United States Environmental Monitoring

Environmental Protection

EPA/600/X-89/05 0

Systems Laboratory

Agency

March 1989

P O Box 93478

Las Vegas NV 89193-3478

Research and Development

Ground-Water Monitoring

in Karst Terranes

Recommended Protocols &

Implicit Assumptions
EPA 600/X-89/050

February 1989

GROUND-WATER MONITORING IN KARST TERRANES:

RECOMMENDED PROTOCOLS AND IMPLICIT ASSUMPTIONS

by

James F. Quinlan

National Park Service

Mammoth Cave, Kentucky 42259

IAG No. DW 14932604-01-0

Project Officer

Eric N. Koglin

Environmental Monitoring Systems Laboratory

Las Vegas, Nevada 89193-3478

This study was conducted

in cooperation with

National Park Service

U.S. ENVIRONMENTAL PROTECTION AGENCY

ENVIRONMENTAL MONITORING SYSTEMS LABORATORY

LAS VEGAS, NEVADA 89193-3478
NOTICE

This document is intended for internal Agency use only.

Mention of trade names or commercial products does not constitute

endorsement or recommendation for use.

i
CONTENTS

Foreword. . . . . . . . . . . . . . . . . . . . . . . . . .

Abstract. . . . . . . . . . . . . . . . . . . . . . . . . .

Figures . . . . . . . . . . . . . . . . . . . . . . . . . .

Acknowledgements. . . . . . . . . . . . . . . . . . . . . .

1.

2.

3.

4.

5.

6.

Introduction . . . . . . . . . . . . . . . . . . . .

Objectives of this document . . . . . . . . . .

Characteristics of ground-water flow in

karst terranes. . . . . . . . . . . . . . . .

Design of Monitoring Systems in Karst Terranes . . .

Where to monitor for pollutants . . . . . . . .

Where to monitor for background water quality .

Discussion of where to monitor for pollutants

and background. . . . . . . . . . . . . . . .

Relevance of off-facility monitoring to RCRA

regulations . . . . . . . . . . . . . . . . .

When to monitor for pollutants and background .

Applicability of Recommended Traced-Spring,

-Cave-Stream, and -Well Monitoring Strategy . . . . .

Regulatory Aspects of Dye-Injection into Wells . . .

How to Determine Where, How, and When to Monitor

Ground Water in Karst Terranes Reliably and

Economically . . . . . . . . . . . . . . . . . . . .

Introduction. . . . . . . . . . . . . . . . . .

Dye-tracing . . . . . . . . . . . . . . . . . .

Selection of dye-injection sites . . . . .

Selection of dye-monitoring sites . . . . .

Dye-tracing methods . . . . . . . . . . . .

Dye-test analysis . . . . . . . . . . . . .

Surveying of cave rivers . . . . . . . . . . . .

Mapping of potentiometric surface . . . . . . .

Implicit Assumptions of Recommended Traced-Spring,

-Cave-Stream, and -Well Monitoring Strategy, with

Examples of Exceptions . . . . . . . . . . . . . . .

Exceptions to assumption #1: Nonpoint

discharge . . . . . . . . . . . . . . . . . .

Exceptions to assumption #2: Nonlocal flow . .

Exceptions to assumption #3: Slow movement

in local-flow system . . . . . . . . . . . . .

Exceptions to assumption #4: Fuzzy and

overlapping basin boundaries . . . . . . . . .

Exceptions to assumption #5: Noncontiguous

ground-water basins . . . . . . . . . . . . .

Exceptions to assumption #6: Springs that

derive some of their flow from more than

one basin . . . . . . . . . . . . . . . . . .

ii

iv

vi

viii

ix

1

1

3

6

6

9

11

18

20

25

26

28

28

28

28

31

32

35

38

39

40

42

43

43

44

46

47
Exception to assumption #7: Diversion of

ground water to the surface . . . . . . .

Exception to assumption #8: Diversion of

surface water to a different ground-water

basin . . . . . . . . . . . . . . . . . .

Assumption #10, the most insidious . . . . .

Introduction . . . . . . . . . . . . .

Procedural causes of falsely negative

tracer test results . . . . . . . . .

Discussion . . . . . . . . . . . . . .

7. Sisyphean Lament . . . . . . . . . . . . . . . .

8. Checklist for Design of Monitoring Systems in

Karst Terranes . . . . . . . . . . . . . . . . .

9. Summary. . . . . . . . . . . . . . . . . . . . .

10. Epilegomenon . . . . . . . . . . . . . . . . . .

. . 49

. . 49

. . 50

. . 50

. . 50

. . 57

. . 59

. . 59

. . 60

. . 62

References. . . . . . . . . . . . . . . . . . . . . . . . . 64

iii
FOREWORD

This document was written for four major reasons: 1) The

hydrology of karst terranes is significantly different from that

of terranes characterized by granular and fractured rocks--flow

velocities in karst may be several orders of magnitude higher

than in other ground-water settings; Darcy’s Law describing flow

is rarely applicable; 2) For monitoring to be relevant and reliable in karst terranes, monitoring procedures must be radically

different from those in non-karst terranes; 3) There is a need

for a practical guide that tells engineers, geologists, hydrologists, and regulators what the monitoring problems are in karst

terranes and how to solve them; and 4) Create awareness of the

state-of-the-art in monitoring in karst terranes--and provoke

thought and discussion about the subject and its implications for

ground-water protection strategy. Some of the conclusions may be

controversial, but I believe it is better to be aware of problems

and to try to solve them than to ignore them.

Approximately 20 percent of the United States (and 40 percent

of the country east of the Mississippi River) is underlain by

various types of karst aquifers. The vulnerability of these

aquifers to contamination--and the consequent threats to public

health and safety, as well as to the environment--make it

imperative that monitoring of these aquifers be reliable. This

document tells how to achieve such needed reliability.

This is a synthesis of the results of research and practical

experience by the author and others. My research has been sponsored by the National Park Service, the U.S. Environmental

Protection Agency, and the Kentucky Water Resources Research

Institute. Much of my research has been done in collaboration

with Dr. Ralph Ewers of Eastern Kentucky University.

Reliability, however, is not obtained easily or automatically. Karst is complex. Although general principles are applicable almost universally, reduction of them to absolute rules is

difficult. Therefore, rules to guide monitoring should be based

on goals or performance, not procedure. Stated generally, there

are two goals of monitoring:

1. Acquire and correctly interpret data relevant to the hydrology of a site.

2. Logically monitor water-quality parameters that are indicative of contaminant-related changes in the hydrology.

iv
There are two problems with the attainment of these goals:

It is easy to assume that one has achieved them, and the design of

a monitoring system is a research project, not the automatic

installation of a few wells downgradient and upgradient from a

site followed by collection of water samples on a regular

schedule. This document discusses reliable procedures for

attaining these goals. Following internal EPA review of this

document, the report will be upgraded to a Project Report, for

general release and use buy three groups of people:

1. Administrators who must evaluate existing or proposed

networks for monitoring water quality in karst

terranes, but have minimal experience in karst

hydrogeology;

2. Consultants and others who must design monitoring

networks but may or may not have extensive experience

in karst terranes; and

3. The well-experienced tracer who is already familiar with

the hydrology and geomorphology of karst terranes but has

minimal familiarity with monitoring problems.

Accordingly, some sections of this document may be of

greater or lesser interest to one group than to another.

This document is a highly revised, nearly threefold expansion

of a review by Quinlan (1989a). It includes a summary of six

other papers published in the proceeding of various conferences

(Quinlan, 1988a; Quinlan and Alexander (1987); Quinlan, Aley and

Schindel, 1988; Quinlan and Ewers (1984, 1985); and Quinlan,

Ewers, and Field, 1988), but most of this document is in no other

publication or report. It is a unique synthesis and summary of

the experience of the writer and his peers who are gratefully

acknowledged.

v
GROUNDWATER MONITORING IN KARST TERRANES:

RECOMMENDED PROTOCOLS AND IMPLICIT ASSUMPTIONS

by: James F. Quinlan

National Park Service

Mammoth Cave, Kentucky 42259

ABSTRACT

Reliable monitoring of ground-water quality in any terrane is

difficult: There are many ways in which violation of sound principles of monitoring-network design and good sampling protocol

make it easy to acquire data that are not representative of the

water or pollutants within an aquifer. In karst terranes it is

especially likely that irrelevant data, inadvertently misrepresenting aquifer conditions, will be obtained.

The special problems of monitoring ground water in most karst

terranes can be grouped into four major categories of problems

that are rarely as significant in other terranes. These categories

1.

2.

3.

are:

Where t o monitor for pollutants: At springs, cave

streams, and wells shown by tracing to include drainage

from a facility to be monitored--rather than at wells to

which traces have not been run but which were selected

because of convenient downgradient location. Wells on

fracture traces and fracture-trace intersections and wells

located randomly can be successfully used for monitoring,

but only if traced positively from the facility to them.

Often, the monitoring can only be done several kilometers

away from the facility.

Where t o monitor for background water quality: At

springs, cave streams, and wells in which the waters are

geochemically similar to those to be monitored for pollutants but which are shown by tracing not to include drainage from the facility--rather than at wells selected

because of convenient location upgradient from it. This,

too, may have to be done several-kilometers away from the

facility.

When to monitor: Before, during, and after storms or

vi
meltwater events and also at known base-flow conditions--

rather than regularly with weekly, monthly, quarterly,

semi-annual, or annual frequency.

4. How to reliably and economically determine the answers to

problems 1, 2, and 3: Reliable monitoring of ground water

in karst terranes can be done, but it is not cheap or

easy.

These problems exist because many of the assumptions made for

monitoring ground-water flow in granular media are not valid in

karst terranes. Many axiomatic, implicit assumptions are made

about flow-systems by those who use various tracers and who may

or may not use the traced-spring, -cave-stream, and -well monitoring strategy discussed herein. Many of these assumptions are

valid only about 95 percent of the time. Ten of them are reviewed and exceptions to eight are discussed. Among the ten

reviewed, the most insidious is that the tracing tests have been

well designed, properly executed, and correctly interpreted--and

that they are capable of yielding an unambiguous interpretation.

Additionally, there are at least thirteen major ways to inadvertently obtain falsely negative results in tracer tests. These

too are cited and discussed.

The recommended monitoring strategy is widely applicable in

most U.S. karst terranes. Most of them are characterized by

local recharge that is discharged at springs. The strategy is

not applicable in karst terranes characterized only by recharge,

by diffuse discharge into sediment, or by sublacustrine or submarine discharge.

Tracing agents are fundamental and necessary tools for

studies, model confirmation and calibration, and predictions of

movement of ground water and pollutants. Their legitimate use

needs legal recognition as being benign and harmless in the

concentrations generally employed--rather than regulation as contaminating substances.

Many of the passionate conflicts between designers of monitoring systems and interpreters of their data are philosophical.

Most of these conflicts are between individuals with different

perceptions of reality.

A checklist can be used to guide the general sequence of

operations necessary in the design of a monitoring system for

ground water in karst terranes. The checklist given, however, is

no substitute for acquisition of an understanding of the concepts

espoused in this document and application of them.

Although some state agencies now require ground-water monitoring at springs, official Federal regulatory recognition of

this necessity is needed.

vii
FIGURES

Number Page

1 Diagram showing the relation between four major

types of aquifers and the dominant porosity

geometry in each. . . . . . . . . . . . . . . . . . .

2 Block diagram showing relations between hydrology,

landforms, and stratigraphy of a hypothetical

karst terrane developed in gently-dipping rocks

in a low-relief landscape . . . . . . . . . . . . . .

3 Map showing hydrology of the Turnhole Spring groundwater basin, the major basin draining into Mammoth

Cave National Park, Kentucky . . . . . . . . . . . . .

4 Map showing distributary springs along Green River

in the Mammoth Cave area, Kentucky . . . . . . . . .

5 Map showing ground-water flow routes, potentiometric

surface, and pollutant dispersal in the vicinity

of the towns of Horse Cave and Cave City, Kentucky . .

6 Map showing complex radial flow of ground water near

a proposed landfill in flat-lying residuum-mantled

limestone in the Ozarks of Arkansas . . . . . . . . .

7 Graphs of water quality and discharge during a 1984

storm in a karst aquifer characterized by conduit

flow, Big Spring ground-water basin, Iowa . . . . . .

8 Diagram showing tetrahedral continuum between four

major types of aquifers, dominant porosity geometry

in each type of aquifer, applicability of the tracedspring, -cave-stream, and -well monitoring strategy in

each type of aquifer, and applicability of the traditional monitoring technique (using randomly located

wells) in each type of aquifer . . . . . . . . . . . .

5

7

12

13

14

16

22

25

viii
ACKNOWLEDGEMENTS

Ideas expressed here have been critically reviewed by E.

Calvin Alexander, Jr., Thomas J. Aley, John V. Brahana, Ralph O.

Ewers, Malcolm S. Field, Eric N. Koglin, Arthur L. Lange, Donald

S. Mull, Joseph A. Ray, Geary A. Schindel, Christopher C. Smart,

and A. Richard Smith. Each has critically reviewed drafts of the

manuscript and made numerous valuable suggestions and improvements that significantly enhanced the clarity and content. The

reductional talents of A. Richard Smith and Eric N. Koglin

repeatedly improved organization and coherence. I am greatly

indebted to each of them. Much of the field work in Kentucky

that initially stimulated many of my conclusions was meticulously

done by Joseph A. Ray and more than 30 other highly motivated

National Park Service research assistants, many of whom were

near-fanatical in their dedication and heroic in their efforts.

I am greatly indebted to them also. Some of my conclusions are a

result of work done as a consultant for various clients in

several states and Puerto Rico.

ix
INTRODUCTION

OBJECTIVES OF THIS DOCUMENT

The hydrology of karst terranes is significantly different

from that of non-karst terranes. Accordingly, the monitoring

techniques necessary for reliable, accurate assessment of their

ground-water quality are significantly different. This document discusses these techniques, recommends ways to solve the

special problems posed by monitoring in karst terranes, calls

attention to possible regulatory problems, discusses the assumptions made when monitoring, and makes recommendations throughout

the text.

Recent comprehensive reviews of the hydrology and geomorphology of karst terranes are the well-illustrated texts by White

(1988) and Ford and Williams (1989). These are the most readable

introductions to the subject. A comprehensive analysis of the

physics, chemistry, and geology of karst aquifers has been

written by Dreybrodt (1988). None of these books is specifically

concerned with environmental problems, but they are useful,

experience-based syntheses of current knowledge. Present and

future trends in karst hydrologic research have been briefly

reviewed by Atkinson (1985). A useful guide to the nature and

distribution of North American karst aquifers is given by Brahana

et al. (1988).

To many hydrologists, geologists, and engineers, the flow of

ground water in karst terranes is mysterious, capricious, and

unpredictable. Many others recognize the term, karst, but do not

understand its significance. Few publications adequately discuss

predictive aspects of environmental hydrologic problems of karst

terranes or offer practical, experience-based insights for

solving them. Notable exceptions are Alexander et al. (1988),

Aley (1977, 1988), Aley & Thompson (1984), Bonacci (1987),

Crawford (1984), Milanovi& (1979), Palmer (1986), Quinlan (1986b,

1988a, 1988b), Quinlan & Alexander (1987), Quinlan and Ewers

(1985, 1986), Quinlan & Ray (1981, 1989), Smart & Hobbs (1995),

Smoot et al. (1987), and White (1988). Both the actual problems

and the policy problems of flow prediction in karst terranes are

real. Yet they are totally ignored in nearly all of the

environmental monitoring literature and by most current U.S.

Environmental Protection Agency (EPA) and state ground-water

monitoring regulations. These problems can, nevertheless, be

solved. The National Water Well Association regularly offers

practical course in karst hydrogeology to professionals who

to learn about these problems and attempt to solve them.

1
Although some EPA publications take note of the problems of

monitoring in karst terranes and endorse the traced-spring,

-cave-stream, and -well monitoring concepts first espoused by

Quinlan and Ewers (1984, 1985) and analyzed in this document

(Office of Ground-Water Protection, EPA, 1987, 1988; Mull et al.,

1988), most EPA documents, and specifically the Technical

Enforcement Guidance Document concerning ground-water monitoring

at sites governed by the Resource Conservation and Recovery Act

(RCRA), ignore these concepts. The significance of the proposed

strategy for monitoring in karst terranes was strongly endorsed

in letters to EPA from numerous authorities who urged, albeit

unsuccessfully, that they be included in the revision of the

Draft TEGD on monitoring for RCRA sites (Office of Waste Programs

Enforcement and Office of Solid Waste and Emergency Response,

EPA, 1986).

Most current regulations concerning ground-water monitoring

disregard the manifold problems of doing it reliably in karst

terranes. In part, this is because they are performance-based,

and rightly so. The true goal of ground-water monitoring should

be to detect the nature and magnitude of changes, if any, in

ground-water quality, as a result of natural processes or human

activity-- rather than just to comply with the letter of the law

and regulations. Monitoring can be reliable, and it can achieve

performance goals, but sometimes it needs the aid of technical

guidance in order to achieve its goals.

Proposed EPA regulations concerning municipal solid waste

landfills are some of the first to specifically address problems

of karst terranes and to prohibit siting them in such areas (U.S.

Environmental Protection Agency, 1988, p. 33333-35). This prohibition could be waived if an owner or operator can demonstrate

structural stability of the facility. The regulation (40 CFR

Section 258.15) is an important official recognition of a significant problem in karst terranes, sinkhole development, but it

ignores the most widespread problems--those associated with

reliable monitoring of ground-water quality as affected by

leakage from a facility located in a karst.

This document discusses where and when to take relevant water

samples and how to get the data essential for making the where

and when decisions; it also discusses implicit assumptions made

when monitoring with the strategy recommended herein. No attempt

is made to discuss the design and construction of monitoring

wells, protocol for sample custody, or quality assurance/quality

control (QA/QC) for dye-tests. All but the last of these topics

have been adequately addressed by others; many topics are being

reviewed and codified by the American Society for Testing and

Materials (ASTM) Subcommittee D-18.21 on Ground-Water and Vadose

Zone Investigations. Section .09 of this ASTM subcommittee,

chaired by the writer, is developing standards for dye-tracing

and ground-water monitoring in karst terranes.

2
Two major types of ground-water flow occur in karst

aquifers-- conduit flow and diffuse flow, each of which is an endmember of a continuum. Springs and cave streams in conduit-flow

systems are “flashy”, as expressed by high ratios between their

maximum discharge and base-flow discharge, typically 10:1 to

1000:1. Discharge responds rapidly to rainfall. Flow is generally turbulent. The waters possess low but highly variable

hardness; turbidity, discharge, and sometimes temperature also

very widely. Where a karst aquifer is less developed and is

characterized primarily by diffuse flow, its behavior is less

flashy; the ratio between maximum discharge and base-flow discharge of major springs is low (4:1 or less) , and the response of

their discharge and water quality to rainfall is slower than in

conduit-flow springs. Flow is generally laminar. Hardiness is

higher than in conduit-flow springs, but hardness, turbidity,

discharge, and temperature have low variability (Quinlan and

Ewers, 1985). The variations in and relations among these

properties and their variability as a function of aquifer flow,

storage, and recharge have been described in a significant paper

by Smart and Hobbs (1986).

Two important and seemingly contradictory points need to be

made about diffuse flow:

1. Movement of water

aquifer is similar

aquifers. Darcy’s

& Howie, 1988).

through most parts of a diffuse-flow

to movement of water through granular

law is operative (Hickey, 1984; Wailer

2. Although water from a diffuse-flow spring may be

discharged from an obvious conduit, perhaps 3 m (10 ft) in

diameter, the geometry and configuration of the “plumbing

system” that feeds it near the orifice is trivial. Wilson

and Skiles (1988) and Stone (1989) have published maps of

different cave systems with more than 11 km (6.5 mi) of

braided passage that feeds diffuse flow springs.

The most significant controls on flow-type are the types

of recharge and storage, as discussed by Smart & Hobbs

(1986). These most influence the degree of variability of

water chemistry and the magnitude, timing, and duration of

response of springs and wells to storms. For very large

ground-water basins there is additional dampening of

response to storms as a consequence of their sheer size

and the greater time necessary to transmit the storm input

to their spring (White, 1988, p. 186-187). Also, individual storms will tend to overlap and seasonal trends will

comprise the most obvious part of the annual record.

A quick, inexpensive way to distinguish between a conduitflow spring and a diffuse-flow spring is to observe its turbidity

4
and to measure its specific conductivity before, during, and

after several large storms. If the spring is characterized by

conduit flow, the water will be turbid; the coefficient of

variation (standard deviation ÷ mean x 100) of its specific

conductivity will be 10 to more than 25 percent. If the spring

is characterized by diffuse flow, its water will always be clear

to slightly turbid; the coefficient of variation of its specific

conductivity will be less than 5 percent (Quinlan and Ewers,

1985). Remember, however, that conduit flow and diffuse flow are

end-members of a continuum. Many springs are fed by a mixture of

both types of flow (White, 1988, p. 183-187).

Ground-water flow in conduits and fractures of karst aquifers

differs radically from flow in other aquifers. Most commonly, it

is to springs by way of caves. Such flow is generally faster

than in other aquifers; extreme velocities of 2300 m/hr (7500

ft/hr) have been observed, while a range of 10 to 500 m/hr (30 to

1500 ft/hr) is typical for many conduits (Quinlan & Ewers, 1985,

p. 202). [The latter two flow velocities are equivalent to

approximately 90 and 4400 km/year (55 and 2700 mi/year).] Thus

the effects of leakage or a spill of hazardous material on water

quality in a karst aquifer can be sensed at great distances in

less than a day. Conceptual and porosity relations between

conduit-flow aquifers, fractured aquifers, granular aquifers, and

diffuse-flow aquifers are shown in Figure 1.

Figure 1. Relation between four major types of aquifers and the

dominant porosity geometry in each. This is a tetrahedral

continuum. (after Quinlan & Ewers, 1985)

5
Waste-disposal facilities should not be located within a

karst terrane unless one is willing to risk sacrificing the use

of at least part of the subjacent karst aquifer as a source of

potable water. This is a high risk, almost a certainty. Nevertheless, many facilities already exist within a karst environment, and hazardous materials are disposed of without supervision

by a karst-aware hydrologist. Also, hydrologists are rarely consulted in the selection of a proper site for an accidental spill

of hazardous material.

Immediately beneath the soil and overlying a domain of

horizontal flow in diffuse-flow and conduit-flow systems, is the

epikarst (also known as the subcutaneous zone). It comprises the

strongly karstified water-bearing rock in the vadose zone beneath

the soil and above the phreatic zone, but it may be separated

from the phreatic zone by tens to hundreds of feet of dry,

inactive, waterless bedrock that is locally breached by vadose

percolation. Commonly, the epikarst is 3 to 10 m thick (10 to 30

feet). Williams (1983) estimates that 50 to 80 percent of the

dissolution done by recharging waters takes place in this interval. Thus its storativity and transmissivity are significantly

greater than in the subjacent, less karstified, relatively dry

rock in the vadose zone. The peculiar hydrology of the epikarst,

and how it differs from that of the diffuse-flow and conduit-flow

part of karst aquifers, has been discussed by Williams (1983),

Friedrich (1981), Gouisset (1981), Headworth et al. (1982) and

Walsh and Puri (1985).

The complex relations between hydrology, landforms, and

stratigraphy in karst developed in gently-dipping rocks in a lowrelief terrane are shown schematically in Figure 2. The cave

stream is recharged through the soil, by sinking streams, seepage

through bedrock, and underground tributaries. Flow is integrated

at a trunk stream that ultimately discharges at a spring. Many

analogies can be made between the hydrology of cave rivers and

surface rivers. The terrane shown is hypothetical, but representative of karst terranes in much of the Midwestern United States.

Ten different types of ground-water flow are shown; most of them

are discussed by Gunn (1986) and White (1988). A comprehensive

discussion of this figure is beyond the scope of this document.

Additional concepts of karst hydrology will be introduced in

the discussion of where to monitor for pollutants and background.

Case histories will be cited to illustrate these concepts.

DESIGN OF MONITORING SYSTEMS IN KARST TERRANES

WHERE TO MONITOR FOR POLLUTANTS

The ground-water monitoring regulations specified by RCRA for

6
treatment, storage, and disposal facilities (TSDF’s) (40 CFR Part

264, Subpart F; and Part 265, Subpart F) prescribe a minimum of

four wells to be installed at a facility to be monitored--one

well upgradient and three wells downgradient, commonly near the

facility boundary and almost always on the facility property.

Sampling is to be performed annually, semi-annually, and perhaps

quarterly, but rarely as often as monthly. Monitoring wells

installed and sampled in this manner in most karst terranes will

generate great amounts of carefully collected, expensive data,

most of which are useless because the wells usually fail to intercept the contaminants they are intended to detect. They fail

to do so because the wells do not encounter the flow lines (cave

streams) draining from a site; if they do intercept them, an

insufficient number of samples is taken at inappropriate intervals and at the wrong time (Quinlan & Ewers, 1985; Quinlan and

Alexander, 1987).

The easiest and most reliable sites at which to monitor

ground-water quality in a karst terrane are springs and directly

accessible cave streams shown by dye-tracing to drain from the

facility being evaluated (Quinlan & Ewers, 1985). It is, however, naive, erroneous, and dangerous to assume that all or most

of the springs that must be regularly sampled during tracer tests

and ground-water monitoring are indicated on U.S. Geological Survey (USGS) 7.5-minute topographic quadrangles. Experience in

numerous karst areas has shown that only about 5 percent of the

springs discharging non-isolated local flow are shown on maps.

Inclusion of a spring on a topographic map is not necessarily an

indicator of the significance of its discharge. Many trivial

springs are included on a map because of their cultural associations. Field work to discover springs is mandatory; there is no

substitute!

The preferred alternative to the use of springs as monitoring

sites is a suite of wells that intercept cave streams shown by

tracing to flow from the facility. Cave streams may be difficult

or impossible to find with traditional geophysical techniques.

New techniques such as the use of streaming potential (measurement of the electrical potential-gradient caused by displacement

of ions from fissures and rock grains as water moves from a

recharge area to a discharge area; Lange, 1988; Lange & Quinlan,

1988) and acoustic detection (measurement of sound waves caused

by the knocking of pebbles against one another during saltation,

by cavitating water, or by cascading, riffling, and dripping

water in partly air-filled cavities) have been tested with

encouraging results by Lange (1988). Stokowski (1987) has also

tested streaming potential as a guide to discovery of caves, but

his results are not supported by his data. Other non-traditional

geophysical techniques also have promise (Lange, 1988).

A second alternative for monitoring sites in karst terranes

is a suite of wells located on fracture traces or on fracture8
trace intersections. These wells are usable only if tracer tests

show a connection with the facility under base-flow as well as

flood-flow conditions. Although some cave passages are coincident with various types of fracture traces and lineaments, not

all fracture-related features are vertical and therefore directly

above cave passages (Werner, 1988). Many cave streams are

developed along bedding planes and are unaffected by vertical

fractures. This fact lessens the probability that a well drilled

on a fracture trace, a lineament, or the intersection of such

linear features will intercept a cave stream. This fact does not

challenge the well-known correlation of such linear features with

increased water yields (Parizek, 1976).

As a third alternative, randomly located wells could also be

used, but only if tracing has first proven a connection from the

facility to each of the monitoring wells under various flow conditions. Domestic, agricultural, industrial, and monitoring

wells are generally sited where convenient, but for a specific

purpose. Even though monitoring-well locations are designed for

detection of contaminants or to intercept contaminant plumes

oriented parallel to the hydraulic gradient or to be upgradient

from them, I consider most such wells to be randomly located--

unless they were deliberately sited along fracture traces or

fracture-trace intersections. Throughout this document I use the

phrase randomly located wells in this sense. Most randomly

located and nonrandomly located wells were not intended for

aquifer-testing or dye-tracing, but they can be so used. Such

wells can not be used as monitoring wells for a facility in a

karst terrane unless they have been shown to be positive for

tracer released at or very near the facility to be monitored.

At most locations in karst terranes, proper and reliable

monitoring can only be done at sites outside the boundary of a

facility. Frequently this monitoring can only be done at places

up to several kilometers away from the facility.

WHERE TO MONITOR FOR BACKGROUND WATER QUALITY

Springs, cave streams, and wells in settings geochemically

and culturally similar to the traced monitoring sites are the

only suitable places to monitor background water quality. This

is true, however, only when these places have been shown by carefully designed, repeated dye-traces done under conditions ranging

from base flow and flood flow not to drain from the facility.

There may be no obvious place that can be monitored for background water quality. This is especially true for a facility

located on a hill that is also a potentiometric high and characterized by radial flow of ground water. If randomly located,

non-traced wells are irrelevant for monitoring possible contaminants in karst ground water, they are equally irrelevant for

monitoring background for contaminants.

9
One must be exceedingly cautious in interpreting negative

results of a tracer test. Only some springs and randomly located

wells that are negative for dye in a judiciously designed, properly executed trace are in settings geochemically and culturally

similar to those of the monitoring wells. Use of such springs

and wells for monitoring background when no other monitoring

sites are available is rational, but not specifically addressed

by current regulations.

Background data at a new facility can also be obtained from

wells that are monitored before, during, and after storms.

Ideally, this will be for at least a year before operations

begin. This duration would probably detect storm-related and

seasonal trends in water quality. Analysis of continuous records

of stage in these wells can be used for selection of stormrelated sampling frequencies and for possible differentiation of

wells into several categories that are based on hydrographic and

chemical responses to storms. Traces should be run during that

year-long period in order to demonstrate the presence or absence

of connections between the various wells and the facility to be

monitored. If a well does not test positive for dye after allowance is made for probable flow velocity, it can not be considered

to be an effective part of a monitoring system. If none of the

wells, or an insufficient number of wells to which traces are

attempted, tests positive for dye, an effective monitoring system

does not exist; one probably does not understand the hydrogeology

of the facility. New hypotheses about flow must be devised and

tested in order to verify the correctness of the understanding of

its hydrogeology.

There are problems associated with the validity of statistical comparison of data from background wells with purportedly

relevant data from monitoring wells. No matter what sophisticated statistical tests "prove", if common sense and field

observations demonstrate that samples from monitoring sites and

background sites are taken from the equivalent of two different

populations, the statistical analyses are invalid. EPA, in

recognition of problems of comparability, has been seeking ways

to make objective, valid comparisons. It has proposed the

concept of “ground-water trigger-levels” for determining if a

facility is in compliance with regulations. State agencies are

to set trigger-levels for compounds and metals after critical

review of data on their toxic effects on health of people and

aquatic life. If the trigger-level of a compound or metal is

exceeded in a monitoring well for a facility, an investigation of

possible sources must be made (U.S. Environmental Protection

Agency, 1988, p. 33370-71). No matter what concept of background

is used, and no matter what the concept of background evolves to,

there is still the problem, addressed herein, of validly

comparing two sets of water-quality data which may not be related

in any way.

10
DISCUSSION OF WHERE TO MONITOR FOR POLLUTANTS AND BACKGROUND

All six types of monitoring sites and background sites--

springs, directly accessible cave streams, wells drilled to cave

streams, wells drilled on fracture traces, wells drilled on

fracture-trace intersections, and wells drilled randomly--must be

tested by tracing, not only during moderate flow but also during flood flow and base flow, in order to prove the usefulness of

these sites for monitoring. This must be done during the

extremes of expected flow conditions because flow routing in

karst terranes commonly varies as stage changes. During flood

conditions the water level in conduits will rise. Some of the

water may temporarily move through conduits that are dry during

low-flow conditions and be switched (decanted) into adjacent

ground-water basins, thus being temporarily diverted by them. An

example of such hydraulic switching is depicted in Figure 3.

During moderate and flood flow, water draining from Park City and

from the west boundary of Cave City in the Turnhole Spring

ground-water basin may flow via intermediate-level and high-level

crossover routes (shown by the north-trending dashed lines with

arrows) to as many as three other basins (#4, Sand Cave; #6, Echo

River; and #7, Pike Spring, as indicated by the short, northtrending arrow that crosses the boundary of the ground-water

basin at the south end of Roppel Cave). These three subsurface

diversion routes are also shown more schematically in the western

third of Figure 4. They are part of the distributary with 12

springs. Other more complex examples of hydraulic switching are

known (Smart & Ford, 1986; Smart, 1988a, 1988b).*

Another peculiarity of water movement in many karst aquifers

is distributary flow. An underground distributary is a dispersal

route analogous to the distributary at the mouth of a major river

that empties into a sea. However, its origin is different, as

discussed by Quinlan & Ewers (1985, p. 205, 207) in their

description of the similarities between surface rivers and large

underground rivers. Figure 4 shows numerous distributaries in

the Mammoth Cave area. The positions and geometry of the underground branching shown are schematic, but their existence has

been confirmed by cave-mapping and by dye-tracing. Some of these

underground flow paths have been mapped. Knowledge of the occurrence and functioning of distributaries is important in the de-

*It would be interesting to make an analysis of the similarities

between hydraulic switching in caves and the principles of

fluidics (fluid logic circuits that can be used for sensing,

logic, memory, timing, and interfacing; Esposito, 1980, p. 338-

367). The whimsical idea of a cave as a giant fluidic computer

is reminiscent of Douglas Adams’s proposal that the Earth is a

giant organic computer designed to calculate the Question to the

Ultimate Answer (Adams, 1980, p. 181-183).

11
Figure 3. Hydrology of the Turnhole Spring ground-water basin,

the major basin draining into Mammoth Cave National Park,

Kentucky. Beds dip north at about 11 to 40 m/km (60 to 210

ft/mi). (after Quinlan & Ewers, 1985)

12
sign of a monitoring system in a karst terrane because pollutants

from a point-source in the headwaters or mid-reaches of a groundwater basin or subbasin may flow to all springs in its distributary system or subsystem during periods of high stage. For

example, pollutants reaching ground water beneath a point-source

in the east-central area of Figure 4 south of Green River would,

depending upon flow conditions, disperse to as many as 52 springs

in 19 isolated segments along a 19 km (12 mi) reach of Green

River. Part of this distributary is shown also in Figure 5. Not

all springs between the extremities of a distributary are

necessarily a part of the distributary.

The extent to which ground-water basins and their subsurface

flow routings can be deciphered by tracing, mapping of the potentiometric surface (water table), and mapping of caves is shown in

Figures 3 and 5. The east half of Figure 5 summarizes some of

the results of a study of the dispersal of heavy metals from a

Figure 4. Distributary springs along Green River in the Mammoth

Cave area, Kentucky. The numbers indicate the total number of

springs in a given distributary system or subsystem. The westernmost large distributary, with 12 springs, is a schematic

representation of part of the Turnhole Spring ground-water basin

shown in Figure 3. (after Quinlan & Ewers, 1985)

13
Figure 5

14
metal-plating plant; the metals had been discharged into the

ground (in concentrations of more than 10 mg/l via a municipal

sewage treatment plant at the town of Horse Cave. [The outline

of the town is just east of the center of the map.] A study of

the chemistry of water from springs, wells, cave streams, and the

west-flowing Green River showed that effluent from the sewage

treatment plant flows underground 1.6 km (1 mi) northeast to the

unmapped cave beneath the town and then, depending upon stage, to

a total of as many as 46 springs in 16 segments along an 8 km (5

mi) reach of the river, about 7 km (4 mi) to the north (Quinlan &

Rowe, 1976). None of the 23 wells monitored for heavy metals

during base flow showed concentrations higher than background

levels, but they should have been sampled also during flood-flow

conditions, when water movement might have been reversed from the

trunk conduit to some of the wells (Reeker et al., 1988).

Figure 6 shows complex, radial flow of ground water in faulted, flat-lying rocks in the Ozarks of Arkansas. The divergent

results of six dye-tests, four of which are within a mile of a

proposed landfill, are summarized in this figure. Although the

flow routing in Figure 6 is more complex than that shown in

Figures 3 and 5, the routing illustrates the advocated maxim,

"Monitor the springs!" Monitoring wells that could be drilled on

fracture traces at the proposed landfill site might detect

seepage of leachate from it, but there is no means--other than by

tracing--to identify correctly and conclusively the places to

which leachate would (and probably would not) flow. Stated

another way, no matter how superbly efficiently the hypothetical

monitoring wells on the landfill property were able to detect

leakage from the landfill--assuming, for sake of discussion, that

they would function so reliably--there is no way, other than by

dye-tracing or by monitoring of numerous springs and wells off

the property, that one could discover the many consequences of

leakage from the proposed landfill!

All four wells at the town of Pindall (Figure 6), which were

pumped continuously during the dye-trace from the east boundary

of the landfill site (the first trace run), tested positive for

dye. By inference, if other wells in the town had been pumped

continuously, many of them, perhaps all, would have been positive. Note that none of the four wells immediately east and

southeast of the first dye-injection site and neither of the two

wells immediately west of the western dye-injection site were

positive for dye during the first test. [None of these six wells

was sampled for dye during the five subsequent tests.] During

the second and third dye-tests, in which dye was injected 1.6 km

(1 mi) east and west of the proposed landfill site, respectively,

only two wells at Pindall were pumped continuously. The southernmost well there was positive for dye in both tests, as shown;

the easternmost well was ambiguous for dye from the east and

negative for dye from the west. The high-yield well on the

fracture trace (at F) was positive for dye from the second,

15
Figure 6

16
third, and sixth dye-tests, but negative for dye from the first,

fourth, and fifth. Dye recovery from this well during the first

test was probably hampered by the effects of chlorine added to

the well water; chlorine may react with tracer dyes and can

destroy low concentrations of any that are present (Smart & Laidlaw, 1977). The difference in tracer results at the high-yield

well may also have been a consequence of greater efficiency in

sampling and analysis during these latter tests and of the use of

the writer’s pumped-well dye-sampling device (Aley et al ., 1989).

The results of the fourth and fifth dye-tests are consistent with

the hypothesis that Mitch Hill Spring is fed by deep, complex

circulation of ground water which is recharged near faults,

lineaments, and fracture traces (Aley, 1988). The implications

of these dye-test results for wellhead and springhead protection

are discussed by Aley (1988) and Quinlan, Aley, and Schindel

(1988). [Dye-trace results from the first test were instrumental

in the 1987 defeat of a proposal to use this site as a landfill.]

Dye for the six traces near Pindall was injected near the

headwaters of several surface streams and at low- to moderateflow conditions. The results might be different if the dye-tests

were run during base-flow and flood-flow conditions, but flow

would still be radial.

Radial flow occurs in many karst terranes and has also been

documented in them at waste disposal facilities (Quinlan & Ewers,

1985, p. 214-219; Aley, 1988); it tends to be associated with

locations on topographic highs. When evaluating a facility by

dye-tracing, one must keep an open mind and place dye-detectors

at not only the logical, obvious places, but also at the illogical, the unlikely, and the “No, it couldn’t ever go there”

places. The cliche "Expect the unexpected" applies, no matter

how experienced one is in tracing.

It is tempting and all too easy to take what I think is false

comfort in the interpretation of negative tracing results to a

well to mean that a waste disposal facility and its liner (if

present) are functioning as designed. More specifically, it is

easy to say that a facility is either not leaking or is adequately attenuating everything put into it, but I think it is naive to

say so unless one is extremely and justifiably confident in the

validity of the tracer-test results (Quinlan, Ewers, & Field,

1988). Alternatively, the negative tracing results could just as

logically be a consequence of monitoring for dye in wells that do

not intersect that part of the aquifer in which ground water

actively circulates. In most karst terranes, the latter explanation is more likely to be correct. If water is not standing on

the ground in pools or ponds, infiltration must be flowing someplace. If infiltration is flowing someplace, an uncapped facility is probably leaking to somewhere.

Sometimes there is resistance on the part of owners/operators

17
to the performance of a dye-test from their facility or its

immediate vicinity. This resistance is based on their fear that

environmental administrators and the public would judge that the

rate of flow of dye from an adjacent sinkhole, for example, to a

spring would be interpreted to be the same as the rate of leakage

from the facility. The rates are not necessarily the same. But

it is essential to perform the trace so that the consequences of

possible leakage become known before leakage actually occurs.

The sinkhole is used as an injection-site because flow velocities

from it will probably be more rapid than through a liner and

because of the economic constraints of investigation time. In

such a situation, the most important monitoring question about

leakage is where to, not when or how fast.

In some epikarsts and soil zones characterized by diffuse

flow, it may be possible to determine the optimal location for

reliable monitoring wells by applying standard, topography-based

principles of well-siting developed for non-karst terranes.

Springs may be unreliable monitoring sites because of the long

residence time of pollutants “hung-up” in the epikarst. Not

enough research has been done on monitoring of epikarsts to

formulate specific protocols for optimally doing so.

RELEVANCE OF OFF-FACILITY MONITORING TO RCRA REGULATIONS

Although no EPA regulations specifically sanction the monitoring protocols recommended herein for karst terranes, these

protocols are not inconsistent with any EPA regulations. The

following paragraphs discuss how the existing regulations can be

interpreted as relevant to these protocols and to off-facility

monitoring.

Conventional monitoring practice and EPA regulations for

RCRA, 40 CFR Section 264.95, require that ground-water monitoring

be done at the compliance boundary of a treatment, storage, and

disposal facility (TSDF). Whenever hazardous constituents from a

regulated unit exceed concentration limits in ground water (set

under Section 264.94) in wells located between the compliance

point (under Section 264.95) and the property boundary downgradient from a facility, the owner or operator must institute a

corrective action program (under Section 264.100). The concept

of off-facility monitoring is not mentioned or implied, but the

RCRA TEGD for ground-water monitoring recommends that “Geologic

environments with discrete solution channels such as Karst

formations must have detection monitoring wells located in those

solution channels likely to serve as conduits for contamination

migration.” (Office of Waste Programs Enforcement and Office of

Solid Waste and Emergency Response, EPA, 1986, p. 47).

It has been shown that analysis of off-facility water samples

from traced-springs, -cave-streams, and -wells may be the only

18
way to reliably detect and monitor whether the facility is

leaking (Quinlan & Ewers, 1985; Quinlan, 1988a; Aley, 1988).

Achievement of the recommendation of the TEGD is most practically

and most economically achieved by following the strategy these

authors advocate. The non-traditional geophysical techniques

described by Lange (1988) and Lange and Quinlan (1988) may

expedite finding the conduit to be monitored, but these techniques are still in a developmental stage and there are limitations

on their applicability. Also, even if a large water-filled

conduit is found by geophysical techniques, it must be tested by

tracing.

I do not believe that lack of specific mention in the RCRA

regulations of monitoring at springs, cave streams, and traced

wells prevent their use. Why? Section 264.97(a) of RCRA regulations requires that a ground-water monitoring system consist of

a sufficient number of wells, installed at appropriate locations

and depths to yield ground-water samples that represent background water quality not affected by the regulated unit represent

the quality of water passing the point of compliance. I contend

that unless downgradient and upgradient wells intercept a cave

stream (not necessarily the same cave stream) they are incapable

of doing what they were intended to do; there is an insufficient

number of wells unless one or more of them intercepts a conduit

that is traced from the facility to be monitored.

I believe that dye-traced springs and cave streams must be

recognized as valid, essential components of a ground-water monitoring system in a karst terrane. Section 264.97(d) requires

that the ground-water monitoring program include consistent

sampling and and analysis procedures designed to ensure monitoring results that provide a reliable indication of ground-water

quality below the waste-management area. In karst terranes such

reliable indications are most easily and best obtained from

springs, cave streams, and wells drilled to intercept them. The

relevance of such interceptions must be proven, however, by

tracing.

Additional support for monitoring at traced-springs and

-cave-streams, rather than at randomly located wells, is implictly but strongly given by other RCRA regulations such as 40 CFR

Section 270.14(c), which is concerned primarily with site characterization. Subpart (c)(2) of this regulation requires identification of the uppermost aquifer and aquifers hydraulically

interconnected beneath a facility property, determination of

ground-water flow direction and flow rate, and statement of the

basis for the conclusions of hydrogeologic investigations of the

facility. In most karst terranes, an approximation of the

required flow direction can be obtained from potentiometric data

but it should be confirmed by tracing. Flow-rate data can be

obtained only by tracing.

19
Section 270.14(c) (4) requires description of any plume of

contamination that has entered the ground water from a regulated

unit and delineation of it on a topographic map. A contaminated

cave stream is a plume, albeit a confined plume; the only way to

delineate it is by potentiometric and/or geophysical mapping

confirmed by drilling (as discussed herein) and by dye-tracing.

Although Section 270.14(b) (19) states that the required topographic map should show the area around a facility to a distance

of 1000 ft (305 m) from its boundary, it would be reprehensibly

legalistic and short-sighted to map a plume to that boundary and

stop. Such a map would imply that the plume ceased to exist at

the boundary or ceased to be the responsibility of the owner/

operator. The plume should be mapped farther, to its end at a

spring, when it is technically feasible to do so. The polluted

spring will contaminate surface waters.

Section 270.14(c) (5) requires detailed plans and an engineering report describing the proposed ground-water monitoring

program to be implemented to meet the requirements of Sections

264.97(a)(2) and 264.98(b). These two Sections require that the

monitoring system represent the quality of ground water passing

the point of compliance. Since the water flowing from the

facility is in a conduit or in a zone of diffuse flow that may

include a conduit, wells must be drilled to intercept and monitor

it. Substitution of monitoring of a spring for monitoring of a

well drilled perhaps 30 m (100 ft) from its point of discharge

would be logical, cost-efficient, and not inconsistent with

existing regulations.

WHEN TO MONITOR FOR POLLUTANTS AND BACKGROUND

Current conventional monitoring protocol generally requires

well sampling annually, semi-annually, occasionally quarterly,

and rarely as often as monthly. This is reasonable in most nonkarst terranes. In karst terranes, however, even at springs and

cave streams judiciously and correctly selected as monitoring

sites by the dye-tracing procedures recommended herein, by Quinlan & Ewers (1985), and by Aley et al . (1989), the analytical results of such regularly collected data can be inadvertently

misleading. The net result is a waste of time and money. Why?

If a karst aquifer is characterized by conduit flow, the chemical

quality of water at a spring to which it drains can be greatly

affected by the effects of storms and meltwater events (Quinlan &

Alexander, 1987; Libra et al ., 1986; Hallberg et al., 1985). In

contrast, the water quality of diffuse-flow systems is generally

only slightly affected (Quinlan & Ewers, 1985) . Sampling must

also be conducted during base-flow conditions and analyses should

be compared with water quality during and after storms and meltwater events. Water quality is dependent upon the type of pollutant source (point-source vs. nonpoint-source), its volume, and

its concentration.

20
In order to reliably characterize the natural, storm-related

variability in water-quality of a spring in a conduit-flow

system, sampling must be done much more frequently than was

customary in the past. Characterization of water quality, when

using typical semi-annual sampling for study of a conduit-flow

system is analogous to estimating annual rainfall of an area

solely on the basis of rainfall data collected on the same two

days of each year.

The effect of sampling frequency on the accuracy of characterizing and depicting storm-related variability in water quality

affected by non-point agricultural pollutants is illustrated in

Figure 7. Data are from the Big Spring basin, Iowa. Figure 7A

is a composite of three figures published by Libra et al . (1986).

Discharge was recorded continuously. Pesticides were sampled up

to 6 times per day, nitrate was sampled up to 20 times per day,

and suspended sediment was sampled up to 17 times per day,

depending upon stage (Bernard E. Hoyer, Iowa Geological Survey,

Iowa City, oral communication, 1987) . For the sake of discussion, however, assume that these parameters were monitored

continuously for the n-day interval shown. Assume also that

apparent variations in water quality as a result of its natural

variability, the statistics of sampling and analysis, and

analytical error are trivial--even though they possibly are not.

The most important thing to be seen in Figure 7A is that pesticides and suspended sediment have a maximum that approximately

coincides with that of the discharge. Nitrate, however, has its

minimum when the others are at a maximum, and reaches its own

maximum several days later when the others are approaching their

low pre-storm values. The reasons for this peculiar lack of

synchrony are discussed elsewhere (Libra et al ., 1986; Hallberg

e t al., 1985; Quinlan & Alexander, 1987). Note also that

pesticide concentrations increased. by more than an order of

magnitude during and after the storm. Such storm-related

variation in nonpoint-source pollutants is common, but longer

term precipitation-related variation in what can falsely be

called “background” for pesticide concentration may range over

more than three orders of magnitude at some sites in just a few

months (Quinlan & Alexander, 1987).

Figures 7B through 7F are based on the data represented in

Figure 7A, but they assume “sampling” at intervals of 12, 24, 48,

and 96 hours. For the storm event that occurred, sampling at 48-

hour and 96-hour intervals (Figures 7E and 7F) is totally incapable of suggesting any significant change in water quality. The

24-hour sampling frequency (Figures 7C and 7D) is better, but the

midnight samples happen to miss the decrease in nitrate and most

of the increase in pesticides.

(Figure 7B) is only a crude

sampling represented by Figure

Even a 12-hour sampling interval

approximation of the continuous

7A.

21
Figure 7. Water quality and discharge during a 1984 storm in a

karst aquifer characterized by conduit flow, Big Spring groundwater basin, Iowa. The data shown in A are assumed to be

accurate and continuous. Figures H through F show the same data

which are "sampled" at the indicated intervals and are derived

from A which is composite from 3 figures in Libra & ~. (1986).

22
In September 1987 the EPA announced proposed standards for

the concentration of various pesticides in ground water. For

atrazine, a herbicide used to control weeds in corn, sorghum,

sugarcane, pineapple, and citrus groves, the maximum allowable

concentration is 3 parts per billion (ppb) (Office of Drinking

Water, EPA, 1987). About 90 percent of the pesticides found in

Big Spring basin are atrazine. Therefore, the peak pesticide

concentrations shown in Figures 7B, 7C, and 7A are about 4 to 6

times higher than this 3 ppb limit. Midnight sampling barely

detected violation of the 3 ppb maximum (Figure 7D) ; the 48-hour

and 96-hour sampling (Figures 7E and 7F) totally failed to do so.

It is to be stressed that the “optimal sampling frequency",

however defined and determined, will vary with the event,

pollutants to be monitored, discharge, flow dynamics, and flow

type (conduit flow versus diffuse flow) of the karst aquifer

studied, as well as whether one is trying to sample for pointsource or nonpoint-source pollutants. For example, the proper

sampling frequency necessary to accurately characterize changes

in the chemical composition of spring discharge affected by

nonpoint agricultural pollution in the Big Spring ground-water

basin (and in many other conduit-flow basins in the Mammoth Cave

area and elsewhere) will, for similar storms and similar antecedent moisture conditions, be far more than is necessary to give

the same accuracy of characterization at the diffuse-flow springs

draining the Edwards aquifer in Texas, the Floridan aquifer, and

much of the Ozarks of Missouri and Arkansas. Another example:

For a given spring and a given set of antecedent precipitation

conditions, the optimal sampling frequency for detection of

leakage from a landfill would be affected by the thickness and

integrity of its cover and the distance of the landfill from the

spring. Therefore, one would be wise to allow for a possible lag

of pollutants behind maximum discharge in response to rain

storms. The optimal sampling frequency at such a landfill could

be different from that for point-source contamination such as a

spill of hazardous materials along a highway.

Perhaps the most economical solution to water quality problems caused by storm-related or meltwater-related pulses of water

that exceed proscribed limits or guidelines for one or more

pollutants is to temporarily divert the potentially polluted

water out of the water-supply circuit (Quinlan & Alexander,

1987). Study would first be needed to identify which pollutants

are present in such waters, when (relative to the hydrography

peak) they reach their maximum concentration, whether it is

practical to continuously monitor for them or a surrogate, or

whether it is best instead to divert stormwater and meltwater

when stage and other flow-related parameters reach certain

critical values.

Statistical procedures for designing and evaluating sampling

strategies are available (Sanders et al ., 1983; Gilbert, 1987;

23
Makridakis et al ., 1983; Chatfield, 1984; Bendat & Piersol, 1986;

Gibbons, 1987; Montgomery et al ., 1987; Rouhani & Hall, 1988),

but they are complex.

Until an acceptable, economically realistic, reliable procedure for sampling ground water in karst terranes is developed

and tested, probably the best protocol is that proposed and

discussed by Quinlan & Alexander (1987, p. 281). In brief,

sampling, especially for nonpoint-source pollutants, should start

at base flow, before the beginning of a storm or meltwater event,

and continue until 4 to 30 times the time from the start of

hydrography rise to the time of its crest, depending upon the

extent to which an aquifer is characterized by conduit flow or

diffuse flow. Sampling may have to be done as often as at 1- to

6-hour intervals in the early part of a precipitation event and

at 4- to 24-hour intervals in the recession limb of its hydrography. Appropriate sampling frequency can be determined by

analysis of either continuous records of stage and specific conductivity or hourly readings for them.

After an event, the decision about which samples to analyze,

if any, should be based on a careful evaluation of the significance of the event, interpretation of data from previous events,

and an estimation of the data needed to characterize the monitoring site. Many samples, sometimes all, can be rightly discarded.

Hard, judicious decisions must be made. These analytical data

must be compared with those for samples taken several times per

year during base flow, storms, and meltwater events. Only then

can one possibly make a reliable assessment or characterization

of the true quality of water draining from a facility. After

several years of data have been accumulated and the aquifer

behavior is understood, sampling frequency may be decreased.

On the basis of conceptual models and limited data, I believe

that monitoring in karst terranes characterized by diffuse flow

is easier and cheaper than in those characterized by conduit flow

because fewer samples are required. Also, the more a karst aquifer is characterized by diffuse flow, the higher the probability

that fracture-trace-sited wells and randomly located monitoring

wells can be used reliably. All wells proposed for monitoring

use, however, must still be tested positively by tracing. This

tracing adds time and cost to the design of a monitoring system,

but in the long run, it is cheaper to design it properly.

Much remains to be learned about when to sample ground water

in karst terranes. The likely possibility for either deliberate

or inadvertent acquisition of falsely negative or falsely

positive data from them makes it imperative that people in charge

of-sampling and officials

have an understanding of

scrutinize all analytical

in charge of evaluation of sample data

karst problems. They should carefully

results from such terranes.

24
APPLICABILITY OF RECOMMENDED TRACED-SPRING, -CAVE-STREAM,

AND -WELL MONITORING STRATEGY

The applicability of the traced-spring, -cave-stream, and

-well monitoring strategy and the conventional (randomly located

well) monitoring strategy in various types of aquifers is shown

in Figure 8. Both monitoring strategies may be applicable in

some fractured aquifers, some diffuse aquifers, and some karst

aquifers. It is to be stressed that there are some diffuse-flow

karst aquifers in which the traditional randomly located well

strategy works efficiently and is better (Beck, 1986; Wailer &

Howie, 1988; Benson et al ., 1988).

The recommended monitoring strategy is not applicable universally. But it is applicable in all diffuse-flow and conduit-flow

aquifers that drain to springs which discharge on land or along

the shores of streams, rivers, lakes, or seas. Accordingly, the

strategy is applicable in most karsts of the following 16 states

that have significant amounts of karst: New York, Pennsylvania,

Maryland, West Virginia, Virginia, Tennessee, Georgia (Appalachians), Alabama, Kentucky, Indiana, Arkansas, Missouri, Iowa

Minnesota, Texas, and Oklahoma. Their karst terranes are characterized by local recharge and discharge. Many of these states

include some of the more densely industrialized and populated

areas of the U.S. Karst comprises approximately 25 to 30 percent

the total area of these states.

Figure 8. Tetrahedral continuum between four major types of

aquifers, dominant pore geometry in each type of aquifer, applicability of the traced-spring, -cave-stream, and -well monitoring

strategy in each type of aquifer, and applicability of the

traditional monitoring technique (using randomly located wells)

in each type of aquifer. The direction of increasing hydraulic

conductivity of each aquifer is shown by the arrows. (modified

from Quinlan & Ewers, 1985)

25
The recommended monitoring strategy is applicable only

locally in parts of the Floridan aquifer of Florida and in Puerto

Rico. In both areas there is significant discharge at springs.

The strategy is not applicable in karst terranes that are merely

recharge areas of regional aquifers such as the Upper Floridan

aquifer of Florida, Georgia, and South Carolina. It is also minimally applicable in karsts mantled by glacial sediments and in

which discharge is diffuse, into sediment and/or bodies of water,

rather than at discrete springs. I estimate that these karsts

comprise about 30 percent of the total area of these states.

The strategy would be applicable in most of the Edwards

aquifer of Texas, much of the Upper Floridan aquifer, and part of

the North Coast limestone aquifer of Puerto Rico where flow is to

springs at the surface, and where most springs are diffuse-flow

rather than conduit-flow. Although much of each of these

aquifers is characterized by sponge-like permeability, many of

their springs are fed by conduits that are commonly braided

(anastomosed) (Veni, 1988; Beck, 1986, p. 240; Wilson & Skiles,

1988; Stone, 1989). Accordingly, one might use geophysical

techniques for trying to find the main conduit and drill to

intercept it, but probably it would be missed. Nevertheless,

monitoring could be successfully accomplished at springs. Wells

in these sponge-like aquifers could be used as monitoring sites

only if there were a positive trace to them from the vicinity of

a facility or from the facility itself during low-, moderate-,

and high-flow conditions.

There are numerous small areas of karst in the western U.S.,

but nearly all of them are in isolated, non-industrialized, unpopulated terranes.

Research is needed on the distribution of and criteria for

applicability of the recommended strategy in karsts of the states

cited in the four preceding paragraphs.

REGULATORY ASPECTS OF DYE-INJECTION INTO WELLS

As discussed under the next major heading, dye-tracing

strategy might include proposals for injection of tracers into

wells. Strictly speaking, injection of dye or other non-toxic

tracers into a well, no matter how noble the reason for doing so,

makes it possible to construe the well to be a Class V injection

well (a well that is not included in Classes I through IV and

that generally injects non-hazardous fluid into or above an

underground source of drinking water; 40 CFR Section 144.6(e))

and is thus subject to State and Federal regulations governing

its use. Also, some state agencies have interpreted tracing

agents used in ground-water investigations to be pollutants or

contaminants. If the tracer is toxic, the well would be a Class

IV injection well (a prohibited type of well which disposes of

26
hazardous or radioactive wastes into or above a formation within

1/4 mile (0.4 km) of an underground source of drinking water; 40

CFR Section 144.6(d)).

Well-meaning as the above interpretation of regulations for

injection wells may be, it is not justifiable in terms of

potential benefits for environmental protection, intent of the

law-makers, or risk of exposure to pollutants. Like boats put

into a lake, tracing agents are used in the water for a good and

definite purpose, not put in it for disposal. And like boats,

dyes generally used for tracing ground water are benign and

harmless in the concentrations commonly employed (Smart, 1984).

Tracing agents are fundamental tools for discovery and

prediction of the velocity and dispersal-path of pollutants in

ground water and surface water. Interpretation of data from

tracer studies makes it possible to protect water quality, public

health, and aquatic life. Such data are crucial to the development of wellhead and springhead protection strategies and can be

essential for the calibration of computer models of water flow

and pollutant movement. Tracing is cost-efficient and is often

the only way to get critically needed data.

A further analogy describing the use of tracing agents can be

made. Doctors use vaccines and a wide range of diagnostic techniques to prevent and treat illnesses. Some of these vaccines

and techniques have definite risks associated with their use.

These risks are assumed by an informed patient because the consequences of not preventing or not diagnosing an illness far

outweigh the slight risk from use of the vaccine or diagnostic

technique.

If and when state officials establish regulations governing

the use of dyes or any other ground-water tracer, they should

require their use by knowledgeable, experienced professionals.

Additional discussion of regulatory problems concerning the

use of dyes for tracing ground water, and a recommended solution

to these problems, is given by Quinlan and Field (1989).

Many Federal and State agencies have sanctioned the use of

dye-tracing studies in the study of ground-water pollution and

time-of-travel of pollutants in rivers. Guidance manuals for

tracing techniques exist and have been sponsored by EPA (Davis et

al., 1985) and by the Societe G4010giwe Suisse (Parriaux et al.,

1988) . Updated manuals on ground-water tracing have been written

by the U.S. Geological Survey (under contract to the EPA: Mull et

al., 1988) and are in preparation for the National Water Well

Association (Aley et al ., 1989). Several manuals on the use of

dyes for measurement of discharge, time of travel, and dispersion

in surface streams have been written by the U.S. Geological

Survey (Wilson et al ., 1986; Kilpatrick & Cobb, 1985; Hubbard et

27
al ., 1982).

The following section on dye-tracing and the design of

monitoring systems in karst terranes is written on the assumption

that it is legally permissible to inject dye into wells. Qualifying statements that would be necessary if tracer-injection into

wells were illegal are omitted for clarity.

HOW TO DETERMINE WHERE, HOW, AND WHEN TO MONITOR GROUND WATER

IN KARST TERRANES RELIABLY AND ECONOMICALLY

INTRODUCTION

Where to monitor is typically best determined with the aid of

three types of field investigation: dye-tracing, surveying of

cave rivers that are shown by tracing to drain from a facility,

and detailed mapping of the potentiometric surface.

DYE-TRACING

Selection of Dye-Injection Sites

Under ideal circumstances, one can run the dye-tests necessary for the design of a monitoring system from a perennial

sinking stream on the facility. Often no stream is available for

dye-injection. In that situation, one can use tank-trucks of

water and inject dye at (in decreasing order of desirability) a:

1. Sinkhole with a hole at its bottom.

2. Sinkhole without a hole at its bottom; excavation may reveal a hole that can be used.

3. Losing-stream reach with intermittent flow.

4. Class V stormwater drainage well.

5. Well drilled on a fracture trace or a fracture-trace

intersection.

6. Abandoned domestic, agricultural, or industrial well.

7. Well randomly drilled for dye-injection.

Alternatively, it is sometimes practical (or necessary) to

trace by injection through the drain field of a septic tank.

This is a difficult trace to perform: it should be done with

great care and by experienced personnel.

Before going to the trouble and expense of dye-tests from

28
injection-site types 2 through 7, they should first be given

either a percolation test or a slug test by injection of potable

water or a cylinder of known volume in order to determine if the

well is open to the aquifer and to see how rapidly they drain.

Alternatively, a pumping test can be run. An electric tape or a

pressure transducer can be used to determine the rate of waterlevel decline in a well during a slug test; plots of such data

can be used to select the site that probably has the most direct

and open connection to the conduits that are the aquifer drains.

If the percolation test of an injection site shows little or no

drainage, do not inject dye into it. The procedures for dyetracing with trucked water are discussed elsewhere (Quinlan &

Ewers, 1985, p. 222; Aley et al ., 1989). Wells drilled for dyeinjection should probably extend about 8 m (25 ft) below the

potentiometric surface or upper bed that may confine a karst

aquifer, but this is a site-specific determination.

It is almost economically impossible to design a dye-trace

that simulates the conditions beneath a landfill in a terrane

characterized, for example, by 20 m (60 ft) of residual soil.

Why? Residuum is anisotropic: The distribution of macropores

within it is unpredictable. Their permeability may be several

orders of magnitude higher than that of the bulk of the residuum

(Quinlan & Aley, 1987). The larger macropores, through which

fluids move most rapidly, may be several centimeters to several

meters apart. The problem of intercepting them with a drill hole

intended for dye-injection is analogous to the problem of searching for orebodies with a drill hole; only the size of the targets

and therefore the necessary spacing between holes for the same

probability of target interception is different. In addition, it

is likely that the act of drilling or coring through clayey

residuum, no matter how carefully it is done, will smear many of

the macropores shut, thus masking their presence. Also, a hole

drilled for dye-injection is likely to miss the subjacent conduit

system. A randomly-drilled hole tests only a tiny percentage of

a facility’s surface and subsurface.

The relative suitability of different drill holes in residuum

or bedrock can be evaluated by slug tests, as discussed above.

If the holes are very close to one another it would be prudent to

have an electric tape or a pressure transducer in the adjacent

holes--just to be sure that the rapid fall of the water level in

the tested hole is not a consequence of leakage into one or more

of the adjacent holes.

Instead of using drill holes for dye-injection, one might

excavate a 10 m x 10 m (30 ft x 30 ft) pit to the depth of the

bottom of a proposed landfill in the example terrane, say 5 m (15

ft), and carefully construct either a simulated compacted or

lined bottom of a cell--or try to make an “undisturbed” bottom.

Even then, one could not be sure of having simulated or tested

the long-term permeability conditions at the bottom of the land29
fill. A similar carefully excavated and constructed pit, if dug

above a solutionally enlarged joint in the subjacent limestone,

might leak during the test or several years later. Its failure

could be hastened by leakage accompanied by synergistic subsidence or collapse of soil that bridged the joint.

Whether using a sinkhole, a drill hole, or a pit for dyeinjection, when designing a dye-trace to evaluate a proposed

landfill site, it is necessary to:

1.

2.

Assume that the soil or liner has or will have differential permeability (leakage) that can not be remediated

by economically justifiable construction methods. This

assumption is supported by an extensive literature, as

discussed by Quinlan & Ewers (1985, p. 199). The longterm permeability of any kind of liner may be affected by

chemical changes induced by leachate (Quinlan & Ewers,

1985, p. 199; Hettiaratchi & Hrudey, 1987), aging and

degradation of polymers used in liners (Segrestin &

Jailloux, 1988), and flaws in construction and installation (Rogowski, 1985; Jaywickrama et al., 1988). The

urgent questions about leakage are when, to where and how

fast. [This is consistent with the EPA policy of assuming

a worst possible case scenario.]

Test for the consequences of the leakage that is certain

to occur. The argument that a dye-trace is irrelevant to

evaluation of a particular landfill because the dye was

not injected at the bottom of the actual landfill and

precisely at its location is specious.

Leakage may reasonably assumed to be a certainty. Therefore,

a dye-test should be designed to maximize the probability of getting the dye through the soil or residuum as rapidly as possible.

Few consulting firms or their clients can afford to wait a year

or more for test results that, until the dye is recovered, remain

negative.

The monitoring system for a facility and the consequences of

leakage from it should be tested by tracing from the facility

itself. This may not be possible. If it is not, traces should

then be run from sites adjacent to the facility, preferably from

opposite sides of it, and at points lying on a line approximately

perpendicular to the suspected flow direction. This increases

the probability of discovering if a facility is near the boundary

of a ground-water basin (or well away from a boundary) and

whether the facility consistently drains to the same spring (or

springs). Such tracer results are relevant to objective evaluation of the facility.

Tracing from sites adjacent to a facility is justified by the

principle of hydrologic juxtaposition. This principle is easily

30
explained. If the geology of a dye-injection site is similar to

that of a site immediately adjacent, it is highly probable that

tracer results from the two sites will be to the same spring.

Obviously, this may not be so in the immediate vicinity of the

boundary between two ground-water basins, but it is the reason

why a second test is recommended for the opposite side of a

facility. In fact, a higher degree of confidence would be

achieved for predictions concerning ground-water flow and

contaminant transport if they were based on the results of a

tracer test run from each side of a facility (on or off it)

rather than a single tracer test run from the middle of it.

Selection of Dye-Monitoring Sites

All springs within a radius of perhaps 8 to 25 or more

kilometers (5 to 15 or more miles) from a facility, especially

those within ±90 degrees of the likely vector of the hydraulic

gradient from it, should be found and monitored during dyetraces. At the beginning of an investigation, a prudent designer

of a dye-trace will generally assume the possibility of radial

flow and will have 360 degrees of coverage with dye-detectors--if

only to defend the test design from criticism of inadequacy. The

radius of spring search is determined by evaluating stratigraphy,

structure, and physiography, and by proposing various tentative

hypotheses about possible flow routes and resurgences. These

working hypotheses must be tested for each area.

Dye-traces to springs at the bottom of sinkholes are especially important for recognizing segments of the plumbing system

of a karst aquifer. These segments between a facility and the

spring to which it drains can be used for monitoring. They offer

two advantages over the use of springs: less dilution of pollutants or surrogate compounds, and earlier detection of them.

Imagine how extremely different the tracing results in Figure

6 would appear if the designer of the dye-tests shown had followed a hunch (or perhaps the dip of the beds beneath the facility)

and monitored only the springs in one particular direction from

the landfill rather than in all directions! At the administrative hearing on whether or not to grant a permit for construction

of the proposed landfill, held after dye-test #1 (from a point

adjacent to the site) but before the other five dye-tests were

run, the state’s witnesses alleged that it was impossible for dye

(or pollutants) to flow in opposite directions. They vigorously

but erroneously impugned the validity of the dye-test. Each of

the five subsequent tests resolved all questions about the

alleged impossibility of radial flow.

Prediction of flow within a ground-water basin characterized

by local flow is usually very much like prediction of flow within

a surface-water basin. It moves "downhill" (downgradient) to the

31
trunk drain at local base level. If the boundaries of either

basin are known, it can be confidently stated that although

small-scale local flow from an area (or facility) may be in

almost any direction, macro-scale local flow will be to tributaries and ultimately to the trunk that drains the basin. For

example, in Figure 3, any dye (or pollutant) injected south of

Mill Hole and west of both Park City and the subbasin boundary

shown as a dotted line can confidently be predicted to flow to

Mill Hole. Similar predictions can be reliably made for potential facilities elsewhere in the Turnhole Spring ground-water

basin and for anywhere in Figure 5.

Dye-Tracing Methods

Ground-water tracing can be done with many different tracers,

but in general the cheapest, most efficacious ones are fluorescent dyes such as fluorescein* (CI Acid Yellow 73), Rhodamine WT

(CI Acid Red 388), CI Direct Yellow 96, and optical brighteners

such as CI Fluorescent Whitening Agents 22 and 28. Brief summaries of practical techniques for dye-tracing have been published (Quinlan & Ewers, 1985; Davis et al., 1985; Quinlan, 1981,

1982, 1986a, 1987; Mull et al ., 1988). [A comprehensive, plainEnglish guide to the use of dyes as tracers is planned for

publication during 1989 (Aley et al ., 1989). It will include

practical hints and suggestions that will enhance the rigor of

test-design with any tracer and will increase the reliability of

test results. It will also include discussion of appropriate

QA/QC procedures.] A useful review of the chemistry of dyes has

been published by Zollinger (1987).

Three types of dye-tracing can be used for evaluating the

suitability of springs, cave streams, and wells for ground-water

monitoring. They are:

1. Qualitative tracing, using either of the following:

A. Visual observation of the dye-plume. Generally this is

wasteful of dye and may cause aesthetic and public relations problems. Also, there is great risk of missing

the dye-pulse when it arrives at the monitoring site,

especially at night.

*Fluorescein is the name generally (but erroneously) used in

America for sodium fluorescein (C20Hlo05Na2) which is highly

soluble in water and sometimes confused with fluorescein

(C20H1205) which is insoluble in water. Both dyes are CI Acid

Yellow 73. This water-soluble dye is known in Europe as

uranine. In order to minimize confusion, I follow the American

convention of using fluorescein when I actually mean sodium

fluorescein.

32
2.

3.

B. Passive detection (with passive detectors consisting of

activated charcoal or cotton, depending upon the tracer

used) plus either visual observation of dye eluted from

charcoal or ultraviolet observation of cotton (Aley et

al., 1989; Quinlan, 1981, 1982, 1987).

Qualitative tracing is sufficient for most dye-tests; when

done with passive detectors, it is generally the most

cost-efficient tracing technique.

Semi-quantitative tracing, using passive detectors and instrumental analysis of dye with a filter fluorometer or a

scanning spectrofluorophotometer of cotton or elutant from

activated charcoal (Aley et al ., 1989: Duley, 1986;

Thrailkill et al., 1983; Behrens, 1982, 1987, 1988; VODinh, 1981). Although instrumental analysis can identify

dye-concentrations several orders of magnitude smaller

than those detectable visually, the many variables

associated with changes in spring or stream discharge,

with reaction kinetics of sorption of dye onto passive

detectors, and with elution of dye from charcoal all make

it impossible to precisely quantify the varying concentrations of dye that passed any specific monitoring site

during a given period of time.

Quantitative tracing, using instrumental analysis of dye

in water samples (either grab-samples or those taken with

an automatic sampler) or of water continuously flowing

through a filter fluorometer, preferably (for either

option) with continuous measurement of discharge. Instrumental analysis enables more precise determination of flow

velocity, the breakthrough tune characteristic of a

tracer’s arrival and retardation, and aquifer dispersivity. It also allows calculation of dye-recovery (the mass

balance relation between the amount of dye injected and

recovered). Many inter-well traces are done with this

type of quantification; some evaluations of wells as

potential sites for monitoring can only be done with

inter-well traces (Molz el al ., 1986).

For a given trace, quantitative tracing is the most expensive procedure, but it can give answers not available by

any other technique (Smart & Ford, 1986; Mull et al .,

1988; Smart, 1988a, 1988b; Aley et al ., 1989.

If quantitative tracing results are needed in the design

of a monitoring system--generally, they are not--it is

commonly far more cost-efficient-to first do a qualitative or semi-quantitative study. This eliminates substantial costs of sampling and analysis of numerous sites

to which no dye travels.

33
There is a need to publish the results of tracer tests in

which quantitative field data, rather than laboratory

data, are compared with semi-quantitative field data that

is obtained concurrently.

Each of these three types of dye-tracing is sufficient and

satisfactory for establishing a hydrologic connection between two

points. Semi-quantitative and quantitative tracing techniques

are more sensitive to detection of small concentrations of dye;

for litigation, they are more convincing. Quantitative tracing

techniques are most sensitive to detection of small, temporary

changes in dye concentration, and they are sensitive indicators

of accidental contamination. If quantitative tracing indicates

that nearly all dye is recovered, one can be confident that the

monitoring system will operate effectively.

No matter which of the three types of dye-tracing is used for

an investigation, it is important to avoid two of the more common

mistakes of neophytes: not sampling at enough sites at which dye

could possibly be recovered (generally using passive detectors)

and not sampling long enough. If not enough sites are sampled or

if sampling is stopped too soon after the first positive results,

one would fail to detect dye at the other places to which it also

goes (if, indeed, it goes elsewhere) at either the same velocity

or a different velocity. Also, one would fail to discover if

some of the dye were stored in the epikarst (as discussed subsequently under EXCEPTIONS TO ASSUMPTION #3) and released over a

long period of time. Any of these consequences of inadequate

sampling for dye will prevent discovery of aquifer properties

that adversely influence the adequacy of the design for a

monitoring system. Other procedural errors that can result in

falsely negative results are discussed on pages 50-58 and by

Quinlan, Aley, and Schindel (1988).

A "reliable?" tracer test can be defined as one that is a

product of careful, prudent design and execution that leaves no

significant doubt about its validity. Commonly there will be at

least one background sample taken at all sites monitored for dye.

Such a test unambiguously demonstrates that there is a hydrologic

connection between A and B. There are degrees of reliability.

The ultimate standard in a reliable test would produce a smooth,

well defined breakthrough curve that accounts for 100 percent of

the injected tracer at concentrations that are greatly above well

defined background levels, and tracer identification would be

done with a scanning spectrofluorophotometer. For most investigations, however, such an ultimate standard is not only unnecessary, but also impossible to attain. There is a need, however,

to establish practical standards to ensure the reliability of

tracer tests.

The need for tracing during low-, moderate-, and flood-flow

34
conditions for many studies, but not all, is stressed repeatedly

in this document. Moderate flow is the ideal time for tracing

because generally it can be done efficiently and one can rapidly

obtain a tentative understanding of the hydrology. Low flow is

difficult because of the extreme duration that may be required

and because many flow routes may cease to function. But some of

the more difficult traces to perform are those attempted during

flood flow. There are three reasons why this is so: Tracer

dilution is extreme, and access to monitoring sites may range

from difficult to impossible (because of flooding). Also, one

must have an intimate familiarity with the monitoring site during

non-flood conditions so that, during flood conditions, dyedetectors can be placed in a spring orifice or channel where the

amount of dye contacting the detector is maximal, and where the

detector will not be left high and dry when the water stage goes

down. This can be very difficult to do when a site is under 3 to

8 m (10 to 25 ft) of floodwater and almost unrecognizable.

Dye-Test Analysis

Ideal tracers are conservative. They do not react with soil,

bedrock, or ground water, and they do not undergo microbial

decay. However, most tracers, including dyes, are slightly

reactive and may undergo adsorption-desorption and cationexchange reactions. . Organic pollutants may undergo similar

reactions which affect their rate of migration. This rate is

predictable and correlated with their octanol-water partition

coefficients (Fetter, 1988, p. 397-405; Winters & Lee, 1987);

organic pollutants may also undergo microbial decay. Depending

upon their mobility, pollutants may travel faster or slower than

a dye. Although tracer velocities can be used as a reliable

guide for prediction of flow velocities of pollutants under

similar antecedent moisture conditions, especially in conduitflow aquifers, tracer velocities in diffuse-flow aquifers (as

well as in granular aquifers) will be significantly higher than

those of most pollutants. In conduits, where reactions of both

tracers and pollutants with the rock matrix will be minimal,

there will be less difference in their velocities.

Interpretation of dye-recovery curves from quantitative

tracer tests can yield much information about the nature of

ground-water flow in a karst aquifer and the structure of its

conduit system, as discussed by Brown and Ford (1971),

Maloszewski and Zuber (1985), Zuber (1986), Lepiller and Mondain

(1986), Gaspar (1987), Mull et al. (1988), Quinlan (1988b), and

Smart (1988a, 1988b). Similarly, much can be learned from the

interpretation of discharge hydrography of springs, as shown by

Wilcock (1968), Brown (1972), Sara (1977), Podobnik (1987),

Meiman et al . (1988), White (1988, p. 183-186), and others.

Mull et al . (1988) have described an empirical technique for

35
studying the travel time of surface streams to the study of cave

streams and spring discharge. The application is more relevant

to spill-response rather than to monitoring of a facility, but it

is very useful. It makes possible a good approximation of the

time of travel, peak concentration, and flow duration of contaminants accidently spilled into a karst aquifer and flowing to a

spring or well; it does so for various discharge conditions. The

technique is a powerful predictive tool for the protection of

water supplies, but it is most judiciously employed after most of

the boundary of a ground-water basin or wellhead protection area

has been determined by dye-tracing. It is not applicable to monitoring of continuous leakage from a facility. The implicit assumptions of their practical application of dye-recovery analysis

are discussed by Quinlan (1988b).

If dye-tracing is needed in order to respond to a spill (to

determine, for example, that it drains to a given spring that

must be monitored for assessment of the long-term effects of the

spill) , then laboratory investigation is needed to determine the

effects of the spilled material on the tracer to be used. This

should be done before the tracer test. For example, the spilled

material could react with the tracer dye, causing it to form a

non-fluorescent compound or causing significant quenching of the

dye's fluorescence. Either reaction could cause trace results

that could be falsely negative. Limited data have been published

on such effects, except for chloride ions and chlorine (e.g.,

Smart & Laidlaw, 1977). A filter fluorometer would be inadequate

for such an investigation; synchronous scanning with a spectrofluorophotometer might identify traces of the dye or detect

exotic organic compounds that might be formed. [This instrument

is also known as a scanning spectrofluorometer, a scanning spectrophotometer, and a fluorescence spectrometer.]

Although it has no bearing on proving if a site will or will

not leak, rigorous tracing protocol requires that the design of a

test include determination of tracer background at all tracerrecovery sites. This determination will influence selection of

the tracer to be used and its quantity. Acquisition of background data is good protocol in any scientific investigation, but

it is also highly desirable if there is any potential for litigation involving the site.

It is sometimes possible to detect with a fluorometer or a

spectrofluorophotometer trace quantities of what appears to be a

green dye in background samples. The background can be derived

from dyes in various foods, household products, antifreeze,

crack-detection penetrant, etc., but such background is extremely

rare, except in urban areas and near leaking landfills. It is

easy to identify the compound(s) contributing to such background.

A synchronous scan of the samples makes it possible to differentiate between various dyes. Coloration of foods and various

products is imparted by mere trace concentrations of dyes, quant36
ities that are usually four or more orders of magnitude smaller

than commonly used in tracing ground water.

Agricultural runoff may also contribute to background for

compounds with green emission. Jones (1989), using a filter

fluorometer, has discovered green emission in a water extract of

horse manure. This green fluorescence, when sensed with a

fluorometer, could be mistaken for fluorescein. I have used a

spectrofluorophotometer and found similar fluorescence in cow

manure, but these background values are trivial and equivalent to

only a few parts per trillion (ppt) of fluorescein.

It should be noted that the green fluorescence peak of cow

manure in water (518 nanometers [rim]) does not correspond to that

of a water extract of crushed grass (509 nm) or that of fluorescein in water (513 rim). An alcohol extract of crushed grass

(saturated solution of KOH in a 70% solution of isopropyl

alcohol, a solution commonly used for eluting dye from activated

charcoal) also fluoresces weakly in green, but strongly in redviolet. This is because the chlorophyll, which fluoresces redviolet, is highly soluble in alcohol and relatively insoluble in

water. Manure and grass is mentioned here in order to establish

that their probable contribution to background in dye-tests is

detectable, but trivial. Nevertheless, this is a topic worth

investigating in moderate detail.

.

Qualitative evaluation of background cannot be relied on for

distinguishing between various green dyes. Also, investigators

with little experience in qualitative identification of dyes can

be easily fooled by green algae and exotic hydrocarbons.

Fluorometric analysis will distinguish between background

samples which include only a green dye such as pyranine (CI

Solvent Green 7; D&C [Drug and Cosmetic] Green No. 8;

fluorescent) and those samples that consist of a common mixture

of a blue dye such as Brilliant Blue FCF (CI Acid Blue 9; FD&C

[Food, Drug, and Cosmetic] Blue No. 1; non-fluorescent) and a

yellow dye such as tartrazine (CI No. 19140; FD&C Yellow No. 5;

non-fluorescent). Dye nomenclature and other dyes are discussed

by Quinlan and Smart (1977), Quinlan (1989b), Aley et al . (1989),

Marmion (1984), Zuckerman and Senackerib (1979), and Zollinger

(1987). A distinction between fluorescein and various other

fluorescent green dyes cannot be made reliably with a filter

fluorometer because their emission spectra overlap one another; a

scanning spectrofluorophotometer must be used (Duley 1986, Aley

et al., 1989).

Success in tracing to randomly located wells that are to be

regularly sampled during a dye-test can be maximized if they are

pumped continually to discharge at a low rate, say 4 l/rein (1

gpm), through a passive dye-detector (such as cotton or activated

charcoal) which is regularly changed once or twice a week for

37
weeks or months. In most settings this small amount of water can

be wasted onto the ground at a reasonable distance from any

building or structure with no adverse effect. Pumpage at high

rates, say 400 l/rein (100 gpm) or more, may distort the flow

field near a well, which is acceptable if recognized. A high

pumpage rate for weeks to months is expensive and wasteful;

furthermore, disposal of the pumped water can be a problem,

especially if it is contaminated. Selection of the rate of pumping must be based on knowledge of what constitutes a low rate in

the aquifer being investigated.

Although a dye-test is like the birth of a baby--no matter

how many men and women are put on the job, it will take however

long is necessary to complete the task--it is also quite

different. The birth of a baby can confidently be predicted to

probably occur about 9 months after conception. A dye-test, however, may be completed within a few hours or days after injection, but it could as easily be weeks, months, or even years.

One must be patient while waiting for tracing results, or risk

malpractice and loss of valuable data. It might take just a few

weeks to do the dye-traces necessary to design and test the

monitoring system for a facility, but it is likely to require 6

to 9 months of intermittent, careful tracing. Warning: it could

take even longer.

I believe that one well-designed tracer test, properly done

and correctly interpreted, is worth 1000 expert opinions . . . or

100 computer simulations of ground-water flow. The only disagreement that colleagues have expressed with this statement is

to jocularly suggest that the two numbers should be reversed.

In dye-tracing, wisdom is knowing what essential questions

need to be asked, and asking them; experience is knowing the most

expedient, most prudent way to get the answers to these essential

questions.

The dye-tracing strategies discussed in this document are

also applicable to delineation of wellhead and springhead protection areas (Quinlan, Aley, and schindel, 1988; Office of Drinking

Water, 1987, EPA, 1988). They should be used.

SURVEYING OF CAVE RIVERS

The surface location of a drill hole sited to intercept a

known cave stream can be determined by any of many conventional

surveying techniques using a transit, compass, tape, or electronic distance-measurement equipment. Such surveys should

include closed loops. It is easier and more accurate, however,

to use low-frequency electromagnetic induction equipment, a socalled “cave radio”, which transmits a signal to the surface from

an accessible cave passage with a stream (Reid, 1984; Cole,

38
1988). Such equipment has been used successfully in the Mammoth

Cave area to drill to cave streams at depths as great as 143 m

(470 ft). The horizonal error in one well drilled to 41 m (135

ft) was only 18 cm (7 in); it was attributed to deviation of the

drill hole rather than to error in the survey.

If cave rivers are not accessible from the surface, geophysical studies such as mapping of streaming potential of descending

waters and mapping of acoustic emissions from cave rivers (Lange,

1988; Lange & Quinlan, 1988) may follow mapping of troughs on the

potentiometric surface. Such geophysical studies should precede

drilling of wells planned to intercept cave rivers which are

inferred from the potentiometric troughs and which are to be

tested by tracing for a connection to a facility.

MAPPING OF POTENTIOMETRIC SURFACE

The potentiometric surface of a karst aquifer should be

mapped with as many control points (water levels) as possible.

For basin analysis, a minimum of 1 well per square kilometer (2.5

wells per square mile) is recommended for most aquifers. A

facility analysis could require more than 40 times this well

density. A carefully contoured potentiometric map, if based on

valid measurements of an adequate number of wells, can be used

to:

1. Predict flow routes of dye (or pollutants);

2. Judiciously select dye-injection sites;

3. Minimize the number of dye-traces necessary for definition

of boundaries of a ground-water basin or evaluation of a

facility;

4. Interpolate flow routes in the areas between dye-traces;

and

5. Detect the possible influence of shale beds and other low

permeable rocks on perching and confinement of water

within or below a karst aquifer.

Water levels for mapping the potentiometric surface can be

measured with an electric tape or an acoustic well probe (Quinlan

& Ewers, 1985; Quinlan, 1981, 1982), but such measurement must

include a QA/QC program using a vertically hanging steel tape as

a standard before, after, and preferably during periods of use,

especially if water levels more than 30 m (100 ft) below the

ground surface are being measured to the nearest tenth of a foot

(0.03 m). Permanent tape-stretch of as much as 2 percent occurs

in electric tapes employing copper twin-lead wire similar to

those used for TV antenna lead-in. Whenever possible, all water39
level measurements should be made with the same electric tape--or

each tape should be calibrated against a standard. Measurements

should be made of water levels unaffected by pumping, and during

low flow or base flow, rather than during the rainy season or

after storms when water levels can locally be significantly higher and the potentiometric surface can have a different configuration.

It is logical, correct, and conventional to interpret the

flow direction of ground water to be perpendicular to the potentiometric contours and downgradient. Sometimes, however, flow

lines appear to be parallel to the contours rather than perpendicular to them, as has been demonstrated in the Edwards aquifer

in Texas (Maclay & Small, 1984; Waterreus & Hammond, 1989). The

flow lines are actually perpendicular, but a lack of sufficient

well data in areas characterized by extreme heterogeneity in

aquifer properties precludes demonstration of such orientation.

Mapping of the potentiometric surface does not eliminate the

need for dye-tracing nor does it not replace tracing. The two

techniques are complementary, but data from a well-designed,

properly executed, and correctly interpreted dye-trace or series

of traces are less ambiguous than a potentiometric map that is

based on water-level data.

IMPLICIT ASSUMPTIONS OF RECOMMENDED TRACED-SPRING, -CAVE-STREAM,

AND -WELL MONITORING STRATEGY, WITH EXAMPLES OF EXCEPTIONS

Ground water has been recognized to circulate in three

different types of flow systems: local, intermediate, and regional (Tbth, 1962; Freeze & Witherspoon, 1967), as reviewed by

Fetter (1988, p. 217-258). Most monitored flow is in local-flow

systems. Indeed, this is the flow system most practical to

monitor and most needful of monitoring.

Many implicit flow-system assumptions will be made by those

who may use the monitoring strategy advocated here and by Quinlan.

and Ewers (1985), but they should realize what these assumptions

are. The accuracy of the first eight of the following ten

assumptions is affected by the hydrogeology of the karst beneath

and adjacent to a facility. Assumption #9 involves logic, and

#10 is influenced by the rigor of the design and execution of

tracing tests--as well as by one’s understanding of the first

eight assumptions and one’s ability to recognize the exceptions

to them. The major implicit assumptions, stated axiomatically,

are:

1. Ground-water discharge is concentrated at a point (a

spring or group of springs) rather than diffused over a

broad area or concentrated along a line (such as a

stream).

40
2. Most flow systems to be monitored in karst terranes are

characterized by local flow, in the sense of T6th (1962)

and Freeze & Witherspoon (1967).

3. Ground-water flow velocities in karst terranes characterized by local flow are the high values already cited

herein. Flow velocities in karst terranes characterized

by intermediate-flow systems and by regional-flow systems

tend to be several orders of magnitude slower than in

local-flow systems.

4. A ground-water basin is a discrete entity having a specific, well-defined boundary.

5. Ground-water basins are contiguous.

6. All the discharge of a spring is from the same groundwater basin.

7. Storm-related diversion of ground water out of a basin, if

it occurs, is via intermediate- and high-level overflow

routes (conduits) leading to adjacent ground-water basins.

8. Temporary, storm-related diversion of surface water within

a karst ground-water basin is not a significant problem

because all water will remain within the ground-water

basin.

9. A positive trace from a facility proves that it is a

source or the source of pollutants discovered at the dyerecovery site.

10. The tracing tests done for investigation of a facility

were properly designed and executed and correctly

interpreted.

Each of the first eight of these implicit assumptions is

correct about 95 percent of the time--often enough to be fairly

assumed until or unless data imply otherwise, but not so often as

to be a certainty. One must always ask, “What are my conceptual

models of the flow system and their assumptions?" and critically

review the validity of each. Systematic analysis and review of

examples of probable exceptions to the above assumptions could be

the subject of another document. Only one or two exceptions to

each of the first eight is cited and briefly discussed below,

along with the relevance of these exceptions to a monitoring

program.

The ninth assumption is patently false. it is a flaw in

logic. A positive trace proves that the tested facility may be a

source or the source of some or all of the pollutants but not

41
that it is a source or the source. All the pollutants could have

come from one or more other facilities in the same ground-water

basin. Proving that pollutants could be derived from a facility

is not the same as proving that they are derived from it. Nevertheless, if there is no other plausible source for the pollutants

in a ground-water basin, the tested facility may be--and probably

is--the source of them. For example, gasoline found in a spring

could have leaked from any of 50 service stations within a basin,

each of which could be shown by tracing to be the possible

source. Tank tests, product audits, soil-gas analysis, suction

lysimeters, and contaminant fingerprinting could be used to find

the probable source (or sources) of gasoline. In contrast, in

the average ground-water basin there are relatively fewer

possible sources of an exotic organic compound. Conversely, if

well-designed, properly executed, and correctly interpreted set

of tracer tests conducted during both low-flow and high-flow

conditions showed that a facility drains to a basin other than

the one drained by a contaminated spring, there is sufficient

proof that the facility is not the source of the contaminants.

Again, this is simple logic, not hydrogeologic subtlety.

The tenth assumption, like the others, is made honorably, but

it is the most insidious. It will be discussed after the following exceptions to the first eight assumptions are cited and

described.

EXCEPTIONS TO ASSUMPTION #1: NONPOINT DISCHARGE

Although some discharge in the karsted dolomite of the Door

Peninsula, Wisconsin, is to springs (Wiersma et al., 1984) that

can be considered as point discharge, most of the discharge is

through sediment and over a broad area beneath Green Bay and Lake

Michigan (Bradbury, 1982; Cherkauer et al ., 1987). Discharge

from a karst aquifer through sediment over a broad area occurs in

the Cafio Tiburones area north and west of Barceloneta, Puerto

Rico; it occurs along a line in the valley of the Rio Grande de

Manati, south of Barceloneta. These Puerto Rican terranes are in,

Tertiary limestones of the alluvium mantled and paludal sediment

mantled shallow aquifer on the north coast; the aquifer seems to

be characterized by diffuse flow, but springs and conduit flow

are locally important.

Monitoring in terranes characterized by areal discharge can

best be done with randomly located wells, perhaps along fracture

traces, but they might not intercept the relevant flow lines.

Monitoring in terranes characterized by seepage along a line is

best done by sampling at intervals along the line and upgradient

from it. Although ground-water flow velocities may be very low,

tracing must nonetheless be done in each hydrologic setting if

confidence in a monitoring effort is desired. If a well proposed

for monitoring a facility in a karst terrane does not have a

42
positive trace to it (or to its site), it is not a monitoring

well for ground water draining from the facility.

Discharge along the length of a stream is an example of nonpoint discharge. Smart (1988b) has described this in the Maligne

karst aquifer in the Rocky Mountains of Canada. Another example

of discharge along a line occurs in the eastern Snake River

Plains aquifer in Idaho (Lindholm, 1986; Wood and Low, 1988).

Admittedly, this basalt aquifer is not a karst aquifer, but it

was for the hydrology of such rocks that the term pseudokarst was

first proposed more than 80 years ago by von Knebel (1906, p.

182-183). Many of the monitoring principles advocated herein may

be applicable to monitoring in this basalt aquifer and in other

highly fractured rocks.

EXCEPTIONS TO ASSUMPTION #2: NONLOCAL FLOW

There are many examples of non-local flow (intermediate flow

and regional flow) that must be monitored in karst aquifers.

Some of the better known examples of such karsts are the Edwards

aquifer (Maclay & Small, 1984; Campana & Mahin, 1985) the Great

Basin carbonate aquifer (Mifflin & Quade, 1988; Fetter, 1988, p.

233-237); and the Floridan aquifer (Fetter, 1988, p. 237-243,

359-361; Miller, 1984; Beck, 1986). Ground-water flow velocities

in them are likely to be much slower than in most conduit-flow

aquifers and more like the low velocities characteristic of most

diffuse-flow aquifers. Flow-paths are likely to be braided

(anastomosed) and dispersive rather than convergent, but most are

still to springs.

EXCEPTIONS TO ASSUMPTION #3: SLOW MOVEMENT IN LOCAL-FLOW SYSTEM

The epikarst, briefly described in the introduction, is a

zone of karstification below the soil profile and above the

phreatic zone; horizontal flow is dominant and storage is

significant (Quinlan & Ewers, 1985: Smart & Hobbs, 1986: Bonacci,

1987, p. 28-35; Williams, 1985; Friederich & Smart, 1981, 1982;

Smart & Friederich, 1986; and Ford & Williams, 1989). Commonly,

it is 3 to 10 m (10 to 30 ft) thick. Although tracer studies in

a British epikarst have shown that vertical flow velocities (to

caves below) locally exceeded 100 m/hr (330 ft/hr), dye was still

detectable in the epikarst 13 months later (Friederich & Smart,

1981, 1982; Smart & Friederich, 1986). Similar results were

found by Even et al . (1986) in an Israeli epikarst. This flow

dichotomy is actually a continuum, but it suggests that the

common adage that ground-water pollution in karst areas is not a

long-term problem because the aquifer is rapidly self-cleaned, is

wrong--or at least unreliable. The duration of retention of a

pollutant in a karst terrane is a function of soil thickness and

sorptive capacity, the efficiency of connection between the

43
ground surface and the epikarstic bedrock, and the efficiency of

flow between the epikarst and the phreatic zone. Indeed, if one

is “lucky” and the volume of a spill on a thick soil is minimal,

all or most of the spilled substance may be retained and/or

sorbed by the soil. Aquifer remediation, therefore, may be

limited to excavation of the soil.

Most dyes are injected into sinking streams, sinkholes, or

wells which bypass the soil and epikarstic zone. In contrast,

contaminants are rarely injected purposely and directly into a

karst aquifer. Pollutants in the epikarst may be monitored using

traditionally sited, randomly located monitoring wells at the

point source to be monitored and they may obey Darcian flow laws,

but such a monitoring system may not intercept them. Nevertheless, contaminants may actually be detected and recovered with

traditional or innovative techniques before they enter the

conduits of a flow system. Once a contaminant has entered the

subjacent conduits and solutionally enlarged joints, Darcian flow

conditions usually do not apply; detection and recovery are

difficult and the protocols recommended in this document are

applicable.

EXCEPTIONS TO ASSUMPTION #4: FUZZY AND OVERLAPPING BASIN BOUNDARIES

It is quite logical to assume that a ground-water basin has a

specific, well-defined boundary. Sometimes, however, the boundary temporarily migrates or ceases to exist during response to

moderate- to flood-flow conditions; some water is diverted to

adjacent ground-water basins, as already discussed. Nevertheless, this assumption is reasonable and generally correct; it can

be extended to allow for temporary, slight shifts in the boundary

between two basins, also in response to storm-related changes in

flow conditions, In contrast, some diffuse-flow aquifers show

significant exceptions to this implicit assumption; the basin

boundaries may be nebulous and gradational during all flow

conditions. For example, all dye injected near the center of a

ground-water basin in the Great Oolite Limestone (in the Bath

district of southwestern England) flows to only one spring.

Successive tests at increasingly greater distances from the

central axis of the basin show that less dye goes to its spring;

the balance of the dye goes to the spring draining the adjacent

ground-water basin. [As the dye-injection point gets successively farther from the central axis, more and more dye goes

to the major spring in the adjacent basin.] Such fuzzy

boundaries are characteristic of basins in this aquifer. Smart

(1977) has recommended that the gradational “boundary” between

any two such ground-water basins be chosen to coincide with

tracer injection points from which the dye divides evenly between

the adjacent springs. Similar results occur elsewhere in England

(Atkinson & Smart, 1981) and probably in other places where

44
diffuse flow predominates. Nebulous, gradational boundaries

between ground-water basins have recently been recognized in

Missouri (Thomas Aley, Ozark Underground Laboratory, Protein,

Missouri, oral communication, 1987) and undoubtedly will be

recognized in other karst terranes of the United States.

Although Smart’s 50 percent boundary criterion is reasonable

for hydrologic budgeting (Smart, 1977), for the delineation of a

wellhead or springhead protection area it is necessary to know

the entire area that contributes recharge. Therefore, for some

hydrologic studies, his valid criterion must be ignored.

Gradational basin boundaries are suggested by the results of

dye-injection into a swallet (a sinkhole into which a stream

empties) at the west boundary of the Turnhole Spring ground-water

basin (Figure 3, on the 520-ft contour). This swallet drains to

two different ground-water basins; it flows both to the northeast

(within the Turnhole Spring basin) and to the west, to a second

major drainage basin (Quinlan, Ewers, et al ., 1983, p. 48-49).

This particular boundary of the ground-water basins has not been

studied, but I believe that the extent of overlap is less than

half a kilometer.

One should be prepared to encounter fuzzy and overlapping

boundaries between ground-water basins. The possible existence

of such boundaries makes it necessary to be extremely thorough in

the design of dye-tracing investigations and confirms the already

recognized need for monitoring for dye at springs in ground-water

basins adjacent to a proposed facility (Aley et al ., 1989).

Overlap of ground-water basins is convincingly illustrated by

the Bear Wallow basin in Kentucky (Quinlan & Ray, 1981). It

occupies 500 km2 (190 mi2) and its three subbasins, Hidden River,

Uno, and Three Springs, resemble a Venn diagram (a diagram

employed in symbolic logic; it uses circles and their relative

position to represent sets and their relationships (Gardner,

1982; Miller, 1986; Edwards, 1989). The Hidden River and Uno

subbasins occupy 80 percent, and 35 percent, respectively, of the

total basin. The paradox of these two subbasins totaling more

than 100 percent of the area of the Bear Wallow basin is explained by the fact that the Three Springs subbasin comprises 15 percent of the entire Bear Wallow basin and is part of the headwaters of each the other two, subbasins; it is common to each of

them. This overlap is significant because the 97 km2 (37 mi2) of

the Uno basin is the size of the area alluded to earlier in this

document as the terrane in Figure 4 from which pollutants could

flow to a total of 52 springs in 19 isolated segments along a 19

km (12 mi) reach of Green River. Probable consequences of leakage from a facility would be significantly fewer if it were

located someplace other than in a ground-water basin in which the

overlap of the headwaters is analogous to a Venn diagram and

discharge is via distributary flow. This overlap is also signif45
icant because adequate monitoring of a facility in the Uno subbasin would require more sites (and more expense) than monitoring

the same facility in a basin that was not part of two other

basins.

Similar overlap of recharge areas of springs occurs also in

the Mendip Hills of southwestern England (Drew, 1968, 1975; Atkinson, 1977).

EXCEPTIONS TO ASSUMPTION #5: NONCONTIGUOUS GROUND-WATER BASINS

Thrailkill and his students have shown that there are two

physically distinct spring types in karst of the Inner Bluegrass

of Kentucky: local high-level springs discharging from shallow

flow paths and major low-level springs discharging from a deep,

integrated conduit system (Scanlon & Thrailkill, 1987; Thrailkill, 1984, 1985). The major low-level springs are characterized

by larger catchment areas (>10 km2; 4 mi2

) and higher discharges

(10-2700 l/see; 2.6-700 gal/see); the local high-level springs

are characterized by smaller catchment areas (<2 km2: 0.8 mi2

)

and lower discharges (0.1-0.8 l/see; 0.025-0.2 gal/see). The

lack of integration of the local high-level spring catchments

into the major low-level spring catchments can be explained by

the impermeability of numerous interbedded shales and the lack of

fractures passing through them. The catchment of each major lowlevel ground-water basin is interpreted to be near-elliptical,

but there is minimal data to support this shape. Each of the

catchments is unrelated to surface drainage, isolated from the

nearby major basins, and commonly separated from its nearest

similar neighbor by 1 to 4 km (0.6 to 2.5 mi) , thus making them

noncontiguous. The reasons for the noncontiguity of these major

basins are not yet understood but may be related to the nature of

the epikarst between conduits and the inhibition of hydraulic

integration by clay and shale in the bedrock (Ralph O. Ewers,

Department of Geology, Eastern Kentucky University, oral communication, 1988).

Noncontiguous basins within a karst terrane in which the beds

dip uniformly at a low angle also occur in the Nashville Basin,

Tennessee (Geary M. Schindel, ATEC Environmental Consultants,

Nashville, Term., oral communication, 1989); they probably exist

elsewhere. The possibility of their occurrence in various

settings should be anticipated and can be detected when traces

are run from each side of a suspected boundary of a ground-water

basin.

It is possible that traditional, randomly located wells may

provide effective monitoring in the interbasin areas between noncontiguous basins. To use these wells reliably in such a karst,

however, the tracing procedures advocated herein must first be

employed and competently shown to yield negative results at

46
springs and positive results at the wells.

EXCEPTIONS TO ASSUMPTION #6: SPRINGS THAT DERIVE SOME OF THEIR

FLOW FROM MORE THAN ONE BASIN

If one or more tributaries to the trunk conduit of a groundwater basin join it at a distance upstream from the orifice that

is equal to or greater than the length of the mixing zone for a

given velocity and difference in water quality (Fischer et al .,

1979, p. 105-147; Hubbard et al ., 1982; Kilpatrick & Cobb, 1985;

Mull et al ., 1988, p. 43-45), monitoring of water quality at its

spring during base flow will not detect the existence of the

subbasin, even if there was a significant difference between the

chemical and physical properties of the two basins or subbasins.

But if the confluence of a basin or subbasin is at less than the

length of the mixing zone for a given velocity and difference in

water quality--or if the confluence is at the surface, possibly

because erosion and slope-retreat progressed upstream--a series

of measurements of a physical or chemical property such as

specific conductivity may show variation across the width of the

spring's stream channel. Many examples of this dual-basin

phenomena are known. They include:

1. Seven Springs, in Fillmore County, Minnesota. Two

distinct ground-water basins discharge from the same

conduit orifice. Dye from one basin discharges from the

left side of the spring, but not the right; dye from the

other basin is discharged on the right side, but not the

left (Mohring and Alexander, 1988). At Seven Springs,

however, the discharge of water from the second basin

occurs only during flood flow, not during base flow.

Seven Springs is part of a distributary.

If water quality of the two basins or subbasins is

similar, the only sure way one could discriminate between

them--or discover the existence of the second is to

measure dye concentrations on both sides of a channel and

in its middle. Prospecting for variations in water quality or dye concentration as an indicator of the nearby

confluence of a subbasin must be done under low-,

moderate-, and flood-flow conditions. It must be shown

that these differences in dye-concentration vary with

position, not with time. [Meiman et al . (1988) have shown

that continous monitoring at the same point during flood

flow detects significant short-term variations in the

quality of waters that have not been mixed. This is

interpreted to be a consequence of piston-flow from

tributaries.]

2. St. Dunstan’s Well, in the Mendip Hills of southwestern

England. The flow of a spring-fed stream is from two

47
adjacent orifices, nly 2 m (6 ft) apart. Water quality

of one of the springs is significantly different from that

of the other. One is usually supersaturated with respect

to CaC03; the other is usually undersaturated. Tracing

has shown that each spring has a separate recharge area

but a partial overlap of these areas is common to both

springs (Drew, 1968, 1970), like the Venn diagrams mentioned previously.

3. Aquas Frias Spring, on the west bank of the Rio Grande de

Manati, Puerto Rico. A dye-trace has shown that a major

tributary joins the inclined throat of the spring that

rises from a depth of 27 m (90 ft). This tributary joins

the throat that is the trunk passage at a depth of 9 m (30

ft) below the spring’s surface. Most of the discharge is

from the trunk passage that is part of the Rio Encantado

Cave System, more than 19 km (12 mi) long (Wes C. Skiles,

Karst Environmental Services, High Springs, Fla., oral

communication, 1989).

The occurrence of dual-basin flow has been attributed above

to be a consequence of distributary flow. As such, it may be the

reoccupation of high-level flow routes that ceased to function

perennially as base level was lowered. Dual-basin flow can also

be a consequence of a spring location being base level for a

large area that was never integrated into a single ground-water

basin.

Of what relevance is this seemingly nit-picking discussion to

the monitoring of ground water? If prospecting a channel for

position-related and time-related variations in specific conductivity or dye-concentration during constant-rate injection shows

that the variations are position-related, one can select the best

place in the channel to monitor for dye and/or pollutants. Such

data may also enhance the probability of more accurate delineation of the recharge area for each subbasin.

There is another continuum here. If the distinct orifice of

two or more spring outlets feeding a stream can be seen at or

from the ground surface, I would call the recharge area of each a

separate basin, but with a common discharge point. If the distinct orifices can not be seen at the surface and if the existence and position of underground confluences must be inferred

from measurements of water quality or position-related dyeconcentration, or by diving, the recharge area of each is a

subbasin. St. Dunstan’s Well represents on end of this

continuum; Seven Springs is near the same end of it. Aguas Frias

Spring is farther from that end. The opposite end of the

continuum is exemplified in Figure 3 by the confluence between

the Mill Hole subbasin (3A in the index map) and the Procter Cave

subbasin (3B, 3B1, and 3B2). The full spectrum of this dual- or

multi-basin continuum is represented by various karst springs in

48
the Hot Springs area of South Dakota (Alexander et al ., 1988).

EXCEPTION TO ASSUMPTION #7: DIVERSION OF GROUND WATER TO THE

SURFACE

The Poorhouse Spring ground-water basin in Kentucky nicely

illustrates the diversion of ground-water flow to a surface water

basin (Quinlan & Ray, 1981, 1989). The basin drains 70 kii,2 (27

mi2) and most flow is to the southeast, through the 3 x 6 m (10 x

20 ft) trunk stream passage that is Steele’s Cave. After very

heavy rains (approximately once a year), some of the water in

this trunk rises about 21 m (70 ft) above its normal level, out

of the sinkhole entrance, and overflows onto the surface. From

there it flows in the opposite direction to the northwest and

west where it augments the flow of a surface stream which drains

an area outside of the Poorhouse Spring ground-water basin. Such

flow, if contaminated, could give spurious values at surface

sites used to monitor pollutants or background in adjacent

ground-water and surface-water basins.

EXCEPTION TO ASSUMPTION #8: DIVERSION OF SURFACE WATER TO A

DIFFERENT GROUND-WATER BASIN

Sometimes surface waters overflow during storms from one

ground-water basin to another. This is illustrated by the behavior of Cayton Branch in Kentucky (Quinlan & Ray, 1981, 1989;

Quinlan, Ewers, et al., 1983, p. 22). Little Sinking Creek (the

southwesternmost surface stream in Figure 3) drains north to the

Green River and contributes to the discharge from the Turnhole

Spring ground-water basin. At a point about 2 km (1.2 mi) south

of the creek’s northernmost swallet, where the south fork of the

stream bends north at the 600-ft potentiometric contour, the

creek goes out of its banks during floods and diverts some of its

discharge westward. The diverted surface water flows about 600 m

(2000 ft) west to a swallet that comprises part of a larger

ground-water basin that drains westward to the Barren River, and

in which the headwaters have been captured by the Turnhole Spring

ground-water basin.

This diversion of surface water from the Turnhole Spring

basin to another ground-water basin could be relevant to a monitoring effort in the second basin if there were significant

quantities of pollutants in the surface waters of Little Sinking

Creek and if they were “exported” to the second basin where they

could adversely affect the reliability of data from cave streams

used for monitoring of background.

The above exceptions to the eight implicit assumptions stated

as hydrologic axioms are uncommon but important. Their existence

justifies a thoroughness in the design of tracer tests and in the

49
interpretation of tracer results that, to some people, might seem

paranoid. Their existence, however, emphasizes the need for

facility-related field work as a prerequisite for the design of a

reliable monitoring system in a karst terrane.

ASSUMPTION #10, THE MOST INSIDIOUS

Introduction

Tracing for evaluation of a facility is easy to do: All one

does is inject tracer such as a dye at one point and recover it

at another. so, too, is well drilling: All one does is set up

the rig, start the motor, and count the money. Both techniques

yield good, reliable results when performed by those trained in

their art. But there are so many things that can go wrong when

trying to do either that it is more prudent and cost-efficient to

have each done by an experienced professional.

It is easy to obtain falsely positive results in tracing--

chiefly as a result of contamination that can be prevented by a

proper QA/QC protocol. [The same can be said for errors in chemical analysis or almost any other analytical technique.] It is

even easier to obtain falsely negative results. This is why I

have called the naive and sometimes erroneous belief that tracer

results and their interpretation are valid “the most insidious

assumption”.

Procedural Causes of Falsely Negative Results of Tracer Test

Results

In addition to occurrence of negative results as a result of

any of the eight hydrogeologic assumptions discussed above, there

are many procedural ways to inadvertently get a false negative in

a tracer test and thus “prove” that leakage of harmful materials

from a facility will not (or does not) occur. Thirteen of the

more common ways are:

1. Inadequate field survey to locate springs or wells to be

monitored for tracer.

2. Injection of dye in a well or at another site that has

minimal hydraulic connection to the subjacent aquifer.

3. Sampling in only one or two directions from an injection

site-- rather than in all directions when radial or multidirectional flow is possible, as described in the

discussion of Figure 6.

4. Sampling at only a few sites--rather than at all sites

possible for recovery of tracer.

50
5.

6.

7.

8.

9.

10.

11.

12.

13.

In

Not sampling often enough to detect the tracer in either

grab samples or on activated charcoal that becomes loaded

with other organic compounds before the dye arrives.

Premature cessation of the tracer test--before there is

enough time for the tracer to reach any monitoring site or

those sites that would be reached after recovery of tracer

at the first site. (Premature cessation is most common

during the dry season, when flow velocities are slowest.)

Sampling only at randomly-located drill holes--rather than

at springs, cave streams, or wells that become turbid

after heavy rains and wells drilled on photolineaments.

Use of an inadequate amount of tracer--an amount so small

that it is likely to be diluted or sorbed to concentrations below the limit of detection.

Use of a tracer inappropriate for the system under study,

one that is likely to be totally or greatly sorbed by

sediment or rock through which it passes, one that is

reactive with a dissolved constituent in water, or one in

which the fluorescence is quenched or enhanced by a

dissolved constituent or by pH.

Use of too high a value for the acceptable threshold of

detectability of dye.

Use of organizations and individuals inexperienced in the

design, execution, and interpretation of tracer tests.

Sampling for dye at only one side (the wrong side) of a

spring with dual-basin flow.

Placement of a dye-detector where contact with spring

water or stream water is not continuous for the time

between placement and recovery.

more detail, the procedural errors likely to yield falsely

negative tracing results are described below. Some of these

errors have already been been mentioned briefly in this document.

1. Inadequate field survey to locate springs or wells to be

used for monitoring the presence of tracer. Published

USGS topographic maps can not be relied upon as the sole

source for data on spring locations. Field work is

essential because generally fewer than 5 percent of the

base-level springs are shown on topographic maps. An

interpretation of regional hydrology based on dye-traces

made only to springs shown on the USGS 7.5-minute

topographic maps is likely to be a distorted, incomplete

caricature of reality. Obtaining this caricature will

51
cost more in terms of time lost while waiting for dye to

be recovered, than doing the tracing investigation

correctly from the beginning.

Although most springs occur along the banks of a stream or

river, some occur in channels. Therefore it is wise to

also set dye-detectors in streams, rivers, and tributaries

--just to sense discharge from unknown springs that might

be in channels, from other springs that may not have been

found, and from reaches characterized by diffuse seepage.

Field work will usually show that there are many domestic

wells, both in service and abandoned, that can be used to

monitor for dye during a test. For many states most of

the wells will not be included in official State or

Federal files; inclusion of only 10 percent is common.

Probably no dye will be recovered in any of these wells,

but it is always worth trying to recover dye in them,

especially if their water gets turbid after heavy rains.

2. Injection of dye in a well or at another site that has

minimal hydraulic connection to the subjacent aquifer.

Dye injected into a well that does not intercept solutionally enlarged fissures, bedding planes, or conduits

may remain there for a long time and not reach the

drainage system of an aquifer except via intergranular

flow. The same

excavated in soil

easily prevented

tests, as already

3. Sampling in only

tion site--rather

can be said of dye put into a trench

or on a quarry floor. This error can be

by first running slug or percolation

discussed.

one or two directions from a dye-injecthan in all directions when radial flow

or multi-directional flow is possible. Radial flow,

although not common, frequently occurs near topographic

divides. Two excellent examples of radial flow at waste

disposal sites are illustrated in Figure 6 and by Quinlan

and Ewers (1985, p. 214-219).

If one has ignored the possibility of radial flow in the

design of a dye-test or a series of dye-tests, and then

gets positive results in those tests which are run, one

can be easily lulled into a false sense of security. For

example, if a hydrologist’s best professional judgement

suggested that ground-water flow in the vicinity of dyeinjection point #1 in Figure 6 was to the north or east

and if he set dye-detectors (bugs) only in those directions and not at springs and wells in the other directions, he might be rather impressed by his perspicacity.

He would also be professionally embarrassed and legally

vulnerable when leachate was subsequently detected at

springs or wells in the south and west.

52
4. Sampling at only a few sites--rather than at all sites

possible for the recovery of tracer. One can not afford

the “economy” of minimal sampling. Aside from the fact

that one does not get an understanding of regional or

basin hydrology without monitoring the sites to which dye

is carried by ground water, months of valuable time can be

lost while waiting for dye to be recovered at sites to

which it will not go. Also, until and unless the dye from

a test is recovered somewhere or it is determined where it

went, that dye can not be used a second time in the basin.

The reason is obvious: Dye recovered after a second dyeinjection could be interpreted to be from the first.

5. Not sampling often enough to detect the tracer. For

example, if a pulse duration is 24 hours and sampling is

weekly, it is highly probable that grab samples will not

detect the pulse. This is not a problem if activated

charcoal is used for dye-detection, but a better understanding of the flow dynamics is achievable if one detects

tracer at a site several times during a test (after at

least one time when the detectors and/or water samples are

negative for dye) and if one can recognize the approximate

time of maximum concentration and monitor the decay of the

tracer-concentration curve. Problems of sampling frequency in karst aquifers characterized by conduit flow and

nonpoint pollutants are discussed by Quinlan and Alexander

(1987).

As a second example, if one is using charcoal detectors

for dye-detection and is doing weekly sampling in waters

highly polluted by organic waste and if the adsorption

sites on the charcoal are thus totally occupied by organic

compounds after 24 hours, elution of detectors changed

weekly is incapable of detecting dye or a representative

sample of dye--unless one is extremely lucky and happens

to set a detector at a time when the dye-cloud is passing

the monitoring site. Even such luck is not enough.

Organic compounds (and possibly associated bacterial

reactions) in streams and ground-water can also remove dye

from charcoal detectors. In a well-designed test in which

a cloud of Rhodamine WT was visually seen to be flowing by

several adjacent detectors in a stream laden with organic

pollutants, the amount of dye recovered on the detectors

was inversely proportional to the duration of their

exposure to polluted stream water; dye-recovery from

detectors with the longest exposure was miniscule (Thomas

Aley, Ozark Underground Laboratory, Protein, Me., oral

communication, 1988).

6. Premature cessation of the tracer test--before there is

enough time for the tracer to reach any monitoring site or

those sites that would be reached after recovery of tracer

53
at the first site. An excellent example of the wisdom of

continuation of detector recovery is shown in Figure 6.

Dye injected next to the landfill site, at #1, arrived in

3 to 5 days at Cannon Spring, about 3.5 km (2.2 mi) away,

having traveled at least 700 to 1200 m/day (2300 to 3800

ft/day): dye was detected in 26 to 33 days at Keith

Spring, a shorter straight-line distance, having traveled

at least 80 to 100 m/day (260 to 330 ft/day). This latter

velocity range is consistent with the straight-line flowvelocity of about 120 m/day (390 ft/day) for dye that

traveled 7.3 km (4.5 mi) southwest to Mitch Hill Spring in

61 days (Aley, 1988).

An extreme example of premature cessation known to the

writer occurred when fluorescein put into an open sinkhole

at a hazardous-waste disposal site during a drought took

approximately 30 days to travel 3.0 km (1.8 mi) to a

spring at a mean maximum velocity of 4 m/hr (13 ft/hr).

Subsequently dye flowed from this spring for at least 30

days. The consultant doing the tracer test terminated it

before dye ceased flowing from the spring. He did not

allow enough time for dye to also reach more distant sites

where recovery was possible, and he refused to have the

detectors that had been changed weekly analyzed for dye.

After he was persuaded to have them analyzed, it was

learned that flow from the facility was radial; dye had

traveled approximately twice as fast in almost the

opposite direction to a domestic well!

Many examples of distributary flow (Quinlan & Ewers, 1985,

p. 205, 207-208) would not have been detected if sampling

for dye had not continued well beyond the time of first

recovery of dye. During the dry season, when ground-water

flow velocities are slowest, it is easy to err by premature cessation of a dye-trace. This is one of several

reasons why the most efficient times for initial tracing

are during moderate flow conditions and the recession of

Storm flow.

Flow times significantly longer than anticipated (or even

negative results) can also be a consequence of injection

of dye at poor sites in which flow is significantly slower

than in the subjacent drainage system. This is another

reason to continue sampling longer than may appear to be

necessary for dye-recovery.

7. Sampling only at randomly located drill holes rather than

at springs, cave streams, and wells that become, turbid

after heavy rains and wells drilled on photolineaments.

The rationale for this statement is given and illustrated

by Quinlan (1989a) and Quinlan and Ewers (1985). In

brief, the probability of randomly located wells

54
8.

intercepting a conduit conveying waste from a site in a

karst terrane is about equal to that of a dart thrown at a

wall map of the U.S. hitting the Mississippi River. Both

events are possible, but the probability of each is

extremely low. One can not afford to prospect blindly for

cave streams by random drilling.

Most randomly located wells in karst terranes are not

suitable for monitoring the quality of ground water

draining from a given site (Quinlan and Ewers, 1985).

Some wells can be used as monitoring points, but only if

dye-tests at high stage and low stage have shown that they

are recharged from the site to be monitored. Each well

that is to be dye-tested for suitability as a monitoring

site should be pumped during the test. at a rate that

adequately senses flow in an aquifer but doesn’t distort

the flow field. Pumping of domestic wells to yield a

continuous discharge of 4 to 8 liters (1 to 2 gallons) per

minute has been found to be quite satisfactory. A device

for maximizing the efficiency of dye-recovery from pumped

wells has been developed by the writer (Aley et al .,

1989) .

Use of an inadequate amount of tracer--an amount so small

that it is likely to be diluted or sorbed to concentrations far below the limit of detection. I know of

situations where either corporate parties or a regulatory

agency--for reasons ranging from fear of potentially

adverse public reaction to problems of alleged toxicity to

strong desire not to discover the truth--tried to prevent

proposed dye-traces from having the slightest chance of

success by deliberately limiting the amount of dye that

could be used. The probable dilution of the dye proposed

for use would have been to a concentration far lower than

the limit of detection of the instrument employed to

analyze for it. Investigators in other situations have,

through ignorance, used too small an amount of dye. One

cannot routinely expect a few ounces of dye to be

unequivocally detectable 16 km (10 mi) away.

When starting a tracing investigation in an area, One

should always, if there is a choice, start with the

simplest, most obvious trace, the one in which the results

are most easily anticipated. This “calibration” procedure

enables a better estimate of the amount of dye needed for

that trace and other traces in the adjacent area.

9. Use of a tracer inappropriate for the system under study,

one that is likely to be totally or greatly sorbed by

sediment or rock through which it passes, one that is

reactive with a dissolved constituent in water, or one in

which the fluorescence is quenched or enhanced by a

55
dissolved constituent or by pH. Until the "ideal" dye is

synthesized and economically available, we must live with

problems of sorption of dyes. As a generalization, the

least sorbed dye commonly used for tracing is fluorescein;

it is superior to Rhodamine WT in most settings where

photodecomposition is not a problem (Smart and Laidlaw,

1977) . Traces through coal mines are more likely to be

successful if CI Acid Red 52 is used; other conventionally

used dyes have a nigher affinity for sorption by ferric

hydroxide (Aldous and Smart, 1988).

Quenching of dye fluorescence by chloride ions and decomposition of dyes by chlorine have been briefly described

on p. 36. Graphs showing the quenching and enhancement of

dye fluorescence by pH have been published by Smart and

Laidlaw (1977) and Behrens (1982, 1987, 1988). Indeed,

use of pH-control for selective suppression or enhancement

of excitation and/or emission spectra can be used to

minimize the interference effects of one dye upon another

when a mixture of two or more of them are used simultaneously and recovered at the same site (Behrens, 1982, 1987,

1988) .

Many fluorescent dyes are suitable for tracing groundwater. Before beginning a dye-test, the characteristics

of the site and the recovery areas must be evaluated and

properties of various possibly suitable dyes must be

compared (Quinlan, 1989b).

10. Use of too high a value for the acceptable threshold of

detectability of dye. A practical value for the detection

limit of a fluorometer or a spectrofluorophotometer is the

concentration of analyte which gives rise to an analytical

signal equal to twice the background noise (Rendell, 1987,

p. 136). Some investigators informally define the acceptable threshold of detectability, however, as many times

this limit of detection. Such a practice is foolish and

can prevent the recognition and acknowledgement of radial

flow, distributary flow, and flow to a sole recovery site

or multiple sites when a less than optimal amount of dye

is used or recovered. One should always be skeptical and

paranoid about anomalous results, especially those near

the limit of detection, but they should be critically

evaluated for their consistency with results that are more

certain.

Definition of background fluorescence and its variation as

an aid to reliable determination of limit of detection is

yet another reason why pre-trace background determinations

are important. An interesting study of daily natural variations in background fluorescence has been made by Jones

(1989). He found no variation for orange (Rhodamine WT),

56
11.

12.

13.

moderate variation for green (fluorescein) [possibly

related to leaching of horse manure], and extreme variation in blue (optical brightener) background. This is not

to say that any of these dyes were present as background.

Rather, there were traces of unidentified substances in

the water which a filter fluorometer detected as though

they were dyes.

Use of organizations and individuals inexperienced in the

design, operation, and interpretation of tracer tests.

Dye-tracing, like neurosurgery, can be done by anyone.

But when either is needed, it is judicious and most costefficient to have it done by experienced professionals,

those who have already made the numerous mistakes

associated with learning or those who have trained under

the tutelage of an expert and learned to avoid numerous

procedural errors that could have economically and

physically fatal consequences.

Sampling for dye at only one side (the wrong side) of a

spring with dual-basin flow. This procedural error has

been explained in the discussion of EXCEPTIONS TO ASSUMPTION #6.

Placement of a dye-detector where contact with spring

water or stream water is not continuous for the time

between placement and recovery. The most common way for

this to occur is to set the detector too high in the

spring or stream channel rather than as low as possible.

When stage falls, the detector is left high and dry, but

it rises before the detector is recovered. A pulse of dye

could come by after the detector was out of the water and

be gone by the time it was again submerged. Another way

of having the same effect is to have a curious passerby

remove the detector and return it to the water a few days

later. [This has actually happened!]

Discussion

These procedural errors are listed and discussed not as a

“knave’s guide to duplicity” but as an aid for recognizing

inadequate investigations.

Quinlan, Ewers, and Field (1988) admitted having mixed

feelings about telling how to get spurious results from tracer

studies. Nevertheless, they believed that administrators and

others who evaluate hydrogeologic studies of facilities must be

able to differentiate between skilled, thorough, rigorous work

and shoddy or inadequate work. If review of the evaluation

report for a facility shows that one or more of the thirteen

procedural deficiencies described above are present or that one

57
or more of the exceptions to the first eight routinely made

assumptions may occur, the validity of the report is

questionable. The ground-water traces are probably incomplete.

Interpretations based upon them are unreliable.

It is easy to conduct poorly-designed tracer studies that

yield indeterminate results. For example, when the tracer is not

recovered, what do the results, more specifically, the lack of

positive results, mean? Both investigators and report evaluators

who are inexperienced with tracer-test design are not likely to

recognize that poor recovery of tracer may be a result of either

poor design of the test or inept execution of it--or both. More

commonly, both parties erroneously tend to accept a lack of

tracer recovery as an indicator of diffuse flow, non-radial flow,

or the alleged unreliability of tracers for characterizing the

hydrology of a site. Investigators and report evaluators may

then develop a false sense of security, believing that leakage

has never occurred, or that ground-water flow velocities are very

slow and like those in granular aquifers, or that a facility can

be monitored by randomly drilled wells.

There is another reason for listing the above causes of

falsely negative tracer tests. The consultant or agency employee

who knows well what constitutes bad tracer-test design and protocol knows better what constitutes good design and protocol. Such

knowledge can improve the quality of dye-test performance and

evaluation.

As repeatedly implied and stated throughout this document,

short-cutting on rigorous design and protocol of water-tracing is

a false economy. The environmental consequences, the ethical

consequences, and the legal consequences of malpractice may be

far too high to be ignored.

Tracer tests are not intrinsically unreliable. They are

reliable. It is tracer tests done by the inexperienced who are

unaware of the assumptions they have made and unaware of how one

can inadvertently get falsely negative and falsely positive

results that are unreliable.

One can learn much about dye-tracing by reading; one should.

But before attempting complex traces that can affect the

expenditure of hundreds of thousands or millions of dollars, it

is judicious to learn also by experience on simple traces and by

working with someone who has extensive dye-tracing experience.

With dye-tracing, as with love-making, there is only so much one

can learn from a book. In the quest for success, there is no

substitute for experience.

58
SISYPHEAN LAMENT

People concerned with monitoring are usually afflicted with

one of two syndromes: “nothing leaks” or “everything leaks”.

The "nothing-leaks syndrome” is dominated by a rationale that

“We designed it not to leak, therefore, it won’t.” In the early

stages this syndrome is characterized by a conviction that there

is no problem with a facility unless an anomalous value for a

parameter is detected by its monitoring system. In the advanced

stages it is characterized by a conviction that there is no

problem unless the anomalous parameter exceeds the proscribed

limits for drinking water. In the terminal stages this syndrome

is characterized by a conviction that there is no problem unless

the anomalous parameter repeatedly exceeds drinking water standards. By then it may be too late to prevent the consequences of

pollution.

The “everything-leaks syndrome” is dominated by the rationale

that few systems work as designed and Murphy’s Law is always

operative. In the early stages this syndrome is characterized by

a conviction that pollutants detected by a well-designed monitoring system are proof that the facility leaks. A deadly Catch-22

logic rapidly brings the syndrome to a terminal stage in which

those afflicted fanatically proclaim that if pollutants are not

detected, the monitoring system is defective and needs to be

improved.

A symptom common to zealots afflicted with the pernicous terminal stage of each syndrome is intolerance for the irrationality

of those afflicted by the terminal stage of the other.

When people with each of these syndromes, especially in their

terminal stages, must work together on a project, each side tends

to keep the other honest--but often at the expense of a harmonious working relationship. Religion and politics are safer,

less controversial topics of conversation.

Would that reality

for all. This topic is

CHECKLIST FOR DESIGN

were perceivable objectively and the same

eloquently discussed by Alexander (1989).

OF MONITORING SYSTEMS IN KARST TERRANES

The following checklist can be used as a guide to the general

sequence of operations necessary for the design of a monitoring

system for ground water in most karst terranes, especially those

characterized by conduit flow. The checklist is useless,

however, without an understanding and application of the concepts

espoused in this document. One should do the following:

1. Review the geologic and hydrologic literature.

59
2. Study and interpret topographic and geologic maps.

3. Make a spring survey.

4. Review this document.

5. Map the potentiometric surface, if possible.

6. Design the dye-tracing study.

7. Do the first trace, preferably during moderate flow conditions.

8. Evaluate results of the first trace and modify the design

of the tracing study, if necessary.

9. Determine whether springs to which a facility drains are

characterized by conduit flow or diffuse flow.

10. Run other dye-traces during moderate flow, always

modifying the tracing plan, as necessary, in the light of

the results of each trace. For many facilities it is

necessary to perform only two traces during moderate flow.

11. Repeat selected traces during base flow and flood flow.

12. Synthesize tracing results, available potentiometric data,

and conductivity and turbidity data used to discriminate

between conduit flow and diffuse flow into a monitoring

plan.

13. Have results of field studies and the proposed monitoring

plan peer-reveiwed.

SUMMARY

In order to be relevant to detection monitoring for pollutants from waste-disposal facilities, water-quality data from

karst terranes must be from springs, cave streams, and wells

shown by tracing to drain from the facility to be monitored.

Tracing should typically be done at least three times: when first

convenient (during moderate-flow conditions, to quickly give a

preliminary, tentative understanding of local movement of ground

water) and later during both base flow and flood flow. The

general prudence of tracing during all three types of flow conditions can not be overemphasized.

Sites for monitoring background should be selected on the

basis of:

60
1. Negative results of these tracing tests but from settings

in which the rocks and waters are geochemically similar to

those of the locations where the tracer tests were positive.

2. Cultural similarity.

A map of the potentiometric surface, if it is based on enough

data from an aquifer not complicated by aquicludes and aquitards,

will greatly enhance one’s ability to efficiently design the

necessary tracing tests, interpret them with greater confidence,

and design an effective, reliable monitoring network.

Sampling for water quality must be frequent and done before,

during, and after storm and meltwater events. Base flow should

be sampled between such events.

Ground-water monitoring in karst terranes can be done

reliably, but the analytical costs are likely to be significantly

higher than those for other terranes. It could be far less

expensive to locate a proposed facility in a non-karst terrane.

There are many ways to design the placement and sampling

frequency of a ground-water monitoring network in a karst terrane

so that it inadvertently yields falsely negative results for the

chemical compound(s) being sought. Accordingly, environmental

consultants and regulators must be ever vigilant to be sure that

negative results are not falsely negative and that positive

results are not falsely positive--either accidentally or

intentionally.

There is need for Federal and State regulations to officially

recognize the utility and wisdom of ground-water monitoring at

springs, cave streams, and traced wells.

Numerous plausible, axiomatic rules (assumptions) can be

stated about ground-water flow in karst terranes, but they are

not absolutes. Exceptions are known for all’ but two of the ten

rules cited; more exceptions will be discovered.

It is easy to inadvertently get falsely negative results with

any tracer, thus "proving" that a facility will not (or does not)

leak and will not affect "x". If an investigator knows what

these pitfalls are, he can avoid them. If a report evaluator

knows them, he can recognize flawed investigations, reject them,

and have them redone properly.

Conflicts between designers, evaluators, and regulators over

whether a monitoring system works or doesn’t work and how to

interpret its data are often a result of irreconcilable philosophical convictions. Some people sincerely believe that a

facility designed not to leak will not leak. Opponents are

61
equally convinced of the impossibility of long-term prevention of

leakage and place great importance on judicious siting and

skeptical interpretation of monitoring data.

The monitoring strategies advocated here and discussed in

more detail by Quinlan and Ewers (1985) work in most karst

terranes characterized by recharge and by discharge at springs.

They are also applicable to the delineation of wellhead and

springhead protection areas. They are a Significant advance over

traditional monitoring strategies and have been recognized as

such (Beck et al., 1987), but this document is far from the last

word on the subject. Much remains to

There is a need for publication of more

EPILEGOMENON

Dye-tracing is essential for the

system in a karst terrane. But why

be learned and described.

case studies.

design of a monitoring

is it that professional

geologists and engineers who would not think of venturing into

the design of a building foundation, a landfill, or a well field

without first having obtained some initial experience with such

matters, and without a review of its design by a competent peer,

all too often assume that anyone can do professional-quality dyetracing on the first attempt? If one’s tracing experience is

limited or nonexistent, the most astute, most ethical, and least

expensive way to minimize the risks of costly litigation for

tracing-related malpractice are to choose one of the following:

1. Gain experience with one of the fewer than 10 individuals

in North America who are adept at tracing with various

types of dyes in complex hydrologic settings.

2. Arrange for one of these individuals to do the tracer

investigation --with the understanding that he will give

some rudimentary training.

3. Hire one of them to design the tracer investigation.

4. At an absolute minimum, use one of them to review the

design of the proposed tracer investigation.

A thoroughly-experienced, adept tracing-consultant potentially saves his client many times his fee and greatly enhances the

reliability of both tracer results and a proposed monitoring

system.

Some readers may choose to interpret the preceding part of

this epilegomenon as a self-serving solicitation offering

consultation services. That would be unfortunate. It is not.

Rather, it is an objective assessment of facts as viewed in the

light of extensive experience.

62
Karst hydrology is foreign to the training and experience of

most American geologists, hydrologists, and engineers. Indeed,

only one of the commonly used hydrology texts gives more than

lip-service recognition to the significant differences between

flow in karst aquifers and flow in other aquifers (Fetter, 1988,

p. 285-295, 233-246).

Only three of the dozens of hocks, monographs, and manuals

published on ground-water hydrology and/or monitoring during the

past eight years include either a discussion or implication of

the recommended strategy of monitoring springs, cave streams, and

traced wells (Office of Ground-Water Protection, EPA, 1987, 1988;

Mull et al., 1988), but the authors of at least one hydrology

text in preparation strongly endorse it. Two recent textbooks on

karst geomorphology and hydrology also endorse it (White, 1988;

Ford & Williams, 1989). No journal papers even hint at the

strategy, but there are numerous recent papers, chiefly by karst

specialists, that do so in conference proceedings (Alexander et

al., 1988; Aley, 1988; Beck, 1986; Field, 1988, 1989; Jennings,

1988; Quinlan, 1986b, 1988a, 1989a; Quinlan & Alexander, 1987;

Quinlan, Aley, & Schindel, 1988; Quinlan, Ewers, & Field, 1988).

The review by Field (1988) summarizes evolving EPA thought on

monitoring in karst terranes, but it is not an official statement

or a draft of proposed agency policy. In brief, the scientific

community has just begun to recognize the effectiveness of the

traced-spring, -cave-stream, and -well monitoring strategy.

It would seem that many pollution control agencies are about

to recognize, at last, the “peculiarities” of karst and to

promulgate their regulations accordingly.

63
REFERENCES

Adams, D. 1980. The Hitchhiker’s Guide to the Galaxy. Harmony,

New York. 215 pp.

Aldous, P.J., and Smart, P.L. 1988. Tracing ground-water movement in abandoned coal mined aquifers using fluorescent

dyes. Ground Water. 26:172-178.

Alexander, E.C., Jr. 1989. Karst hydrogeology and the nature of

reality: Philosophical musings of a frustrated curmudgeon.

Keynote address given at Environmental Problems in Karst

Terranes Conference (2nd, Nashville, Tennessee). National

Water Well Association, Dublin, Ohio. [in prep; to be

published separately from the proceedings]

Alexander, E.C., Jr., Davis, M.A., Alexander, S.C., and Lively,

R.S. 1988. Thermal springs of the southern Black Hills.

In: Schilberg, G., and Springetti, D., ed. Caves and Associated Features of the Black Hills (1988 NSS Convention

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