Groundwater Inventory, Monitoring, and Assessment Technical Guide - DRAFT
Contents
5.0
Groundwater Evaluation and Assessment .................................................................................1
5.1
Introduction........................................................................................................................... 1
5.2
Groundwater Evaluation and Assessment Methods ................................................................ 2
5.2.1
Aquifer Tests............................................................................................................................ 2
5.2.2
Assessment of Groundwater Vulnerability ............................................................................. 4
5.2.3
Analyzing Groundwater-Level Data......................................................................................... 7
5.2.4
Geophysical Methods ............................................................................................................ 19
5.3
Groundwater – Surface Water Analysis and Assessment ....................................................... 19
5.3.1
Methods for Evaluating Groundwater-Surface Water Interactions...................................... 20
5.3.2
Understanding the Role of GDEs .......................................................................................... 30
5.3.3
Determining Environmental Flows and Levels for GDEs ....................................................... 32
5.3.4
Groundwater Age-Dating Methods ....................................................................................... 33
5.4
Groundwater Contamination Investigation Concepts ............................................................ 33
References .................................................................................................................................... 36
Appendix 5-A: Assessment of Groundwater Vulnerability using the DRASTIC Methodology ............. 41
Appendix 5-B: Flow Net Case Study: Wilshire Fen, Fremont-Winema National Forest, Oregon ......... 47
Appendix 5-C: Determining Environmental Flows and Levels for GDEs ............................................ 50
Appendix 5-D: Principles of Solute Fate and Transport ................................................................... 58
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5.0
Groundwater Evaluation and Assessment
Data analysis and evaluation are critical components of any inventory and monitoring program.
Evaluation is also a key phase of the adaptive management process and often overlooked or ineffective in
practice. Regular review and evaluation of data during collection can help ensure that any errors are
identified and corrected quickly. Analysis and evaluation of data between phases of data collection and
at the conclusion of an inventory or monitoring program provides for identification of issues or concerns
and enables appropriate interpretation of the results.
A subsequent assessment provides an interpretation of the data from an inventory or monitoring
program to assist managers’ understanding of relationships among existing conditions, trends in
conditions, or conservation actions (or inaction), and their consequences relative to desired outcomes.
5.1
Introduction
An important consideration in selecting an evaluation and assessment methodology is an understanding
of what methods are available for data analysis and how they can be applied to meet business
requirements. Few management questions are readily addressed directly by data collected in the field.
In general, field and laboratory data must be manipulated and interpreted to be useful for decision
making, such as comparison with a standard or a baseline. Often multiple measurements and
observations over time are required to develop trend or baseline information. In some instances data
collected must be compared with other datasets from other inventory programs to develop correlations
or investigate cause-effect relationships.
Cause-and-effect relationships are much more difficult to determine and require a structured
investigation to isolate causal factors. Groundwater monitoring systems provide only part of the
information needed to determine cause-and-effect relationships. A well-designed monitoring program,
however, can narrow the range of potential causes and serve as an indicator of the need to invest in a
more detailed, expensive administrative study.
The information in this section and associated documents is not intended to provide complete
instructions on how to perform the described evaluation and assessment activities. Although Forest
Service personnel may be able to perform some of them, most activities will require contractors with the
appropriate skills and equipment. This information is primarily intended to provide Forest Service
personnel with a basic understanding of the common activities and options associated with these types
of projects, and a base of knowledge to facilitate communication and project management.
This section includes:
Section 5.2 - Groundwater Evaluation and Assessment Methods – Modeling methods and their
application for addressing management needs.
Section 5.3 - Groundwater-Surface Water Analysis and Assessment – Conceptual frameworks,
methodologies for investigating these interactions, and concepts and approaches for determining
environmental flows and levels needed to sustain groundwater-dependent ecosystem function.
Section 5.4 - Groundwater Contamination Investigation Concepts – Processes that affect a
contaminant as it moves through the system, guidelines and considerations in the design of
remediation programs, and a particular method used to characterize contaminant loadings.
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Data and information derived from methods described in Sections 2 through 4 of this technical guide
support the evaluation and assessment methods described in this section.
5.2
Groundwater Evaluation and Assessment Methods
A wide range of groundwater evaluation and assessment methods are available to support agency
decision making. This subsection discusses some commonly used methods and provides guidance on
their application. It necessarily provides a simplified explanation of key hydrogeological concepts and
methods and omits a number of additional factors that may control the behavior of specific groundwater
systems and appropriate evaluation and assessment methodologies; the details are beyond the scope of
this document.
5.2.1
Aquifer Tests
This section provides an overview of the purpose of aquifer testing, hydraulic characteristics that can be
estimated from aquifer testing, and the general methodology for analysis of aquifer test data. For details
on key scientific concepts, assumptions and limitations, planning and design, field procedures, and
methods of analysis, see Appendix A, Forest Service Groundwater Technical Notes on Aquifer Testing
and on Instantaneous Change in Head (slug) Testing. The purpose of the technical notes is to familiarize
Forest Service personnel with a technique’s general theory, application, and overall procedures, and to
facilitate work with contractors.
An aquifer test (a pumping test is a particular type of aquifer test) is conducted to evaluate aquifer
hydraulic characteristics by stimulating the aquifer through constant pumping and observing the aquifer's
response (drawdown) in observation wells. Aquifer testing is a common tool used to characterize a
system of aquifers, confining units and flow system boundaries. A slug test is a type of aquifer test where
an instantaneous change in water level (increase or decrease) is made, and the effects are observed in
the same well. This is often used for geotechnical or engineering purposes to get a quick estimate
(minutes instead of days) of the aquifer properties immediately around the well.
Aquifer tests are typically interpreted by using an analytical solution of aquifer flow (one of the most
common being the Theis solution; Theis 1935) to match the data observed in the real world. The
parameters from the idealized solution are then generally assumed to apply to the real-world aquifer, as
in using short-term pumping test results from a new water supply well to estimate the extent of
drawdown at full operation. In more complex cases, a numerical model may be used to analyze the
results of an aquifer test, but the added complexity does not necessarily ensure better results.
Aquifer testing differs from well testing in that the performance of the well is the primary interest in the
latter, while the characteristics of the aquifer are goal of the former. Aquifer testing often utilizes one or
more monitoring wells, also referred to as observation wells, in addition to the well where the test is
taking place. A monitoring well is simply a well that is not being pumped but is used to monitor the
hydraulic head in the aquifer. Typically, monitoring and pumping wells are screened across the same
aquifers. An aquifer test is usually conducted by pumping water from one well at a steady rate and for at
least one day, while carefully measuring the water levels in the monitoring wells. When water is pumped
from the pumping well the pressure in the aquifer that feeds that well declines. This decline in pressure
results in lowering of the water level, drawdown, in an observation well. Drawdown decreases with
radial distance from the pumping well and increases with the length of time that the pumping continues
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because groundwater flow to the well is impeded by the geological materials it flows through. The
measure of this impedance is called hydraulic conductivity (see fig. 5-1).
Figure 5-1—Pumping a single well in an idealized unconfined aquifer. Dewatering occurs in a cone of
depression of unconfined aquifers during pumping by wells (Alley et al. 1999, fig. A-2).
The aquifer characteristics evaluated by most aquifer tests are:
Hydraulic conductivity: The rate of flow of water through a unit cross sectional area of an aquifer,
at a unit hydraulic gradient.
Specific storage (or storativity): A measure of the amount of water a confined aquifer will give up
for a certain change in head.
Transmissivity: The rate at which water is transmitted through a unit thickness of an aquifer
under a unit hydraulic gradient. It is equal to the hydraulic conductivity multiplied by the
thickness of the aquifer.
Additional aquifer characteristics sometimes evaluated, depending on the type of aquifer, include:
Specific yield or drainable porosity: A measure of the amount of water an unconfined aquifer will
give up when completely drained.
Leakage coefficient: Some aquifers are bounded by aquitards (or leaky confining units) which
slowly give up water to the aquifer, providing additional water to reduce drawdown.
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Aquifer boundaries: The presence of aquifer boundaries (recharge or no-flow) and their distance
from the pumped well and observation wells.
What is commonly referred to as well yield or specific capacity on well driller reports for new water
supply wells does not correspond to any of the above measures. It is a measure of well production
efficiency. Ideally it represents the sustainable production rate of the well, although this is not always
the case. Sometimes it can be used to estimate hydraulic conductivity (Bradbury and Rothschild, 1985).
An appropriate model or solution to the groundwater flow equation must be chosen to fit to the
observed data. There are many different choices of models; the one chosen depends on what factors are
deemed important. These factors include:









Leaky aquitards,
Unconfined flow (delayed yield),
Partial penetration of the pumping and/or monitoring wells,
Finite wellbore radius — which can lead to wellbore storage,
Dual porosity aquifers (typically in fractured rock),
Anisotropic aquifers,
Heterogeneous aquifers,
Finite aquifers (the effects of physical boundaries are seen in the test), and
Combinations of the above situations.
Nearly all aquifer-test solution methods are based on the Theis solution (Theis 1935), which is built upon
a series of simplifying assumptions. Other methods relax one or more of the assumptions the Theis
solution is built on and produce a more flexible (and more complex) result.
Designing, conducting, and analyzing an aquifer test can be complex and expensive, requiring extensive
training and experience. The type of test to be conducted depends on the hydrogeologic conditions of
the site, the purpose of the study, and the availability of funds and personnel. Consultation with
someone who is experienced in conducting aquifer tests is strongly recommended, especially in the
design stage. In addition, permits may be needed from State regulatory agencies in some states. If a
contractor conducts or analyzes the test, the Forest Service representative for the contract should be
familiar enough with aquifer-test procedures and analyses to ensure that contract requirements are met
and that test results are valid and scientifically defensible.
5.2.2
Assessment of Groundwater Vulnerability
The vulnerability of a groundwater resource to contamination depends on both, its intrinsic vulnerability
and the locations and types of naturally occurring and anthropogenic contamination, relative locations of
receptors, and the fate and transport characteristics of contaminant(s).
The intrinsic vulnerability of a groundwater system depends on the aquifer properties (hydraulic
conductivity, porosity, hydraulic gradients) and the associated sources of water and stresses for the
system (recharge, interactions with surface water, travel through the unsaturated zone, depth to the
water table, and well discharge). Intrinsic vulnerability assessments do not target specific natural or
anthropogenic sources of contamination but instead consider only the physical factors affecting the flow
of water to, and through, the groundwater system. As described in section 5.4 there are additional
factors dependent on aquifer and contaminant characteristics that also affect the vulnerability of a
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groundwater system to contamination. Karst aquifers typically have a high intrinsic vulnerability due to
the ease and speed with which contaminants can enter and move within the system (Zwahlen 2003).
Some volcanic and fractured aquifers are similarly vulnerable, particularly if they are at the surface or
overlain by highly permeable materials.
The National Research Council (1993) summarizes the broad array of definitions and approaches that are
used by government, private and academic organizations in assessing the vulnerability of groundwater to
contamination. The publication defines groundwater vulnerability as "the tendency or likelihood for
contaminants to reach a specified position in the groundwater system after introduction at some location
above the uppermost aquifer."
Assessments of the vulnerability of groundwater to contamination range in scope and complexity from
simple, qualitative, and relatively inexpensive approaches to rigorous, quantitative, and costly methods.
Tradeoffs must be carefully considered among the competing influences of the cost of an assessment,
the scientific and legal defensibility, and the amount of uncertainty in meeting the objectives of the
water resource decision maker.
Subjective rating methods focus on policy or management objectives. Relative degrees of groundwater
vulnerability usually delineated as low, medium, and high are common endpoints for all subjective rating
methods. This broad definition includes the “index” methods described below and is distinct from the
more costly and involved statistical and process-based methods that do not include subjective ratings.
Index methods (and closely associated "overlay methods") assign numerical scores or ratings directly to
various physical attributes to develop a range of vulnerability categories. The index method is one of the
most commonly used categorical rating methods and was among the earliest methods used (National
Research Council 1993). The most widely used index method is DRASTIC (Aller et al. 1987). Index
methods are a popular approach to groundwater vulnerability assessments because they are relatively
inexpensive, straightforward, use data that are commonly available or estimated, and produce an end
product that is easily interpreted and incorporated into the decision-making process.
Groundwater vulnerability assessments demand large volumes of data that are not practical to work with
in a non-computerized setting. To handle these large data organizational needs, a geographic
information system (GIS) is required. A GIS is capable of performing analysis routines such as proximity
analysis or raster surface derivation. Its primary advantage is that it integrates data layers from a
multitude of sources and scales them into one system. These traits make GIS an excellent tool for
managing the groundwater modeling process, analyzing the results, and updating and archiving spatiallyreferenced datasets (Richards et al. 1993). Base data layers required for an index method GIS analysis
typically include:
Bedrock Geology
Surficial Geology
Hydrography
Digital Elevation Model
Precipitation
Depth to Initial Groundwater
Soils
Recharge
Slope
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DRASTIC Model Overview
Many models have been designed to assess groundwater pollution hazard, but DRASTIC (Aller et al. 1987)
and DRASTIC-like models are the most widely used for such efforts (U.S. EPA 1993). The DRASTIC method
has been used to produce maps in many parts of the United States (Durnford et al. 1990), and has been
used to develop maps at a variety of scales, including national, (Kellogg et al. 1997, Lynch et al. 1994),
statewide (Hamerlinck and Ameson 1998, Seelig 1994, University of Wyoming 1998), and individual
counties and townships (Shukla et al. 2000). DRASTIC has also been used for assessing groundwater
vulnerability to oil and gas development (Colorado Geological Survey 2010, USDA Forest Service 2012).
DRASTIC was originally designed as an easy-to-use model that would allow a user with a basic knowledge
of hydrogeology to assess the relative potential for groundwater contamination. The model was neither
designed nor intended to replace on-site inspections or to specifically site any type of facility or practice.
Scientists designed the system to generalize the pollution potential for areas of 100 acres or larger.
Because pollutants vary widely in their mobility and attenuation characteristics, DRASTIC developers
assumed a generic pollutant with the travel properties of water. Because it was neither practical nor
feasible to obtain quantitative evaluations of the many micro-scale processes that affect contaminant
transport and distribution from a regional perspective, it was necessary to look at the broader physical
parameters that incorporate the many processes. When these processes are coupled with an evaluation
of the hydrogeology of the area, a realistic estimation of groundwater vulnerability is possible. The
DRASTIC method assesses aquifer sensitivity based on seven independent parameters that form its
acronym:
Depth to groundwater
Recharge (net annual)
Aquifer media
Soil media
Topography (slope)
Impact of the vadose zone (unsaturated area below soil but above water table)
Conductivity (saturated hydraulic).
These seven parameters constitute the pollution potential equation
PP = DrDw + RrRw + ArAw + SrSw + TrTw + IrIw + CrCw
where PP is the pollution potential, r is the rating, and w is the weight for each factor. The point rating
system for DRASTIC was determined by the best professional judgment of the original method
developers. Ratings for each of the parameters can range from 1 to 10, based on the relative role that
the unit plays in pollution potential. Higher numbers indicate greater potential for pollution. DRASTIC
weights can range from one to five and reflect the relative importance of each of the seven parameters.
Users alter weights based on the particular land use in question to adjust for differences in impacts of the
parameters on each use.
The index pollution potential calculated by the model is a relative indicator of the potential to
contaminate groundwater. The value has no real quantitative meaning other than to describe in relative
terms which regions within the study area have a higher potential for contamination than others. This
index value can only be applied meaningfully within its hydrogeologic setting or area of similar hydrologic
characteristics (Aller et al. 1987).
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Two examples of how the DRASTIC methodology can be applied to develop groundwater vulnerability
indices are presented in Appendix 5-A. The first case study demonstrates how groundwater vulnerability
can be assessed by using a “table-base” approach for a portion of the Beaverhead-Deerlodge National
Forest in Montana. The second case study demonstrates how groundwater vulnerability can be assessed
by using a “model-base” approach for the Pawnee National Grassland in Colorado.
The Forest Service GIS Data Dictionary contains information about Forest Service Data Standards for the
collection and storage for geology map unit feature classes at national, broad, mid, and base levels,
related aquifer vulnerability and domains of valid values. It also contains sample geodatabase designs
with standard metadata templates. The structure also contains standard feature-level metadata fields to
document changes of individual or groups of features consistent with this technical guide. The Forest
Service GIS data dictionary is available on the Forest Service FSWeb at
http://fsweb.datamgt.fs.fed.us/index.shtml.
5.2.3
Analyzing Groundwater-Level Data
Groundwater-level data are routinely collected for hydrologic studies that provide information necessary
to manage groundwater resources. These data may be needed to assess baseline hydrologic conditions,
to evaluate impacts of development on the natural system, to establish hydrologic trends due to climate
change, or to determine the direction and rate of movement of contaminants. The data themselves
cannot address these issues, so that some interpretation of the data is required for managers to make
informed decisions. This subsection describes three common approaches to using groundwater-level
data to understand groundwater systems and refine a conceptual model.
The Forest Service Groundwater Technical Note, “Groundwater Level Measurements,” provides
procedures on how to collect water-level measurements (see Appendix A). Water levels are the most
common indicator monitored in groundwater resource investigations, followed by water quality
parameters.
There are also other related Forest Service Groundwater Technical Notes providing detailed guidance on:
Well Construction and Development, and Groundwater Monitoring Well Installation for Shallow Water
Tables and Wetlands.
When analyzing water-level data, the following should be considered:

Quality of the data: Water-level data are subject to errors. These errors may be due to errors in
measurement (e.g., misread numbers, false positives), errors in recording the data on a field
sheet (e.g., transposition of numbers), or equipment problems (e.g., transducer failure or drift).
Quality checks should be done to ensure reliable data. Comparing water-level measurement to
historical levels for the well, or graphically plotting the data, may help to detect such errors.

Vertical hydraulic gradients: Within an aquifer, groundwater levels can differ, depending on the
location of the well with respect to recharge or discharge areas, and depending on the length and
placement of the well screen. For example, in a groundwater discharge area, water levels are
higher in the deeper parts of an aquifer than in the shallower parts of the aquifer. If a well is only
open to the deep part of the aquifer, then the water level will reflect this higher level. Also, if the
well has a long screen or open interval, the water level will represent an average water level for
the entire length of the screen.

Multiple aquifers: Water levels can vary greatly in different aquifers. Care should be taken in the
analysis to know in which aquifer the well is completed or screened. Also, the presence of low-
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permeability geologic formations can result in the presence of perched water that will not
represent the regional water table or potentiometric level.

Pumping wells: Pumping of groundwater can lower groundwater levels, and these levels can take
a considerable amount of time to recover to pre-pumping levels. Data obtained from a pumping
well should be noted as such. Any pumping wells near the monitoring well should be noted and
information for the pumping well, such as pumping rate, duration of pumping, and wellconstruction data, should be obtained.

Surface water: Groundwater and surface water can be well connected and one can influence the
other (Winter et al. 1998). The nature of the connection between ground and surface water is
often a key question to be answered in hydrologic studies. Surface water levels and flow data
should be obtained and considered in a groundwater-level analysis.
Time-trend analysis
Water levels, either as depth to water or water-level elevation, are plotted against time in a time-trend
analysis. This can be done manually or by using computer software, such as Excel. Meteorological data,
such as rainfall and barometric pressure, flow data for a nearby stream, or pumping data from a nearby
well should be plotted along with the water levels. Such a plot readily allows the user to see trends in
water levels with time, or to quickly detect anomalous data points. Anomalous data could represent a
measurement error or an outside influence on the water level such as pumping, rainfall, or a rise in stage
of a nearby stream (fig. 5-2). Statistical data for water levels at the well, such as daily mean level or
monthly maximum level, can be plotted and compared to the current measurement to look for climate
trends (fig. 5-3).
Figure 5-2—Transducer hydrograph from a piezometer constructed in fractured bedrock, San Bernardino National
Forest, California. The hydrograph shows water-level changes from several different stressors, including an
earthquake, storm recharge, and nearby tunnel construction.
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Figure 5-1—Hydrographs showing (A) continuous record of daily water levels for about 10 years, (B)
comparison between water-level measurements made in a single year to historical high and low levels,
and (C) statistical distribution (boxplots) of water levels for each month (from Taylor and Alley 2001).
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Potentiometric Surface Maps
Potentiometric surface maps are among the most basic and useful tools available for characterizing
groundwater flow systems. A potentiometric surface map depicts the pressure potentials of aquifers. A
water table map is a specific type of potentiometric surface map that depicts the hydraulic conditions of
the water table in an unconfined aquifer. In a confined aquifer, the potentiometric surface depicts the
pressure head at a position within the aquifer (e.g., the top of the hydrostratigraphic unit). In practice,
the terms potentiometric and piezometric are used interchangeably.
Water-table and other potentiometric surface maps are constructed to help determine directions of
groundwater flow, assess groundwater-surface water relationships, and identify recharge and discharge
areas. Such maps can be constructed manually or with computer software (e.g., Surfer). Software should
be used with care, as the programs do not use hydrologic insight to draw the contours. Computercontoured maps often have “bulls-eye” patterns, which result from the software’s interpolation routines,
and have no bearing on hydrologic conditions. Additionally, computer contouring software may not take
into account geologic features, faults for example, that can influence groundwater levels. If contouring
software is used, information such as software name and version, assumptions included in the software,
and the interpolation method needs to be documented. Regardless of the contouring method, data
points used to draw the contours should be shown on the map (fig. 5-4). The U.S. Geological Survey has
published many potentiometric surface maps that can serve as examples.
An accurate potentiometric surface map requires enough groundwater-level observations to develop
contours that do not miss important features of the flow system. Considerable interpretation and
judgment may be required in developing contours when well data points do not seem to fit into a
coherent pattern. For example, if water-level data are drawn from multiple sources, measurements in
nearby wells may have been taken at different times of the year and may not be directly comparable. On
the other hand, if all the data have been collected so as to minimize effects of short-term or seasonal
fluctuations, examination of individual monitoring point characteristics may yield explanations for
anomalous data. For instance, a single well data point that is far out of line with nearby wells may be
tapping a different aquifer than the other wells. Even wells screened in the same aquifer and measured
at the same time may give seemingly inconsistent readings if they are screened at different elevations in
the aquifer. If an anomalous data point cannot be readily explained as being unrepresentative for any
reason after verifying the values through re-sampling, then further field investigation may be required to
determine whether any localized hydrogeologic conditions are causing the anomaly.
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Figure 5-2—Using water-level measurements in wells (A) to construct a potentiometric surface map (B)
to determine groundwater flow directions (C) (from Winter et al. 1998, fig. A-2).
The starting point for a potentiometric surface map is a base map. The base map identifies well locations
and other key hydrologic features, such as springs, streams, and lakes. Drawing equipotential contours
requires some skill and judgment. Errors in contouring fall into two general categories:
1. Failure to exclude data points that are not representative; and
2. Failure to take into account hydrologic features that change the distribution of potentiometric
head, such as aquifer heterogeneity or surface features.
The following are common situations in which contouring errors might occur:
a) Extending contour lines more than a reasonable distance beyond control points. A common error is
to draw contour lines to fill the available area of the base map, even if only a portion of the base
map contains data points.
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b) Failure to exclude measurements from wells cased below the water table surface in recharge and
discharge areas for water table maps. For example, only well c in figure 5-5 gives an accurate reading
of the water table surface.
c) Failure to adjust contour lines in areas of topographic depressions occupied by lakes. Figure 5-6a
illustrates incorrect and correct interpretations in this situation.
d) Failure to recognize locally steep gradients caused by fault zones. Figure 5-6b illustrates how
conventional contouring methods erroneously portray the groundwater flow systems on the two
sides of a fault.
Figure 5-3—Cross-sectional diagram showing the water level as measured by piezometers located at
various depths. The water level in piezometer c is the same as well b since it lies along the same
equipotential line (after Mills et al. 1985).
e) Failure to consider localized mounding or depression of the potentiometric surface from
anthropogenic recharge or pumping. Pumping wells create a cone of depression around the well,
with steepened hydraulic gradients. Agricultural irrigation, artificial recharge using municipally
treated wastewater, and artificial ponds and lagoons usually cause a mounding of water tables.
When the source of recharge is confined to a relatively small area, a localized mound develops with
elevations increasing toward the center, rather than decreasing as in a pumped well. Area-wide
recharge will reduce hydraulic gradients compared to natural aquifer conditions. These features are
especially significant when they are located near a groundwater divide, because small shifts in the
location of a divide may have a major impact on the direction in which contaminants flow.
f)
Failure to consider seasonal and other short-term fluctuations in water levels. If an aquifer
experiences seasonal high and low water tables, well measurements are not comparable unless they
are taken at the same time of year. Other factors, such as dramatic changes in atmospheric pressure
and precipitation events, might reduce the comparability of well measurements even if they are
taken at the same time of year.
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Figure 5-4—Common errors in contouring water table maps: (a) topographic depression occupied by
lakes and (b) fault zones (from Davis and DeWiest 1966).
g) Use of measurements from wells tapping multiple aquifers. Wells in which the screened interval
includes multiple aquifers generally yield inaccurate water level or potentiometric measurements for
any of the sampled aquifers because the measured head reflects the interaction between heads of
the intersected aquifers. Figure 5-7 illustrates how the failure to differentiate measurements from
wells completed in two aquifers, combined with a well that connects the two, results in an apparent
depression in the potentiometric surface.
Other factors could affect water-level measurements and maps constructed from them but often are not
easy to recognize. These include improperly constructed or developed wells, improperly performed or
recorded measurements, and inconsistencies in elevation reference points used to convert depth to water to
elevation. A review of field documentation and well construction reports, if available, may identify these
sources of errors and how they can be corrected.
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Figure 5-5—Error in mapping the potentiometric surface due to mixing of two confined aquifers with
different pressures (from Davis and DeWiest 1966).
Contours of water-level change over time are also constructed to determine impacts of drought,
pumping, flooding, or recovery from these impacts. Examples of these types of maps are provided for
the High Plains regional aquifer by McGuire et al. (2003, figs. 19 and 21). Other examples of analyses of
water-level data are provided by Taylor and Alley (2001).
Anisotropy
Groundwater flows from areas of higher pressure to lower pressure. In general, that corresponds to flow
from higher elevation to lower elevation. In an isotropic aquifer (where hydraulic conductivity are the
same in all directions, Khorizontal=Kvertical), the direction of groundwater flow is perpendicular to
potentiometric contour lines, downgradient toward the next lower contour line. In many aquifers, such
as bedded sedimentary rocks, hydraulic conductivity is higher in the plane of the beds than perpendicular
to the bedding (Kh>Kv). In some aquifers, such as highly fractured crystalline rock, the horizontal
hydraulic conductivity is directionally variable and may be more permeable parallel to the direction of the
fracture system than perpendicular (Kh>Kh). Aquifers with unequal directional hydraulic conductivity
are termed “anisotropic.” In anisotropic aquifers, the groundwater flow direction will be at an angle
other than perpendicular to the contour, and the amount of this deviation will vary with the degree of
anisotropy and with the orientation of the contour relative to the orientation of the preferential
hydraulic conductivity (Freeze and Cherry 1979, pp. 174-178). Figure 5-8 illustrates how anisotropy in a
fractured rock aquifer alters the direction of groundwater flow compared to that expected in an isotropic
aquifer.
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Figure 5-6—Effect of fracture anisotropy on the orientation of the zone of contribution to a pumping
well (after Bradbury et al. 1991). In the figure, Kx and Ky are the two orthogonal components of
horizontal hydraulic conductivity (Kh).
Flow Nets
A set of intersecting equipotential lines (or potentiometric contours) and flow lines constructed according
to a strict set of rules is called a flow net, which can be a powerful tool for the analysis of groundwater
flow (Freeze and Cherry 1979, p. 168). A discussion of the rules governing the construction of flow nets is
beyond the scope of this section, and the reader is referred to Freeze and Cherry (1979, chap. 5) for a
detailed description of flow-net construction. Once a flow net is properly constructed, the amount of
groundwater flow under steady-state conditions through the area represented by the flow net can be
calculated if the hydraulic conductivity of the aquifer is known. Figure 5-9 shows an example flow net for
a simple system (modified from Freeze and Cherry 1979, fig. 5.2), in which groundwater is flowing from
the left side of the figure to the right side.
Groundwater flow through porous media (sand, silt, etc.) is described by Darcy’s Law (Darcy 1856) as long
as the flow remains slow enough to be laminar (non-turbulent). Assuming laminar groundwater flow has
been found to be reasonable in most subsurface geologic settings other than highly developed karst and
psuedokarst. The general form of Darcy’s Law is:
Q=-KiA,
where Q is the groundwater discharge across a cross-sectional area A, K is the hydraulic conductivity, and
i is hydraulic head gradient (change in head between two points divided by distance between those two
points.
By reformulating the equation for flow nets, Darcy’s Law can be used to calculate the amount of
groundwater flow through the area represented in figure 5-9.
Q = (mKH)/n
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Q is the groundwater flow rate, K is the hydraulic conductivity, H is the total change in hydraulic head
across the flow net, m is the total number of flow tubes (the area between the flow lines), and n is the
number of divisions of head in the flow net.
Figure 5-7—Example of a flow net for a simple flow system. For the figure, m = 3, n = 6, H = 60 ft, K =
0.001 ft/d, so that Q = 0.03 ft3/d per square foot of section perpendicular to the flow lines.
A standard flow net assumes that the aquifer is isotropic in the direction of flow. Though this assumption
is commonly not completely correct, it can be a useful assumption when anisotropy is low (often the case
when the dominant direction of flow is along sedimentary beds). When an aquifer is markedly
anisotropic, commonly the case in areas of vertical flow in unconsolidated and sedimentary aquifers (e.g.,
recharge and discharge areas), the actual direction of groundwater flow will not be perpendicular to the
equipotential lines. Instead, the direction of flow will deviate from the perpendicular at an angle that
depends on the ratio of the horizontal to the vertical hydraulic conductivity.
A potentiometric surface map can be developed into a flow net by constructing flow lines that intersect
the equipotential lines at right angles. Flow lines are imaginary paths that trace the flow of water
particles through the aquifer. Although there are an infinite number of both equipotential and flow lines,
the former are constructed with uniform differences in elevation between them, while the latter are
constructed so that they form, in combination with equipotential lines, a series of squares. A flow net
carefully prepared in conjunction with Darcy's Law allows estimation of the quantity of water flowing
through an area, and of the variability of transmissivity and hydraulic conductivity. Plan and cross-section
views of flow nets drawn for a gaining stream and a losing stream are shown in figures 5-10 and 5-11.
Plan view flow nets are a valuable tool in delineating the zone of contribution to a well, or for boundary
conditions for pumping wells.
Even without constructing a formal flow net, valuable information on potential groundwater flow
directions and patterns can be gained by constructing a few flow lines as described above. This can also
be used as a way of checking the reasonableness of potentiometric map contouring.
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Figure 5-8—Plan view of flow net and cross-section view through losing stream segment.
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Figure 5-9— Plan view and cross-section of flow net for gaining stream (from Heath 1983).
Appendix 5-B provides a case study illustrating the construction of a flow net for the Wilshire Fen on the
Fremont-Winema National Forest in Oregon.
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5.2.4
Geophysical Methods
Most geophysical methods applicable to groundwater investigations can be generally described in two
broad categories: surface methods and borehole methods (U.S. EPA 1993b, see Appendix A, Forest
Service Groundwater Technical Note on Geophysical Methods). Borehole geophysical methods have the
greatest utility in groundwater studies, but data collection can be problematic after wells are completed.
Surface geophysical methods are generally used to interpret geological conditions and their possible
controls on groundwater. In addition, surface methods can be used to map contamination under some
conditions. Recently, considerable technology and methodology have been developed for use in
fractured-rock settings. Weight and Sondregger (2001) summarize geophysical techniques commonly
used in hydrogeology. It is important to keep in mind that many geophysical methods yield non-unique
results that are best interpreted in combination with other lines of evidence, especially physical and
geological sampling.
Surface geophysical methods help in areal reconnaissance of geology and shallow groundwater
conditions (Zohdy et al. 1974). Three techniques are widely applicable to a variety of geological settings,
and can be useful in hydrogeological studies: (1) electrical resistivity, (2) electromagnetic conductivity (or
“EM”), and (3) seismic refraction. These methods are generally employed in hydrogeological applications
for four broad objectives: (1) evaluating groundwater quality, (2) determining the depth to the water
table, (3) determining the depth to the bedrock surface, and (4) evaluating subsurface lithology and
physical properties. Electrical methods, including square-array and azimuthal resistivity surveys and EM
surveys, are particularly useful for determining the surface location and orientation of potential waterbearing fractures (Slater et al. 1998, Taylor and Fleming 1988).
Borehole geophysics involves recording and analyzing continuous or point measurements of physical
properties made in wells or test holes (Keys 1990). The terms borehole and downhole are used
interchangeably to refer to such measurements. Most specific borehole geophysical techniques have
long been in use by the petroleum industry, in which holes being logged are usually deep and filled with
drilling muds or saline water. Many of these techniques are not suitable, or require adaptation, for use in
freshwater aquifers, which are the focus of most near-surface hydrogeological investigations.
Nevertheless, suitable borehole geophysical methods can greatly enhance the geological and
hydrogeological information obtained from water supply or monitoring wells. The development of
logging tools specifically designed for use in freshwater wells, such as the EM39 borehole conductivity
meter (McNeill 1986), and high-precision thermal and electromagnetic borehole flowmeters (Paillet
1994, 2000) should contribute to greater use of downhole methods in the future.
Borehole and core logging can provide data on the geology of the borehole, individual fractures, and the
fluid in the hole. Commonly used borehole logging methods include caliper, fluid, resistivity, and gamma
logs. Optical and acoustic imaging methods and heat pulse flowmeters are particularly useful for
detecting and evaluating individual fractures. Newer technologies that are not yet in common use
include digital borehole imaging, borehole radar, and seismic and resistivity tomography.
5.3
Groundwater – Surface Water Analysis and Assessment
The dynamic interchange of water between the surface and subsurface is a key part of the hydrologic
cycle. With about two-thirds of the available freshwater on Earth in groundwater, perennial surface
waters must be supported by the discharge of groundwater. Groundwater systems in turn, particularly in
areas of low rainfall, are often sustained by seepage from surface waters sourced in areas of higher
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precipitation. In addition, this dynamic exchange between groundwater and surface water drives critical
biogeochemical processes both at the interface between the two (the hyporheic zone in streams and
rivers and the hypolentic zone in lakes and ponds) and within the aquifers and surface waters
themselves. Therefore, delineating and describing groundwater-surface water interaction and the
ecosystems and ecological process supported by that interaction meets an important resource
management need.
5.3.1
Methods for Evaluating Groundwater-Surface Water Interactions
Depending on the study purpose, methods that are appropriate for the respective spatial and temporal
scale must be used (Kalbus et al. 2006). Different methods are better suited for characterizing or
measuring flow over large and small study areas (Rosenberry and LaBaugh 2008).
For example, determining where measurable groundwater discharge is occurring at the watershed scale,
remote sensing using aerial infrared imagery can be an effective reconnaissance tool. For a watershedscale study, small-scale flow phenomena are likely to be of little importance to the overall study goal. In
such large-scale studies, the net flux integrated over an entire stream reach, lake, or wetland often is the
desired result. Watershed-scale flow modeling, groundwater flow modeling, flow-net analysis, or dyeand geochemical-tracer tests are often on the order of hundreds of meters to a kilometer or more.
If the goal of a study is to identify or delineate specific zones or areas of surface-water-to-groundwater
flow, or groundwater-to-surface-water flow, smaller scale spatial and temporal variations in flow become
important. That will require measurement tools that provide results over an intermediate scale, tens to
hundreds of meters, or smaller scale should be selected. Measuring surface water flow at two places
some distance apart in a stream segment enables calculation of gains or losses in flow in the segment, as
appropriate for these intermediate scale studies. For local studies in which flow to or from surface water
may be focused, tools such as seepage meters, minipiezometers, and buried temperature probes may be
most appropriate.
Most measurements for the purpose of quantifying exchange between surface water and groundwater
are obtained at points within a short distance of the edge of the surface water body. Shorelines and
banks represent the horizontal interface between groundwater and surface water, a zone that is highly
dynamic spatially and temporally. Determination of how the interaction of surface water with
groundwater changes over time is made possible by using data-recording devices (“data loggers”) in
conjunction with pressure transducers, thermistors, and water-quality probes. In Rosenberry and
LaBaugh (2008, pp. 8–12) the characteristics of near-shore environments in detail and present numerous
examples are discussed. The movement of water between surface water and groundwater can occur in a
variety of settings or landscapes (fig. 5-12), each of which can be related to a break in slope of the water
table defined by “an upland adjacent to a lowland separated by an intervening steeper slope” (Winter
2001).
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Figure 5-12— Generalized hydrologic landscapes: A, narrow uplands and lowlands separated by a large,
steep valley side (mountainous terrain); B, large, broad lowland separated from narrow uplands by
steeper valley sides (playas and basins of interior drainage); C, small, narrow lowlands separated from
large, broad uplands by steeper valley side (plateaus and high plains); D, small, fundamental hydrologic
landscape units nested within a large, fundamental hydrologic landscape unit (large riverine valley with
terraces); E, small, fundamental hydrologic landscape units superimposed on a larger fundamental
hydrologic landscape unit (coastal plain with terraces and scarps); F, small, fundamental hydrologic
landscape units superimposed at random on large, fundamental hydrologic landscape units (hummocky
glacial and dune terrain) (Rosenberry and LaBaugh 2008, fig. 4).
Once the purpose of the study and the scale of the investigation have been established, methods of
investigation can be selected to determine most effectively where exchanges between surface water and
groundwater are taking place, the direction of flow, the rate or quantity of that flow, and whether the
rate and direction of flow changes over time. Rosenberry and LaBaugh (2008) discuss each of these
determinations and measurements in some detail. Fonseca (2008) discusses monitoring groundwaterdependent ecosystems in southeastern Arizona, and presents a generalized monitoring plan for these
systems.
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Common methods for examining the exchange of water between ground- and surface water bodies are
described in the following subsections. Some of these methods involve using previously installed
hydrological instruments and existing data rather than making new measurements. In some cases,
however, the investigator may also install new wells, stream-gaging equipment, or rain gauges to obtain
sufficient data to make the analysis more robust and the results less uncertain. Other methods require
that the investigator make additional, specific measurements or observations of hydrological, physical,
chemical, or biological characteristics (Rosenberry and LaBaugh 2008, p. 16). Much of the discussion in
this subsection is taken from Rosenberry and LaBaugh (2008).
Stream Discharge Measurement Method
The simplest way to measure the amount of flow between a stream and the surrounding groundwater
system is by means of a “seepage run” (Harvey and Wagner 2000). This method requires the
measurement of streamflow at several locations and any flow in tributary streams or diversion channels
in between the measurement points. Typically, these measurements are made during times of low or
base flow. The amount of groundwater contribution to the stream or loss of streamflow to the
groundwater system is simply the residual of the stream-flow measurements. This method works well in
small streams, but for larger streams and rivers the errors associated with the measurement of flow in
the channel often are greater than the net exchange of water to or from the stream or river. If there are
wells pumping within close proximity of the seepage run, the pumping could affect the results and if
possible the pumping rates should be determined and included in the interpretation.
Flow-Duration Curve Interpretation
The use of “flow-duration” curves (Searcy 1959) is another simple method for estimating groundwater
contribution to a stream. A flow-duration curve is a cumulative frequency curve that shows the percent
of time specified discharges were equaled or exceeded during a given period. Comparison of established
flow-duration curves can provide valuable insights into the drainage characteristics of different streams
or different reaches of a stream. The U.S. Geological Survey (USGS) has estimated flow-duration curves
for most of the established USGS stream gauges, and the information can be accessed on the USGS
Waterwatch website (http://waterwatch.usgs.gov/index.php). Steep curves are indicative of a high
degree of runoff; flat curves are indicative of a high degree of surface or subsurface storage in the
drainage basin. The point at which the curve flattens, typically between 50 to 90 percent flow-duration
rate, often indicates the groundwater flow rate to the stream.
Analytical Methods
A fundamental assumption in the analytical methods is that streamflow is an integrated response to
rainfall, groundwater recharge, and groundwater discharge over the stream’s watershed, and that
groundwater discharge to the stream provides the steady flow, commonly referred to as baseflow, in the
stream between rainfall events. Analytical models generally determine baseflow through hydrographseparation techniques (see subsection on hydrograph separation below). Several automated routines
have been developed to assist in this approach (Rutledge 1992, 1998, 2000).
Barlow (2000) developed a computer program, STRMDEPL, to calculate stream-flow depletion by wells
using analytical techniques. This program was updated and extended by Reeves (2008).
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Numerical Modeling Method
Two basic types of numerical models currently are popular for evaluation of groundwater-surface water
(GW/SW) interactions – rainfall-runoff models and groundwater flow models. Several rainfall-runoff
models have been developed; these models aerially subdivide watersheds and subwatersheds and
calculate hydrologic parameters for each subarea. Rainfall-runoff models generally are calibrated to
match river flow at the outlet of a watershed or subwatershed. Some models include the groundwater
component of flow in each area. The current trend is to couple distributed-area watershed-scale rainfallrunoff models with groundwater flow models to better assess the temporal and spatial variability of the
interaction between groundwater and surface water (e.g., GSFLOW, Markstrom et al. 2008).
The most popular software used to simulate groundwater flow is the USGS program, MODFLOW
(Harbaugh et al. 2000). This modeling software can simulate groundwater flow to or from a stream,
spring, reservoir, or lake by using one of several “modules” or “packages.” Many of these packages
require determination of hydraulic conductivity and thickness of the sediments or rock that underlies the
surface water body. Groundwater flow models are calibrated to match measured groundwater levels
and flows (e.g., groundwater discharge to streams or springs) throughout the model area.
Recommendations for application of numerical models in National Forest System assessments are
discussed in detail in the Forest Service Technical Note on Modeling of Groundwater Systems.
Well and Flow-Net Methods
The flow-net analysis method, often called the “Darcy approach,” is probably the most frequently used
method for quantifying flow between groundwater and surface water, especially on a whole-lake or
watershed scale. In this method, a combination of measurements of water levels in near-shore water
table wells and measurements of water stage in adjacent surface water bodies are used to calculate
water table gradients between the wells and the surface water body. There are two ways typically used
to generate the flow net: in one the investigator segments the shoreline of the surface water body,
depending on the number and location of nearby wells; the other is to generate equipotential lines based
on hydraulic head and surface water stage data, by using flow-net analysis to calculate flows to and from
the surface water body. Both are described in detail, including a discussion of sources of error, by
Rosenberry and LaBaugh (2008, pp. 43–49). A general discussion of flow net construction is included in
section 5.2.3
Seepage Meter Method
The seepage meter is one of the most commonly used devices for making a direct measurement of the
flux of water across the sediment-water interface. Lee (1977) developed an inexpensive and simple
meter that has changed little during the decades since its inception (fig. 5-13). The meter consists of the
cut-off end of a 55-gallon (208-Liter) drum, which is attached via tube to a plastic bag partially filled with
a known volume of water. The drum (chamber) is submerged in the surface water body and placed in
the sediment to contain the seepage that crosses that part of the sediment-water interface. The bag
then is attached to the chamber for a measured amount of time, after which the bag is removed and the
volume of water contained in the bag is re-measured. The change in volume during the time the bag was
attached to the chamber is the volumetric rate of flow through the part of the bed covered by the
chamber (volume/time). The volumetric rate of flow then can be divided by the approximately 0.25
square meter area covered by the chamber to express seepage as a flux velocity (distance/time). Flux
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velocity is useful because it normalizes the area covered by the seepage meter and allows comparisons of
results with other studies (and other sizes of seepage meters). Seepage flux velocity is multiplied
typically by a coefficient that compensates for inefficiencies in flow within the meter, restrictions to flow
through the connector between the bag and the chamber, and any resistance to movement of the bag as
it fills or empties. Other types of seepage meters include a few with automated data collection that allow
measurement of temporal variability of seepage flux. Some of these meters are based on modifications
to borehole flowmeters, such as the heat-pulse and electromagnetic flowmeters, or are based on
acoustic-velocity technology. Rosenberry and LaBaugh (2008, pp. 54–66) provide a detailed discussion of
seepage meters and their use, including sources of error and best practices.
Figure 5-13—Example of a half-barrel seepage meter (modified from Lee and Cherry 1978). The top
panel shows a typical installation with the bag connected to a tube inserted through a stopper. The
bottom panel shows installation in shallow water with a vent tube to allow trapped gas to escape.
Minipiezometer Method
The minipiezometer method uses a portable probe (a minipeizometer) connected to a manometer, which
together is sometimes referred to as a hydraulic potentiomanometer (fig. 5-14). The manometer
provides a comparison between the stage of a surface water body and the hydraulic head beneath the
surface water body at the depth to which the screen at the end of the probe is driven (Winter et al.
1988). The difference in head divided by the distance between the screen and the sediment-water
interface is a measurement of the vertical hydraulic-head gradient (VHG). Driven to different depths
beneath the sediment-water interface, the minipiezometer can provide information about variability in
vertical hydraulic-head gradient with depth.
Vertical hydraulic gradient is a unitless measure that is positive under upwelling (or discharge) conditions
and negative under downwelling (or recharge) conditions (Baxter et al. 2003). Specifically, VHG = h/l,
where h is the difference in head between the water level in the minipiezometer and the level of the
stream surface (cm) and l is the depth from the streambed surface to the first opening in the
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minipiezometer sidewall (for minipiezometers with multiple perforations, the location of the middle of
the perforated interval is also often used).
To get accurate and precise estimates of the height of the stream surface relative to the water level in
the minipiezometer, it was found necessary to use a ‘‘stilling well’’ (Baxter et al. 2003). The stilling well
was simply a hollow tube (same diameter as the piezometer) open at both ends that was attached to the
side of the minipiezometer via a pair of plastic clips. The stilling well was always placed alongside of the
minipiezometer in a line perpendicular to any current. The top of the stilling well extended above the
stream’s surface (but not above the level of the minipiezometer), and the bottom opened near the
substratum but was not driven into the streambed.
Quantification of groundwater-stream water exchange can be accomplished by measuring vertical
hydraulic gradient (VHG) along with streambed hydraulic conductivity in minipiezometers. Baxter et al.
(2003) describes a method for conducting slug tests in minipiezometers for determining the hydraulic
conductivity of the streambed and equations to calculate the vertical flux through the streambed.
The device does not give a direct indication of seepage flux, but when used in combination with a
seepage meter, which does measure water flux, the two devices can yield information about the
hydraulic conductivity of the sediments (e.g., Kelly and Murdoch 2003, Zamora 2006). Because this
device provides a quick characterization of the direction and magnitude of the vertical hydraulic gradient,
it is useful as a reconnaissance tool in lakes, wetlands, and streams. It also is useful in areas where nearshore water table wells or piezometers do not exist, are sparsely distributed, or are impractical to install
and maintain. A detailed discussion of the use of minipiezometers and other similar devices, along with a
discussion of sources of error is provided by Rosenberry and LaBaugh (2008, pp. 49–54).
Figure 5-14—Components of a minipiezometer system (modified from Winter et al. 1988).
Selecting the appropriate methods of measurement and calculation is one of the most important
decisions when quantifying exchange between groundwater and surface water (Rosenberry and LaBaugh
2008, p. 66). Each method has advantages and disadvantages that may be relevant to the study area of
interest. Table 5-1 provides a general guideline to conditions or situations in which flow nets, seepage
meters, and minipiezometers for quantifying GW/SW interactions are particularly well- or ill-suited.
Using more than one method to quantify the exchange between groundwater and surface water can be
informative and valuable in increasing the confidence in the flux values estimated or calculated.
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Table 5-1—Conditions for which methods of quantifying flow between groundwater and surface water
are well- or ill-suited (modified from Rosenberry and LaBaugh 2008, table 4).
Method
Well-suited for:
Ill-suited for:
Calculations from water levels in
network of wells and surface
water stage
- Basin-scale quantification
- Distinguishing areas of inflow from
areas of outflow
- Determining large-scale aquifer
characteristics
- Relatively homogeneous aquifers
Minipiezometers and well-probe
measurements
- Fine sand to medium gravel
sediments
- Quick reconnaissance for
qualitative determination of
direction of flow
- Determining variability of vertical
hydraulic gradient with depth
- Collection of water quality samples
- Direct measurement of seepage
flux
- Areal distribution of seepage flux
- Sediments ranging from clayey-silt
to fine- medium gravel
- Calm-water settings
- Shallow-water settings
- Determining flux of some chemicals
that enter or leave a surface
water body
- Steep and (or) rocky shorelines
where installation of wells is
difficult or impossible
- Low-lying terrain where shoreline
migration is large and
evapotranspiration is a
significant factor
- Areas with complex geology or
vertical flow regimes where
effective depth of aquifer is
nearly impossible to determine
- Fine-grained sediments
- Rocky shorelines or bedrock
- Surface water body with any
appreciable wave action
- Fast-flowing water
- Organic, gas-rich sediments
Seepage-meter measurements
- Surface water body with any
appreciable wave action
- Areas with strong currents or fastflowing water
- Very soft, low-density sediments
- Rocky sediment beds
- Bed areas with dense vegetation
Hydrograph Separation Methods
The objective of hydrograph separation is to split the total stream-flow or spring-flow hydrograph into
two parts: (1) the surface runoff or quick-response flow that is related to storm or snowmelt events and
(2) the groundwater (baseflow) contribution. Baseflow generally occurs through groundwater discharge
to streams and springs. Baseflow will augment surface runoff during storms or snowmelt, but tends to be
the primary source of flow during periods of stream-flow recession when there is no precipitation.
Estimates of baseflow and surface runoff may be affected by upstream flow diversion into or out of a
stream or flow regulation that changes the natural flow of a stream, and should be taken into account
when considering use of this method. Hydrograph separation provides a useful tool for estimating the
proportion of discharge that is derived from groundwater verses surface water for different time periods.
The “curve-shape” method uses a consistent approach to apportion discharge between baseflow and
runoff across each flow event (storm or snowmelt). Figure 5-15 illustrates this concept.
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Figure 5-15–Hydrograph separation using the curve-shape method to apportion streamflow into
baseflow and runoff components.
Efforts have been made to develop automated approaches that can be applied consistently and allow
calculation of baseflow estimates for large datasets (for examples, see Sloto and Crouse 1996 and
Rutledge 1998).
In general, average annual estimates of baseflow and surface runoff using curve-shape-based baseflow
separation techniques are more reliable than monthly or daily estimates. The accuracy of estimates of
baseflow characteristics depends on the period of record used in the analysis. Records for stations with
periods of extreme climatological conditions, such as mostly dry or mostly wet years, will exhibit a bias
toward the extreme. Long-term stations that experienced long-term climatological conditions will
provide a more reliable average baseflow estimate (Sloto and Crouse 1996).
The hydrochemical method (Freeze and Cherry 1979) of hydrograph separation is commonly used to
determine the relative amounts of surface water and groundwater in the total discharge of a spring or
stream. The contribution of groundwater to the total flow is calculated as follows:
æ C -C ö
d
÷÷
Qg = Q çç
C
C
è g
d ø
where Qg is the discharge from groundwater, Q is the total discharge, C is the constituent concentration
in the total discharge, Cd is the constituent concentration in surface runoff, and Cg is the constituent
concentration in local groundwater.
To illustrate the method, relative amounts of surface water and groundwater in the total discharge of the
Anaconda Job Corps Center public water supply spring and adjacent Foster Creek in southwest Montana
are shown in figure 5-16. Separations were performed by using dissolved calcium concentrations. The
surface water contribution to the spring varies seasonally from almost none in late winter and early
spring to nearly 90 percent of the total flow during spring runoff (fig. 5-16 A). Figure 5-16 B shows the
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large contribution of groundwater to the flow of Foster Creek during the baseflow season (late summer,
fall, and winter).
Figure 5-16—Chemical hydrograph separation from (A) the Anaconda Job Corps Center Spring and (B)
Foster Creek into surface water and groundwater discharge components using dissolved calcium
concentration.
Chemical Budget Method
Calculation of chemical budgets for a stream, lake, or wetland is another way in which solutes can be
used to make quantitative estimates of surface water exchange with groundwater (Rosenberry and
LaBaugh 2008, p. 24). Many conservative constituents present either naturally or artificially in
groundwater have been used in the calculation of chemical budgets, including calcium, sodium, and
chloride. This approach involves characterizing the average concentration of the selected constituent in
the surface water body and in each of the other components of the water budget for that surface water
body.
The ratios of the isotopes of oxygen and hydrogen present in water have also been used to distinguish
sources of water, including groundwater discharge to surface water bodies (Dincer 1968, Kendall and
McDonnell 1998). These isotopes are useful because they are part of the water and not solutes dissolved
in the water. The method works well when the degree of isotopic fractionation of the water is different
for different sources of water. The process of evaporation tends to remove lighter isotopes, leaving the
heavier isotopes behind. Thus, the ratio of lighter to heavier isotopes will change over time in the water
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and the water vapor. More detailed explanation of the use of stable isotopes in catchment hydrology is
given in the Forest Service Technical Note on Groundwater Tracing Methods.
Tracer Methods
Dyes and other soluble tracers can be added to water and then “tracked” to provide direct, qualitative
information about groundwater movement to streams or other surface waters (see the Technical Note
on Groundwater Tracing Methods). Fluorescent dyes that are readily detected at small concentrations
and pose little environmental risk make a useful tool for tracing groundwater flow paths, particularly in
karst terrain. Thus, dye-tracer studies can be used to determine the time-of-travel for groundwater to
move to and into surface water, as well as hydraulic properties of aquifer systems (Mull et al. 1988). A
Forest Service Groundwater Technical Note on Water Tracing in Karst Terrains with Fluorescent Dyes,
describes procedures for implementing a dye trace. The use of dyes as tracers is also described in more
detail by Rosenberry and LaBaugh (2008, pp. 71–114).
Commonly, a reconnaissance of a karst groundwater basin is made to identify likely areas of potential
surface water flow into groundwater or groundwater flow to the surface. An inventory is made of
springs, sinkholes, boreholes or screened wells, and sinking streams. Appropriate sites are picked for dye
injection, and the potential discharge areas, springs, and stream reaches are monitored over an
appropriate period of time, hours or days, for appearance of the dye.
Tracer Injection/Synoptic Sampling Method
A traditional approach to quantifying hydrologic interactions has been to measure discharge and
chemistry at a catchment outlet. That approach, however, may not provide the needed spatial detail to
support detailed decision making. A catchment or mass-loading approach provides spatial detail by
combining tracer-injection and synoptic-sampling methods to quantify loading (Kimball 1997). Synoptic
sampling is the collection of samples from many locations during a short period of time, typically a few
hours. It is like a "snapshot" of the water quality changes along a stream at a given point in time.
Traditional discharge measurement methods work well where the channel bottom and banks are
smooth. In mountain streams, however, the stream bottom typically is covered with cobbles, allowing
much of the water to flow through the cobbles of the streambed where it cannot be measured by a
flowmeter. Thus, accurate discharge measurements are difficult to obtain in mountain streams, even
under the best of conditions. To address these questions, the tracer-dilution method coupled with
synoptic sampling can be used.
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 the concentration of the injected tracer and the rate at which it is added to the
stream is known, the mass that is added to the stream is also known. By measuring the concentration of
the tracer upstream and downstream from the injection point, the discharge in the stream downstream
from the injection point can be calculated. Mathematically, this is written:
Qs = (Ci * Qi) / (CB-CA),
where Qs is the discharge of the stream, Ci is the tracer concentration in the injection solution, Qi is the
rate of injection into the stream, CB is the tracer concentration downstream from the injection point, and
CA is the tracer concentration upstream of the injection point.
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Thermal Methods
Fiber optic distributed temperature measurements can reveal groundwater-surface water interaction in
riverine, lacustrine, and wetland environments (e.g., Stonestrom and Constantz 2003, Lapham 1989;).
Recent advances in temperature sensor technology have renewed interest in the use of temperature to
monitor the exchange of water between groundwater and surface water.
Fiber-optic distributed temperature sensing (FO DTS) technology developed and commercialized in the
last decade supplements conventional discrete-point and remote sensing technologies for temperature
measurement. FO DTS systems use laser light traversing optical telecommunication fibers to measure
temperature continuously along the entire fiber length. Although instrumentation capabilities are
improving rapidly, single- and multi-channel systems are currently capable of measuring the temperature
of fibers as long as 30 kilometers with thermal resolution ranging from 0.1 to 0.01 ˚C at a spatial
resolution of 1 meter and temporal resolution on the order of seconds to minutes, subject to a tradeoff
between fiber length and spatial resolution effects resulting from measurement averaging or “stack”
time.
FO DTS involves propagation of a laser pulse along an optical fiber and the measurement and analysis
of light backscattered by the fiber. Because the propagation velocity of light in a fiber is known (or can
be measured), the location of a measurement “point” can be determined by careful timing of the
scattered light arrivals relative to the incident pulse; the temperature at the point is determined
through analysis of the backscattered light in terms of temperature-sensitive mechanisms (including
Brillouin and Raman scatter). Both optical time-domain reflectometry (OTDR) and optical
frequency-domain reflectometry (OFDR) methods are used for data analysis in commercially available
systems. The spatial and thermal resolution of the measurements can be improved through
averaging (or “stacking”) of multiple measurements. For additional details, see Selker et al. (2006).
5.3.2
Understanding the Role of GDEs
Three principal forms of analysis and assessment are generally conducted in association with
groundwater-dependent ecosystems (GDEs):
The characterization and function of an individual GDE feature.
The assessment and evaluation of a feature’s relationship to and support of ecological
systems within a larger area.
3. Determining environmental flows and levels needed to sustain groundwater dependent
ecosystem function.
1.
2.
Groundwater-dependent ecosystems are the ecological outcome of groundwater-surface water
interactions. Section 3.3 discussed approaches to identifying and characterizing GDEs as a key first step
in determining management options.
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Landscape-Scale Relationships
Another important type of analysis for GDEs involves understanding the context of individual GDE
features and how they and other GDEs function and sustain ecological systems over larger land areas.
Assessment of the biological significance and role of a feature in sustaining biological diversity is a
common business need within the National Forest System.
The Nature Conservancy (TNC) has developed a methods guide for integrating groundwater needs of
ecosystems and species into conservation planning (Brown et al. 2007). TNC’s Methods Guide identifies a
variety of data and information inputs needed to identify the types and locations of GDEs at a coarse
scale across the landscape.
Collection of data needed to conduct assessments using TNC’s Methods Guide or similar approaches will
strengthen understanding of GDEs and their role in the larger landscape. These assessments are usually
conducted within a defined watershed or groundwater system. Table 5-2 illustrates principal data inputs
identified in TNC’s Methods Guide.
Many of these data are best analyzed within a geographic information system, which requires analytical
skills beyond traditional expertise in ecology and hydrogeology to execute.
A variety of condition assessment methods exists and can be used to compare conditions between
different GDE features of the same type. Observed or measured data are the preferred inputs to these
condition assessments vs. interpreted classes or ratings because they can be repeated more accurately
over time and between different sites.
Table 5-2–Data inputs needed for assessment of GDE relationships to sustaining biological diversity
using TNC’s Methods Guide.
Data Category
Hydrologic
Regime
Water Chemistry
& Temperature
Hydrogeologic
Setting
Attribute Description
-
Quality and quantity (timing, location, and duration) of water delivery
-
Water quality or specific water chemistry
Water temperature regime
Topography and slope of land surface in the watershed
Composition, stratigraphy, and structure of subsurface geological materials in
the watershed and underlying the ecosystem
Position of the ecosystem in the landscape with respect to surface and
subsurface-groundwater flow patterns
Aquifer recharge zones
Wetlands
Springs
Lakes
Streams
Groundwater-dependent species present
Habitat restricted to locations with groundwater discharge or maintained or
associated with groundwater discharge or a shallow water table
Water chemistry or quality conditions provided by or influenced by
groundwater
-
Ecological
Setting
GroundwaterDependent
Species
-
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5.3.3
Determining Environmental Flows and Levels for GDEs
The Forest Service and The Nature Conservancy are developing a method to determine the
environmental flows and levels (EFL) for groundwater-dependent ecosystems. The current Forest Service
working definition of environmental flows and levels is:
“Environmental flows and levels describe the quantity, quality, timing and range of variability of
water flows and levels required to sustain or restore freshwater and estuarine ecosystems and
the functions and services they provide. Environmental flows and levels include instream flows,
geomorphic and flood flows, groundwater levels, and lake and wetland levels established for
environmental purposes.”
It is based on the definition of environmental flows in the 2007 Brisbane Declaration to protect
freshwater systems.
However, EFL includes concepts important for groundwater-dependent ecosystems, such as static water
levels. Within the context of groundwater dependent ecosystems, the concepts associated with EFL
were previously referred to as “Environmental Water Requirements” or “Ecological Water Requirements
(EWR)” in various Forest Service working documents. These concepts are now collectively referred to as
environmental flows and levels.
GDEs, by definition, are supported by aquifers. Any activity that lowers or raises the water table,
piezometric surface, or groundwater discharge rate or timing, or alters the groundwater chemistry can
affect the integrity of GDE habitat. Drought, major development, mining within a GDE’s recharge zone,
and leaky septic systems can significantly alter groundwater discharge to a GDE. As stated above, the EFL
approach can be applied to either changes in water quantity or quality. However, at this point the
emphasis of the TNC and Forest Service partnership has been on establishing EFL methodology for
changes in water quantity, which will be reflected in the discussion from this point forward.
Determining quantitative EFL is important for ensuring adequate water to sustain GDEs in cases where
management activities, such as mining, wells for municipal water supply, and water withdrawals for
livestock operations, affect groundwater flow to GDEs. EFL takes into account the ability of GDEs to
adjust to changes in the water regime. Therefore, EFL allows for an acceptable level of change to occur
relative to the existing hydrologic conditions. When use of water resources shifts the hydrologic
conditions below that defined by the EFL, irreversible ecological harm may occur, including impairment
or loss of ecological structure and function.
An EFL establishes the limit to water-level change for an aquifer or water body. It is the limit at which
further withdrawals or flooding would be significantly harmful to the GDE. The establishment of an EFL
for a discharge wetland or phreatophyte community generally defines a limit on water-table drawdown
or increase, or a reduction in discharge feeding the feature from the aquifer. If monitoring indicates the
actual water levels or flows are below the defined EFL, a recovery plan can be developed to reach
acceptable levels.
The EFL seeks to ensure adequate groundwater discharge to protect the most sensitive environmental
values of a GDE. By protecting the most sensitive environmental values, the GDE as a whole is assumed
to be protected. For example, in discharge wetlands where the most sensitive environmental values may
be the presence of rare plants and peat substrates, a recommended EFL would reflect the groundwater
level that would allow for no loss of rare plants and no net decomposition of peat. In a phreatophyte
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community, on the other hand, the most sensitive environmental value may be the presence of particular
species that are closely adapted to specific water table elevations, and so the EFL would reflect those
requirements.
Additional background information on EFLs and details about application of the methodology can be
found in Appendix 5-C.
5.3.4
Groundwater Age-Dating Methods
Age dating of groundwater can be useful in evaluating groundwater-surface water interactions.
Additional detail on specific methods and their applicability can be found in the Forest Service Technical
Note on Groundwater Age-dating Methods.
5.4
Groundwater Contamination Investigation Concepts
Groundwater contamination investigations and remediation activities have been conducted on
administrative units in all regions in the past due to legacy activities, such as mining, and past land
management. Assessment and remediation of groundwater contamination requires not only a detailed
understanding of the groundwater flow system, but a thorough knowledge of the processes that affect
the contaminant as it moves through the system. The methods for groundwater contaminant work are
complex in nature and tend to be highly site-specific in their application, and are outside the scope of this
technical guide. Therefore, this section focuses on key concepts supporting groundwater contaminant
investigations. Appendix 5-D - Principles of Solute Fate and Transport describes these processes,
including their relative importance and the parameters that are required to quantify them. The rest of
this section describes guidelines and considerations in the design of remediation programs.
Remediation of Contaminated Groundwater
Once groundwater is contaminated, it is difficult and typically very expensive to restore to natural or precontamination conditions. The broad range of chemical, physical, and biological characteristics of the
thousands of potential groundwater contaminants coupled with the complex heterogeneities of
subsurface flow and contaminant transport make it very difficult to determine the exact nature and
extent of groundwater contamination in a given area or aquifer. If the value of the contaminated
groundwater is great enough or the risk to potential receptors (for example, drinking water wells or
surface waters), it is very important to determine the nature and extent of the contaminant or
contaminants. The results of this investigation are then used to evaluate options for groundwater
remediation. Unfortunately, inadequate investigations, often resulting from efforts to save money and
time, are a major reason for many failed remediation attempts. The importance of a thorough
investigation cannot be overemphasized in assuring the success of the subsequent action to restore
groundwater quality and in minimizing overall costs.
Remedial investigation or RI is the formal term used for investigations performed under the
Comprehensive Environmental Response Compensation and Liability Act (CERCLA or Superfund)
regulations. Other terms may be used for similar efforts, often depending on the regulatory program
under which the work is performed. These include site investigation (SI), Resource Conservation and
Recovery Act (RCRA) remedial facility investigation (RFI), and, less commonly, Phase 3 environmental site
assessment. While the names are different, they represent the same type of activity.
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There are similar data collection activities that may precede full investigations as well. Common names
for these are site inspection under Superfund, RCRA facility assessment (RFA), and Phase 2 environmental
site assessment for real property transactions. These are intended to assess the presence of
contamination that may require further investigation and possible remediation, but not to define its
degree and extent as required for a properly designed remediation program. There has been tendency
or pressure in recent years to use the results of these less thorough investigations as the basis for
implementing remediation. This may result in remediation that is ultimately inadequate or excessively
expensive.
Strategies and technologies that are typically used to remediate contaminated groundwater include the
following general categories:



Removing or controlling the source of contamination;
Hydraulically controlling the contaminant plume(s) to isolate the contaminated groundwater; and
Treatment of the contaminated groundwater, either in situ or by collecting, treating and
returning the groundwater to the aquifer.
The decision regarding which category is appropriate for a given situation depends largely on the
following factors:



The compatibility of the remedy to the hydrogeologic setting;
The ability to achieve the remediation goals; and
The cost and time required for implementing the remedy.
A performance-monitoring program must be developed and implemented at groundwater contamination
sites that are being, or have been, remediated. The performance-monitoring plan should be designed to
determine whether the remedies utilized have achieved the established water quality goals.
Groundwater quality goals for any particular groundwater contamination site are established based on
legal and political requirements, use requirements, or the constraints of remediation technology. When
developing a performance-monitoring program, the following factors should be considered:








The extent of the groundwater contamination;
The potential receptors of the contaminated groundwater, such as a stream or water-supply well;
The applicable regulatory requirements;
The hydrogeologic setting;
The sampling frequency and methodology;
Appropriate parameters;
Sample collection, transport, and analysis; and
Quality assurance and quality control procedures.
In general, performance-monitoring programs should include the following four features:
1.
2.
3.
4.
Clearly established compliance locations;
Clearly established compliance limits and schedules;
Early warning and trigger-level limits and locations; and
Appropriate contingency measures to be implemented in the event compliance cannot be
achieved.
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When developing plans for managing a groundwater contamination site, it is important to allocate an
appropriate budget, staff time, field time, and lab time.
Remedial activities, including remedial investigations, are almost always completed by consultants or
contractors who specialize in this type of work. Forest Service personnel may be involved in these
activities either by providing some contractor oversight, or by reviewing and commenting on remedial
activities completed on behalf of another party on National Forest System lands or other property that
the Forest Service has an interest in. This work may include developing a scope of work for a request for
bid, reviewing documents prepared by a consultant or contractor, or overseeing fieldwork performed for
a remedial activity.
While performing remedial activities requires specialized training and experience, understanding the
general concepts that underlie them does not. The preceding information and the associated appendix
provide a basic understanding of what to be aware of when overseeing and reviewing these types of
activities. Specific circumstances may require more specific information.
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Shukla, S., Mostaghimi, S., Shanholt, V.O., Collins, M.C., and Ross, B.B., 2000. A county-level assessment
of ground water contamination by pesticides. Ground Water Monitoring and Review, v. 20, no. 1, p. 104119.
Sloto, R.A.; Crouse, M.Y. 1996. HYSEP: A computer program for streamflow hydrograph separation and
analysis: U.S. Geological Survey Water-Resources Investigations Report 96-4040, Lemoyne, PA, 46p.
Slater, L.; Sandberg, S.; Jankowski, M. 1998, Improvement in the azimuthal EM method – the value of
signal processing: Environmental and Engineering Geophysical Society, Symposium of the Application of
Geophysics to Environmental and Engineering Problems 1998 Conference Proceedings, pp. 177-186.
Stonestrom, D.A.; Constantz, J.E., eds. 2003. Heat as a tool for studying the movement of groundwater
near streams: U.S. Geological Survey Circular 1260, 96 p.
http://pubs.usgs.gov/circ/2003/circ1260/pdf/Circ1260.pdf.
Taylor, C.J.; Alley, W.M. 2001. Ground-water-level monitoring and the importance of long-term waterlevel data: U.S. Geological Survey Circular 1217, 68 p. (http://pubs.usgs.gov/circ/circ1217/)
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U.S. Environmental Protection Agency (EPA). 1993a. A Review of Methods for Assessing Aquifer
Sensitivity and Ground-water vulnerability to Pesticide Contamination. EPA 813-R-93-002. U.S.
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Winter, T.C., Harvey, J.W., Franke, O.L., and Alley, W.M., 1998, Groundwater and surface water: a single
resource: U.S. Geological Survey Circular 1139, 79 p. (http://pubs.usgs.gov/circ/circ1139/)
Zamora, C. 2006. Estimates of vertical flux across the sediment—water interface by direct measurement
and using temperature as a tracer in the Merced River, California: Sacramento, California State University
Sacramento, Master’s thesis, 90 p.
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Appendix 5-A: Assessment of Groundwater Vulnerability using the DRASTIC Methodology
Case Study – Beaverhead-Deerlodge National Forest, Montana
This case study provides an example of how the DRASTIC methodology can be applied to National Forest
System (NFS) lands by using a “table-based” approach. This information in figure 5-A-1 was constructed
by using the following GIS layers: topography, slope, aspect, and geology. Estimates of the depth to
water were made by examining water well information stored in a statewide well database, and
estimates of hydraulic conductivity, recharge, and soil thickness were made by consulting a
hydrogeologist and soil scientist familiar with the area. Index value computations for the various
hydrogeologic settings are presented in table 5-A-1. Geologic units were combined into hydrogeologic
settings based on similar hydrogeologic properties as suggested by Aller et al. (1985).
Figure 5-10—Vulnerability map of a portion of the Beaverhead-Deerlodge National Forest constructed
using the DRASTIC Method. Aquifers are rated from high to low vulnerability based on hydrogeologic
factors.
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Table 5-A1. – DRASTIC computation matrix showing methods for computing index values for various
hydrogeological settings in the Pioneer, Mountains, MT
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Case Study - Pawnee National Grassland, Colorado
This case study provides an example of how the DRASTIC method can be applied to National Forest
System (NFS) lands by using a “model-based” approach. The process is similar to the “table-based”
approach in concept but the analysis and results are developed within a GIS-based model.
A Groundwater Vulnerability Model was developed for an oil and gas leasing environmental impact
statement (EIS) on the Pawnee National Grassland (PNG) for the purpose of assessing the vulnerability of
groundwater to surface spills from oil and gas operations. The DRASTIC methodology developed for the
U.S. EPA by Aller et al. (1987) was the methodology chosen. DRASTIC has often been modified to better
address local issues or better represent a local hydrogeologic setting (Merchant 1994). A sensitivity
analysis was performed to assess the relative influence of each DRASTIC parameter. Those that
influenced the results little were eliminated from the model; others were combined into a single
modified parameter.
The PNG assessment is modified from the original DRASTIC methodology in a number of aspects. The
PNG procedure only uses three mapping layers including soil texture, aquifer media, and depth to water.
DRASTIC uses map layers for saturated hydraulic conductivity and soil media while the PNG procedure
incorporates these two layers into a soil texture layer and the aquifer media layer the incorporates both
variables. The aquifer media map layer also incorporates the impact of the vadose zone in the
assessment of ratings for each geologic unit. Topography is eliminated as a model layer because the area
is relatively flat. Recharge is also eliminated because the precipitation amounts are uniform across the
model area. The following describe the development of each of these model parameters:
1) Depth to groundwater interpolation: Well data was obtained from the Colorado Division of
Water Resources interactive mapping online database. Using the geostatistical wizard, waterlevel data in wells showed a non-normal distribution, therefore a logarithmic transformation was
required. A radial basic function interpolation (spline tension) was generated to display the
depth to groundwater layer. The interpolation created a polygon layer of the depth to
groundwater values.
2) Aquifer Media ratings were based on geologic map unit hydrogeologic properties assigned from
geologic maps ranging in scale from 1:1,000,000 to 1:100,000. Numeric ratings between 1 and 7
were assigned.
3) Soil texture: Soil texture ratings were assigned by using the Natural Resource Conservation
Service (NRCS) – Soil Survey geographic database (taxonomy/particle size attribute). Numeric
ratings between 1 and 5 were assigned. All layers were converted to a raster data format to run
the model. Using GIS (Model Builder and raster calculator), all layers (ratings) were added
generating a groundwater vulnerability value per every 10 meter x 10 meter cell.
The PNG procedure does not adhere to the DRASTIC method for assigning weighting values to predefined
map classes but instead uses equal weights for each of the sensitivity parameter maps based on the lack
of scientific evidence to support any weight assignment. While it may be appropriate to assume that a
relationship exists, it seems more conservative to apply equal weights until further research can better
define the relationships between each of the parameters.
The final sensitivity map (fig. 5-A-2) is created by using a GIS to overlay the three individual parameterrating maps and sum their rating values. The ratings shown on the final sensitivity map reflect the
contribution of each individual parameter. Higher ratings depict areas where the groundwater is
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inherently more sensitive to contamination; lower ratings highlight areas that are less sensitive. Figure 5A-3 shows the map construction methodology using ArcMap GIS.
In general the information displayed in figure 5-A-2 shows the very high vulnerability class is where
Quartetnary alluvium and eolian deposits occur; the high vulnerability class is where the Ogallala
Formation outcrops in the northern part of the east unit and where permeable soils occur on the west
unit. The low vulnerability class only occurs where the low permeability Pierre Shale outcrops.
Figure 5-A-2–Groundwater Vulnerability map of the Pawnee National Grasslands constructed by using
the DRASTIC Method.
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Figure 5-A-3—Methodology for constructing the groundwater vulnerability map of the Pawnee
National Grasslands using the DRASTIC Method and ArcMap GIS.
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References
Aller, L.; Bennett, T.; Lehr, J.H.; Petty, R.J. [et al.]. 1985. DRASTIC- A standardized system for evaluating
ground water pollution potential using hydrogeologic settings. U.S. Environmental Protection Agency,
Ada, OK, EPA/600/2-85/018, 163 p.
Aller, L.; Bennett, T.; Lehr, J.H.; Petty, R.; Hackett, G. [et al.]. 1987. DRASTIC: a standardized system for
evaluating ground water pollution potential using hydrogeologic settings. U.S. Environmental Protection
Agency. EPA/600/S2-87/035. Cincinnati, OH.
Merchant, J.W. 1994. GIS-Based Groundwater Pollution Hazard Assessment: a critical review of the
DRASTIC model. Photogrammetric Engineering and Remote Sensing. 60(9).
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Appendix 5-B: Flow Net Case Study: Wilshire Fen, Fremont-Winema National Forest, Oregon
Figures 5-B-1 and 5-B-2 show both plan view and cross-sectional flow nets constructed to illustrate
horizontal as well as vertical groundwater flow within Wilshire Fen. When viewed in cross-section (fig. 5B-2), groundwater flow paths are shown to emerge from the bedrock and curve towards a seepage face
at the ground surface where groundwater is discharging to the surface. Pronounced upward gradients in
the nested well/piezometer clusters in areas with peat at the surface indicate that the peat layer acts as a
confining layer.
Water-level elevations from shallow wells screened within the peat layer were used to map the
groundwater flow system within Wilshire Fen (fig. 5-B-1). The map shows the configuration of the water
table and groundwater flow direction. Heads from the water table wells were plotted on a plan-view
map and contour lines drawn. Flow lines were then drawn perpendicular to the equipotential lines. The
total flux of groundwater through the fen was estimated by using a flow-net analysis. A sufficient
number of flow lines were drawn so that the resulting rectilinear shapes form approximate squares. The
areas between the flow lines are called stream tubes. The intervals between equipotential lines are
termed “head drops.” Once the flow net is constructed, the amount of groundwater flow through the
area represented by the flow net, under steady-state conditions, can be calculated if the hydraulic
conductivity of the aquifer is known. Several slug tests were performed to estimate the hydraulic
conductivity of the peat and pumice hydrostratigraphic units. A form of the Darcy equation was used to
approximate flow through the fen:
Q
mKHb
n
(1)
where
Q = the quantity of groundwater flowing through the wetland,
m = the number of stream tubes across a flow net,
K = the hydraulic conductivity of the aquifer,
H = total head drop across the area of interest,
b = the effective thickness of the aquifer, and
n = number of equipotential head drops over the area of interest
The K used to calculate the flux of groundwater through the fen is an average for the pumice and peat
layers weighted by thickness, using the following equation:
K wt 
K peat b peat  K pumiceb pumice
b peat  b pumice
(2)
The average thickness of each layer was determined from the auger boring logs.
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N
0
10 m
Contour interval, 20 cm
Well location
Equipotential line
Flow line
Spring box
Note: Groundwater flow direction (groundwater is flowing from the top side of the figure to the bottom) is
indicated by flow lines (blue lines), and lines of equal hydraulic head (equipotential lines) are shown with
dashed lines. The average K for peat is 0.2 m/d, and the average K for pumice is 12.0 m/d. Using equation (2),
the average K for the fen is 7.7 m/d. Using equation (1), where m = 14, n = 9, H = 1.8 m, K = 7.7 m/d, and b = 2.2
m, the total flux (Q) of groundwater through the fen is 47.4 m3/d (8.7 gpm).
Figure 5-B-1—A flow net generated to indicate flow of water through Wilshire Fen, Fremont-Winema
National Forest, Oregon.
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32
N
Drilled well
Groundwater Inventory, Monitoring, and Assessment Technical Guide - DRAFT
S
31
W5
30
W4
29
W6
Relative Elevation, meters
Water table
Pumice
28
Vegetation plot wells
Peat
27
Bedrock
26
Discharge from bedrock
25
0
10
20
30
40
50
60
70
80
Distance, meters
Note: Groundwater flow direction is indicated by flow lines (blue lines), and lines of equal hydraulic head
(equipotential lines) are shown with dashed lines.
Figure 5-B-2–Hydrogeologic cross-section and flow net generated to illustrate flow of water through
Wilshire Fen, Fremont-Winema National Forest, Oregon.
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Appendix 5-C: Determining Environmental Flows and Levels for GDEs
The Forest Service and The Nature Conservancy are developing a method to determine the
environmental flows and levels (EFL) for groundwater-dependent ecosystems (GDEs). The current Forest
Service working definition of environmental flows and levels is:
“Environmental flows and levels describe the quantity, quality, timing and range of variability of
water flows and levels required to sustain or restore freshwater and estuarine ecosystems and
the functions and services they provide. Environmental flows and levels include instream flows,
geomorphic and flood flows, groundwater levels, and lake and wetland levels established for
environmental purposes.”
It is based on the definition of environmental flows in the 2007 Brisbane Declaration to protect
freshwater systems.
However, EFL importantly also includes concepts important for groundwater-dependent ecosystems,
such as static water levels. Environmental flows and levels has previously been termed “Environmental
Water Requirements,” “Ecological Water Requirements,” and “EWR” in Forest Service working
documents. That older terminology will be replaced by environmental flows and levels.
Background
GDEs, by definition, are supported by aquifers. Any activity that lowers or raises the water table,
piezometric surface, or groundwater discharge rate or timing, or alters the groundwater chemistry can
affect the integrity of GDE habitat. Drought, major development, mining within a GDE’s recharge zone,
and leaky septic systems can significantly alter groundwater discharge to a GDE. As stated above, the EFL
approach can be applied to either changes in water quantity or quality. However, at this point the
emphasis of the TNC and Forest Service partnership has been on establishing EFL methodology for
changes in water quantity, which will be reflected in the discussion from this point forward.
Determining quantitative EFL is important for ensuring adequate water to sustain GDEs in cases where
management activities, such as mining, wells for municipal water supply, and water withdrawals for
livestock operations, affect groundwater flow to GDEs. EFL takes into account the ability of GDEs to
adjust to changes in the water regime. Therefore, EFL allows for an acceptable level of change to occur
relative to the existing hydrologic conditions. When use of water resources shifts the hydrologic
conditions below that defined by the EFL, irreversible ecological harm may occur, including impairment
or loss of ecological structure and function.
An EFL establishes the limit to water-level change for an aquifer or water body. It is the limit at which
further withdrawals or flooding would be significantly harmful to the GDE. The establishment of an EFL
for a discharge wetland or phreatophyte community generally defines a limit on water-table drawdown
or increase, or a reduction in discharge feeding the feature from the aquifer. If monitoring indicates the
actual water levels or flows are below the defined EFL, a recovery plan can be developed to reach
acceptable levels.
The EFL seeks to ensure adequate groundwater discharge to protect the most sensitive environmental
values of a GDE. By protecting the most sensitive environmental values, the GDE as a whole is assumed
to be protected. For example, in discharge wetlands where the most sensitive environmental values may
be the presence of rare plants and peat substrates, a recommended EFL would reflect the groundwater
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level that would allow for no loss of rare plants and no net decomposition of peat. In a phreatophyte
community, on the other hand, the most sensitive environmental value may be the presence of particular
species that are closely adapted to specific water table elevations, and so the EFL would reflect those
requirements.
The tradition in hydrology is to separate groundwater and surface water into two disciplines, with little
consideration of the interactions between these two pieces of the hydrological cycle. Much research has
gone toward deriving EFL for rivers, but few studies have addressed water requirements for GDEs (but
see, e.g., Eamus et al. 2006). Environmental flow research grew out of surface water hydrology and
ecology, so it is not surprising that very few studies have considered groundwater processes with respect
to environmental flows and levels. Furthermore, strongly groundwater dependent systems occupy only a
small percentage of the land area (Bedford and Godwin 2003), especially in arid areas (Springer et al.
2008). Groundwater-fed systems, however, are often very biologically diverse and harbor a
disproportionate number of rare and endemic species relative to their total area (Bedford and Godwin
2003, Moore et al. 1989). Furthermore, GDEs are often an important component of a basin’s
headwaters, and the National Forest System includes many headwater areas.
Key Concepts
The most important parameters to which GDEs and dependent species respond are: (a) maintenance of a
water table level within a natural range of variation; (b) groundwater flux to the GDE; (c) head
distributions in the aquifer; and (d) water quality. The low variability in the hydroperiod in springs,
discharge wetlands, and phreatophyte communities results in habitats with specific community
composition and relatively static spatial extent. This differs from EFLs for rivers, where the energy in
environmental flows is essential for the maintenance of geomorphic processes and where the ecosystem
is more dynamic. While the determination of environmental flows for rivers generally is done at a larger
spatial scale, such as an entire watershed, EFLs for GDEs generally occur at much smaller spatial scales
like individual wetlands.
To evaluate an EFL, information must be gathered to quantify (1) the ecological attributes of the GDE,
and (2) hydrological attributes of the groundwater system supporting the GDE’s water requirements.
Ecological data includes the species, ecosystems, and ecological processes that are groundwaterdependent and the nature of their dependency. The hydrologic data includes a description of the water
regime requirements of those groundwater-dependent species, ecosystems, and ecological processes.
These data are combined into a series of groundwater-ecology relationships.
The ecological dataset directly addresses current ecological values and issues, for example: use by
aquatic species, presence of endangered species, tree deaths associated with altered hydrology, and the
presence of invasive species (Davis et al. 2001). This means that EFLs developed this way should be easily
defensible in tradeoff situations. Conceptual models can be developed by using the ecological data to
indicate the consequences of various water regime scenarios for groundwater-dependent species. The
ecological requirements of species can be measured with relatively low-tech methods. Unlike
hydrological elements of key significance (i.e., depth-to-water table), data on the distribution and
abundance of species of interest can be collected rapidly by using standardized sampling techniques.
Hydrologic data collection and interpretation can be quite complex and require a high level of expertise.
Thus a groundwater specialist must be engaged in this part of the project. Furthermore, gathering the
hydrologic data can be resource-intensive, especially if a hydrologic model is necessary.
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As described below in the conceptual framework, each of the datasets informs the other. In some cases,
EFLs can be developed by using only one type of data (ecology or hydrology) by making informed
assumptions about the other. However, this analysis will be less robust than a joint analysis using both
kinds of data. For example, if the management objective is to maintain a healthy wetland plant
community and a strategy is devised to provide a water regime to achieve this objective, monitoring
water levels and vegetation health should provide an indication of whether the managed water regime is
achieved and whether the biological objective is met. But if a hydrologic model is also included in the
analysis, land managers will be able to estimate how much water can be withdrawn from the ecosystem
without passing critical water thresholds. The coupling of hydrological and biological monitoring
programs can therefore provide information regarding species response to changes in water depth.
Detailed procedures for determining environmental flows and levels will be presented in a separate
Environmental Flows and Levels Methods Guide that is currently under development. The following
conceptual framework was developed during a pilot study in the Fremont-Winema National Forest in
Oregon.
Conceptual Framework
The framework described here focuses on assessing environmental flows and levels (EFLs) for
groundwater-dependent wetlands receiving significant groundwater discharge and lacking significant
channelized surface flow (also known as “discharge wetlands”).
Figure 5-C-1 depicts the procedures for applying the framework for quantifying EFLs. It was developed
during an EFL pilot study in the Fremont-Winema National Forest and was adapted from other
conceptual methods (e.g., Eamus et al. 2006). In this method, the hydrogeology of the GDE (left hand
side of diagram) and groundwater-ecology relationships (right hand side of diagram) are measured
simultaneously by using a variety of well-established methods. This process is interactive: results from
each step are used to inform subsequent steps; results from the hydrogeologic analyses are used to
inform the groundwater-ecology relationships, and vice versa. All of the hydrogeologic approaches used
here are described elsewhere in this Technical Guide; therefore, they will not be described here. Many of
the ecological methods are found in the GDE Level II Inventory Field Guide (USDA Forest Service 2012).
Evaluation Procedures
Table 5-C-1 provides a detailed description of how to apply the evaluation procedures described in figure
5-C-1.
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Figure 5-C-1—Framework for assessing environmental flows and levels of GDEs.
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Table 5-C-1–EF/L evaluation procedures.
Step 1
Determine if the conceptual hydrogeologic model indicates the proposed management activity may
affect groundwater-ecology relationships?
(HG1) Develop a conceptual hydrogeologic model of the GDE.
The conceptual model can be based on field reconnaissance, best professional judgment, information
from the scientific literature, and field data if available. A conceptual model is often displayed as a
cross section through the landscape showing groundwater flow paths from recharge area to a GDE,
and the location of the management activity that will affect those flow paths (see section 2.4).
Below are some elements of a hydrogeologic conceptual model:
-
Geologic materials, structural geology, and geomorphology that affect where groundwater
may be found and the direction it moves.
Location and area of groundwater recharge
Groundwater flow paths delivering groundwater to the GDE
Fluctuation in position of the water table throughout the year (i.e., the wetland hydroperiod)
Soil stratigraphy within the wetland (i.e., the local unconfined aquifer), indicating any
confining layers
Vertical and horizontal head gradients within the wetland.
Geomorphic history of wetland development, for example from recent glacial or pyroclastic
events (last 5,000–20,000 years).
Local and/or regional climate
The conceptual model must also explicitly include a quantitative description of the ways in which
groundwater flow to the GDE may be affected by the proposed management activity.
(E1) Identify groundwater-ecology relationships for key groundwater-dependent species & ecosystem functions
or processes.
The identification of relationships is based on field reconnaissance, best professional judgment, and
information from the scientific literature. In this step, species and processes that are partially or
entirely dependent on groundwater are identified. This is done by answering the following questions
(Eamus et al. 2006):
-
Which populations or species of an ecosystem are groundwater dependent?
What ecosystem processes are groundwater dependent?
For those populations, species, or processes determined to be groundwater dependent, what
is the degree of dependency?
What attributes of groundwater (level, flux, hydraulic head, quality) are important to the
dependent populations/species?
Does the hydrogeologic conceptual model indicate that the proposed management activity may impact the
groundwater-ecology relationships?
NO
No further action required
YES
Go to STEP 2
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Step 2
Determine if the water budget indicates groundwater alteration is of the same or greater order of
magnitude as groundwater inflow to the GDE and/or directly affects the groundwater-ecology
relationships?
((HG2) Develop a water budget from field data and/or available information from the scientific literature.
A generalized wetland water budget is expressed as follows:
Total Inflows (Precip + GWin + SWin) – Total Outflows (ET + GWout + SWout) = Change in Storage
Each of the terms can be measured in a number of ways. Climate parameters (precipitation, ET) can
be obtained from local weather stations or measured on-site (described in section 2.4.4).
Groundwater inflow and outflow are measured by using a flow-net analysis (described in subsection
5.2.3). Surface water inflow and outflow can be measured by using weirs (see Forest Service
Groundwater Technical Note – Measurement of Discharge at Springs and Wetlands).
Water budget components are then compared to the expected groundwater alteration from the
proposed management activity. If the expected change in groundwater discharge to the wetland is of
the same order of magnitude (or greater) as other components of the water budget (particularly
groundwater inflow), then the management activity could alter the groundwater-ecology
relationships.
(E2) Quantify groundwater-ecology relationships for key species and ecosystem processes.
The quantification of the relationships is done by using a combination of field data and information
from the scientific literature. Quantitative groundwater-ecology relationships are developed for
select groundwater-dependent species and processes. Examples of groundwater-ecology
relationships include:
-
A specific, narrow range in depth-to-water table required by obligate wetland plant species
The persistence of open cool water pools fed by groundwater that are used by amphibians for
nesting or over-wintering habitat
Discharge of calcareous groundwater that supplies calcium to invertebrates
Persistently high water table that promotes peat accretion by preventing organic matter
oxidation
Supply of base cations in groundwater that bind phosphorus and impose nutrient limitation on
the ecosystem
Field data and data from the published literature are used to develop quantitative relationships, for
example between specific plant species and their minimum tolerated depth to the water table, or
between an amphibian and its maximum water temperature.
Does the water budget indicate groundwater alteration is of the same or greater order of magnitude as
groundwater inflow to the GDE and/or directly impacts the groundwater-ecology relationships?
NO:
Monitor to confirm no effect
YES:
Go to STEP 3
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STEP 3
Do analytical solutions indicate that groundwater alteration (flooding or drawdown) may extend
beyond the species or ecosystem process thresholds?
(HG3) Use analytical approach to quantify groundwater-level change associated with groundwater management
activity in the aquifer discharging to the GDE.
In the case where the management activity involves withdrawals from the GDE’s source aquifer, an
analytical approach can be used to quantify the amount of drawdown from extraction points. The
Theis equation can be used for a confined aquifer and the Neuman solution for an unconfined aquifer
(see section 5.2.1). Another approach is to perform an aquifer test (also known as an area of
influence test) to quantify drawdown (see sections 5.2.1).
Do analytical solutions indicate that groundwater alteration (flooding or drawdown) may extend beyond the
species or ecosystem process thresholds?
YES
And results from the analytical solutions are adequate to meet management needs: make a
management decision
YES
But results from analytical solutions are not adequate to make a management decision: go to STEP 4
NO
But analytical model results are too uncertain relative to the importance of the species or ecosystem
processes: go to STEP 4
NO
And analytical model results are adequate to conclude limited effect of management activity: monitor
to confirm no thresholds are exceeded.
STEP 4
Do analytical solutions indicate that groundwater alteration (flooding or drawdown) may extend
beyond the species or ecosystem process thresholds?
(HG4) Develop a numerical model to increase the accuracy and/or confidence of relationships between
groundwater management and GDE ecological thresholds.
For applications with complex hydrogeologic settings or where more precise estimates of effects are
imperative, a numerical model (e.g., MODFLOW) that incorporates aquifer geometry and accurately
represents the flow system is the tool of choice (see Forest Service Technical Note on Modeling of
Groundwater Systems).
(E3/E4) Define area of influence and quantitative groundwater thresholds for key species and processes
identified in the previous step. Quantify habitat change (loss or increase) by combining groundwater-ecology
thresholds and expected groundwater-level change.
Based on results from the hydrogeologic models, the locations and extent of drawdown below the
thresholds established in step E2 can be estimated. For example, the ecological data are combined
with the cone of depression curve from the numerical model to determine the wetland area affected
and potential losses of species or processes in that area.
Outcome Make a management decision about acceptable level of species or habitat loss associated with the
management activity.
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References
Bedford, B. L.; K. S. Godwin. 2003. Fens of the United States: Distribution, characteristics, and scientific
connection versus legal isolation. Wetlands 23: 608-629.
Davis, J.A., R.H. Froend, D.P. Hamilton, P. Horwitz, A.J. McComb and C.E. Oldham. 2001. Environmental
Water Requirements to Maintain Wetlands of National and International Importance, Environmental
Flows Initiative Technical Report Number 1, Commonwealth of Australia, Canberra.
Eamus, D., R. Froend, R. Loomes, G. Hose, and B. Murray. 2006. A functional methodology for
determining the groundwater regime needed to maintain the health of groundwater-dependent
vegetation. Australian Journal of Botany 54: 97-114.
Moore, D.R.J.; Keddy, P.A.; Gaudet, C.L.; Wisheu, I.C. [et al.]. 1989. Conservation of wetlands: do infertile
wetlands deserve a higher priority? Biological Conservation 47: 203–217.
Springer, A. E., L. E. Stevens, D. E. Anderson, R. A. Parnell, D. K. Kreamer, L. A. Levin, and S. P. Flora. 2008.
A comprehensive springs classification system: Integrating geomorphic, hydrogeochemical, and
ecological criteria. In: Stevens, L.E.; Meretsky, V.J., eds. Aridland springs in North America: Ecology and
conservation. University of Arizona Press and The Arizona-Sonora Desert Museum, Tucson, AZ. pp. 49-75.
U.S. Department of Agriculture (USDA), Forest Service. 2012. Groundwater-dependent ecosystems: level
II inventory field guide inventory methods for project design and analysis. Gen. Tech. Report WO-86b.
U.S. Department of Agriculture, Forest Service. 131 p.
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Appendix 5-D: Principles of Solute Fate and Transport
Several mechanisms influence the spread of a solute or contaminant in a groundwater flow field.
Dispersion and differences in density and viscosity may accelerate solute movement, while various
retardation processes slow the rate of movement compared to that predicted by simple advective
transport. Fetter (1999) presents a comprehensive discussion of contaminant hydrogeology. The major
mechanisms of solute fate and transport in the subsurface are summarized below.
Groundwater Advection
In its natural state, groundwater generally moves very slowly, but continuously. Advection is the process
by which dissolved solutes are carried along with the flowing groundwater. Advecting solutes are
traveling at the same rate as the average linear velocity of the groundwater if the solutes are not subject
to any sort of reactions with the porous media. These movement patterns are generally governed by the
space occupied by the mass of the liquid flowing through the media and the rate(s) of flow encountered
within these spaces. The hydraulic conductivity of a geological formation depends on a variety of
physical factors within the formation, such as effective porosity; particle size, arrangement, distribution,
and shape; and secondary features, such as fracturing and dissolution. Generally, hydraulic conductivity
values for unconsolidated porous materials vary with particle size. Fine-grained clayey materials exhibit
lower values than those of coarse-grained, sandy materials. Figure 5-D-1 displays the relative hydraulic
conductivity of different geologic materials.
Figure 5-D-1–Relative hydraulic conductivity of different geologic materials (Heath 1983).
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Effective Porosity
Effective porosity is the portion of the total porosity that is interconnected and contains flowing water.
Depending on the nature of the hydrogeologic medium, effective porosity may range from little to
essentially all of the total porosity. Effective porosity is basically an estimated parameter, because the
actual volume of interconnected pore spaces in most porous media is unknown. Effective porosity is
estimated from the ratios of saturated and dry porous materials. In coarse-grained materials that drain
freely, effective porosity is essentially equal to total porosity and is generally defined as the ratio of the
volume of water that drains by gravity to the total volume of saturated porous material. Effective
porosity governs calculations of how rapidly water and contaminants move through the subsurface via
advection. Furthermore, the amount of effective porosity affects the ease of removing a contaminant
from a groundwater system. Low effective porosity means more pore spaces where contaminants can
collect and be difficult to remediate.
Diffusion
Diffusion is the process by which a solute moves from areas of higher chemical potential (high
concentration) to areas of lower chemical potentials (low concentration). This process is also known as
molecular diffusion. Diffusion occurs in the absence of any bulk hydraulic movement of the solution; that
is, solutes diffuse (spread) regardless of whether the bulk mass of liquid is static or moving through the
hydrogeologic medium.
Dispersion
Groundwater molecules move at different rates depending on position within the aquifer and within the
interconnected pores in the aquifer; some are faster than the average linear velocity while some are
slower (Mills et al. 1985). There are three causes for this phenomenon: (1) friction on pore walls, (2)
variations in pore sizes, and (3) variations in path length.
As groundwater moves through the pores, it will move faster at the center of the pore than along the
walls because of friction. In cases where pore size varies, groundwater will move through larger pores
faster. Groundwater molecules have tortuous flow paths (called tortuosity) and some will travel longer
pathways than others. Because the invading solute-containing water is not all moving at the same rate,
mixing occurs along the flow path. This mixing is termed mechanical dispersion. The mixing that occurs
along the direction of fluid flow is termed longitudinal dispersion; mixing that occurs normal to the
direction of fluid flow is termed transverse dispersion. Because molecular diffusion cannot be readily
separated from mechanical dispersion in flowing groundwater, the two are combined into a parameter
called hydrodynamic dispersion. Because of hydrodynamic dispersion, the concentration of a solute will
decrease over distance along the flow path. Generally speaking, the solute will spread more in the
direction of groundwater flow than in the direction normal to the groundwater flow because longitudinal
dispersivity is typically substantially higher than transverse dispersivity. Because dispersion anisotropy is
often difficult to measure, a default value of a factor of 10 for higher longitudinal relative to transverse is
often used. In fact, most solute plumes tend to be long and thin.
Quantifying dispersion may be important in fate assessment, because contaminants can move more
rapidly through an aquifer by this process than by simple plug flow (uniform movement of water through
an aquifer with a vertical front). In other words, physical conditions, such as the presence of more
permeable zones where water can move more quickly, and chemical processes, such as movement by
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molecular diffusion of dissolved species at greater velocities than the water, result in more rapid
contaminant movement than would be predicted by groundwater equations for physical flow, which
assume average values for hydraulic conductivity.
Chemical Reactions
There are many types of chemical reactions that can be important in groundwater systems including ion
exchange, sorption, biotransformation and biodegradation, oxidation-reduction, acid-base, dissolution,
precipitation, and complexation. A detailed discussion of the chemistry of natural waters is beyond the
scope of this document. Some additional information is provided below on the most common reactions
in contaminant transport. More information on aqueous geochemistry is available in Chapelle (2000),
Drever (1997), Langmuir (1997), Stumm and Morgan (1996), Morel and Hering (1993), and Hem (1985).
Ion Exchange - Ion exchange processes exert an important influence on retarding the movement
of chemical constituents in groundwater. In groundwater systems, ion exchange occurs when
ions in solution displace ions associated with geological materials. This process removes
constituents from the groundwater and releases others to the flow system. One major
consideration in ion exchange is that the exchange capacity of a given geological material is
limited. A measure of this capacity is quantified in a term called “ion exchange capacity” and is
defined as the amount of exchangeable ions in milliequivalents per 100 grams of solids at pH 7.
Typically, clay materials such as montmorillonite exhibit greater cation (positively charged ions)
exchange capacities than other minerals such as quartz, which is the primary component of sand.
This difference is attributable to the often much greater surface area of clays than other
minerals.
Anionic (negatively charged ions) exchange in aquifer systems is not as well understood as
cationic exchange. Anions such as sulfate, chloride, and nitrate are not retarded significantly by
anion exchange because most mineral surfaces in natural water systems are negatively charged.
Chloride ions may often be regarded as conservative or non-interacting ions, which move largely
unretarded with the advective velocity of the groundwater mass.
It is important to recognize that the ion-exchange capacity of a geological material may retard
contaminant movement from a waste or other source for years or even decades. However, if the
source continues to supply a strongly ionic leachate, it is possible to exceed the exchange
capacity of the geological material, eventually allowing unretarded transport of the contaminant.
Changes in environmental conditions or groundwater solution composition can also cause the
release of constituents formerly bound to the geological materials.
Sorption - Sorption involves the surface interaction of a dissolved constituent with a solid
material. More specifically, the term encompasses both adsorption-desorption reactions and
absorption. The former refers to a buildup or a release of a constituent on the surface of a solid
as a result of molecular-level interactions, while the latter implies a more or less uniform
penetration of the solid by a solute. In many environmental settings, this distinction may serve
little purpose as there is seldom information about the specific nature of the interaction. A
number of factors control the interaction of a contaminant and the surfaces of soil or aquifer
materials. These include chemical and physical characteristics of the constituent, composition of
the surface of the solid, and the fluid media encompassing both. By gaining an understanding of
these factors, logical conclusions can often be drawn about the impact of sorption on the
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movement and distribution of constituents in the subsurface. The failure to take sorption into
account can result in a significant underestimation of the amount of a contaminant at a site, the
time required for it to move from one point to another, and the cost and time involved for
remediation. The properties of a contaminant that have a profound effect on its sorptive
behavior include water solubility, polar/ionic character, octanol/water partition coefficient,
acid/base chemistry, and oxidation/reduction chemistry.
Precipitation/Dissolution -The migration of dissolved metals in groundwater can be significantly
influenced by the thermodynamic relationships between species in solution and aquifer solid
phases. Attenuation of trace metals in groundwater can be caused by precipitation or coprecipitation of mineral phases. Determination of whether or not a metal species is soluble, or is
likely to remain in solution at concentrations of concern is a key consideration when
contemplating the metals in groundwater.
Complexation - Complexation is the sorption of a dissolved ion to a mineral surface to form a
surface complex. Common hydrous oxides of iron, aluminum, manganese and silicon are the
dominant sorbents in nature because they are finely dispersed and coat other particles and are
electrostatically active. Sorption of inorganic ions onto these oxide coatings or other surfaces is
dependent on pH, ionic strength and competition from other ions in solution.
Biotransformation and Biodegredation -The transformation of both organic and inorganic
chemicals by microorganisms readily occurs in many subsurface environments, including landfills
and septic systems. Microbial processes may be a major factor in the transformation of both
natural and anthropogenic organic materials present in groundwater. These transformations can
result in the formation of CO2, CH4, H2, H2S, N2, NH3, and NO gases, among other compounds.
Under the appropriate circumstances, pollutants can be completely degraded to harmless
products; whereas, under other circumstances, they can be transformed to new substances that
are more mobile or more toxic than the original contaminant. Quantitative predictions of the
fate of biologically reactive substances are primitive in comparison with predictions for other
processes that affect pollutant transport and fate.
Biotransformations in groundwater were previously thought to mimic those known to occur in
surface water bodies, but detailed fieldwork has demonstrated the fallacy of this assumption.
With the relatively long residence times and stable environments in groundwater systems, watertable aquifers are now known to harbor appreciable numbers of metabolically active
microorganisms distinctly different from those in surface waters. These groundwater organisms
frequently can degrade organic contaminants in the subsurface that would not be degraded on
the surface. Thus, it is necessary to consider biotransformation as a process that affects
pollutant transport and fate.
Radionuclides in Groundwater
Most groundwater sources have very low levels of radioactive contaminants (radionuclides). Natural
radionuclides in groundwater are referred to as primordial radionuclides and have exceptionally long
half-lives. These very low levels are not considered to be a public health concern. Of the small
percentage of drinking water systems with radioactive contaminant levels high enough to be of concern,
most of the radioactivity is naturally occurring. Certain rock types have naturally occurring trace amounts
of “mildly radioactive” elements (radioactive elements with very long half-lives) that serve as the
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“parent” of other radioactive contaminants (“daughter products”). These radioactive contaminants,
depending on their chemical properties, may accumulate in drinking water sources at levels of concern.
The “parent radionuclide” often behaves very differently from the “daughter” radionuclide in the
environment. Because of this, parent and daughter radionuclides may have very different drinking-water
occurrence patterns. For example, groundwater with high radium levels tends to have low uranium
levels and vice versa, even though uranium-238 is the parent of radium-226.
Most parts of the United States have very low “average radionuclide occurrence” in groundwater
sources. However, some parts of the country have, on average, elevated levels of particular
radionuclides compared to the national average. For example, areas in the Midwest have significantly
higher average combined radium-226/radium-228 levels, while some Western States have elevated
average uranium levels compared to the national average. There are other radionuclides that have been
known to occur in a small number of drinking water supplies, but their occurrence is thought to be rare
compared to radium-226, radium-228, and uranium. Uranium is present in groundwater in amounts
ranging from 0.05 parts per billion (ppb) to 10 ppb (the median is about 1.5 ppb).
Radon-222, another naturally occurring radionuclide of concern in groundwater, has a half-life of 3.8 days
and is produced continuously in aquifers by the disintegration of the parent nuclide radium-226.
Radioactivity in groundwater is normally measured in the units of microcuries per milliliter (μCi/ml).
Normal groundwater contains from less than 1 x 10-7 μCi/ml to about 3 x 10-5 μCi/ml radon; the median
is about 2 x 10-6 μCi/ml.
Groundwater has been contaminated with radionuclides beyond background levels through the mining,
refinement, and processing of uranium ore; releases of oil and gas well-produced formation waters or
brines; improper disposal of commercial, industrial, and laboratory waste materials; initial production of
nuclear fuels and explosives; reprocessing used reactor elements; discharge of cooling water that has
been exposed to nuclear activation; escape of volatile material from evaporation and burning; the
release of radionuclides used in science and medicine; and disposal of wastes from reactor operations
and fuel reprocessing.
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References
Chapelle, F. H. 2000. Groundwater microbiology and geochemistry. New York: John Wiley and Sons. 496
p.
Drever, J.I. 1997. Geochemistry of natural waters, the surface and groundwater environments. Upper
Saddle River, NJ: Prentice Hall. 436 p.
Fetter, C.W. 1999. Contaminant hydrogeology. Englewood Cliffs, NJ: Prentice Hall. 500 p.
Heath, R.C. 1983. Basic Ground-water hydrology. U.S. Geological Survey Water-Supply
Paper 2220. 84p.
Hem J. D., 1985, Study and Interpretation of the Chemical Characteristics of Natural Water, U.S.
Geological Survey, Water Supply Paper 2254, 263p.
Langmuir, D. 1997. Aqueous environmental geochemistry. Upper Saddle River, NJ: Prentice Hall. 600 p.
Mills, W.B.; Borcella, B.B.; Ungs, M.J.; Gherini, S.A.; Summers, K.V.; Lingsung, M.; Rupp, G.L.; Bowie, G.L.;
Haith, D.A. [et al.]. 1985. Water Quality Assessment: A screening procedure for toxic and conventional
pollutants in surface and groundwater, Parts 1 and 2. Report EPA 600/6-85/002a,b. Athens, GA:
Environmental Research Laboratory, U.S. Environmental Protection Agency.
Morel, F.M.M.; Hering, J. G. 1993. Principles and applications of aquatic chemistry. New York: WileyInterscience, 608 p.
Stumm, W.; Morgan, J.J. 1996. Aquatic chemistry, chemical equilibria and rates in natural waters. 3rd ed.
New York: John Wiley & Sons, Inc. 1022 p.
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