) was studied in a contaminant

plume that discharges from a sand aquifer to a freshwater tidal wetland at Aberdeen Proving Ground,

Maryland (Lorah, et al., 1997). 

Biodegradation processes of two of the major contaminants, trichloroethylene (TCE) and 1,1,2,2-

tetrachloroethane (PCA), are the focus of this combined field and laboratory study. The fate of PCA in

the wetland is particularly emphasized in this research because the occurrence and dominant pathways

of PCA degradation in ground water or soil were largely unknown. The few previous studies on PCA

degradation were laboratory experiments that were constructed with anaerobic mineral medium or

glass beads and seeded with mixed cultures from municipal sludge waste, or with abiotic aqueous

mixtures of transition-metal coenzymes (Lorah, et al., 1997). The general field approach used in this

study included (1) installing nested drive-point piezometers to characterize the ground-water chemistry

and contaminant distribution along two transects through the wetland (one shown in Figure 1), and (2)

using porous membrane sampling devices (peepers) to obtain centimeter-scale resolution of

contaminant distribution in the wetland porewater. The general laboratory approach included

conducting batch microcosm experiments with wetland sediment and porewater under methanogenic,

sulfate-reducing, and aerobic conditions to confirm field evidence of biodegradation pathways,

investigate potential controlling factors on biodegradation, and estimate biodegradation rates.

Field evidence collected along the two ground-water flowpaths shows that anaerobic

biodegradation of TCE and PCA is enhanced in the wetland compared to the aquifer sediments (Lorah,

et al., 1997; Lorah and Olsen, in press). The enhanced biodegradation is associated with the natural

increase in dissolved organic carbon concentrations and decrease in redox state of the ground water

along the upward flow direction in the wetland sediments. The aquifer typically is aerobic. Ironreducing conditions are predominant in the lower wetland sediment unit composed of clayey sand and

silt, and methanogenesis was predominant in an upper unit composed of peat (Figure 2). A decrease in

concentrations of TCE and PCA and a concomitant increase in concentrations of anaerobic daughter

products occurs along upward flowpaths through the wetland sediments (Figures 1 and 2). The

daughter products 1,2-dichloroethylene (1,2-DCE), vinyl chloride (VC), 1,1,2-trichloroethane (1,1,2-

TCA), and 1,2-dichloroethane (1,2-DCA) are produced from hydrogenolysis of TCE and from PCA
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

127

Figure 2. Vertical distribution of (A) the parent contaminants TCE and PCA and possible anaerobic daughter

products and (B) selected redox-sensitive constituents at site WB-26, June-October 1995. 

Figure 1. Concentrations of TCE in ground water along section A-A, June-October 1995.

degradation through hydrogenolysis and dichloroelimination pathways. Total concentrations of TCE,

PCA, and their degradation products, however, decrease to below detection levels (generally less than

0.5 μg/5) within 0.15 to 0.30 m of land surface. Natural attenuation in the wetland sediments seems to

be effective even where relatively high concentrations of VOCs are discharging upward through very

thin (less than 2 m) layers of wetland sediment and when microbial activity probably decreases during

cooler seasons (Lorah, et al., 1997; Lorah and Olsen, in press). Field evidence indicates that highly

reducing conditions are not necessary for dichloroelimination of PCA to 1,2-DCE, or for

hydrogenolysis of PCA to 1,1,2-TCA and then to 1,2-DCA. Maximum concentrations of VC, however,
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

128

Figure 3. Degradation under methanogenic conditions in microcosms amended with 3.0

)mol/L of TCE or 2.9 )mol/L of PCA at day 0: (A) TCE and the daughter products cis-12DCE,

trans-12DCE, and VC in the TCE-amended microcosms; (B) PCA and the sum of daughter

products in the live PCA-amended microcosms and formalin-amended controls; (C) daughter

products 12DCE (total of the cis and trans isomers), VC, 112TCA, and 12DCA in the PCAamended microcosms; and (D) cis-12DCE and trans-12DCE in the PCA-amended

microcosms. [The concentrations shown are the average measured in duplicate microcosms

sacrificed at each time step.]

coincided with the highest concentrations of methane in the wetland porewater, indicating that

continued anaerobic degradation of 1,2-DCE to VC and of VC to the non-toxic end-product of

ethylene may require the highly reducing conditions of methanogenesis.

Laboratory microcosms that were prepared using wetland sediment and ground water from the site

confirmed field evidence of biodegradation pathways and allowed estimation of biodegradation rates

(Lorah, et al., 1997). TCE biodegradation occurred through hydrogenolysis to 1,2-DCE (predominately

the cis isomer) and VC under methanogenic (Figure 3) and sulfate-reducing conditions. For PCA

degradation under methanogenic conditions, the cis and trans isomers of 1,2-DCE and VC were the

predominant persistent daughter products in one set of microcosm experiments (Lorah and Olsen,

1999) (Figure 3). In two sets of later experiments, however, 1,2-DCA was the predominant persistent

daughter product from PCA and 1,1,2-TCA degradation under methanogenic conditions. The

differences between the experiments seemed to result from differing pathways of 1,1,2-TCA

degradation, rather than from differences in the initial PCA degradation pathway. 1,1,2-TCA was

produced simultaneously with 1,2-DCE early in the time course in all experiments. The 1,1,2-TCA

produced from hydrogenolysis of PCA was degraded by dichloroelimination to VC in the first

experiment, whereas it was degraded by continued hydrogenolysis to 1,2-DCA in the second and third

experiments. Natural temporal or spatial variations in the microbial populations in the wetland

sediments may have caused the differing degradation pathways in these experiments. 

In all PCA-amended microcosms, 1,1,2-TCA occurred simultaneously with 1,2-DCE early in the

time course, indicating that hydrogenolysis and dichloroelimination of PCA can occur simultaneously

(Lorah and Olsen, 1999). Significantly lower ratios of cis-1,2-DCE to trans-1,2-DCE were produced

by dihaloelimination of PCA than by hydrogenolysis of TCE (Figure 3). Only one other study, which

was conducted in the laboratory using anaerobic municipal sludge, has reported evidence for both of
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

129

these PCA degradation pathways (Chen, et al., 1996). Recognition of the dichloroelimination pathway

for PCA is important because TCE is a common co-contaminant with PCA at this site and other

hazardous-waste sites. The 1,2-DCE isomer distribution potentially could assist in determining whether

PCA degradation is occurring at sites where TCE is a co-contaminant. In the PCA-amended

microcosms with the wetland sediment, TCE production commonly was less than 5 % of the initial

PCA concentration in the live and sterile microcosms, showing that abiotic dehydrochlorination is not

a significant degradation pathway for PCA in this environment. 

Under methanogenic conditions, first-order biodegradation rates of TCE ranged from 0.30 to 0.37

day-1 (half-life of about 2 days), showing extremely rapid biodegradation in these organic-rich wetland

sediments. Although the TCE biodegradation rate was an order of magnitude slower under sulfatereducing conditions (0.032 day-1) than methanogenic conditions, the rate was still two orders of

magnitude higher than those reported in the literature for anaerobic TCE biodegradation in microcosms

constructed with sandy aquifer sediments (Rifai, et al., 1995). The slow production and degradation of

1,2-DCE and VC in TCE-amended microcosms that were incubated under sulfate-reducing conditions

or with the addition of an inhibitor of methanogenic activity confirmed field evidence that

methanogenic activity is important in continued anaerobic degradation of these daughter products. 

First-order rate constants for anaerobic degradation of PCA and 1,1,2-TCA ranged from 0.15 to

0.58 day-1 (half-lives of 1.2 to 4.6 days), again showing that biodegradation of highly chlorinated

VOCs is extremely rapid in the wetland sediments. Similar PCA degradation rates were observed

under methanogenic and sulfate-reducing conditions, although methane production and sulfate

reduction occurred simultaneously during the first 15 days of incubation in the microcosms that were

amended with sulfate to stimulate sulfate-reducing conditions. The addition of an inhibitor of

methanogenic activity to PCA-amended microcosms decreased the methane production rates by a

factor of 10 and caused a nearly 50% decrease in the PCA degradation rate compared to microcosms

without the inhibitor. In addition, the production and subsequent degradation of the daughter products

1,2-DCA, 1,2-DCE, and VC were slower when methanogenic activity was inhibited in the PCAamended microcosms. Both the TCE-amended and PCA-amended microcosm experiments, therefore,

indicate that complete anaerobic degradation of the chlorinated VOCs is most rapid when

methanogenic activity is high. 

Although the wetland sediments have predominantly anaerobic conditions, aerobic conditions may

be present in surficial sediments near the air-water interface and in subsurface sediments near plant

roots, providing a suitable environment for methanotrophs that can degrade chlorinated VOCs through

cometabolic oxidation (Lorah, et al., 1997). Although coupling of anaerobic and aerobic degradation

processes has been suggested as the best possible bioremediation method for chlorinated VOCs such as

TCE, few studies have investigated the degradation of TCE under both anaerobic and aerobic

conditions for a natural subsurface setting. In aerobic microcosm experiments with the wetland

sediment, biodegradation of cis-1,2-DCE, trans-1,2-DCE, and VC only occurred if methane

consumption occurred, indicating that methanotrophs were involved. Aerobic biodegradation rates for

cis-1,2-DCE, trans-1,2-DCE, and VC were in the same range as those measured for TCE and PCA

under anaerobic conditions. Production of these anaerobic daughter products of TCE and PCA,

therefore, could be balanced by their consumption where methanotrophs are active in the wetland

sediment, including near land surface and in the rhizoplane and root tissues of aquatic plants in

wetlands (Lorah, et al., 1997). 

In summary, biodegradation through both anaerobic and aerobic processes is a significant natural

attenuation mechanism for chlorinated hydrocarbons in these wetland sediments, causing a reduction
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

130

in contaminant concentrations and toxicity before surface-water receptors are reached. This combined

field and laboratory study provides a general approach and scientific basis for investigating the

feasibility of natural attenuation as a remediation alternative for other sites where ground-water plumes

discharge to wetlands and similar organic-rich environments at ground-water/surface-water interfaces.

Because conditions in the wetland sediments are naturally conducive to biodegradation of the

chlorinated VOCs in the discharging ground water, biodegradation could potentially be sustained

indefinitely, unlike many ground-water environments where biodegradation commonly is limited by

the supply of organic substrates. The results of this study also increase our understanding of the fate of

chlorinated VOCs in the environment, providing information beneficial in characterizing contaminant

behavior and in implementing bioremediation systems for other ground-water contaminant plumes.

REFERENCES

Chen, Chun, Puhakka, J. A., and J.F. Ferguson, 1996. Transformations of 1,1,2,2-tetrachloroethane

under methanogenic conditions, Environmental Science and Technology, v. 30, no. 2, pp. 542-547.

Lorah, M.M., and L.D. Olsen, 1999. Degradation of 1,1,2,2-tetrachloroethane in a freshwater tidal

wetland: Field and laboratory evidence, Environmental Science and Technology, v. 33, p. 227-234.

Lorah, M.M., and L.D. Olsen, in press. Natural attenuation of chlorinated volatile organic compounds

in a freshwater tidal wetland: Field evidence of anaerobic biodegradation: Water Resources

Research.

Lorah, M.M., Olsen, L.D., Smith, B.L., Johnson, M.A., and W.B. Fleck, 1997. Natural attenuation of

chlorinated volatile organic compounds in a freshwater tidal wetland, Aberdeen Proving Ground,

Maryland: U.S. Geological Survey Water-Resources Investigations Report 97-4171, 95 pp.

Rifai, H.S., Borden, R.C., Wilson, J.T., and C.H. Ward, 1995. Intrinsic bioattenuation for subsurface

restoration, in Intrinsic Bioremediation, R. E. Hinchee and others (eds.), Columbus, Ohio, Battelle

Press, pp. 1-29.

AUTHOR INFORMATION

Michelle M. Lorah and Lisa D. Olsen, U.S. Geological Survey, 8987 Yellow Brick Road, Baltimore,

MD 21237.
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Discharge of Contaminated Ground Water to

Surface Water: An Ecological Risk Assessment

Perspective

By Mary Baker Matta and Tom Dillon

INTRODUCTION/OVERVIEW

The ecological risk associated with the discharge of contaminated ground water to surface water

may be difficult to evaluate at complex hazardous waste sites. However, at many sites, groundwater

discharges provide a significant pathway for contamination to reach ecological receptors. Although

EPA has guidance for generic risk assessment, and has specific guidance for ecological risk

assessments at CERCLA sites, specific guidance on how to evaluate ecological risk of groundwater

discharges is lacking.

The hyporheic zone connects the ecological communities in ground water, sediment, and surface

water. Although benthic macroinvertebrate communities are often considered in ecological risk

assessments (and they may be adversely affected by contamination in the hyporheic zone), microbial

communities of the hyporheic zone are rarely considered receptors of concern in ecological risk

assessments. Because groundwater discharges contribute to sediment and surface water contamination

they also have the potential to adversely affect fish and other aquatic species, particularly where

bioaccumulative compounds are released or where sensitive life stages are in close proximity to

groundwater discharge points. 

The nature and extent of the contamination present in groundwater and characteristics of the

ecosystem will determine what and where to sample, and what tests should be conducted to evaluate

ecological risk and develop protective cleanup levels if risk is significant. Natural attenuation is

increasingly considered as a remedial option at many sites. Natural attenuation is influenced by

microbial activity and physico-chemical characteristics of the groundwater and its movement.

Therefore, if natural attenuation is to be considered as a viable remedial option, the microbial

community in the hyporheic zone should be protected, and its role in chemical fate and transformation

should be assessed. At sites where ecological risk is significant, more active cleanup measures may be

necessary to protect natural resources.

NOAA’s experience in aquatic ecological risk assessment is applied in this poster to provide some

recommendations for evaluating ecological risk of groundwater discharge to surface waters. This

poster focuses on data needs and how to answer them, structured around the EPA risk assessment

process.

RISK ASSESSMENT STEPS 1 AND 2: SCREENING LEVEL ECOLOGICAL RISK ASSESSMENT

NOAA prepares Coastal Hazardous Waste Site Reviews using available information to screen sites

based on:

 • proximity of the site to the coast or inland water bodies that support anadromous fish populations
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132

 • site history, contaminants that are likely to be or have been detected at the site

 • potential contaminant migration pathways (including groundwater)

 • presence of sensitive habitats or species near the site

This same information is required in screening level risk assessments.

Maximum contaminant concentrations in groundwater are screened by NOAA using a value ten

times greater than chronic ambient water quality criteria to account for potential dilution when

groundwater enters surface waters and to protect all aquatic species. Therefore, detection limits for

chemical analysis of unfiltered groundwater should be less than 10 times chronic ambient water quality

criteria to conduct this assessment. The most recent edition of Hazardous Waste Site Reviews (NOAA,

1997) indicates that groundwater discharges are a widespread problem for natural resources:

 • At 10 of the 11 sites evaluated, groundwater is a potential pathway for contaminants to reach

natural resources.

 • At 8 of the 11 sites evaluated, concentrations in groundwater exceed screening levels.

Sites Reviewed in December, 1997

Region 1 Beede Waste Oil, Plaistow, NH

Region 2 V&M/Albaladejo Farms, Vega Baja, Puerto Rico

Region 3 Fort George G. Meade, Anne Arundel Co, MD

Norfolk Naval Base, Norfolk, VA

Salford Quarry, Lower Salford Townshp, PA

Region 4 Brunswick Wood Preserving, Brunswick, GA

MRI Corp, Tampa, FL

*Terry Creek Dredge Spoil/Hercules, Brunswick, GA

Tyndall Air Force Base, Bay Co, FL

Region 6 Madisonville Creosote Works, Madisonville, LA

Region 10 Oeser Company, Bellingham, WA

*Groundwater not a significant pathway to natural resources

RISK ASSESSMENT STEP 3: BASELINE RISK ASSESSMENT PROBLEM FORMULATION

Steps:

 • Refine list of contaminants based on screening

 • Summarize toxicological effects

 • Consider likely fate and transport of contaminants

 • Determine receptors likely to be at risk

 • Determine complete exposure pathways

 • Develop conceptual site model with risk questions

 • Develop assessment endpoints 

Considerations specific to sites with groundwater contamination:

 • Potential biogeochemical alterations of contaminants.

-persistent organic contaminants might be released as an LNAPL or DNAPL and 

bind to sediments as they are released to surface water. 
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- Metals might be released in dissolved form and complex with other compounds at 

the discharge point, groundwater discharge to marine water may alter solubility of metals. 

- Groundwater discharge may leach metals from sediments.

- Organic contaminants released in a solvent plume may be more bioavailable

• Conceptual model for the site should include pathways for exposure and mechanisms of toxicity

for the contaminants and receptors specific to the site

• Consider characteristics of discharge specific to the regime: 

-groundwater discharge into intertidal zone at marine sites; 

- discharge into tidal creeks in salt marshes;

- hyporheic zone in alluvial rivers as a function of channel morphology, bed 

roughness, and permeability (Triska, et al. 1989). 

• Consider hyporheos as receptors of concern (macrofauna and microbes)

Assessment Endpoints of General Concern to NOAA (with some specific examples)

• Protection of benthic community structure and function

(Protection of stonefly populations from direct toxic effects)

• Protection of aquatic community structure and function

(Protection of fish eggs and larvae from direct toxic effects, including ecologically relevant

sub-lethal effects)

(Protection of hyporheic microbial community from direct toxic effects)

• Protection of fish populations and communities

(Protection of fish from reproductive effects)

(Protection of fish from immune system disruption)

(Protection of fish from reductions in survival and growth)

• Protection of specific habitat functions (for example, nutrient cycling)

(Protection of hyporheic microbial community from direct toxic effects)

• Protection of fishery resources from contamination

(Protection of human health from exposure to bioaccumulated contaminants)

RISK ASSESSMENT STEPS 4, 5, AND 6: STUDY DESIGN/DATA QUALITY OBJECTIVES/FIELD

VERIFICATIONS/SITE INVESTIGATIONS

Steps

• Evaluate uncertainty in existing data

• Develop measurement endpoints to evaluate assessment endpoints

• Develop work plan to evaluate exposure and effects

-methods

-statistical considerations

-sampling locations, timing, frequency
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• Verify sampling design

• Site investigation

• Data analysis

Considerations specific to sites with groundwater contamination:

•Evaluating exposure potential is a major concern; need to determine nature and extent 

of contamination and factors that may control bioavailability and toxicity:

-Discharge locations

-Flux of contaminants

-Form/speciation of contaminants

-Presence of non aqueous phase layers (NAPL)

-Potential contaminant interactions, for example enhanced transport of hydrophobic

(biomagnifying) contaminants in a dissolved solvent plume

• NOAA recommends a weight of evidence approach based on:

-Groundwater models predicting a vector discharge to surface water

-Remote sensing and geological analysis: identification of paleochannels using groundpenetrating radar; location of surface depressions, abandoned meander channels to locate

springbrooks through aerial photography (Stanford and Ward, 1993).

-Developing a water budget - upstream/downstream gauging, tracer studies

-Changes in water quality parameters due to groundwater discharge (temperature, pH,

conductance, nutrients, DOC, oxygen (Triska, et al., 1989).

-Direct measurements

•Specific effects measurements will be a function of receptors of concern; contaminants; and

exposure pathway

-Benthic macroinvertebrates-generally a concern for direct toxicity via surface water or

sediment contamination as a result of groundwater discharges

-Microbes-generally a concern for direct toxicity of groundwater, pore water, or surface water

-Fish- may include a concern for direct toxicity of pore water or surface water to early life

stages (for example, where dissolved metals are discharged) or a concern for indirect

toxicity after contamination of sediment and food organisms.

• Groundwater and NAPL can contaminate sediments and surface water

• Ecological risk associated with sediment contamination is usually evaluated via 

-chemical analysis of sediment (focus on persistent contaminants)

-toxicity testing 

-benthic community evaluations 

-chemical analysis of biota for bioaccumulative contaminants

-comparisons of tissue concentrations to literature effects thresholds

-food web modeling for persistent bioaccumulative contaminants 

• Ecological risk associated with water contamination is usually evaluated via

-chemical analysis of water (focus on persistent contaminants or continuous releases of less

persistent contaminants)
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135

-aquatic toxicity tests 

-comparisons of water concentrations to literature or AWQC benchmarks.

• Sampling and analysis concerns:

-number and locations of samples should be adequate to evaluate exposure potential

and minimize uncertainty (plumes should be characterized, and locations where sensitive

receptors would be exposed should be sampled)

-seasonal considerations affecting groundwater flow or the seasonal presence of 

sensitive receptors may require multiple sampling events

-analyze filtered and unfiltered water (organisms are exposed to both)

-analyze pore water where appropriate for the receptor (fish eggs, for example)

-collect ancillary data that can be used to interpret bioavailability, toxicity, and 

potential for natural attenuation (sediment grain size, sediment total organic carbon, acid

volatile sulfides, dissolved organic carbon, nutrients, alkalinity, dissolved oxygen).

-use detection limits low enough to compare to benchmarks

-consider metal speciation and effects on bioavailability and toxicity

-selection of reference sites is critical in evaluating exposure and effects (reference 

sites should be similar to the study site, but located away from point sources of

contamination)

RISK ASSESSMENT STEP 7: RISK CHARACTERIZATION

Steps:

• Interpreting data

• Tracking sources

• Modeling mass flux

• Modeling food web effects

• Reducing and incorporating uncertainty

• Applying protective assumptions 

• Interpreting the weight of evidence

• Developing cleanup levels

Considerations specific to sites with groundwater contamination:

• Interpreting data

-agree on interpretive methods before collecting data

-compare results to benchmarks, control samples, and reference site results

-consider normalizing data to nutrients, grain size, or other factors

• Source tracking (relative contribution of groundwater and surface releases) is important to ensure

success of the remedy (in terms of reducing risk to ecological receptors). Consider the potential

for recontamination through ongoing groundwater discharge.

• Models on groundwater discharge rates and contaminant loading are usually

highly uncertain, therefore, to be protective, conservative assumptions must be made until

better techniques, data, and models are available.

•Conduct specific studies to determine the potential for natural attenuation, which is driven largely

by microbes and their physico-chemical environment (consider the specific types of microbes

required to degrade contaminants at the site; some require aerobic).
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

136

•Water quality and sediment goals for receptors can be empirically derived from toxicity tests but

back calculating for groundwater quality goals requires a good model

Examples:

At the Metal Bank of America site in Philadelphia, PA, the primary pathway for contaminants to

reach natural resources was through discharge of contaminated groundwater and non-aqueous phase

layer. PCBs accumulated in sediment, clams, and fish near the site to concentrations that were

determined to pose significant risk. 

RISK ASSESSMENT STEP 8: RISK MANAGEMENT 

Considerations specific to sites with groundwater contamination:

•Potential for recontamination- uncontrolled groundwater discharges have the

potential to interfere with a sediment remedy

•Monitoring-the effects that drove the selection of the remedy should be incorporated into 

monitoring, significant ongoing exposures should be monitored over time and re-evaluated

CONCLUSIONS/RECOMMENDATIONS

• Potential for groundwater to provide a contaminant pathway to aquatic species should be

considered, it seems to be a problem throughout the country in all regions examined.

• NOAA screens GW for potential concern using 10 times chronic AWQC values

• Detection limits for unfiltered groundwater should be low enough to conduct this screening

• Consider potential biogeochemical alterations and complex interactions between contaminants.

• Consider characteristics of discharge specific to the regime: 

• Consider hyporheos as receptors of concern (macrofauna and microbes)

• Use multiple methods and a weight of evidence to evaluate exposure potential (discharge 

locations and contaminant fluxes)

• Evaluate potential for bioaccumulation and food web effects

 • Specific sampling and analysis methods will depend on characteristics of contaminants and

receptors at the site (but consider seasonal variations and sample in such a way as to provide

information relevant for the behavior of the organisms at the site)

 • A critical need is to reduce uncertainty in risk conclusions: better methods are needed to

locate and quantify contaminant flux from groundwater to surface water

 • Until better data is available, protective assumptions should be used to evaluate risk to natural

resources

REFERENCES

NOAA, 1997. Coastal hazardous waste site reviews, December 1997, Seattle, WA, 109 pp.

Stanford, J.A. and J.V. Ward., 1993. An ecosystem perspective of alluvial rivers: connectivity and the

hyporheic corridor, Journal of the North American Benthological Society, v. 12, pp. 48-60.
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

137

Triska, F.J., V.C. Kennedy, R.J. Avanzino, G.W. Zellweger, and K.E. Bencala, 1989. Retention and

transport of nutrients in a third order stream: Hyporheic processes. Ecology, v. 70, pp. 1893-1905.

U.S. EPA, 1997. Ecological Risk Assessment Guidance for Superfund: Process for Designing and

Conducting Ecological Risk Assessments. Interim Final, Edison, NJ, EPA 540-R-97-006.

U.S. EPA, 1992. Framework for Ecological Risk Assessment, Washington, DC, EPA/630/R-92/001.

AUTHOR INFORMATION

Mary Baker Matta and Tom Dillon, NOAA Coastal Protection and Restoration Division, Seattle,

Washington.

(NOAA Hazardous Waste Site Reviews are available from John Kaperick, NOAA Office of Response

and Restoration Bin C15700 Seattle, WA 98115).
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

138

Defining Groundwater Outcrops in West Neck

Bay, Shelter Island, New York Using Direct

Contact Resistivity Measurements and Transient

Underflow Measurements

By Ronald Paulsen

ABSTRACT

Mapping out groundwater discharge zones can be a formidable task due to the highly variable nature of

the discharge zone, temporal and spatial variability of seepage rates within the zone, and complications

from tidal loading. An integrated approach incorporating direct contact resistivity logging and transient

seepage rate measurements was undertaken to delineate the seepage zone and monitor discharge on a

bay-wide scale. Conductivity values for the saturated sediments ranged from 280 S/cm in freshwater

zones to 12,800 S/cm in zones with high salinity. The discharge zone at West Neck Bay, Shelter Island

was observed to extend to 10-75 feet offshore. The groundwater seepage within the discharge zone was

measured using a time transient seepage meter that was developed with ultrasonic technology. Seepage

velocities in the study area ranged from 1.27 x 10-3 cm/s to 3.94 x 10-5 cm/s, equivalent to a mean value

of 16 5/m2

/d. Integrating over the horizontal extent of the seepage zone, the total daily discharge was

estimated to be 1.72 x 106 5/day for the north-east section of West Neck Bay. This estimate of the total

discharge due to underflow is comparable to the recharge in the contributing area, estimated to be 1.50

x 106 5/day for this section of the bay. 

DIRECT CONTACT RESISTIVITY MEASUREMENTS

To characterize near-shore sediments in a simple and rapid manner, we used geophysical logging

to determine the electrical resistivity of surface sediments off-shore. Electrical measurements have

been used for some time to characterize the lithology and hydraulic characteristics of geological

structures. The basic concept of resistivity logging dates back to 1927 when C.M. Schlumberger made

the first well log near Paris (Goldberg, 1997). Conductivity generally increases with increasing

porosity (Gueguen and Palciauskas, 1994). Archie (1942) invoked laboratory measurement of

conductivity to infer amounts of water and hydrocarbons in the pore space. The electrical conductivity

of saturated sediment is commonly analyzed in terms of the formation factor F as a function of the

porosity where s is the electrical conductivity of the saturated bulk sediment, w is that of the interstitial

solution, and the Archie exponent n ~ 1-2. Archie’s law is applicable when the conductivity of the

interstitial solution is much higher than that of the sediment particles, so those surface conduction

phenomena are insignificant.

At off-shore locations where groundwater discharge is negligible, resistivity measurements of the

sea water and sediments saturated by water of identical salinity can be used to determine the formation

factor and infer the porosity from Archie’s law (Aller 1982). In sediments where freshening of the pore

spaces has occurred due to groundwater discharge, the measurements usually show a decrease of

electrical resistivity with depth, which provides important qualitative constraints on the increase of

salinity in the pore fluid and the depth range over which the transition from fresh to sea water occurs. 
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

139

Several different electrode configurations are commonly used to measure electrical resistivity. Our

project employs a direct contact probe (Figure 1) arranged in a Wenner array, which is a non-linear

array with the potential electrodes placed close together and evenly spaced. This configuration

provides discrete measurements even if good contact is not always maintained. Electrode A and B are

the positive and negative electrodes that measure the current I, and M and N are the electrodes that

measure the voltage drop V (Figure 2).

Figure 1.

Tapered 

enhance soil 

Contact 

Electrode B

Electrode N

Electrode M

ElectrodeA

Engineering type 

provides mechanical 

and electrical 

1” (25mm) O.D. steel 

Shielded cable 

signal 

Figure 2.

The off-shore horizontal extent of the interface was delineated by direct contact resistivity

measurements. The resistivity probe was driven manually into the bay bottom by scuba divers at sixinch increments. The unit’s string pot (that was originally designed for use with a Geoprobe percussion

drill keeps track of the depth measurement automatically and also trigger the electrical measurement)

had to be modified accordingly. The string pot was mounted on a jig and manually moved along a

displacement that would coincide with the depth that the probe was being driven into the bottom

sediments. Resistivity measurements were also simultaneously triggered manually. After the resistivity

was logged, the diver then drove the probe to the next six-inch level. This continued until a freshwater

zone was contacted or the probe had been driven to a maximum depth of 4 ft. The diver then moved on

to the next off-shore position at a horizontal spacing of ~30 feet, and the manual probing and logging

operations were repeated. 

Cross-sectional plots as shown in Figure 3 where prepared from the field measured resistivity of

bay bottom sediments off shore. The blue (dark) areas indicate the location of fresh water outcrops offshore and the red (light) areas indicate only saltwater is present.

DESCRIPTION OF GROUNDWATER UNDERFLOW DEVICE

In recent years ultrasonic flow meters have been developed and used to measure relatively low

flow rates in a variety of water and wastewater industries. In this study we take advantage of this

advancement in ultrasonic technology to develop a seepage meter for continuous measurement of

submarine groundwater discharge. 
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

140

Fresh/ Salt W ater Interface at W est Neck Bay:

cross-sectional profile from electrical logging

DISTANCE FROM MEAN TIDE MARK, ft

t

sea level

10500

8400

3000 1600

electrical conductance of

sea water: ~ 33,000 μ S /cm

data collection at horizontal spacing

of 25’ and depth interval of 0.5’

12000

1600 12000

Figure 3.

Flow Inlet Flow Outlet

Flow Integral

Transducer

Integral

Transducer

Bi-Directional Sonic Beam Path

(faster from left due to flow)

Figure 4.

A transient-time flow meter uses the effect of the flow on the travel time of an ultrasonic signal as

the bases for determining the flow rate. Figure 4 shows a cross section of the meter with the path of the

flow tube sonic beam and non-intrusive transducers. A multi-pulse sonic signal is transmitted through

the flow tube in both directions by transducers located at opposite ends. When there is no flow the

signal will arrive at each transducer at the same time. However, when there is flow in the tube the

upstream flow will cause the signal to arrive ahead of the signal traveling downstream. The difference

in transit times (t) between the two signals is proportional to the liquid’s flow velocity Vf

. The constant

of proportionality depends on the average of the upstream and downstream transit times and length L

of the tube. The specific discharge from the

seepage surface q is inferred from the flow

velocity by multiplying Vf

 by the ratio between

the areas of the flow tube and the collection

funnel.

Our seepage meter system is based on the

widely used technique of placing a funnel (24” x

24”) into the seepage surface on the seabed to

capture submarine groundwater discharge (Figure 5). This seepage flow is then directed via tubing

through the ultrasonic meter, which is connected to a data logger. The sampling frequency and duration

are programmed into the logger by the investigator. This meter can resolve seepage rates on the order

of 10-6 cm/s, detect reversals in flow, and it includes a totalizer which acquires data on cumulative

volume of water passing through the meter which (when after normalized by the collection funnel area)

provides the specific discharge. 

CONCLUSION

This study has demonstrated the feasibility of using a methodology that integrates geophysical

logging and transient seepage measurement to map out the spatial distribution of seepage and measure

the underflow discharge in real time. Conductivity values for the saturated sediments were observed to

range from 280 S/cm in freshwater zones to 12,800 S/cm in zones with high salinity. The discharge

zone at West Neck Bay, Shelter Island was observed to extend to 10-75 feet offshore. Electrical
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

141

Figure 6.

CAPTURE

FUNNEL

AND

FLOW

TUBE

DATA

LOGGER

Figure 5.

conductivity profiles of the coastal system obtained by directcontact resistivity logging delineates in cross-section the subset

of pore water that has been subjected to significant freshening,

and in turn they provide important constraints on the geometry

of the fresh/salt water interface and the mechanisms of mixing. 

While the resistivity logging is very effective in identifying key

areas with pronounced seepage, continuous measurements using

the ultrasonic seepage meter provide high-resolution data on the

discharge in real time. Relatively high seepage velocities

ranging from 1.27 x 10-3 cm/s to 3.94 x 10-5 cm/s (with a mean

value equivalent to 16 "/m2

/d) were measured in the study area

(Figure 6). The input of underflow to the hydrological budget

was evaluated. Integrating over the projected area of the

seepage zone, the total daily discharge was estimated to be 1.72

x 106 "/day for the northeast section of West Neck Bay. This

estimate of underflow discharge is comparable to the recharge

in the contributing area, estimated to be 1.50 x 106 "/day for this

section of the bay (Schubert,1998).

The techniques used for this

investigation were effective in defining

the spatial extent of the off shore

groundwater seepage zone and in

measuring the flux within that zone

(Shaw and Prepas, 1989). The

investigator working within this

dynamic and spatially variable zone are

advised to use extreme caution in their

interpretation of any field measurements. Adequate numbers and

replications of groundwater seepage

measurements need to be taken in order

to in order to address the spatial

variability of the bay bottom and

changes in near shore hydraulic gradients associated with tidal flux and precipitation. 

REFERENCES

Aller, R.C., ,1982. Diffusion coefficients in near shore marine sediments, Limnol. Oceanogr., v. 27 no.

3, pp. 552-556.

Archie, G.E., 1942. The electrical resistivity log as an aid in determining some reservoir

characteristics, Trans. Am. Inst. Min. Metall. Pct. Eng., v. 146, pp. 54-62.

Gueguen, Y. and V. Palciauskas, 1994. Introduction to the physics of rocks, Princeton Univ. Press,

Chapter 8.

Paulsen, R.J., C.F. Smith, and T.-f. Wong, 1997. Development and evaluation of an ultrasonic

groundwater seepage meter, in Geology of Long Island and Metropolitan New York, pp. 88-97.
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

142

Shaw, R.D. and E.E. Prepas, 1989. Anomalous short term influx of water into seepage meters,

Limnology and Oceanography, v. 34, no. 7, pp. 1343-1351.

Shaw, R.D. and E.E. Prepas, 1990. Groundwater-lake interactions: I. Accuracy of seepage meter

estimates of lake seepage, Journal of Hydrology, v. 119, pp. 105-120.

Schubert, S. E., 1998. Areas contributing groundwater to the Peconic estuary and groundwater budgets

for North and South Forks and Shelter Island, Eastern Suffolk County, New York, U.S. Geological

Survey Water Resources Investigation ,97-4136.
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

1

The use of tradenames does not imply endorsement by the U.S. Geological Survey. 

143

Influence of Stream Orientation on Contaminated

Ground-Water Discharge

By Don A. Vroblesky

The discharge zones of contaminated ground water are often beneath surface-water bodies. Such

discharge zones have been identified by using analyses of bottom-sediment gas bubbles (Vroblesky

and Lorah, 1991) and have been mapped using diffusion samplers in shallow Coastal Plain aquifers

(Vroblesky and others, 1991; 1992) and in a fractured-rock aquifer (Vroblesky and others, 1996). In

this paper, analyses of diffusion samplers buried beneath a gaining stream were used to investigate the

relation between stream orientation and ground-water contaminant discharge. 

Many environmental factors have the potential to influence the locations of contaminant discharge

to surface water. These factors include preferential ground-water flow from one side of the creek,

channeling of contaminants by fractures or other permeability heterogeneities, or capture of

contaminants by vegetation near the stream. The purpose of this report is to present preliminary data

suggesting that stream-channel orientation relative to the ground-water flow direction also is an

influential factor affecting contaminant discharge from ground water to surface water.

The approach used to delineate areas of contaminated ground-water discharge to surface water

involved the installation of diffusion samplers in the bottom sediment of Huntington Downs Creek, in

Greenville, South Carolina. The diffusion samplers consisted of 40-m5 (milliliter) glass vials enclosed

in sealable polyethylene bags. Preparation, burial, recovery, and analysis of the diffusion samplers is

described in an earlier study (Vroblesky and others, 1996). Samplers were buried approximately 1 ft

(0.3 m) deep in the creek-bottom sediment, which consisted of sand or weathered rock. The origin

point of the diffusion-sampler array in Huntington Downs Creek was a spring at the origin of the creek

(Figure 1). The total length of the array was approximately 400 ft (122 m), ending at the discharge

culvert to Huntington Downs Pond. The distance between samplers in each line was approximately 10

ft (3.048 m). The site identification numbers for the diffusion-sampler locations were the distance, in

feet, along the creek channel downstream from the spring at the origin of the creek.

The diffusion samplers were installed in the creek bottom sediment on May 25, 1994, and

recovered on June 1, 1994. Immediately upon removal of each diffusion sampler from the bottom

sediment, the outer polyethylene membrane covering the vial opening was cut open, leaving the inner

polyethylene membrane intact. A cap was screwed onto the diffusion sampler over the inner

polyethylene membrane (the polyethylene was between the glass vial and the Teflon1

 septum of the

cap). The samples were analyzed for volatile organic compound (VOC) content within 48 hours of

sample collection. A 50-μL vapor sample was slowly withdrawn and immediately analyzed on a

Photovac1

 10S55 field gas chromatograph (GC), equipped with a capillary and a packed column. The

total VOC analysis was accomplished by calibrating the GC against a 10-ppm vapor standard of

trichloroethene (TCE). In this investigation, the difference between total VOC concentrations detected

in diffusion samplers and their respective replicates ranged from 5 to 21 percent.
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

144

Water-level measurements from an observation well near the source of the creek and an

observation well downgradient from Huntington Downs Pond, and extrapolation of water-table

contours based on measurements made in wells approximately 350 to 550 ft south of the creek (Rust

Environment and Infrastructure, 1995) show that ground water near the creek moves generally north to

northwestwardly, discharging to ponds and streams (Figure 1). The water-table distribution derived

from these measurements is used in this investigation to show the probable dominant directions of

ground-water flow in the vicinity of Huntington Downs Creek (Figure 2). 

The shallow aquifer is composed of a silty clay saprolite containing relict metamorphic structures

and rock fabrics. Estimates of hydraulic conductivity in the saprolite aquifer in the study area range

from less than 2 ft/d (Rust Environment and Infrastructure, 1995) to about 3.2 ft/d Kubal-Furr and

Associates, 1996). The estimated rate of ground-water flow in the saprolite aquifer (average thickness

of approximately 40 ft) is about 76 ft/yr (Kubal-Furr and Associates, 1996). Steeply dipping relict

fractures and foliation planes are present in the saprolite and appear to influence the direction of

contaminant transport (Rust Environment and Infrastructure, 1995). Based on lithologic observations

made during an excavation of the spring in August 1994, the spring appears to be the discharge zone of

water-bearing relict fractures in the saprolite.

Analysis of the vapor in the diffusion samplers for total VOCs showed that the concentrations, as

vapor, ranged from 0.3 to 21.2 ppm relative to TCE. Analysis of the vapor in diffusion samplers from 
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

145

locations 21, 72, 110, 200, and 270 ft downstream from the spring, using the gas chromatograph

coupled to a capillary column, indicated the presence of TCE and tetrachloroethene (PCE). A similar

analysis on the vapor in the diffusion sampler from 349 ft downstream from the spring showed no

detectable concentrations of TCE or PCE. Because the only known source for VOC contamination is

south of the creek, the TCE and PCE detected in the diffusion samplers are considered to represent

contaminated ground-water discharge from the southern side of the stream. 

The highest concentrations of VOCs (16.6 to 21.2 ppm) detected in the diffusion samplers beneath

the stream were found near the origin of the creek, approximately 16 ft downstream from the spring

(Figure 2). The presence of a spring indicates that there is an upward hydraulic gradient from the

ground water into the surface water. Therefore, the relatively high concentration of total VOCs in the

ground water beneath the stream immediately downstream from the spring implies that contaminant

discharge to the creek involved seepage of contaminated ground water upward through the stream bed

as well as movement of contaminants into the stream via the spring. 

The distribution of VOC concentrations in the diffusion samplers beneath the stream showed a

general downstream decrease in the maximum concentrations detected (Figure 2). The concentrations

of TCE and PCE in diffusion samplers also decreased downstream. Thus, the downstream decrease in
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

146

maximum detected concentrations of total VOCs probably reflects a decrease in concentrations of

ground-water contaminants along a flowpath away from the source. The overall distribution of VOC

concentrations in the diffusion samplers, however, showed an irregular pattern, with areas of relatively

low concentrations between areas of relatively high concentrations (Figure 2). For example, the

concentrations of total VOCs in diffusion samplers in the reach between sampling points 140 and 190

were consistently less than 3 ppm; however, concentrations of total VOCs in diffusion samplers

immediately upstream and downstream from that reach were greater than 10 ppm. In fact,

concentrations of total VOCs in diffusion samplers between sampling points 140 and 180 were below

the apparent background concentration measured at location 349, implying that there was no

substantial discharge of chlorinated VOCs to the creek in that reach. A unique feature of the creek

reach between sampling points 140 and 180 is that it is oriented approximately parallel to the probable

dominant direction of ground-water flow. In contrast, the creek reach encompassing sampling points

50 to 90 and 200 to 250 are oriented at a sharper angle to the probable dominant direction of groundwater flow (Figure 2B). VOC concentrations in most diffusion samplers from those reaches were

higher than in the reach encompassing sampling points 140 to 180 (Figure 2A). 

A variety of factors potentially affect the distribution of contaminant discharge to a stream. Not all

of the factors were identified in this preliminary investigation. For example, potential influences

include bed-sediment heterogeneities, relict fractures, vegetation, and hyporheic zones. However, the

uniformly low concentrations of total VOCs in the channel reach oriented approximately parallel to the

probable dominant direction of ground-water flow (between diffusion samplers 140 and 190)

compared to adjacent reaches oriented at a sharper angle to ground-water flow, strongly suggest that

orientation of the creek exerts a major influence on contaminant discharge. Reaches of the creek that

transect the axis of the contamination plume receive greater contaminant and ground-water discharge

than the reach oriented along the axis of contaminant transport. 

In conclusion, diffusion samplers placed beneath creek-bed sediments were used to detect the

distribution of VOC contamination discharging from ground water to Huntington Downs Creek, in

Greenville, South Carolina. The uniformly low concentrations of total VOCs in the channel reach

oriented approximately parallel to the probable dominant direction of ground-water flow, between

diffusion samplers 140 and 190, compared to adjacent reaches oriented at a sharper angle to groundwater flow, strongly suggest that orientation of the creek is a major influence on contaminant

discharge. Reaches of the creek that transect the axis of the contamination plume appear to receive

greater contaminant discharge than a reach oriented along the axis of contaminant transport. These data

imply that site investigators attempting to locate zones of ground-water contaminant discharge to

surface water in meandering streams should put particular emphasis on reaches transecting the

dominant direction of contaminated-ground-water transport.

REFERENCES

Kuball-Furr and Associates, 1996. 1995 Annual ground-water monitoring report, Consultants report to

General Electric Company, Greenville, South Carolina, 35 pp.

Rust Environment and Infrastructure, 1993. Analyses of surface water samples from the North End,

October 12, 1993 through April 6, 1994, Consultant's data reported to the General Electric

Company, Greenville, South Carolina.

Rust Environment and Infrastructure, 1994. Data from the April 1, 1994 sampling of the North End,

Consultant's data reported to the General Electric Company, Greenville, South Carolina.

Rust Environment and Infrastructure, 1995. North end off-site area environmental investigation report,

Consultant's report to the General Electric Company, March 1995, 25 pp.
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

147

Vroblesky, D.A. and M.M. Lorah, 1991. Prospecting for zones of contaminated ground-water

discharge to streams using bottom-sediment gas bubbles, Ground Water, v. 29, no. 3, pp. 333-340.

Vroblesky, D.A., M.M. Lorah, and S.P. Trimble, 1991. Mapping zones of contaminated ground-water

discharge using creek-bottom- sediment vapor samplers, Aberdeen Proving Ground, Maryland,

Ground Water, v. 29, no. 1, pp. 7-12.

Vroblesky, D.A., J.F. Robertson, Mario Fernandez, and C.M. Aelion, 1992. The permeable-membrane

method of passive soil-gas collection, in Proceedings of the Sixth National Outdoor Action

Conference: National Water Well Assoc., May 5-13, 1992, Las Vegas, NV, pp. 3-16.

Vroblesky, D.A., L.C. Rhodes, J.F. Robertson, and J.A. Harrigan, 1996. Locating VOC contamination

in a fractured-rock aquifer at the ground-water/surface-water interface using passive vapor

collectors:, Ground Water, v. 34, no. 2, pp. 223-230.

AUTHOR INFORMATION

Don A. Vroblesky, U.S. Geological Survey, Columbia, S.C.
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

148

Factors Controlling Hyporheic Exchange in a

Southern Ontario Stream: Modeling Riffle-Scale

Patterns in Three Dimensions Using MODFLOW 

By R.G. Storey, D.D. Williams, and K.W.F. Howard

INTRODUCTION

The general pattern of exchange flow between stream surface waters and their hyporheic zones

(Fig.1) has been observed in numerous field situations (e.g., Valett, et al., 1994), and in laboratory

flume studies (e.g., Thibodeaux and Boyle, 1987). These latter studies have shown that the

downwelling/upwelling sequence is produced by a rise in hydraulic head of the surface water as it rises

over the surface of a riffle, and a rapid drop in head as surface water flows down the downstream slope

of the riffle. Thus from a surface water point of view, the forces controlling the riffle-scale pattern of

exchange flow have been adequately explained.

However, in the field situation hyporheic zones are subject not only to the forces applied by

surface waters, but also to hydraulic gradients associated with the underlying aquifer. Stream reaches

are rarely hydraulically neutral, but either receive net ground water inflow or export water to the

aquifer. How then do riffle-scale exchange flows persist without being overwhelmed by net ground

water movements?

The aims of this study were first to determine the local geological and hydrogeological conditions

necessary to produce riffle-scale exchange flows within a given regional hydrogeological environment.
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

149

A secondary aim was to determine how hyporheic exchange flows would be affected by natural

changes or artificial disruptions to aspects of the aquifer, stream flow or stream bed.

METHODS

The Speed River, in Southwestern Ontario, is a low-gradient stream, approximately 6 m wide and

0.15-0.35 m deep at the study site. The primary aquifer, a dolomite bedrock, is overlain by 20 m of low

permeability glacial kame deposits. The stream itself lies in a bed of modern alluvium 1-1.5 m deep,

and extending 30-40 m wide on each side of the stream, which has a very high permeability.

An area of the river catchment, 1,000 m x 500 m, was modeled using MODFLOW (USGS), a

three-dimensional finite difference ground water flow model. The two lateral boundaries of the model

were defined by the Speed River catchment boundaries, and the upper and lower boundaries followed

ground water flowlines. The model focused on a single riffle of the stream, a 10 m long section.

Stream stage was defined by constant head boundaries in the top layer of the model, with a rise and fall

of a few cm over the length of the riffle site.

RESULTS AND DISCUSSION

Initially the model was run without inserting the high conductivity alluvial deposits. In these runs

hyporheic flows everywhere were upwelling, and towards the stream laterally, even when aquifer

heads were low in summer.

When the zone of high permeability, representing alluvial deposits, was inserted around the stream,

flow patterns changed dramatically. Hydraulic gradients within the near-stream zone became very low

(<1 cm per m) and small differences of <3 cm in surface hydraulic head between upstream and

downstream ends of the riffle produced downwelling and lateral exchange flows in the subsurface.

This occurred even though hydraulic gradients in the material surrounding the alluvium were strongly

towards the stream.

These results show that in a low gradient stream system with strong hydraulic gradients from the

catchment towards the stream, exchange flows can still occur in a zone of high- permeability alluvium.

The essential feature of this alluvium was that it lowered hydraulic gradients within the hyporheic zone

to within the range of variation shown by the stream surface heads as they flowed between pools and

riffles. Thus these variations were able to alter flow paths up to 1.5 m deep in the stream bed.

In reducing the vertical hydraulic gradient, the alluvium changed the dominant hydraulic gradient

beneath the stream bed from almost vertical to almost horizontal. This meant that flows within the

alluvium were driven by hydraulic heads downstream rather than by those in the underlying aquifer

and surrounding catchment. The large flux of water flowing downstream within the alluvial sediments

was able to supply or withdraw sufficient water to support lateral and vertical exchange flows between

the stream and alluvial sediments, independently of hydraulic heads beneath and to the sides of the

alluvium.

Sensitivity of exchange flows to changes in system conditions

In this simulation, the highly permeable alluvial sediments allowed exchange flows in the

hyporheic zone to operate somewhat independently of the heads in the aquifer. Thus exchange flows

were relatively insensitive to changes in aquifer heads; however field data showed that a large increase
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

150

of several meters in aquifer heads from summer to fall did result in loss of exchange flows. Exchange

flows are sensitive, however, to a decrease in permeability of the alluvium, as may result from siltation

of the river bed.

Importance of alluvial sediments to exchange flows in other stream systems

Do all streams that receive net input of ground water from their catchments, require a zone of high

conductivity alluvium in order for exchange flows to occur? This depends on the steepness of the

stream, which determines the hydraulic gradient between upstream and downstream ends of a riffle,

and the hydraulic gradient in the aquifer. In a steeper stream, or one with weaker hydraulic gradients

between aquifer and stream, exchange flows can occur in less permeable near- stream sediments.

CONCLUSIONS

Modeling on a small scale has shown that surface water head differences of a few centimeters

between riffles and pools can produce exchange flows within permeable alluvial sediments, despite net

discharge of ground water to the stream. This model reveals local interactions between surface water

and ground water which would not be predicted by larger scale models, but which have important

chemical and biological consequences for the stream and ground water systems.

REFERENCES

Thibodeaux, L.J. and J.D. Boyle, 1987. Bedform-generated convective transport in bottom sediment,

Nature, v. 325, pp. 341-343.

Valett, H.M., Fisher, S.G., Grimm, N.B., and P. Camill, 1994. Vertical hydrologic exchange and

ecological stability of a desert stream ecosystem, Ecology, v. 75, pp. 548-560.

AUTHOR INFORMATION

R.G. Storey and D.D. Williams, Division of Life Sciences, University of Toronto at Scarborough, 1265

Military Trail, Scarborough Ontario, M1C 1A4, Canada.

K.W.F. Howard, Division of Physical Sciences, University of Toronto at Scarborough, 1265 Military

Trail, Scarborough Ontario, M1C 1A4, Canada.
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

151

Solute and Solid Phase Relationships in the

Surface Hyporheic Zone of a Metal Contaminated

Stream, Silver Bow Creek, MT

By Johnnie N. Moore and William W. Woessner

INTRODUCTION

Transformations occurring within the hyporheic zone impact the chemistry of both the overlying

surface waters and the underlying ground water systems. The hyporheic zone is a key ecological niche

crucial to the health of stream biota, as well as a major site of exchange, metabolism, and storage of

particulates and solutes in rivers. There is, therefore, a critical need to understand the nature of the

geochemistry that governs the transformation of materials through the hyporheic zone, especially

metals and metalloids that are toxic to aquatic organisms utilizing the hyporheic zone.

The complexities of the substrate in most streams/rivers make determining the relationships

between solute and solid phases difficult. To address this issue we have used ceramic beads as an

artificial substrate to examine what solid phases are in equilibrium within the shallow hyporheic zone.

Combined with “mini-tube wells,” this inexpensive and rapid method can be used to examine solute

and solid phase components in essentially any stream/river.

Site Conditions

Silver Bow Creek , at the headwaters of the Clark Fork River, has received contaminated surface

water and sediments from the Butte gold, silver and base metal mining and processing region for over

a century (Moore and Luoma, 1991). Over 100 million tons of tailings and mining wastes were

released into Silver Bow Creek, a portion of which were deposited at the Miles Crossing Research Site

located 18 Km down stream (Figure 1).

At the Miles Crossing Site, Silver Bow Creek has an average discharge of 850 L/s. Its water is near

neutral, high in dissolved oxygen and relative low in dissolved metals (Table 1). The fluvial plan is

covered with up to 2 m of metal rich mine tailings highly elevated in arsenic, cadmium, copper, iron,

lead, manganese and zinc.

Groundwater flow is generally in the direction of the sloping fluvial plain (Woessner, 1998; Smart,

1995; Shay, 1997) (Figure 1). Groundwater within the fluvial plan is acidic and contaminated (Table

1). On the large scale, groundwater exchange with the creek occurs as the stream stage falls below the

fluvial plain water table. In the study area reaches of flow through channel (Site 1 and 3) and zero

exchange (parallel flow)(Site 2) were observed.
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

152

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Figure 1. Location map of the Miles Crossing Research Site including Silver Bow Creek

and the associated fluvial plain (zone of stream deposited higher hydraulic conductivity

sediments. Groundwater flow is parallel to the fluvial plain and down valley. Bed tubes

and mini tube wells were installed in three portions of the stream bed, Sites 1, 2 and 3.

Definition of the Hyporheic Zone

The hyporheic zone is generally defined as the saturated zone beneath and associated with the

stream channel that shares some biological, chemical or physical characteristics with the surface water

Williams, Triska, et al., 1989; Valett, et al., 1990; Hendricks and White, 1991; Valett, 1993). Our work

uses geochemistry to define this transition zone between 100 % surface water and 100% groundwater.

We further focus part of our effort on the “surface” hyporheic zone defined as the transition zone

within 30 cm of the streambed at our site. Benner (1995) and Benner, et al. (1995) described a

geochemical transition zone extending to a depth of one or more meters at Site 1. Nagorski (1997)

continued work at the Site 1 and two additional sites focusing her effort on conditions within 30 cm of

the streambed.
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

153

Table 1.

Water Quality Silver Bow Creek

(Nagorski, 1997)

Fluvial Plain

Groundwater (Benner,

1995; Shay, 1997)

Constituent Std. Units and mg/5 Std. Units and mg/5

pH 7.7 4.2-4.9

D.O. 7.2 <1

Cond. 0.424 2.0

Alkal. 1.3E+02 0

Cl 14 20

NO3-N 1.3 1-2

SO4 78.4 1487

Al (<0.07) 33

As 0.010 0.010

Ca 47 141

Cd (<0.01) 0.55

Cu 0.136 19

Fe 0.22 365

Mg 10.4 33

Mn 0.90 28

Na 23 42

Pb (<0.1) (<0.06)

Zn 0.645 54

METHODOLOGY

Bead Tube Samplers

Bead tube samplers are 40-175 cm long polycarbonate tubing (1cm OD, 0.6 cm ID) slotted with a

1 mm width ban sawed on two sides at 3 mm intervals were filled with aluminosilicate beads (2 mm

average diameter). Plastic dividers were inserted into the columns at 10-cm intervals to minimize

vertical migration of water in the samplers. Completed bead tubes were then acid cleaned in 20%

reagent grade HCl for two hours and rinsed repeatedly with sterilized deionized water until a pH of 5

was reached. 

Bead tube samplers were inserted into the bed by driving a dual tube steel rod into the stream

sediments, removing the center solid rod, inserting the bead tube and then removing the outer steel

tube. Sediments were allowed to collapse around the sampler. The tubes were installed so that about 10

cm of the sampler extended above the stream bed. Bead tubes were retrieved after 42 to 52 days.

Retrieved tubes were rinsed in the field to remove excess sediment, labeled, photographed,

wrapped in plastic and stored. In the laboratory, columns were oven dried at 70 C and sectioned into 4

to 7 segments, depending on the amount of visible coating. Approximately one gram of beads were

placed in an acid-washed centrifuge tube to which 10 m5 of 40% metal-grade HCl was added. The

sample was shaken for 1 hour and centrifuged for 10 minutes. Solutions were analyzed for major

metals using ICAPES using standard procedures.
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

154

“Mini Tube Wells”

Small diameter tube wells were constructed using 0.95 cm diameter polyethylene with 5 cm of the

tip slotted and covered with a nylon mesh screen. These groundwater sampling devices were installed

as single instruments or nests using the same driven dual tube method used to install the bead tubes. 

These small diameter tubes were sampled using a 60 cc acid-washed syringe after purging at lease

one tube volume and one syringe volume. Tubes were then tightly capped to keep them full of bed

water. The syringe-collected samples were pushed through a 0.45 um acid-washed filter and into two

acid washed sample bottle. Bed water samples for cation analyses were preserved with trace metal

grade HNO3 and samples for anion analyses were untreated. Samples were placed on ice and returned

to the lab for standard IC and ICAPES analyses. DO, pH, specific conductance and alkalinity were

determined in the field.

RESULTS

At Site 1, surface water is alkaline, oxygenated and contains relatively high concentrations of

nitrate and low concentrations of sulfate; ground water is acid and has low concentrations of oxygen,

and high sulfate. The bead tube data allows for a higher resolution of the complexities of the hyporheic

zone (transition zone) than tube wells as interfaces can be resolved by examining the continuous bead

tube. Iron is a major control due to the precipitation of Fe-oxyhydroxides at the interfaces between the

surface water and ground water with the hyporheic zone. Solute Fe concentrations are relatively low in

the surface water and hyporheic zone water, with a concomitant elevation in the solid phase Fe on the

beads. Two sets of bead tube sections show that Fe precipitates at the surface water-hyporheic zone

boundary and at the ground water-hyporheic zone boundary. High values of solid phase Fe seen in the

surface water beads resulted from fine sediment particles attached to the bead surfaces, not from Feoxyhydroxides precipitates.

Certain elements (As, Cu, Mo, P, Pb, Sr) are strongly related to Fe precipitation, but others are

offset (e.g., Mn and Zn). Surface water and ground water contain relatively low concentrations of As

and Fe, however, the surface hyporheic zone has elevated concentrations of As and Fe. We believe

these elevated concentrations result from the dissolution of Fe-oxyhydroxides that contain As at the pH

and dissolved oxygen levels found only in the surface hyporheic zone. At some sampling points Fe

precipitates were either absent of showed a complex interfingering. 

CONCLUSIONS

Bead tubes and tube wells provided detailed geochemical data in the near surface hyporheic zone

(30 cm). The presence of iron hydroxides and co precipitating As and metals was observed and

quantified. The bead tubes provided continuous sampling of the hyporheic zone and were useful in

establishing the presence of geochemical interfaces. These instruments are best suited for sampling of

groundwater and surface water systems with geochemical contrasts.

We can identify relationships at the 1-2 cm scale in the solid phase and at the 10 cm scale in the

solute phase using these methods. We think that improvements in resolution could be obtained by

constructing bead samplers with smaller beads and using high-resolution “peepers” (dialysis membrane

samplers) for solute sampling. This modification should resolve changes within the hyporheic zone at

the cm or possibly the sub-cm scale. However, the use of peepers would increase the cost and
Proceedings of the Ground-Water/Surface-Water Interactions Workshop July 2000

155

complexity of the technique. For many problems the method we present should offer a rapid and

relatively easy technique to examine the fine scale relationships within the hyporheic zone.

REFERENCES

Benner, S.G., 1995. Geochemical processes in a transition zone between surface water and acidic,

metal-rich groundwater. M.S. Thesis, Univ. of Montana, Dept. of Geology, Missoula, 71 pp.

Benner, S.G., Moore, J.N., and E. Smart, 1995. Geochemical processes in a transitions zone between

surface water and acidic, metal-rich groundwater. ES&T, v. 29, pp. 1789-1795.

Moore, J.N. and Luoma, S.N., 1991. Hazardous wastes from large-scale metal extraction. ES&T, 24,

pp. 1279-1284.

Hendricks, S.P. and D.S. White, 1991. Physiochemical patterns within a hyporheic zone of a northern

Michigan river, with comments on surface water patterns. Can. J. Fish. Aquat. Sci., v. 48, pp.

1645-1654.

Nagorski, S.A., 1997. Impacts by acidic, metals-rich groundwater on the hyporheic zone of an

intermontane stream. M.S. Thesis, Univ. of Montana, Dept. of Geology, Missoula, 137 pp.

Shay, D.T., 1997. An Investigation of the hydrogeology and geochemistry of a floodplain aquifer

system impacted by mine tailings, Silver Bow Creek, Montana. M.S. Thesis, University of

Montana, Dept. of Geology, Missoula, 147 pp.

Smart, E.W., 1995. Surface water and groundwater interaction in a shallow unconfined alluvial aquifer

and small mountain stream Silver Bow Creek, Montana. M.S. Thesis, University of Montana,

Dept. of Geology, Missoula, 170 pp.

Triska, F.J., Kennedy, V.C., Avanzino, R.J., Zellweger, G.W., and K.E. Bencala, 1989. Retention and

transport of nutrients in a third-order stream in northwestern California: hyporheic processes.

Ecology, v. 70, pp. 1893-1905.

Valett, H.M., 1993. Surface-hyporheic interactions in a Sonoran Desert stream: Hydrologic exchange

and diel periodicity. Hydrobio., v. 259, pp. 133-144.

Valett, H. M., Fisher, S. G. and E.H. Stanley, 1990. Physical and chemical characteristics of the

hyporheic zone of a Sonoran Desert stream. J. N. Am. Benthol. Soc., v. 9, no. 3, pp. 201-205.

Williams, D.D. and H.B.N. Hynes, 1974. The occurrence of benthos deep in the substratum of a

stream. Freshwat. Biol., v. 4, pp. 233-256.

Woessner, W.W., 1998. Changing views of stream-groundwater interaction. Eds. J. Van Brahana, U.

Eckstein, L. K. Ongley, R. Schneider and J. E. Moore. Proceedings of the Joint Meeting of the

XXVIII Congress of the International Association of Hydrogeologists and the Annual Meeting of

the American Institute of Hydrology: Gambling With Groundwater, Las Vegas, Nevada, Sept. 28-

October 2, 1998, AIH, St. Paul, MN., p. 1-6.

AUTHOR INFORMATION

Johnnie N. Moore and William W. Woessner, Department of Geology, University of Montana,

Missoula, MT 59801.