Processes Affecting the Trihalomethane Concentrations Associated

with the Third Injection, Storage, and Recovery Test at Lancaster,

Antelope Valley, California, March 1998 through April 1999

Prepared in cooperation with the Los Angeles County Department of Public Works and

the Antelope Valley-East Kern Water Agency

U.S. Geological Survey

Water-Resources Investigations Report 03-4062
Processes Affecting the Trihalomethane Concentrations 

Associated with the Third Injection, Storage, and 

Recovery Test at Lancaster, Antelope Valley, California, 

March 1998 through April 1999

By Miranda S. Fram1, Brian A. Bergamaschi1, Kelly D. Goodwin2, Roger Fujii1, 

and Jordan F. Clark3

Prepared in cooperation with the

LOS ANGELES COUNTY DEPARTMENT OF PUBLIC WORKS and 

the ANTELOPE VALLEY–EAST KERN WATER AGENCY

Sacramento, California

2003

Water-Resources Investigations Report 03-4062

U.S. GEOLOGICAL SURVEY

1

U.S. Geological Survey, Placer Hall, 6000 J Street, Sacramento, California 95819-6129

2

Cooperative Institute of Marine and Atmospheric Chemistry

4301 Rickenbacker Causeway, Miami, Florida 33149

3

University of California, Santa Barbara, Department of Geological Sciences

Webb Hall, Santa Barbara, California 92106 7212-58
For additional information write to:

District Chief

U.S. Geological Survey

Placer Hall– Suite 2012

6000 J Street

Sacramento, California 95819-6129

http://ca.water.usgs.gov

Any use of trade, product, or firm names in this publication is for descriptive purposes only and does 

not imply endorsement by the U.S. Government.

U.S. DEPARTMENT OF THE INTERIOR

GALE A. NORTON, Secretary

U.S. GEOLOGICAL SURVEY

Charles G. Groat, Director
Contents iii

CONTENTS

Abstract ................................................................................................................................................................ 1

I. Introduction....................................................................................................................................................... 3

Purpose and Scope ...................................................................................................................................... 5

Previous Studies.......................................................................................................................................... 5

Project Design ............................................................................................................................................. 6

Site Description and Chronology of Third Cycle ....................................................................................... 7

Acknowledgments....................................................................................................................................... 11

II. Formation and Fate of Trihalomethanes during the Injection, Storage, and Recovery Cycle ........................ 11

Variations in Water Quality and Water Levels during the Third Cycle ..................................................... 11

Trihalomethanes................................................................................................................................. 13

Dissolved Organic Carbon and Ultraviolet Absorbance.................................................................... 18

pH....................................................................................................................................................... 18

Residual Chlorine............................................................................................................................... 19

Chloride and Bromide........................................................................................................................ 19

Bromine to Chlorine Ratio in the Trihalomethanes ........................................................................... 20

Water Levels ...................................................................................................................................... 22

Trihalomethane Formation.......................................................................................................................... 23

Trihalomethane-Formation-Potential Experiments............................................................................ 23

Storage Experiment............................................................................................................................ 24

Estimate of Trihalomethane Formation in the Injected Water........................................................... 28

Trihalomethane Fate.................................................................................................................................... 29

Biodegradation ................................................................................................................................... 29

Sorption.............................................................................................................................................. 29

Mixing................................................................................................................................................ 30

Mass Balance of Chloride, Dissolved Organic Carbon, and Trihalomethanes.................................. 31

Mass Balance Calculations ....................................................................................................... 31

Mass Balance Results ............................................................................................................... 34

Chloride ........................................................................................................................... 34

Dissolved Organic Carbon............................................................................................... 34

Trihalomethanes............................................................................................................... 36

Implications for Water Flow in the Aquifer................................................................................................ 37

Conclusions................................................................................................................................................. 40

III. Modeling Dissolved Constituents, Trihalomethanes, and Sulfur Hexafluoride Tracer Concentrations in 

Extracted Water.................................................................................................................................................... 41

Sulfur Hexafluoride as a Tracer .................................................................................................................. 41

Experimental Methods ....................................................................................................................... 42

Tracer Results .................................................................................................................................... 43

Modeling ..................................................................................................................................................... 44

Tracer Mixing Model ......................................................................................................................... 44

Descriptive Mixing Model ................................................................................................................. 47

Potential Implications of Mixing for Repetitive Cycles of Injection, Storage, and Recovery.................... 51

Conclusions................................................................................................................................................. 53

IV. The Potential for Biodegradation of Trihalomethanes by Aquifer Bacteria.................................................. 54

Sediment Microcosm Experiments ............................................................................................................. 55

Water Enrichment Microcosm Experiments............................................................................................... 60

Bacterial Density......................................................................................................................................... 60
iv Contents

Conclusions................................................................................................................................................. 65

V. Summary and Conclusions.............................................................................................................................. 66

References Cited .................................................................................................................................................. 68
Figures v

FIGURES

Figure 1. Map showing locations of sites and geographic features relevant to the third injection, 

storage, and recovery test (March 1998 through April 1999) at Lancaster, Antelope 

Valley, California............................................................................................................................... 4

Figure 2. Generalized subsurface geology and locations of wells and nested piezometers used for the third

injection, storage, and recovery test (March 1998 through April 1999), Lancaster, Antelope 

Valley, California............................................................................................................................... 8

Figure 3. Graph showing average daily pumping rates (A) and cumulative volumes of water injected

and extracted (B) at well 7N/12W-27P2 (well 4-32) and well 7N/12W-27P3 (well 4-34), 

and average daily water levels at piezometers 7N/12W-27P6–8 (C) during the third injection, 

storage, and recovery cycle (March 1998 through April 1999), Lancaster, Antelope Valley, 

California ........................................................................................................................................... 10

Figure 4. Graphs showing trihalomethane (A) and chloride (B) concentrations in water collected from 

well 7N/12W-27P2 (well 4-32) during the three injection, storage, and recovery cycles 

(April 1996 through April 1999), Lancaster, Antelope Valley, California ....................................... 12

Figure 5. Graphs showing trihalomethane concentrations (A), dissolved organic carbon concentrations 

(B), ultraviolet absorbance values at 254 nanometers (C), pH values (D), free and total residual 

chlorine concentrations (E), chloride concentrations (F), bromide concentrations, and molar 

bromine to chlorine ratio in trihalomethanes (H, I) in ground water, injection water, and 

extraction water collected from well 7N/12W-27P2 (well 4-32) and from nested piezometers 

7N/12W-27P6–8 during the third injection, storage, and recovery cycle (March 1998 through 

April 1999), Lancaster, Antelope Valley, California......................................................................... 14

Figure 6. Graph showing trihalomethane concentrations in injection water collected from well 

7N/12W-27P2 (well 4-32) during the third injection, storage, and recovery cycle 

(March 1998 through April 1999), Lancaster, Antelope Valley, California ..................................... 25

Figure 7. Graph showing values of specific trihalomethane formation potential (STHMFP) and specific 

ultraviolet absorbance at 254 nanometers (SUVA254) for injection water and ground water 

collected from well 7N/12W-27P2 (well 4-32) during the third injection, storage, and recovery 

cycle (March 1998 through April 1999), Lancaster, Antelope Valley, California ............................ 26

Figure 8. Graph showing trihalomethane concentrations in injection water collected from well 

7N/12W-27P2 (well 4-32) during the third injection, storage, and recovery cycle 

(March 1998 through April 1999), Lancaster, Antelope Valley, California ..................................... 27

Figure 9. Graphs showing method for calculating mass balance of a conservative constituent without 

(A,B) and with (C,D) mixing between injected water and ground water .......................................... 33

Figure 10. Graphs showing chloride (A), dissolved organic carbon (B) and trihalomethane (C) 

concentrations in ground-water collected before the first, second, and third cycles 

(April 1996 through April 1999), and in injection and extraction water collected from well 

7N/12W-27P2 (well 4-32) during the third injection, storage, and recovery cycle 

(March 1998 through April 1999), Lancaster, Antelope Valley, California ..................................... 35

Figure 11. Graph showing cumulative moles of trihalomethanes (THM) recovered during the extraction 

period of the third injection, storage, and recovery cycle (March 1998 through April 1999) 

Lancaster, Antelope Valley, California ............................................................................................. 36

Figure 12. A simple balloon conceptual model (A) and a more realistic conceptual model (B) for water 

flow in the aquifer during the third injection, storage, and recovery cycle (March 1998 

through April 1999), Lancaster, Antelope Valley, California ........................................................... 38

Figure 13. Graph showing sulfur hexafluoride (SF6) concentrations in injection and extraction water 

collected from well 7N/12W-27P2 (well 4-32) during the third injection, storage, and recovery 

cycle (March 1998 through April 1999), Lancaster, Antelope Valley, California ............................ 43
vi Figures

Figure 14. Graphs showing measured and sulfur hexafluoride (SF6) tracer-derived (A) and chloride 

tracer-derived (B) trihalomethane concentrations in extraction water collected from well 

7N/12W-27P2 (well 4-32) during the third injection, storage, and recovery cycle (March 

1998 through April 1999), Lancaster, Antelope Valley, California .................................................. 46

Figure 15. Graph showing trihalomethane and sulfur hexafluoride concentrations in extraction water

collected from well 7N/12W-27P2 (well 4-32) during the third injection, storage, and 

recovery cycle (March 1998 through April 1999), Lancaster, Antelope Valley, California ............. 47

Figure 16. Diagram of the descriptive mixing model for mixing between injection water and ground water 

during extraction ................................................................................................................................ 48

Figure 17. Graphs showing measured and model-derived sulfur hexafluoride concentrations (A), chloride

concentrations (B), and trihalomethane concentrations (C) in extraction water collected from 

well 7N/12W-27P2 (well 4-32) during the third injection, storage, and recovery cycle 

(March 1998 through April 1999), Lancaster, Antelope Valley, California ..................................... 50

Figure 18. Graphs showing predicted concentrations of conservative injected constituents in an aquifer 

during 10 annual cycles of injection, storage, and recovery for two different injection period 

durations (A,B) and four different amounts of mixing between cycles (B,C,D,E) ........................... 52

Figure 19. Graphs showing mass of chloroform (CHCl3) (A) and bromoform (CHBr3) (B) in sediment 

microcosms containing sediment and ground water from Lancaster, Antelope Valley, 

California ........................................................................................................................................... 57

Figure 20. Graphs showing mass of chloroform (CHCl3) (A) and bromoform (CHBr3) (B) in aerobic, 

enriched sediment microcosms containing sediment and ground water from Lancaster, 

Antelope Valley, California............................................................................................................... 58

Figure 21. Graph showing mass of chloroform (CHCl3) in anaerobic, enriched sediment microcosms 

containing sediment and ground water from Lancaster, Antelope Valley, California ...................... 59

Figure 22. Graph showing mass of bromoform (CHBr3) in anaerobic, enriched sediment microcosms 

containing sediment and ground water from Lancaster, Antelope Valley, California ...................... 59

Figure 23. Graphs showing mass of chloroform (CHCl3) (A) and bromoform (CHBr3) (B) in water 

enrichment microcosms stored for an incubation period and containing extraction water or 

ground water from Lancaster, Antelope Valley, California .............................................................. 61

Figure 24. Graphs showing mass of chloroform (CHCl3) (A) and bromoform (CHBr3) (B) in sterile 

control samples for water enrichment microcosms stored for an incubation period and containing 

sterile filters and extraction water from Lancaster, Antelope Valley, California.............................. 62

Figure 25. Graph showing bacterial densities in water samples collected from well 7N/12W-27P2

(well 4-32) during the third injection, storage, and recovery cycle (March 1998 through 

April 1999), Lancaster, Antelope Valley, California......................................................................... 63

Figure 26. Graph showing bacterial densities in water samples collected from nested piezometers 

7N/12W-27P6, 27P7, and 27P8 during the third injection, storage, and recovery cycle 

(March 1998 through April 1999), Lancaster, Antelope Valley, California ..................................... 64

Figure 27. Graph showing Bacterial densities in water samples collected from wells 7N/12W-27P2 

(well 4-32), 7N/12W-27J4 (well 4-13), 7N/12W-27H3 (well 4-33), and 7N/12W-27J6 

(well 4-42), and the nested piezometers 7N/12W-27P6, 27P7, and 27P8, during the third 

injection, storage, and recovery cycle (March 1998 through April 1999), Lancaster, 

Antelope Valley, California............................................................................................................... 65
Tables vii

TABLES

Table 1. Nomenclature for wells used during the third injection, storage, and recovery cycle

(March 1998 through April 1999), Lancaster, Antelope Valley, California ..................................... 9

Table 2. Trihalomethane concentration data for water collected from nested piezometers 7N/12W-27P6–8 

during the third injection, storage, and recovery cycle (March 1998 through April 1999), 

Lancaster, Antelope Valley, California ............................................................................................. 17

Table 3. Water quality data for water collected from nested piezometers 7N/12W-27P6–8 during the 

third injection, storage, and recovery cycle (March 1998 through April 1999), Lancaster, 

Antelope Valley, California............................................................................................................... 21

Table 4. Average concentrations of constituents in injected water and ground water used in mass-balance, 

tracer mixing model, and descriptive mixing model calculations ..................................................... 31
viii Conversion Factors, Acronyms, Abbreviations, and Water-Quality Information

CONVERSION FACTORS, ACRONYMS, ABBREVIATIONS, AND WATERQUALITY INFORMATION

CONVERSION FACTORS

Temperature in degrees Celsius (°C) may be converted to degrees Fahrenheit (°F) as follows:

°F = (1.8 × °C) + 32

ACRONYMS

AODC acridine orange direct count 

AVEK Antelope Valley–East Kern Water Agency 

DBP disinfection by-products

DOC dissolved organic carbon 

EPA U.S. Environmental Protection Agency 

LACDPW Los Angeles County Department of Public Works 

MCL maximum contaminant level 

RSD percent relative standard deviation 

STHMFP specific trihalomethane formation potential 

SUVA254 specific ultraviolet absorbance at 254 nanometers 

SWP State Water Project 

THM trihalomethane

THMFP trihalomethane formation potential 

USGS U.S. Geological Survey 

UVA254 ultraviolet absorbance at 254 nanometers 

ABBREVIATIONS

μg/L microgram per liter 

μg/kg microgram per kilogram 

/cm per centimeter 

mg/L milligram per liter 

μmol/L micromole per liter 

μS/cm microsiemens per centimeter

mL milliliter 

nm nanometer 

pmol/L picomole per liter 

Multiply By To obtain

foot (ft) 0.3048 meter

gallon per minute (gal/min) 0.06309 liter per second
Conversion Factors, Acronyms, Abbreviations, and Water-Quality Information ix

(Br/Cl)THM the ratio of bromine to chlorine in the trihalomethane

CHBr3 bromoform 

CHCl2Br bromodichloromethane 

CHCl3 chloroform 

CHClBr2 dibromochloromethane 

CO2 carbon dioxide 

KH2PO4 potassium dihydrogen phosphate 

NH4Cl ammonium chloride 

δ18O delta oxygen-18

SF6 sulfur hexafluoride 

Water-Quality Information

Chemical concentration is given in milligrams per liter (mg/L), micrograms per liter (μg/L), or picomoles per liter (pmol/L). Milligrams per liter is a unit expressing the mass of a solute per unit volume

(liter) of water. One thousand micrograms per liter is equivalent to 1 milligram per liter, and 1,000

milligrams per liter is equivalent to 1 gram per liter. The numerical value in milligrams per liter is

about the same as for concentrations in parts per million, and the numerical value in micrograms per

liter is about the same as for concentrations in parts per billion. Micromoles per liter is a unit expressing the number of moles of a solute per unit volume (liter) of water. One million picomoles per liter is

equivalent to 1 micromole per liter, and one million micromoles per liter is equivalent to 1 mole per

liter. 

Ultraviolet light absorbance is given in per centimeter (/cm). Wavelength of light is given in nanometers (nm). Specific ultraviolet absorbance is given in liters per milligram per meter [(L/mg)/m] and is

equal to the ultraviolet light absorbance in per centimeter multiplied by 100 and divided by the dissolved organic carbon concentration in milligrams per liter. Specific trihalomethane formation potential is given in millimoles per mole (mmol/mol) and is equal to the trihalomethane concentration in

micromoles per liter divided by the dissolved organic carbon concentration in millimoles per liter.

Bacterial cell density is given in cells per milliliter (cells/mL).

Specific conductance is given in microsiemen per centimeter at 25°C (μS/cm). Microsiemen per centimeter is numerically equivalent to micromhos per centimeter. Turbidity is given in nephelometric turbidity units (NTU). Gas flow rate is given in milliliters per minute (mL/min). Volume is given in

milliliters (mL) or microliters (μL). One thousand microliters is equivalent to 1 milliliter, and 1,000

milliliters is equivalent to 1 liter. Length is given in meters (m) and micrometers (μm). One thousand

micrometers is equivalent to 1 millimeter, and 1,000 millimeters is equivalent to 1 meter. Mass is

given in micrograms (μg), and one million micrograms is equivalent to 1 gram. Magnetic field

strength is given in millitesla (mT).
x Well-Numbering System

WELL-NUMBERING SYSTEM

Wells are identified and numbered according to their location in the rectangular system for the subdivision of public lands. 

Identification consists of the township number, north or south; the range number, east or west; and the section number. Each section is

divided into sixteen 40-acre tracts lettered consecutively (except I and O), beginning with “A” in the northeast corner of the section and 

progressing in a sinusoidal manner to “R” in the southeast corner. Within the 40-acre tract, wells are sequentially numbered in the order they 

are inventoried. The final letter refers to the base line and meridian. In California, there are three base lines and meridians; Humboldt (H), 

Mount Diablo (M), and San Bernardino (S). All wells in the study area are referenced to the San Bernardino base line and meridian (M). 

Well numbers consist of 15 characters and follow the format 007N012W27P002S. In this report, well numbers are abbreviated and written 

7N/12W-27P2. Wells in the same township and range are referred to only by their section designation, 27P2. The following diagram shows 

how the number for well 7N/12W-27P2 is derived.

R13W R12W R11W

Meridian

San Bernardino

 T8N

 T7N

 T6N

RANGE

TOWNSHIP

 R1W R1E

27

R12W

T7N

Well-numbering diagram

6 5 4 3 2 1

7

18 13

8 9 10 11 12

17 16 15 14

19 24

30 25

31 36

20 21 22 23

29 28 26

32 33 34 35

SECTION 27

D C A

EFGH

M L K J

N P Q R

B

7N/12W-27P2
Abstract 1

Processes Affecting the Trihalomethane Concentrations 

Associated with the Third Injection, Storage, and 

Recovery Test at Lancaster, Antelope Valley, California, 

March 1998 through April 1999

By Miranda S. Fram, Brian A. Bergamaschi, Kelly D. Goodwin, Roger Fujii, and Jordan F. 

Clark

ABSTRACT

The formation and fate of trihalomethanes 

(THM) during the third injection, storage, and 

recovery test at Lancaster, Antelope Valley, 

California, were investigated as part of a program 

to assess the long-term feasibility of using 

injection, storage, and recovery as a water-supply 

method and as a way to reduce water-level 

declines and land-subsidence in the Antelope 

Valley. The program was conducted by the U.S. 

Geological Survey in cooperation with the 

Los Angeles County Department of Public Works 

and the Antelope Valley–East Kern Water Agency. 

The water used for injection, storage, and recovery 

must be disinfected before injection and thus 

contains THMs and other disinfection byproducts. THMs (chloroform, CHCl3, 

bromodichloromethane, CHCl2Br, 

dibromochloromethane, CHClBr2, and 

bromoform, CHBr3) are formed by reaction 

between natural dissolved organic carbon that is 

present in water and chlorine that is added during 

the disinfection step of the drinking water 

treatment process. THMs are carcinogenic 

compounds, and their concentrations in drinking 

water are regulated by the U.S. Environmental 

Protection Agency. During previous cycles of the 

Lancaster program, extracted water still contained 

measurable concentrations of THMs long after 

continuous pumping had extracted a greater 

volume of water than had been injected. This 

raised concerns about the potential long-term 

effect of injection, storage, and recovery cycles on 

ground-water quality in Antelope Valley aquifers. 

The primary objectives of this investigation 

were to determine (1) what controlled continued 

THM formation in the aquifer after injection, 

(2) what caused of the persistence of THMs in the 

extracted water, even after long periods of 

pumping, (3) what controlled the decrease of THM 

concentrations during the extraction period, and 

(4) the potential for natural attenuation of THMs 

in the aquifer.

Laboratory experiments on biodegradation 

of THMs in microcosms of aquifer materials 

indicate that aquifer bacteria did not degrade 

CHCl3 or CHBr3 under aerobic conditions, but did 

degrade CHBr3 under anaerobic conditions. 

However, the aquifer is naturally aerobic and 

CHCl3 is the dominant THM species; therefore, 

biodegradation is not considered an important 

attenuation mechanism for THMs in this aquifer. 

The alluvial-fan sediments comprising the aquifer 

have very low contents of organic matter; 

therefore, sorption is not considered to be an 

important attenuation mechanism for THMs in this 

aquifer. Laboratory experiments on formation of 

THMs in the injection water indicate that 

continued THM formation in the injection water 

after injection into the aquifer was limited by the 

amount of residual chlorine in the injection water 
2 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

at the time of injection. After accounting for 

THMs formed by reaction of this residual 

chlorine, THMs behaved as conservative 

constituents in the aquifer, and the only process 

affecting the concentration of THMs was mixing 

of the injection water and the ground water.

The mixing process was quantified using 

mass balances of injected constituents, the sulfur 

hexafluoride (SF6) tracer that was added to the 

injected water, and a simple descriptive 

mathematical mixing model. Mass balance 

calculations show that only 67 percent of the 

injected THMs and chloride were recovered by the 

time that a volume of water equivalent to 

132 percent of the injection water volume was 

extracted. Pumping 250 percent of the injection 

water volume only increased recovery of injected 

THMs to 80 percent. THM and SF6 concentrations 

in the extracted water decreased concomitantly 

during the extraction period, and THM 

concentrations predicted from SF6 concentrations 

closely matched the measured THM 

concentrations. Because SF6 is a conservative 

tracer that was initially only present in the 

injection water, parallel decreases in SF6 and 

THM concentrations in the extracted water must 

be due to dilution of injection water with ground 

water. The simple descriptive mixing model 

mathematically described mixing of injection 

water and ground water that was displaced by 

injection using a single-zone mixing model. This 

simple model adequately predicted the 

concentrations of conservative constituents (SF6, 

THMs, and chloride) in the extracted water during 

the extraction period, providing further evidence 

to support the conclusion that mixing injection 

water and ground water in the aquifer was the 

primary process controlling the concentration of 

the THMs in the extracted water.

Water-quality data and modeling results 

suggest that injection, storage, and recovery cycles 

will affect aquifer water quality. THM 

concentrations in water samples from piezometers 

located 80 horizontal feet from the 

injection/extraction well remained high and 

variable throughout the extraction period. This 

suggests that water entered regions of the aquifer 

during injection from which it was not efficiently 

recovered during extraction. In particular, water 

may have become stranded in the upper parts of 

the aquifer during injection and therefore could 

not be efficiently recovered during extraction 

owing to drawdown around the well. The 

descriptive mixing model was used to forecast the 

results of repeated annual cycles of injection, 

storage, and recovery. For the scenario of equal 

volumes of injected and extracted water, the model 

forecasts that the concentration of THMs (or of 

any conservative constituent in the injection water) 

in the ground water near the injection/extraction 

well would increase to nearly 100 percent of the 

concentration of THMs in the injection water 

within 10 years. This increase in THM 

concentration would be less if ground water from 

outside the region directly affected by injection 

also mixed with the injected water or if the volume 

of extracted water greatly exceeded the volume of 

injected water. Finally, the model results indicate 

that extraction of all injected constituents is 

difficult using existing extraction methods because 

the volume of water that must be extracted 

increases exponentially as the acceptance criteria 

for the concentration in the water remaining in the 

aquifer decreases.
I. Introduction 3

I. INTRODUCTION

By Miranda S. Fram and Roger Fujii

Ground water is an important source of water 

supply in the Antelope Valley, California (fig.1). Since 

the late 1940s, ground-water pumpage has exceeded 

natural recharge, resulting in as much as 350 feet (ft) of 

water-level declines and more than 6 ft of land 

subsidence in some areas (Ikehara and Phillips, 1994). 

The Antelope Valley augments its ground-water 

supplies with imported water from the California State 

Water Project (SWP). The SWP is a series of storage 

reservoirs and aqueducts that transports surface water 

from northern to southern California (fig.1). Facing 

rapid population growth and increasing demand for 

water supply, water managers in the Antelope Valley 

are seeking ways to maximize the use of available 

water supplies.Using injection, storage, and recovery— 

injecting treated SWP water into the aquifer system 

during periods of greater surface-water availability to 

be used later during periods of surface-water deficit— 

is a potential water-supply method for meeting 

increasing water demands. This water-supply method 

would permit storage of additional imported water 

during the wet season when surface water is more 

available. The U.S. Geological Survey (USGS), in 

cooperation with the Los Angeles County Department 

of Public Works (LACDPW) and the Antelope Valley–

East Kern Water Agency (AVEK), did research and 

monitoring experiments during three test cycles of 

injection, storage, and recovery in Lancaster, Antelope 

Valley, California, from September 1995 through April 

1999 to assess the feasibility of using this water-supply 

method in the Antelope Valley.

The tests were designed to investigate how 

injection, storage, and recovery cycles affect water 

levels, land subsidence, land-surface deformation, 

aquifer water-quality, and regional ground-water flow 

patterns. A cycle consists of three periods: an injection 

period during which water is injected into the aquifer 

through a well, a storage period during which the well 

is idle, and a recovery period during which water is 

extracted from the aquifer by pumping from the same 

well. Surface water used for injection is treated and 

disinfected (usually by chlorination) prior to injection 

into the aquifer system to prevent biofouling and 

potential introduction of microbial contaminants into 

the aquifer.

Water-quality monitoring during the first two 

cycles showed high levels of trihalomethanes (THM) in 

the extracted water during the initial stage of pumping 

(Los Angeles County Department of Public Works, 

2000). THMs are volatile, halogenated organic 

compounds and include chloroform (CHCl3), 

bromodichloromethane (CHCl2Br), 

dibromochloromethane (CHClBr2), and bromoform 

(CHBr3). They are carcinogenic disinfection byproducts (DBP) formed by reaction between natural 

dissolved organic carbon (DOC) present in the water 

and chlorine added during the drinking-water-treatment 

process. The U.S. Environmental Protection Agency 

(EPA) regulates the concentrations of THMs and other 

DBPs in finished drinking water (U.S. Environmental 

Protection Agency, 1998). THM concentrations in the 

extracted water during the initial stage of pumping 

exceeded the EPA maximum contaminant level (MCL) 

of 80 micrograms per liter (μg/L). LACDPW blended 

the extracted water with water from other sources to 

lower THM levels in the water delivered to its 

customers to a level below the MCL. The more serious 

problem, however, was that the extraction water still 

contained measurable concentrations of THMs long 

after continuous pumping had presumably retrieved all 

of the injected water (Los Angeles County Department 

of Public Works, 2000). This observation raised 

concerns about the long-term effect of injection, 

storage, and recovery on aquifer water quality and thus 

indicates a potential problem for the feasibility of using 

this water-supply method in the Antelope Valley. 

Research and monitoring experiments during the third 

test cycle (March 1998 through April 1999) were 

expanded to include investigation of the formation and 

fate of THMs during the cycle. 
4 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

Figure 1. Locations of sites and geographic features relevant to the third injection, storage, and recovery test (March 1998 through April 1999) at 

Lancaster, Antelope Valley, California.

Sacramento

Los Angeles

0 50

0 50 100 miles

100 kilometers

124o 120o

42o

42o

37o 37o

116o

120o

33o

116o

33o

Lake

Isabella

San Diego

Sacramento River

California Aqueduct

Kern River

San Joaquin River

CALIFORNIA

Lancaster

H.O.

Banks

pumping

plant

Delta

State Water Project

San Francisco

Mojave Desert

Antelope Valley
I. Introduction 5

Purpose and Scope

The roles of the USGS in the injection, storage, 

and recovery tests were to collect and analyze 

hydraulic and aquifer-system deformation data, to 

develop a simulation/optimization model to design and 

manage a larger-scale injection program, and to 

determine the factors controlling the formation and fate 

of THMs in the aquifer system. 

This report presents detailed discussions of the 

factors controlling the formation and fate of THMs in 

the aquifer system during the third cycle. The report is 

divided into five sections:

• The introductory section consists of a review of 

previous work and brief descriptions of the project 

design, the Lancaster site, and the chronology of 

the third cycle.

• The second section includes discussion of the 

results of monitoring of water quality in the 

injection/extraction well and a nearby set of nested 

piezometers and of THM formation experiments, 

evaluation of processes potentially controlling 

THM concentrations based on these results, 

presentation of mass-balance calculations for 

THMs and other constituents, and development of 

a conceptual model for water flow in the aquifer 

during the cycle.

• The third section discusses the results of the sulfur 

hexafluoride (SF6) tracer experiment and presents 

mathematical models of mixing between ground 

water and injected water.

• The fourth section discusses the results of 

experiments that assessed the potential for 

biodegradation of THMs in the aquifer.

• The final section summarizes the factors 

controlling the formation and fate of THMs in the 

aquifer system during the third cycle and discusses 

the implications for using injection, storage, and 

recovery as a water-supply method in Lancaster.

A series of companion reports presents the other 

parts of the project. A compilation of the hydrologic 

data and the land-surface and aquifer deformation data, 

and methods for collection are reported by Metzger and 

others (2002). The analytical methods used and data 

collected for the investigation of the formation and fate 

of THMs are reported by Fram and others (2002). A 

description of the use of microgravity surveys to 

determine water-level changes will be presented in a 

forthcoming report by Howle and others (in press). 

Another report will present hydraulic and deformation 

data collected at the Lancaster site and describe a 

simulation/optimization model developed to evaluate 

the feasibility of using injection, storage, and recovery 

as a water-supply method in the Lancaster area 

(Phillips and others, in press).

Previous Studies

The formation and fate of THMs during 

injection, storage, and recovery cycles have been 

investigated in several other locations. The behavior of 

THMs during multiple cycles of injection, storage, and 

recovery has been extensively studied at a site in 

Las Vegas, Nevada. Treated water from Lake Mead 

was injected through new and existing wells into an 

aquifer composed of oxidized, alluvial fan sediments, 

and water was later extracted after a period of storage 

(Brothers and Katzer, 1990). THM concentrations in 

the extracted water at the beginning of the recovery 

period were higher than the concentrations in the 

injection water and the difference was likely due to the 

continued reaction of residual free chlorine that was 

present in the injected water (Miller and others, 1993). 

THM concentrations in the extracted water decreased 

during the recovery period. Geochemical mixing 

models constructed using the major-ion chemistry of 

the injected water, ground water, and extracted water 

indicated that the THM decrease observed during the 

extraction period of the first several cycles could be 

accounted for solely by mixing between injection water 

and ground water during extraction (Miller and others, 

1993; Bernholtz and others, 1995). In subsequent 

cycles, the concentrations of brominated THM species 

(CHCl2Br, CHClBr2, and CHBr3) decreased more than 

predicted for dilution by mixing between injection 

water and ground water. Thomas and others (2000) 

suggested that multiple cycles resulted in acclimation 

of bacteria capable of degrading brominated THMs, 

either under aerobic conditions or in anaerobic micro 

environments developed in the aquifer. However, the 

concentrations of CHCl3 were still completely 

accountable by dilution due to mixing, and experiments 

indicated that aquifer bacteria were not capable of 

degrading CHCl3 under any conditions (Landmeyer 

and others, 2000). 
6 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

Singer and others (1993) also studied the fate of 

THMs in the Las Vegas injection, storage, and recovery 

program and reached a different conclusion concerning 

the importance of mixing between injected water and 

ground water. They postulated that mixing between 

injected water and ground water could not have 

occurred because the regional lateral ground water flow 

rates were too small to have permitted significant 

motion of the ground water during the time period of 

the cycle. Therefore, they attributed the observed 

decrease in THM concentrations during storage and 

recovery to biodegradation. However, they noted that 

the high dissolved oxygen and nitrate concentrations in 

the extracted water samples precluded widespread 

anoxic conditions in the aquifer (biodegradation 

requires anoxic conditions) and suggested that some 

unknown aquifer heterogeneity factor must be 

important for producing conditions conducive to 

biodegradation of THMs (Singer and others, 1993).

The fate of THMs during injection, storage, and 

recovery cycles at a site in Charleston, South Carolina, 

was different than at the Las Vegas site, primarily 

because the nature of the aquifer is different. The 

aquifer beneath Charleston is confined, anoxic, and 

composed of fossiliferous, sandy limestone (Campbell 

and others, 1997; Mirecki and others, 1998). 

Geochemical modeling of the major-ion chemistry of 

injected water, ground water, and extracted water 

showed that the composition of the extracted water 

could be accounted for by mixing between injected 

water and ground water and dissolution of aquifer 

materials (Mirecki and others, 1998). THM 

concentrations in the extracted water decreased during 

the recovery period, but no attempt was made to 

quantitatively account for the decrease (Campbell and 

others, 1997). Examination of the data indicates that 

the concentration of THMs decreased faster than 

predicted for mixing between injected water and 

ground water as defined by variations in the 

concentration of conservative constituents, suggesting 

that a process in addition to dilution from mixing 

contributed to the decrease. Based on the behavior of 

THMs in injection, storage, and recovery tests in other 

anoxic aquifers, the additional process was probably 

biodegradation of brominated THMs. The 

concentration of brominated THMs decreased during 

the storage periods of injection, storage, and recovery 

cycles in anaerobic (denitrifying) aquifers in Florida, 

Colorado, and London, England (Singer and others, 

1993). Experimental data consistently show that 

biodegradation of brominated THMs does occur under 

anaerobic conditions (for example, Bagley and Gossett, 

1995).

THM concentrations in stored water in an 

injection and storage pilot project at the margin of the 

San Francisco Bay in Palo Alto, California, showed 

evidence of sorption and biodegradation of THMs 

(Roberts and others, 1982; Roberts, 1985). Treated 

wastewater was injected into a shallow silty sand 

aquifer confined by clay units and containing ground 

water with low dissolved oxygen concentrations. 

Breakthrough curves measured at nearby observation 

wells yielded estimated retention capacities of 2.5 and 

4.0 micrograms per kilogram (μg/kg) for CHCl3 and 

CHBr3, respectively, compared to the unretained tracer 

constituent, chloride (Roberts and others, 1982). 

Continued monitoring during the storage period 

showed consumption of the dissolved oxygen in the 

injected water and decreases in the concentrations of 

all the THM species, particularly of the brominated 

THM species. Roberts and others (1982) suggest that 

the decreases in THM concentrations during the 

storage period were due to biodegradation.

Project Design

The previous studies of the fate of THMs in 

injection, storage, and recovery projects identified four 

processes that may affect the concentration of THMs in 

water that has been injected, stored, and extracted from 

an aquifer: continued formation of THMs in the 

injected water after injection into the aquifer, 

biodegradation of THMs by bacterial communities in 

the aquifer, sorption of THMs to aquifer materials, and 

mixing between injected water containing THMs and 

ground water containing little or no THMs. Evaluation 

of these four processes would provide answers to the 

primary questions concerning the formation and fate of 

THMs during the injection, storage, and recovery 

cycles:

• What controls the continued formation of THMs 

in the aquifer after injection?

• What causes the continued presence of low 

concentrations of THMs in the extracted water 

after all the injection water presumably has been 

retrieved?

• What causes the decrease in THM concentrations 

as extraction proceeds?
I. Introduction 7

• Are there natural attenuation mechanisms that can 

decrease the THM concentrations in the aquifer?

The continued formation of THMs in the 

injected water was investigated in two types of 

laboratory experiments. THM formation potential 

(THMFP) experiments were done to assess the 

compositional nature of the THM-forming DOC in the 

injection water and to determine the maximum amount 

of THM formation possible from the injection water. 

Injection water was stored for 1 to 16 weeks under 

controlled laboratory conditions to determine whether 

THM formation in the injection water was limited by 

the amount of DOC or the amount of residual free 

chlorine in the injected water. 

The potential for reduction of THM 

concentrations by the natural attenuation process of 

biodegradation of THMs in the aquifer was assessed in 

laboratory experiments. The biodegradation 

experiments included several types of microcosms 

containing ground water or injection water with or 

without sediment from the aquifer, and were incubated 

under aerobic and anaerobic conditions. The potential 

for sorption of THMs to aquifer materials was 

evaluated using the water-quality data obtained during 

the injection, storage, and recovery cycle.

Mixing between injected water and ground water 

may also decrease THM concentrations in the aquifer 

and may account for the continued presence of low 

concentrations of THMs in the extracted water after all 

the injection water presumably has been retrieved. 

Mixing was evaluated by a tracer experiment, by 

mathematical modeling of mixing, and by mass 

balance calculations for dissolved constituents. A 

conservative tracer, sulfur hexafluoride (SF6), was 

added to the injection water, and concentrations in the 

extraction water were monitored so that the proportion 

of injected water in the extracted water could be 

determined. Variations in dissolved constituents in the 

extraction water during the recovery period were 

compared with concentrations predicted using a 

mathematical model of a simple mixing process. 

In addition, water-quality samples were collected 

periodically from the well used for both injection and 

extraction to determine the composition of the injection 

water and, later, the extraction water. Water-quality 

samples also were collected from a nearby set of nested 

piezometers. These samples provided a time series of 

water-quality data used to delineate the behavior of 

THMs and other chemical constituents during the 

cycle.

Site Description and Chronology of Third Cycle

The injection, storage, and recovery test site is 

located in the Antelope Valley near the city of 

Lancaster, California (fig.1). A detailed description of 

the site is given in Fram and others (2002) and Metzger 

and others (2002); only a brief summary will be given 

here. 

Antelope Valley is a topographically closed basin 

at the western end of the Mojave Desert; it is 

subdivided into 12 ground-water subbasins bounded by 

faults and bedrock outcrops (Bloyd, 1967; Carlson and 

others, 1998). The three injection, storage, and 

recovery tests occurred in the Lancaster subbasin at the 

LACDPW’s Avenue L and 5th Street West well field in 

Lancaster (fig. 2).

The Lancaster subbasin contains alluvial and 

lacustrine deposits, which are locally as much as 

5,000 ft thick (Mabey, 1960; Dibblee, 1967; Londquist 

and others, 1993). The alluvial deposits consist of 

interbedded heterogeneous mixtures of silt, sand, and 

gravel (Dutcher and Worts, 1963; Bloyd, 1967); the 

lacustrine deposits primarily consist of thick layers of 

clay, interbedded with thinner sand and silty sand 

layers (Dibblee, 1967). Stratigraphic, hydrologic, and 

water-quality data were used to divide the deposits into 

three aquifers: an upper, a middle, and a lower aquifer 

(Leighton and Phillips, 2003). At the injection, storage, 

and recovery demonstration site, the upper aquifer 

extends from the water table to a depth of about 510 ft 

below land surface, the middle aquifer extends from 

about 510 to about 730 ft below land surface, and the 

lower aquifer extends from about 870 ft below land 

surface to the bedrock (fig. 2). Ground-water flow in 

the upper aquifer is unconfined, flow in the middle 

aquifer is partially confined by fine-grained sediment 

layers, and flow in the lower aquifer is confined by the 

lacustrine deposit that separates the middle and lower 

aquifers. 
8 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

Two wells were used during the third injection, 

storage, and recovery cycle: wells 7N/12W-27P2 

(well 4-32) and 7N/12W-27P3 (well 4-34) (fig. 2,

table 1). (The local names for the injection and 

extraction wells [in parentheses above] are used for the 

convenience of readers more familiar with these 

names.) Wells 4-32 and 4-34 penetrate the upper and 

middle aquifers and are screened from 282 to 717 ft 

and 280 to 710 ft below land surface, respectively 

(fig. 2). Well 4-34 is about 180 ft west of well 4-32. 

In February 1998, a set of four nested 

piezometers, 7N/12W-27P5–8, was installed in a 

borehole about 80 ft east-northeast of well 4-32 (fig. 2, 

table 1). (The local names for the piezometers are not 

used in this report.) Borehole geophysical logs were 

used to determine the most suitable depth for the 

screened interval for each piezometer (Metzger and 

others, 2002). The piezometers were screened at depths 

of 330–370 ft (27P8), 440–460 ft (27P7), 540–560 ft 

(27P6), and 890–910 ft (27P5) below land surface 

(fig. 2). The deepest piezometer, 27P5, was placed in 

the lower aquifer and was not used for this project. A 

well-bore velocity log completed at well 4-32 under 

pumping conditions showed that most of the water 

extracted from the well came from a high flow zone in 

the upper aquifer at about 460 to 510 ft below land 

surface (Phillips and others, in press). Piezometer 27P6 

was installed in the upper part of the middle aquifer, 

and piezometer 27P7 was installed in the lower part of 

the upper aquifer at the approximate depth of the 

maximum flow zone in well 4-32 (fig. 2). Piezometer 

27P8 was installed near the water table at the time of 

construction. The water level fluctuated during the 

injection, storage, and recovery cycles; for example, 

the water level in piezometer 27P8 ranged from 302 ft 

below land surface in mid-May through mid-June 1998 

(during the injection period) to 347 ft below land 

surface in April 1999 (during the extraction period) 

(fig. 3C). Detailed descriptions of cores collected 

during installation of the piezometers are given in Fram 

and others (2002).

Figure 2. Generalized subsurface geology and locations of wells and nested piezometers used for the injection, storage, and recovery test (March 1998 

through April 1999), Lancaster, Antelope Valley, California. 

Local well names are in parentheses.

0

200

400

600

800

1,000

FEET BELOW LAND-SURFACE

Injection/extraction

7N/12W-27P2

(well 4-32)

Extraction

7N/12W-27P3

(well 4-34)

Nested

piezometers

7N/12W-27P5_8

27P8

27P7

27P6

27P5

AVENUE L

5TH STREET WEST

180 feet 80 feet

27P5_8

Well 4-34 Well 4-32

34°39.43'

N 118°08.17'

distances are

approximate

Mean water table

Unsaturated

zone

Upper

aquifer

Middle

aquifer

Lacustrine

unit

Lower aquifer

screened

interval
I. Introduction 9

For all three cycles, the water used for injection 

into the wells was imported from the SWP (fig. 1). 

Existing AVEK pipelines conveyed water from the 

SWP to the West Quartz Hill Water Treatment Plant 

where it was treated with chlorine. This treated water 

was then transported in LACDPW and AVEK pipelines 

to well 4-32.

During the injection periods of the first and 

second cycles (1996 and 1997), the water delivered by 

the SWP originated from the Sacramento–San Joaquin 

Delta and was conveyed by the H.O. Banks pumping 

plant and the SWP’s California Aqueduct (fig. 1). 

However, during the injection period of the third cycle 

(1998), the water delivered from the SWP originated 

from Lake Isabella and the Kern River (fig. 1). This 

source water was chemically different from the Delta 

water used during the first two cycles and resulted in 

some differences in the water-quality patterns observed 

in the third cycle compared with the first two cycles.

Injection at well 4-32 for the third cycle began 

April 15, 1998, and continued through June 16, 1998. 

Water flow into the wellhead was maintained between 

700 and 800 gallons per minute (gal/min) (fig. 3A). 

From the well, the injected water moved into the 

aquifer by gravity flow. The total volume injected was 

58 million gallons (fig. 3B). The injection period was 

followed by 2 weeks of water storage in the aquifer 

during which time no pumping occurred. 

The first phase of extraction from well 4-32 

began June 30, 1998, and ended October 24, 1998, 

during which time water flow was maintained at 

400–550 gal/min (fig. 3A). The volume of water 

extracted during this phase was 77 million gallons, 

which was 1.3 times the volume of water injected 

(fig. 3B). No extraction occurred between October 24, 

1998, and February 22, 1999, due to failure and 

replacement of the pump. The second phase of 

extraction from well 4-32 began February 22, 1999, 

and ended April 29, 1999, during which time water 

flow was maintained at 750–800 gal/min and 

73 million gallons of water were extracted. The total 

volume of water extracted from well 4-32 was 

150 million gallons, which was 2.5 times the volume of 

water injected (fig. 3B). After the well 4-32 pump 

failed, water was extracted from nearby well 4-34 to 

meet water demand. Extraction from well 4-34 began 

December 28, 1998, and continued through April 29, 

1999, and the total volume of water extracted from 

well 4-34 was 154 million gallons (fig. 3B). The 

extracted water was blended with other water and 

incorporated into the LACDPW water distribution 

system.

Water-quality samples were collected from well 

4-32 and the nested piezometers 27P6–8 during the 

entire injection, storage, and recovery cycle (March 

1998 through April 1999). Descriptions of sampling 

intervals and procedures, analytical methods, and a 

compilation of all of the water-quality data obtained 

during the third cycle are reported in Fram and others 

(2002). Water samples also were collected for use in 

the THM formation and biodegradation experiments. 

The methods used for these experiments and the data 

obtained also are reported in Fram and others (2002). 

Table 1. Nomenclature for wells used during the third injection, storage, and recovery cycle (March 1998 through April 1999), Lancaster, Antelope Valley, 

California

State well 

number 

7N/12WLocal well 

number

USGS location 

identification number Aquifer zone screened Use of well

27P5 5L-PZ1 343943118081701 Lower Not used

27P6 5L-PZ2 343943118081702 Middle Piezometer

27P7 5L-PZ3 343943118081703 Upper Piezometer

27P8 5L-PZ4 343943118081704 Upper Piezometer

27H3 4-33 344088118074701 Upper and middle Extraction

27J4 4-13 344002118074701 Upper and middle Extraction

27J6 4-42 344003118074901 Upper, middle, and lower Extraction

27P2 4-32 343943118081801 Upper and middle Injection and Extraction

27P3 4-34 343943118082101 Upper and middle Extraction
10 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

Figure 3. Average daily pumping rates (A) and cumulative volumes of water injected and extracted (B) at well 7N/12W-27P2 (well 4-32) and well 

7N/12W-27P3 (well 4-34), and average daily water levels at piezometers 7N/12W-27P6–8 (C) during the third injection, storage, and recovery cycle (March 

1998 through April 1999), Lancaster, Antelope Valley, California.

V is the cumulative volume of water extracted (positive numbers) or injected (negative numbers) relative to the total volume of water injected at well 4-32. 

Methods for data collection are described by Metzger and others (2002).

Wells

PUMPING RATE,

IN GALLONS PER MINUTE

0

200

400

600

800

1,000

well 4-34

well 4-32

DEPTH TO WATER TABLE,

IN FEET BELOW LAND SURFACE

290

310

330

350

370

390

MAMJ J A S O ND J F MA

1998 1999

Piezometers

DATE

27P8 (330-370 ft)

27P7 (440-460 ft)

27P6 (540-560 ft)

C

CUMULATIVE VOLUME OF WATER

INJECTED AND EXTRACTED,

IN MILLIONS OF GALLONS

0

40

80

120

160

B

A

Wells

well 4-34

well 4-32

-1.0 -0.5 0

0.5 V = 1.0 1.5 2.0 2.5

(pump broken)

Injection Extraction
II. Formation and fate of trihalomethanes during the injection, storage, and recovery cycle 11

Acknowledgments

This study was done by the USGS in cooperation 

with the Los Angeles County Department of Public 

Works (LACDPW) and the Antelope Valley–East Kern 

Water Agency (AVEK). Discussions with Loren 

Metzger, Joshua Berghouse, and Steven Phillips during 

preparation of this report, reviews by Steven Phillips, 

James Thomas, Rick Iwatsubo and Peter Martin, and 

editing by Mary Gibson contributed immensely to this 

work. 

II. FORMATION AND FATE OF 

TRIHALOMETHANES DURING THE 

INJECTION, STORAGE, AND RECOVERY 

CYCLE

By Miranda S. Fram

In this section of the report, concentrations of 

dissolved constituents in the ground water, injection 

water, and extraction water are used to assess the roles 

of mixing, biodegradation, and sorption in determining 

the concentrations of THMs in the extracted water. An 

important part of this assessment is determining factors 

that controlled continued formation of THMs in the 

aquifer after injection.

Studies of other injection, storage, and recovery 

programs have used the geochemistry of the ground 

water, injection water, and extraction waters to quantify 

the amount of mixing of injected water and ground 

water, and to constrain the importance of other 

processes such as biodegradation, sorption, and 

precipitation/dissolution. For example, Mirecki and 

others (1998) used inverse modeling of the major-ion 

concentrations in extracted water from the Charleston, 

South Carolina program to show that extracted water 

compositions represented different mixtures of the 

mean injected water and ground water compositions 

plus dissolution of evaporite and carbonate minerals in 

the aquifer. 

The geochemistry of ground water, injection 

water, and extraction water was used to address the fate 

of THMs in the Las Vegas, Nevada program (Brothers 

and Katzer, 1990; Miller and others, 1993). Mass 

balances were constructed for THMs and the injection 

water during a cycle using the mean compositions of 

ground water and injected water, electrical conductivity 

as a conservative tracer, and measured compositions of 

the extracted water. The percentage of injected THMs 

recovered was similar to the percentage of the tracer of 

the injection water recovered, but both were lower than 

the percentage of the volume of the injected water 

recovered, indicating mixing between injected water 

and ground water. In addition, a THM formation 

experiment compared THM concentrations at three 

points in time: at the time of injection, after 

consumption of any residual chlorine in the water, and 

after addition of excess chlorine. The results indicated 

that the extracted water had the potential to form a 

mean of 25 μg/L additional THMs when re-chlorinated 

after extraction (Miller and others, 1993). Inverse 

modelling of the major-ion geochemistry of extracted 

waters showed that the compositions could be 

adequately explained by mixing of mean ground-water 

and injection-water compositions and that no 

significant mineral precipitation had occurred in the 

aquifer (Brothers and Katzer, 1990). Subsequent 

studies of the Las Vegas program used chloride as a 

conservative tracer to quantify the proportions of three 

components in samples of extracted water: injected 

water, native ground water, and injected water left over 

from previous cycles. Comparison between 

concentrations of chloride and of individual THM 

species indicated that CHCl3 concentrations in the 

extracted water were consistent with conservative 

mixing, but that the concentrations of the brominated 

THM species were decreased by some additional 

process during residence in the aquifer (Thomas and 

others, 2000). 

Variations in Water Quality and Water Levels 

during the Third Cycle

The variations in concentrations of chemical 

constituents throughout the cycle provide insight into 

processes within the aquifer during the cycle. Data 

discussed in this report are given by Fram and others 

(2002), Metzger and others (2002), and tables 2 and 3.
12 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

Figure 4. Trihalomethane (A) and chloride (B) concentrations in water collected from well 7N/12W-27P2 (well 4-32) during the three injection, storage, 

and recovery cycles (April 1996 through April 1999), Lancaster, Antelope Valley, California. 

Data are from Fram and others (2002) and Metzger and others (2002).

DATE

AM J J A S ON D J F M AM J JA S ON D J FMAM J JA S OND J FMA

CHLORIDE,

IN MILLIGRAMS PER LITER

1996 1997 1998 1999

A

B

Cycle 1 Cycle 2 Cycle 3

TRIHALOMETHANES,

IN MICROGRAMS PER LITER

0

20

40

60

80

100

0

20

40

60

80

100

120

140

Injection periods

Extraction and

storage periods
II. Formation and fate of trihalomethanes during the injection, storage, and recovery cycle 13

The pattern of THM concentrations in the 

injected and extracted water from well 4-32 during the 

third injection, storage, and recovery cycle resembled 

those of the two previous cycles—moderate 

concentrations in the injection water, a large increase in 

concentration at the beginning of extraction, and a 

gradual decrease as extraction proceeded (fig. 4A). 

However, there were some important differences in 

water chemistry between the third cycle and the two 

previous cycles. THM concentrations in the injection 

water and in the extraction water during the early part 

of the extraction periods were higher in the first two 

cycles than in the third; the EPA MCL of 80 μg/L was 

exceeded in four extraction water samples in each of 

the first two cycles and none in the third cycle (fig. 4A). 

However, if the THM concentrations are expressed on a 

molar rather than a mass concentration basis, the 

concentrations in the three cycles were similar. The 

THMs in the first two cycles contained much higher 

proportions of brominated species (CHCl2Br, 

CHClBr2, and CHBr3) than did the THMs in the third 

cycle. (Bromine has approximately twice the molar 

mass of chlorine; thus the brominated THMs are more 

massive than CHCl3.)

SWP water used for injection during the first two 

cycles originated primarily from the Sacramento–

San Joaquin Delta, whereas water used in the third 

cycle originated from the Kern River (fig. 1). The mean 

concentrations of DOC in the injection water in the 

second and third cycles were similar—1.80 and 1.76 

milligrams per liter (mg/L), respectively—but the 

concentrations of inorganic constituents were very 

different (Fram and others, 2002; Metzger and others, 

2002). The concentration of chloride in water injected 

into and extracted from well 4-32 during the three 

cycles is shown in figure 4B as an example. The median 

chloride concentration in the injection water in the 

second cycle was 52 mg/L, which was significantly 

higher than the concentration in the ground water, 

whereas the median chloride concentration in the 

injection water in the third cycle was 7 mg/L, which 

was lower than the concentration in the ground water. 

As a result, chloride concentrations in the extracted 

water decreased during extraction in the first two 

cycles, but increased during extraction in the third 

cycle (fig. 4B). 

These differences in injection water chemistry 

between the first two and the third cycles are not 

expected to cause a significant difference in the 

processes affecting the formation and fate of THMs 

during the three cycles. Although this report only 

discusses the third cycle, it is expected that the results 

can be applied to the other cycles at the Lancaster site, 

with few modifications due to changes in injection 

water composition.

The constituents that are important for assessing 

the formation and fate of THMs during the third cycle 

are THMs, DOC, residual chlorine, bromide, the ratio 

of bromine to chlorine in the THM ((Br/Cl)THM), and 

an appropriate conservative tracer such as chloride. The 

variations of these constituents in the ground water, 

injection water, and water extracted from well 4-32 and 

the nested piezometers during the third cycle are shown 

in figure 5 and discussed here.

Trihalomethanes

Third cycle injection water contained 22 to 

40 μg/L THMs (fig. 5A) with a mean concentration of 

27.5 μg/L. The concentration of THMs in water 

sampled from well 4-32 prior to the third injection was 

less than 2 μg/L (Los Angeles County Department of 

Public Works, 2000). THM concentrations in the 

extracted water decreased from about 59 μg/L at the 

beginning of extraction to about 10 μg/L on October 

21, 1998 (fig. 5A), when extraction was halted due to 

failure of the pump on October 24, 1998. When 

extraction resumed on February 22, 1999, THM 

concentrations in the extracted water were 15 to 

26 μg/L and then dropped steadily to 3 μg/L as 

extraction continued through April 28, 1999 (fig. 5A). 

THM concentrations measured by the USGS and 

LACDPW in water samples from the nested 

piezometers showed poor agreement (table 2). Both 

sets of data are plotted in figure 5A. The differences 

may be due to the imperfect sampling techniques used 

to obtain undegassed water samples from the 

piezometers (see Fram and others, 2002). In that case, 

the highest concentration reported for each sample may 

be the most accurate.

On March 12, 1998, prior to the third cycle, 

water from piezometers 27P6 and 27P7 had no 

detectable concentrations of THMs, whereas water 

from piezometer 27P8 contained 2.4 μg/L THMs 

(table 2). Because THMs do not naturally occur in 

ground water at these concentrations, the THMs must 

have been residual from the previous injection, storage, 

and recovery cycles. 
14 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

Figure 5. Trihalomethane concentrations (A), dissolved organic carbon concentrations (B), ultraviolet absorbance values at 254 nanometers (C), pH 

values (D), free and total residual chlorine concentrations (E), chloride concentrations (F), bromide concentrations (G), and molar bromine to chlorine ratio 

in trihalomethanes (H, I) in ground water, injection water, and extraction water collected from well 7N/12W-27P2 (well 4-32) and from nested piezometers

7N/12W-27P6–8 during the third injection, storage, and recovery cycle (March 1998 through April 1999), Lancaster, Antelope Valley, California. 

Data are from Fram and others (2002), Metzger and others (2002), and tables 2 and 3. TRIHALOMETHANES, IN MICROGRAMS PER LITER

0

20

40

60

80

DISSOLVED ORGANIC CARBON,

IN MILLIGRAMS PER LITER

0.0

0.5

1.0

1.5

2.0

2.5

0.00

0.01

0.02

0.03

0.04

DATE

MAM M JJ JF A ASOND

1998 1999

well 4-32

27P6

27P7

27P8

well 4-32

27P6

27P7

27P8

well 4-32

27P6

27P7

27P8

Injection

Injection

Injection

A

B

C

ULTRAVIOLET ABSORBANCE AT 254

NANOMETERS, IN PER CENTIMETER
II. Formation and fate of trihalomethanes during the injection, storage, and recovery cycle 15

Figure 5.—Continued.

DATE

MAM M JJ JF A ASOND

1998 1999

Injection

Injection

Injection

D

E

F

pH,

IN STANDARD UNITS

5

6

7

8

9

RESIDUAL CHLORINE,

IN MILLIGRAMS PER LITER

0.0

0.4

0.8

1.2

1.6

total residual chlorine

free residual chlorine

CHLORIDE,

IN MILLIGRAMS PER LITER

0

5

10

15

20

25

30

well 4-32

27P6

27P7

27P8

43

well 4-32

27P6

27P7

27P8
16 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

Figure 5.—Continued.

DATE

MAM M JJ JF A ASOND

1998 1999

Injection

Injection

Injection

G

H

I

BROMIDE,

IN MILLIGRAMS PER LITER

well 4-32

MOLAR BROMINE-TO-CHLORINE

RATIO IN TRIHALOMETHANE

MOLAR BROMINE-TO-CHLORINE

RATIO IN TRIHALOMETHANE

well 4-32

stored injection water

27P6

27P7

27P8

area shown in panel I

well 4-32

stored injection

water

27P6

27P7

27P8

0.00

0.05

0.10

0.15

0.20

0.25

0.0

0.3

0.6

0.9

0.02

0.04

0.06

0.08

0.10

0.12

0.14
II. Formation and fate of trihalomethanes during the injection, storage, and recovery cycle 17

Table 2. Trihalomethane concentration data for water collected from nested piezometers 7N/12W-27P6–8 during the third injection, storage, and 

recovery cycle (March 1998 through April 1999), Lancaster, Antelope Valley, California

[Samples analyzed at two laboratories: U.S. Geological Survey (USGS), Sacramento District and Los Angeles County Department of Public Works 

(LACDPW). CHCl3, chloroform; CHCl2Br, bromodichloromethane; CHClBr2, dibromochloromethane; CHBr3, bromoform; THM, trihalomethane. 

μg/L, microgram per liter; nd, not detected; nr, not reported; <, less than]

Sampling date Analyzing

agency

CHCl3

(μg/L)

CHCl2Br

(μg/L)

CHClBr2

(μg/L)

CHBr3

(μg/L)

Total THM

(μg/L)

Piezometer 27P6

3/12/98 LACDPW nd nd nd nd nd

6/15/98 USGS 47.8 8.8 1.0 <0.2 57.6

6/15/98 LACDPW 47.0 9.3 nd nd 56.3

8/4/98 USGS 43.9 11.0 1.4 <0.2 56.3

8/4/98 LACDPW 36.9 7.7 1.6 nd 46.2

9/3/98 USGS 45.9 8.9 1.2 <0.2 56.0

9/3/98 LACDPW 32.8 9.1 2.8 nd 44.7

10/7/98 USGS 58.2 9.9 1.7 <0.2 69.8

10/7/98 LACDPW nr nr nr nr 36.5

11/5/98 USGS 39.0 6.2 1.0 <0.2 46.2

11/5/98 LACDPW nr nr nr nr 49.8

12/2/98 USGS 33.8 6.1 1.1 <0.2 41.0

12/3/98 LACDPW nr nr nr nr 60.9

3/24/99 USGS 16.0 4.9 1.4 <0.2 22.3

Piezometer 27P7

3/12/98 LACDPW nd nd nd nd nd

6/15/98 USGS 50.2 9.2 0.8 <0.2 60.2

6/15/98 LACDPW 46.1 9.0 1.1 nd 56.2

8/4/98 USGS 50.5 12.2 1.0 <0.2 63.7

8/4/98 LACDPW 21.1 4.9 nd nd 26.0

9/3/98 USGS 48.6 9.4 1.0 <0.2 59.0

9/3/98 LACDPW 41.5 7.4 1.4 nd 50.3

10/7/98 USGS 62.2 10.7 1.6 <0.2 74.5

10/7/98 LACDPW nr nr nr nr 48.2

11/5/98 USGS 37.0 6.1 0.8 <0.2 43.9

11/5/98 LACDPW nr nr nr nr 63.2

12/2/98 USGS 22.2 4.9 0.9 <0.2 28.0

12/3/98 LACDPW nr nr nr nr 28.6

3/24/99 USGS 11.2 3.4 1.0 <0.2 15.6

Piezometer 27P8

3/13/98 LACDPW 2.4 nd nd nd 2.4

6/15/98 USGS 47.2 8.8 1.0 <0.2 57.0

6/15/98 LACDPW 40.9 9.2 1.2 nd 51.3

8/4/98 USGS 46.6 13.1 <0.5 <0.2 59.7

8/4/98 LACDPW 17.6 4.3 nd nd 21.9

9/3/98 USGS 26.7 5.8 <0.5 <0.2 32.5

9/3/98 LACDPW 23.3 6.0 nd nd 29.3

10/7/98 USGS 44.1 8.7 <0.5 <0.2 52.8

10/7/98 LACDPW nr nr nr nr 45.4

11/5/98 USGS 25.7 6.0 <0.5 <0.2 31.7

11/5/98 LACDPW nr nr nr nr 34.4

12/2/98 USGS 23.8 7.2 2.4 <0.2 33.4

12/3/98 LACDPW nr nr nr nr 49.1

3/24/99 USGS 29.8 9.6 0.5 <0.2 39.9
18 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

THM concentrations in water samples collected 

from the nested piezometers were nearly always higher 

than concentrations in water collected from well 4-32 

on the same date during the injection and extraction 

periods (fig. 5A). At the end of the injection period, 

water from all three piezometers contained about 

60 μg/L THM. THM concentrations in water from 

piezometers 27P6 and 27P7 remained at about 60 μg/L 

through September and then increased to a maximum 

of 75 μg/L on October 4, 1998. These high THM 

concentrations in piezometers 27P6 and 27P7 are 

noteworthy because on September 26, 1998, the 

volume of water extracted from well 4-32—just 80 ft 

away—exceeded the volume of water injected 

(fig. 3B). THM concentrations in piezometers 27P6 and 

26P7 then declined, reaching lows of about 20 μg/L at 

the end of the cycle (fig. 5A). THM concentrations in 

water from piezometer 27P8 were somewhat more 

variable than those in water from the other two 

piezometers. This difference in behavior may be due to 

the location of the three piezometers relative to the 

water table. The screened interval of piezometer 27P8 

is near the pre-injection water table, and during the 

extraction period, the water table declined below the 

top of the screen, whereas the screened intervals for 

piezometers 27P6 and 27P7 were always significantly 

below the water table (figs. 2 and 3C). THMs are 

volatile compounds and may volatilize from pore 

waters into air in the pore spaces in the unsaturated 

zone above the water table.

Dissolved Organic Carbon and Ultraviolet Absorbance

DOC concentrations in the injection water 

ranged from 1.4 to 2.0 mg/L and had a mean value of 

1.76 mg/L. The concentration in the extraction water 

declined from about 1.8 mg/L to about 0.3 mg/L during 

the first part of extraction (June 30, 1998, through 

October 24, 1998), but varied unsystematically 

between 0.5 and 2.5 mg/L during the second part 

(fig. 5B). 

Water samples collected prior to the third 

injection, storage, and recovery cycle, in March 1998, 

from well 4-32 and piezometers 27P6 and 27P7 

contained about 0.2 mg/L DOC and likely represent the 

native ground-water composition. DOC concentrations 

in water from piezometers 27P6 and 27P7 reached a 

maximum of about 1.2 mg/L near the end of the 

injection period (June 15, 1998) and declined gradually 

to about 0.8 mg/L by the end of the first phase of the 

extraction period (fig. 5B). 

Ultraviolet light absorbance values at a 

wavelength of 254 nanometers (UVA254 values) 

(fig. 5C) showed patterns of variation during the cycle 

similar to those shown by DOC concentrations. 

UVA254 values ranged from 0.033 to 0.019 per 

centimeter (/cm) in the injection water and from 0.018 

to 0.007/cm in the extraction water. Many of the 

samples from the piezometers had UVA254 values 

greater than 0.04 (not shown in fig. 5C), which may 

have been due to interference from particulate 

materials during analysis (Fram and others, 2002).

pH

The pH of ground-water samples collected from 

wells 4-32 and 4-34 prior to each of the three cycles 

ranged from 7.2 to 8.3 (Los Angeles County 

Department of Public Works, 2000; Fram and others, 

2002; Metzger and others, 2002), and the pH of 

samples collected from the nested piezometers between 

August and December, 1998, ranged from 7.3 to 8.6 

(fig. 5D). These pH values may reflect equilibration 

between the water and the aquifer sediments. The 

ground-water samples also contained 12 to 26 mg/L of 

calcium (Los Angeles County Department of Public 

Works, 2000; Metzger and others, 2002), and the 

aquifer sediments contained horizons of caliche, a 

calcium carbonate precipitate commonly formed in 

soils in arid environments (Fram and others, 2002). 

Water in equilibrium with the atmosphere and solid, 

pure calcium carbonate has a pH of 8.3 and a calcium 

concentration of 20 mg/L (Stumm and Morgan, 1996). 

Thus, the pH and calcium concentrations in the 

ground-water samples are consistent with those in pore 

waters in equilibrium with oxidized, calcium 

carbonate-bearing sediments.
II. Formation and fate of trihalomethanes during the injection, storage, and recovery cycle 19

In contrast, the pH of the injection water ranged 

from 6.3 to 6.8 (median value = 6.6) (fig. 5D). The pH 

of the extracted waters was 6.0 at the beginning of 

extraction (fig. 5D), which was lower than the pH of all 

of the injected water and ground-water samples. This 

indicates that pH behaved non-conservatively in this 

system and, thus, must have been affected by reactions 

occurring in the system. Continued reaction between 

DOC and free chlorine in the injected water after 

injection into the aquifer may account for some of the 

observed decrease in pH. Oxidation-reduction 

reactions reduce free chlorine to chloride, oxidize 

DOC, and produce hydrogen ions (increased acidity, 

lower pH) (for example, Larson and Weber, 1994). An 

increase in carbon dioxide (CO2) concentration in the 

water upon injection also may account for some of the 

decrease in pH. Possible mechanisms for increasing 

CO2 concentration include entrainment of air during 

injection and microbial respiration of DOC in the 

injected water.

As extraction proceeded from well 4-32, before 

the pumping hiatus due to pump failure, the pH of the 

extracted water steadily increased toward the pH values 

observed in the ground-water samples (fig. 5D). This 

trend could be produced by mixing between injection 

water and ground water, or by gradual equilibration of 

injection water with the carbonate-bearing aquifer 

sediments. It is not possible from this data alone to 

determine the relative importance of mixing and 

equilibration. However, the major-ion chemical 

composition of water collected from the piezometers at 

the end of the injection period (June 15, 1998) closely 

resembles that of the injection water (Fram and others, 

2002; fig. 5F), but the pH values are significantly 

higher than those of the injection water, suggesting that 

some equilibration did occur.

Residual Chlorine

Total residual chlorine concentrations in the 

injection water at the time of injection ranged from 

0.7 to 1.38 mg/L, while free residual chlorine 

concentrations ranged from 0.5 to 1.0 mg/L (fig. 5E). A 

linear regression relation calculated using the seven 

samples in which both total and free residual chlorine 

were measured was used to estimate the free residual 

chlorine concentrations in the remaining samples. The 

mean free residual chlorine concentration for all the 

injection water samples was 0.79 mg/L (relative 

standard deviation [RSD] = 28 percent). Most samples 

of extraction water had undetectable residual chlorine 

concentrations; residual chlorine was only detected in a 

few samples collected during the first three months of 

extraction (fig. 5E).

Chloride and Bromide 

Chloride concentrations in the injection water 

ranged from 5.7 to 10 mg/L except during one short 

period when concentrations ranged from 19 to 26 mg/L 

(fig. 5F). Because only 17 samples were collected 

during the 2-month injection period, estimation of the 

average chloride concentration of the injection water is 

uncertain. Weighted averages were calculated by 

weighting the chloride concentration in the water each 

day by the daily volume of water injected. Daily 

chloride concentrations were interpolated from the 

measured values in several different ways, and the 

resulting average chloride concentration ranged 

between 7.5 and 9.0 mg/L. The chloride concentration 

in the extraction water was less than 9 mg/L during the 

first 10 days of extraction, suggesting that the lower 

end of the range estimated for the injection-water 

chloride concentration may be more accurate.

During the first phase of extraction from 

well 4-32, the chloride concentration in the extraction 

water increased gradually from about 7 to about 

20 mg/L, and during the second phase of extraction 

chloride concentrations in the extraction water 

decreased from about 20 to about 10 mg/L (fig. 5F).

Chloride concentrations in the ground water of 

the upper and middle aquifers near well 4-32 were 

heterogeneous. The chloride concentrations in water 

collected from well 4-32 shortly before the first, 

second, and third cycles were 14, 19, and 11 mg/L, 

respectively (fig. 5F), and the concentration in water 

from well 4-34 was 21 mg/L prior to both the first and 

second cycles and 36 mg/L prior to the third cycle (Los 

Angeles County Department of Public Works, 2000; 

Fram and others, 2002; Metzger and others, 2002). 

Wells 4-32 and 4-34 are screened continuously through 

the upper and middle aquifers (fig. 2), and water 

collected from them represents a mixture of water 

derived from different levels in the aquifers. Ten other 

wells that also are screened only in the upper and 

middle aquifers and are within 2.5 miles of well 4-32 

had chloride concentrations ranging from 3 to 28 mg/L 

prior to the third cycle (Los Angeles County 

Department of Public Works, 2000). The difference 

between chloride concentrations in wells 4-32 and 4-34 
20 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

indicates that the ground water in the upper and middle 

aquifers is naturally laterally heterogeneous in 

composition. 

The data from the piezometers indicates strong 

vertical heterogeneity as well. Water from piezometers 

27P6 and 27P7 contained 4 mg/L chloride prior to the 

third cycle, whereas water from piezometer 27P8 

contained 43 mg/L chloride (fig. 5F). However, 

because water from piezometer 27P8 prior to the third 

cycle contained THMs (fig. 5A), its high chloride 

concentration probably reflects a contribution from 

injection water from the previous cycle. Water from 

piezometer 7N/12W-27F8 contained 18 mg/L chloride 

(and no THMs) before the second cycle (Metzger and 

others, 2002). Like piezometer 27P8, it is screened 

from 395 to 415 ft in the upper aquifer, but is located 

about 2,200 ft north-northwest. The piezometer data 

suggest that chloride concentrations in the ground 

water may be greater near the top of the upper aquifer 

than at the base of the upper aquifer or the top of the 

middle aquifer where piezometers 27P7 and 27P6, 

respectively, are screened. No information is available 

about the chloride concentrations in ground water at 

the base of the middle aquifer.

During the injection and extraction periods of the 

third cycle, chloride concentrations in samples from 

piezometers 27P6 and 27P7 ranged from 7.5 to 12.4 

mg/L, which was similar to concentrations in the 

injection water and always lower than concentrations in 

samples of extraction water collected on the same day 

(fig. 5F). Dissolved solids, nitrate, and sulfate 

concentrations and specific conductance values in 

samples from piezometers 27P6 and 27P7 also were 

similar to those in the injection water and lower than 

those in samples of extraction water collected on the 

same day (table 3; Fram and others, 2002). Samples 

from piezometer 27P8 generally contained higher 

concentrations of chloride (fig. 5F), dissolved solids, 

sulfate, and nitrate than samples from piezometers 

27P6 and 27P7. 

Bromide concentrations in the extraction water 

were generally well correlated with chloride 

concentrations (fig. 5F,G), and it is presumed that a 

similar correlation existed in the injection water. 

Accurate measurements of the bromide concentration 

in the injection water were not obtained because the 

samples were collected after chlorination, which causes 

incorporation of bromide into halogenated disinfection 

by-products.

Bromine to Chlorine Ratio in the Trihalomethanes

THMs consist of chloroform (CHCl3), 

bromodichloromethane (CHBrCl2), 

dibromochloromethane (CHBr2Cl), and bromoform 

(CHBr3). The molar ratio of bromine to chlorine 

incorporated in trihalomethane species [(Br/Cl)THM] 

reflects the mechanisms and kinetics of THM 

formation and the bromide content of the source water 

from which the THMs form. Brominated THMs are 

formed because the chlorine added during the watertreatment process oxidizes any bromide dissolved in 

the water to form the reactive species hypobromous 

acid. Water-treatment plants generally add chlorine by 

bubbling chlorine gas through the water. Free chlorine 

is almost completely hydrolyzed to hypochlorous acid 

at the chlorine concentrations used in water treatment 

(Larson and Weber, 1994). The bromide oxidation 

reaction is very rapid and scavenges bromide out of the 

water, converting it to hypobromous acid (Morris, 

1978; Rook and others, 1978). Hypobromous acid 

reacts much faster with DOC to form THMs than does 

hypochlorous acid (Rook and others, 1978; Oliver, 

1980; Amy and others, 1985; Symons and others, 

1993); thus the rate of formation of brominated THMs 

is greater than the rate of formation of CHCl3 alone.

Injection water samples collected between 

April 16 and April 23, 1998, and between May 12 and 

June 15, 1998, had (Br/Cl)THM values between 0.025 

and 0.080, whereas samples collected between April 29 

and May 8, 1998, had values between 0.175 and 0.255 

(fig. 5H). (The (Br/Cl)THM values were calculated by 

assigning a concentration of zero to concentrations of 

individual species that were below the method 

detection limit and thus are minimum values.) The 

injection water samples collected between April 29 and 

May 7, 1998, contained higher concentrations of 

chloride (fig. 5F), sulfate, and dissolved solids than did 

the other injection water samples (Fram and others, 

2002) and presumably also had higher bromide 

concentrations. 
II. Formation and fate of trihalomethanes during the injection, storage, and recovery cycle 21

Table 3. Water quality data for water collected from nested piezometers 7N/12W-27P6–8 during the third injection, storage, and recovery cycle 

(March 1998 through April 1999), Lancaster, Antelope Valley, California

[Dissolved organic carbon (DOC) and ultraviolet absorbance at 254 nanometers (UVA254) were analyzed at the U.S. Geological Survey (USGS), Sacramento 

District. All other constituents were analyzed by the Los Angeles County Department of Public Works (LACDPW). mg/L, milligram per liter; /cm, per 

centimeter; μS/cm, microsiemens per centimeter; na, not analyzed]

Sampling 

date

DOC

(mg/L)

UVA254

(/cm)

Specific 

conductance

(μS/cm)

pH

(standard 

units)

Dissolved 

solids

(mg/L)

Chloride

(mg/L)

Nitrate

(mg/L)

Sulfate

(mg/L)

Piezometer 27P6

2/18/98 0.17 0.004 na na na na na na

3/12/98 .24 .003 280 7.76 172 4.20 1.21 23.4

6/15/98 1.11 .017 219 7.68 142 7.24 .65 35.0

8/4/98 1.26 .191 232 8.54 144 8.00 .64 36.5

9/3/98 1.02 .160 219 8.62 138 9.00 .71 38.0

10/7/98 .91 .061 218 8.47 148 7.49 .97 29.6

11/5/98 1.01 .112 222 8.56 168 7.65 .75 29.2

12/2/98 .82 .105 225 8.51 142 9.69 1.74 32.0

3/24/99 1.67 .012 na na na na na na

Piezometer 27P7

2/18/98 0.11 0.003 na na na na na na

3/12/98 .18 .003 247 7.89 152 4.00 .96 21.2

6/15/98 1.28 .019 186 7.24 120 8.30 .55 35.0

8/4/98 1.19 .025 198 8.37 120 7.63 .15 33.5

9/3/98 .93 .017 204 8.28 130 9.01 .79 33.0

10/7/98 .86 .017 208 8.14 140 8.81 .94 29.5

11/5/98 .81 .015 217 8.36 154 9.81 1.03 3.4

12/2/98 .91 na 253 8.35 162 12.40 1.03 34.0

3/24/99 .65 .009 na na na na na na

Piezometer 27P8

2/18/98 0.18 0.002 na na na na na na

3/12/98 .46 .006 456 7.80 304 43.00 1.54 46.0

6/15/98 na na 286 8.16 180 12.70 .77 42.0

8/4/98 1.24 .299 231 8.06 146 11.90 .25 39.0

9/3/98 .60 .021 262 7.37 168 20.00 1.90 41.0

10/7/98 .96 .052 237 7.51 150 14.90 1.39 34.0

11/5/98 .88 .169 251 7.87 174 14.80 1.40 38.4

12/2/98 .73 .033 274 7.31 174 20.20 1.79 41.0

3/24/99 2.13 .015 na na na na na na
22 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

The (Br/Cl)THM values for water extracted from 

well 4-32 increased from about 0.03–0.09 to 0.09–0.14 

between June 30, 1998, and October 24, 1998 

(fig. 5H,I). This increase in (Br/Cl)THM values occurred 

concomitantly with the decrease in THM 

concentrations in the extracted water (fig. 5A). The 

(Br/Cl)THM values in the extracted water varied 

unsystematically between 0.06 and 0.12 during the 

second phase of extraction from well 4-32 between 

February 22 and April 29, 1999. 

Samples from the nested piezometers 27P6–8 

generally had lower (Br/Cl)THM values than did 

samples collected from well 4-32 (fig. 5I). The 

(Br/Cl)THM values for samples from all three 

piezometers were approximately 0.05 at the end of the 

injection period (June 16, 1998) and remained almost 

constant through November 5, 1998 (fig. 5I), while 

THM concentrations in the piezometer samples ranged 

from 35 to 75 μg/L (fig. 5A). Between November 5, 

1998, and March 24, 1999, the (Br/Cl)THM values for 

water samples from all three piezometers rose to a 

maximum value of 0.09. Water from piezometers 27P6 

and 27P7 always had similar (Br/Cl)THM values, 

whereas water from piezometer 27P8 had a higher 

(Br/Cl)THM value in December 1998 but a lower value 

in March 1999. 

Water Levels

Changes in water levels in the aquifer during the 

injection, storage, and recovery cycle may provide 

information about water flow in the aquifer during 

different periods of the cycle. Water levels were 

monitored continuously in piezometers 27P6 and 27P8 

and intermittently in piezometer 27P7 during the third 

cycle (fig. 3C). The depth to the water table in the 

piezometers decreased from 330 ft below land surface 

to a minimum of 290 and 302 ft below land surface in 

27P6 and 27P8, respectively, during the injection 

period. Water levels in piezometers 27P6 and 27P8 

responded to changes in the injection rate; depth to the 

water table increased in both piezometers during the 

short hiatus in injection from April 28 to May 5, 1998 

(fig. 3B,C). The rapid water-level response measured in 

both piezometers to the injection at well 4-32 indicated 

that the aquifer zones screened in the piezometers were 

in direct hydraulic communication with the aquifer 

zones screened in well 4-32 during the injection period. 

In contrast to the injection period, water levels in 

piezometers 27P6 and 27P8 responded very differently 

during the extraction period. After injection ceased, the 

depth to water measured in piezometer 27P8 increased 

sharply to 330 ft below land surface, then increased 

immediately to about 333 ft at the start of extraction at 

well 4-32, and then slowly increased to 347 ft below 

land surface over the 10 months of the extraction 

period (fig. 3C). Cessation of pumping at well 4-32 on 

October 24, 1998, beginning of pumping at well 4-34 

on December 29, 1998, and resumption of pumping at 

well 4-32 on February 22, 1999, caused only minor 

changes in water level in piezometer 27P8. In contrast, 

the water level in piezometer 27P6 responded directly 

to changes in pumping at wells 4-32 and 4-34 (fig. 3C). 

The depth to water increased to a maximum of 357 ft 

below land surface during the first phase of the 

extraction period when water was being pumped from 

well 4-32, and then decreased to a level equal to that in 

piezometer 27P8 (335 ft below land surface) when no 

water was being pumped from well 4-32 or 4-34. 

Initiation of pumping at well 4-34 corresponded to an 

increase in the depth to water measured in piezometer 

27P6, and then resumption of pumping at well 4-32 led 

to another sharp increase in the depth to water to its 

maximum of 380 ft below land surface. Periodic waterlevel measurements made at well 4-32 during the 

extraction period indicate that water levels declined to 

as much as 404 ft below land surface (K. Rosander, 

Los Angeles County Department of Public Works, oral 

commun., 2003), significantly below the screened 

interval of piezometer 27P8 (330–370 ft below land 

surface). Therefore, the aquifer zone screened by 

piezometer 27P8 was not in direct hydraulic 

communication with the aquifer zones screened by 

wells 4-32 and 4-34 during at least part of the 

extraction period, explaining the limited water-level 

response in piezometer 27P8 to the extraction from 

wells 4-32 and 4-34.
II. Formation and fate of trihalomethanes during the injection, storage, and recovery cycle 23

Trihalomethane Formation

This section of the report describes the factors 

that controlled the continued formation of THMs in the 

aquifer after injection and presents an estimate of the 

extent of THM formation in the injected water. THM 

concentrations in samples collected during the third 

injection, storage, and recovery cycle rose from a mean 

of 27.5 μg/L at the time of injection to about 59 μg/L at 

the beginning of extraction (fig. 5A). Similar increases 

occurred during the first two cycles (fig. 4A). THMs 

and other halogenated organic compounds are formed 

by the reaction of DOC and residual chlorine. The 

amount of THM formed depends on several variables 

including the reaction-limiting concentration of 

chlorine or DOC present in the water, the propensity of 

the DOC to form THMs (its quality with respect to 

THM formation), the contact time between the DOC 

and the chlorine, the ratio of the concentrations of 

bromide and DOC, and the pH (for example, Rook, 

1977; Babcock and Singer, 1979; Reckhow and others, 

1990). The effect of some of these variables on THM 

formation in the injection water was systematically 

investigated as part of this study.

DOC is a complex, heterogeneous material 

whose chemical composition and structure varies 

significantly, depending on its sources and the 

biogeochemical processing it has undergone. Chemical 

and structural features of DOC collectively are referred 

to as DOC quality. In the context of this study, DOC 

quality is important because the extent of THM 

formation during chlorination of a water sample 

depends on the quantity (that is, concentration) and the 

quality of the DOC. The quantity and quality of DOC 

in the injected water varied during the injection period; 

thus, variations in the extent of THM formation were 

expected. Two types of experiments were done to 

assess THM formation and DOC quality: 

trihalomethane formation potential (THMFP) 

experiments and a storage experiment. The THMFP 

experiments yielded the maximum amount of THM 

formation possible from samples of injection water 

(given fixed conditions of pH, temperature, time, and 

residual chlorine concentrations) and provided an 

indicator of DOC quality. The storage experiment 

tested whether THM formation in the injected water 

was limited by the concentration of DOC or of residual 

free chlorine. 

Trihalomethane-Formation-Potential Experiments

Trihalomethane formation potential (THMFP) is 

the maximum amount of THM that will form from a 

water sample under controlled conditions of pH, 

temperature, time, and residual chlorine concentration. 

Because the molar yield of THMs from a given amount 

of DOC may vary by more than a factor of 10 (for 

example, see Fujii and others, 1998), no simple relation 

exists between DOC concentration or chlorine 

consumption and the extent of THM formation in 

natural waters. The variability in DOC quality means 

that THMFP is difficult to predict and thus must be 

measured experimentally. The THMFP experiment 

examines the relation between DOC quantity and 

quality and THM formation.

The experimental method for measuring THMFP 

is described in detail by Fram and others (2002). On 

arrival in the USGS laboratory in Sacramento, one 

aliquot of injection water was immediately quenched 

with sodium sulfite and analyzed for THM 

concentration. This concentration represents the 

concentration after one day of storage owing to transit 

time between the well and the laboratory. Another 

aliquot of injection water was quenched with sodium 

sulfite and purged with nitrogen to remove free 

chlorine and THMs. Samples were buffered to pH 8.3 

with a boric acid/sodium hydroxide solution, spiked 

with sufficient sodium hypochlorite to ensure the 

presence of a chlorine residual at the end of the 

experiment, sealed in headspace-free amber glass 

serum vials with Teflon-faced septa, and stored for 

7 days in the dark at 25°C. At the end of the 7 days, the 

THM concentration was measured by purge and trap 

capillary gas chromatography. This concentration 

represents the residual THMFP. The total THMFP for a 

water sample is the 1-day storage concentration plus 

the residual THMFP. Experimental results were 

tabulated by Fram and others (2002). 
24 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

The total THMFP of injection-water samples 

ranged from 120 to 228 μg/L (fig. 6) and averaged 

175 μg/L (RSD = 13 percent). These values are much 

higher than those for the THM concentrations at the 

time of injection. The ground water had a total THMFP 

of 21.1 μg/L, which is much lower than that of the 

injection-water samples. This low value primarily 

reflects the much lower DOC concentration in the 

ground water (0.2 mg/L) in comparison with that in the 

injection water (1.4–2.0 mg/L). The highest 

(Br/Cl)THM value was measured in the THMFP of the 

ground water. The THMFP of the ground-water sample 

was only 21 μg/L THMs, but the (Br/Cl)THM was equal 

to 0.81 (fig. 5H). This (Br/Cl)THM value was much 

higher than the values for all other field and 

experimental samples analyzed in this study and 

reflects the high ratio of bromide to DOC in the 

ground-water sample.

Results from the THMFP experiments also can 

be used to characterize the compositional nature, or 

quality, of the DOC. Specific trihalomethane formation 

potential (STHMFP) is defined as the total THMFP, in 

millimoles of THMs per liter, divided by the DOC, in 

moles of carbon per liter, and has units of millimoles 

per mole (mmol/mol). STHMFP is an indicator of 

DOC compositional quality and represents the molar 

efficiency of the DOC to form THMs. STHMFP values 

for the injection water ranged from 6.9 to 11.3 

mmol/mol (median value was 9.9 mmol/mol) (fig. 7). 

In other words, approximately 10 of every 1,000 

carbon atoms in the DOC formed THMs during 

chlorination under the experimental conditions. The 

STHMFP value for the ground-water sample shows 

that the DOC in the ground water is chemically distinct 

from the DOC in the injection water. The STHMFP 

value of the ground water was 7.3 mmol/mol (fig. 7), 

which is significantly lower than the STHMFP values 

of the injection water sample (parametric prediction 

interval at α = 0.05 and nonparametric prediction 

interval at α = 0.1; Helsel and Hirsch, 1995).

DOC quality also can be quantified using optical 

measurements. For example, aromatic moieties can 

absorb light of wavelengths around 254 nanometers 

(nm). Because light absorbance depends on DOC 

quantity as well as quality, ultraviolet-absorbance data 

are presented normalized to DOC concentration. 

SUVA254 is the ultraviolet absorbance at 254 nm 

divided by DOC concentration and has units of liters 

per meter per milligram. SUVA254 is frequently used as 

an indicator of the aromaticity of the DOC (Traina and 

others, 1990; Chin and others, 1994). Absorbance data 

are tabulated in Fram and others (2002). Because 

THMs are generally thought to form from aromatic 

moieties within the DOC, SUVA254 is considered an 

indicator of the THM formation potential of a water 

sample (Rook, 1977; Reckhow and others, 1990; U.S. 

Environmental Protection Agency, 1998). Thus, a 

correlation between SUVA254 and STHMFP is 

commonly expected. However, for these samples, 

SUVA254 and STHMFP are poorly correlated (fig. 7) 

(coefficient of correlation, r2, for SUVA254 and 

STHMFP in the injection-water samples equals 0.16). 

This observation suggests that for these source waters, 

the aromatic carbon precursor model for THM 

formation is not sufficient. 

Storage Experiment

A storage experiment was done to assess 

whether THM formation in the injected water was 

limited by DOC or chlorine. The injection water 

contained 0.5–1.4 mg/L of free and total residual 

chlorine (fig. 5E), and 1.4–2.0 mg/L of DOC (fig. 5B) 

at the time of injection. This water continued to form 

THMs (and other chlorination by-products) after 

injection because the residual chlorine continued to 

react with the DOC. THM formation over time was 

measured directly by storing samples of injection water 

and measuring THM concentrations as the residual 

chlorine originally present in the samples was 

consumed. 

For the storage experiment, unopened vials of 

injection water were stored for different periods of time 

before measuring the THM concentrations. The first 

analysis was done after 1 day (the transit time between 

the well in Lancaster and the laboratory in 

Sacramento). The other vials then were stored in the 

dark at 25°C for periods of 1, 2, 4, 8, and 16 weeks. At 

the end of each storage period, THM concentrations 

were analyzed by purge and trap gas chromatography 

by the USGS. Fram and others (2002) present detailed 

descriptions of the experimental and analytical 

methods and give the experimental results.
II. Formation and fate of trihalomethanes during the injection, storage, and recovery cycle 25

Figure 6. Trihalomethane concentrations in injection water collected from well 7/N12W-27P2 (well 4-32) during the third injection, storage, and recovers 

cycle (March 1998 through April 1999), Lancaster, Antelope Valley, California

Date are from Fram and others (2002)

DATE, IN 1998

April May June

TOTAL TRIHALOMETHANES, IN MICROGRAMS PER LITER

At time of injection

After 1-day transit

After 1- or 2-week storage

After 4-week storage

After 8- or 16-week storage

Total trihalomethane formation potential

EXPLANATION

0

50

100

150

200

250
26 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

Figure 7. Values of specific trihalomethane formation potential (STHMFP) and specific ultraviolet absorbance at 254 nanometers (SUVA254) for injection 

water and ground water collected from well 7N/12W-27P2 (well 4-32) during the third injection, storage, and recovery cycle (March 1998 through April 

1999), Lancaster, Antelope Valley, California. 

Data are from Fram and others (2002).

SPECIFIC ULTRAVIOLET ABSORBANCE AT 254 NANOMETERS,

IN LITERS PER METER PER MILLIGRAM

1.0 1.2 1.4 1.6 1.8 2.0 2.2

SPECIFIC TRIHALOMETHANE FORMATION

POTENTIAL, IN MILLIMOLES PER MOLE

6

7

8

9

10

11

12

Injection water

Ground water

Median

Error bars
II. Formation and fate of trihalomethanes during the injection, storage, and recovery cycle 27

The concentration of THMs in the stored 

samples increased during the first 4 weeks of storage 

and then remained relatively unchanged (fig. 8), as 

would be expected if one or both of the reactants 

(reactive DOC or free chlorine) were depleted during 

this time frame. Mean THM concentrations ranged 

from 28 μg/L at the time of injection (starting time for 

the storage experiment) to 40 μg/L after 1 day, to 

73 μg/L after 2 weeks, to 89 μg/L after 4 weeks (fig. 8). 

The differences between the mean values at 4, 8, and 

16 weeks were not statistically significant (the Student 

t test of difference between two means, α = 0.05; 

Helsel and Hirsch, 1995). 

The mean THM concentration of all of the 

samples stored for 1, 2, 4, 8, or 16 weeks was 86 μg/L 

(σ = 18 μg/L). This value is significantly less than the 

value determined in the THMFP experiment 

(175 μg/L). Because the injection period spanned 

8 weeks and was followed by 2 weeks of storage before 

extraction began, there probably was sufficient time for 

the reactions to occur. Only one sample from the first 

few days of recovery contained traces of free residual 

chlorine (fig. 5E). Therefore, the THM formation in the 

injected water in the aquifer was limited by the 

concentration of residual free chlorine in the injection 

water. These results also indicate that if the water was 

re-chlorinated, the THM concentration would increase 

to 175 μg/L. 

Figure 8. Trihalomethane concentrations in injection water collected from well 7N/12W-27P2 (well 4-32) dring the third injection, storage, and recovery 

cycle (March 1998 through April 1999), Lancaster, Antelope Valley, California.

Mean concentration at each time is indicated by the white points. data from Fram and others (2002).

SAMPLE STORAGE TIME, IN WEEKS

0 4 8 12 16 20

TRIHALOMETHANES, IN MICROGRAMS PER LITER

0

25

50

75

100

125
28 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

Estimate of Trihalomethane Formation in the Injected Water

THM formation in the injected water, after 

injection and storage in the aquifer, was estimated 

using selected samples from well 4-32 and the 

piezometers. Interpretation of the water-quality data 

indicated that the water in these samples had not mixed 

with ground water and had not been modified by 

biodegradation or sorption. Samples that meet these 

criteria should have the chemical characteristics of the 

injection water. The chloride concentration and the 

(Br/Cl)THM values for the ground-water samples were 

used to select the set of representative samples that 

most closely resembled injection water. The chloride 

concentration of the injection water ranged between 

7 and 9 mg/L; samples that had chloride concentrations 

within this range were selected for further review. 

(Br/Cl)THM values are very sensitive to the bromide 

content of water from which the THM formed and to 

later processes that might selectively affect the four 

THM species. We surmised that the injection water 

after storage in the aquifer and reaction of residual 

chlorine under aquifer conditions would contain THMs 

with (Br/Cl)THM values similar to values in the aliquots 

of injection water from the storage (1–16 week) 

experiments. Thus, samples that met the chloride 

criterion and had a (Br/Cl)THM value between 0.03 and 

0.06 (fig. 5I) were determined to represent the THM 

formation in the injected water after injection into and 

storage in the aquifer. Mixing of injected water with 

ground water before consumption of residual chlorine 

would result in THMs with higher (Br/Cl)THM values, 

and biodegradation or sorption would result in THMs 

with lower (Br/Cl)THM values (discussed in more detail 

in the section “Fate of trihalomethanes”). Four samples 

of extraction water from well 4-32 and eight samples 

from piezometers 27P6 and 27P7 were selected based 

on these criteria. The mean THM concentration in 

these samples, 58 μg/L or 0.48 micromole per liter 

(μmol/L), is assumed to be the THM concentration of 

the injected water after injection and storage in the 

aquifer.

The mean THM concentration of the selected 

ground-water samples (58 μg/L) is less than the THM 

concentration observed in the storage experiment 

(86 μg/L). Possible explanations for this difference 

include the following:

• Water flow during injection was turbulent and may 

have contributed to volatilization of THMs from 

the injected water.

• THM formation is a function of pH; fewer THMs 

form when the pH is lower. The pH values of the 

first several samples of extracted water were lower 

than the pH values of the injected water (fig. 5D). 

This decrease in pH may have been due to 

entrainment of CO2 during the injection process in 

addition to the pH decline that accompanies 

chlorine-consumption reactions.

• The aquifer sediments may have included material 

that reacted with chlorine, thus decreasing the 

amount of chlorine available to react with DOC to 

form THMs.

• DOC in the injection water may have rapidly 

degraded after injection into the aquifer, thus 

decreasing the amount of DOC available to react 

with chlorine to form THMs.

The THM concentration in the injected water 

after injection and storage in the aquifer could be 

reduced by dechlorination treatment of the injection 

water at the well-head. The residual free chlorine in the 

injected water could be removed by reaction with 

dechlorinating reagents such as sulfur dioxide, sodium 

bisulfite, sodium thiosulfite, sodium sulfite, ascorbic 

acid, or sodium ascorbate. However, depending on the 

dechlorination processes used, dechlorination may 

increase the potential for microbial growth, increase 

dissolved solids concentration, or decrease pH in the 

injected water.
II. Formation and fate of trihalomethanes during the injection, storage, and recovery cycle 29

Trihalomethane Fate

The (Br/Cl)THM values in water samples 

collected during the third cycle were used to help 

distinguish the processes that may or may not have 

affected the formation and fate of THMs in the aquifer. 

Biodegradation, sorption, and mixing would each 

produce predictable patterns of variation in (Br/Cl)THM

values and THM concentrations in the extracted water. 

These predicted patterns can be compared to the 

observed patterns to infer the relative importance of the 

processes affecting the THM formation and fate in the 

aquifer. One of the most intriguing features of the data 

presented in figure 5I is the steady increase in the 

(Br/Cl)THM value in the extracted water during the first 

phase of extraction, concomitant with a steady decrease 

in THM concentrations (fig. 5A). Any explanation of 

the observed decrease in THM concentrations in the 

extracted water also must account for the (Br/Cl)THM

values. Three primary processes that may affect THM 

concentrations and (Br/Cl)THM values are 

biodegradation of THMs by aquifer bacteria, sorption 

of THMs to aquifer materials, and mixing between 

injected water and ground water. Each of these 

processes will produce different patterns of variation of 

(Br/Cl)THM in the extracted water, and thus, the 

observed pattern of variation can be used to constrain 

the relative importance of these processes during the 

third injection, storage, and recovery cycle.

Biodegradation

Biodegradation of THMs has been observed in 

field and laboratory studies; however, biodegradation 

has been demonstrated to occur only under anoxic 

conditions and only brominated THMs are consumed 

(see the fourth section of this report for discussion). 

Therefore, if significant biodegradation of THMs had 

occurred during the third cycle, the (Br/Cl)THM of the 

remaining THMs would have decreased because the 

bacteria would have consumed only the brominated 

THM species. The data for well 4-32 show precisely 

the opposite trend: the (Br/Cl)THM of the remaining 

THMs increased with time, indicating that 

biodegradation probably had not occurred (fig. 5I). 

Furthermore, experiments designed to detect 

biodegradation of THMs showed that biodegradation 

of THMs in the aquifer at the Lancaster site was highly 

unlikely (see the fourth section of this report for further 

discussion).

Sorption

Sorption of THMs to sedimentary materials has 

been observed in field and experimental studies 

(Roberts and others, 1982; Curtis and others, 1986; 

Roberts and others, 1986; Walton and others, 1992; 

Peng and Dural, 1998). Sorption of hydrophobic 

organic compounds such as THMs is governed by a 

linear bulk partition coefficient, Kp, that is a strong 

function of the organic carbon content of the 

sedimentary material and the octanol-water partition 

coefficient of the organic compound (Chiou and others, 

1979, 1983; Karickhoff and others, 1979; 

Schwarzenbach and Westall, 1981; Gschwend and Wu, 

1985) (equations 1–4):

(1)

(2)

(3)

(4)

where 

Koc is the organic carbon/water partition 

coefficient,

Kow is the octanol/water partition coefficient,

a and b are empirical constants,

Kp is the linear bulk sediment-water 

partition coefficient,

foc is the fraction of organic carbon in the 

sediment,

Csol is the concentration of the compound in 

the solid at equilibrium,

Cwat is the concentration of the compound in 

the pore water at equilibrium,

Cinit is the concentration of the compound in 

the injected water,

ρ is the density of the sediment, and

ε is the porosity of the sediment.

Koc log a Kow = ( ) log + b

Kp Koc foc = ×

Kp

Csol

Cwat

= ----------

Cinit

ρCsol + εCwat

ε = --------------------------------
30 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

The bulk partition coefficient is defined as the 

ratio between the concentrations of the compound in 

the solid sediment material and the pore water at 

equilibrium. As injected water moves into the pore 

spaces of the aquifer sediment material, a compound 

present at some concentration, Cinit, in the injected 

water will be partitioned between the solid sediment 

material and the pore water. If a large amount of 

sorption occurs, then the difference between Cinit and 

Cwat will be significant.

Combining equations 1–4 and inserting 

appropriate values for Kow, a, b, ρ, and ε yields an 

estimate of the minimum organic carbon content 

required for sorption of THMs to the aquifer sediments 

to result in a significant, measurable difference 

between Cinit and Cwat. An organic carbon content of 

greater than about 0.05 percent would result in enough 

sorption of THMs to the aquifer sediment to noticeably 

affect the concentration of THMs in the water.

Hand-sample examination of core samples 

collected during drilling for installation of the nested 

piezometers yielded estimated organic material content 

of much less than 1 percent in the finer-grained layers, 

and organic material was not observed in the coarsergrained layers (Fram and others, 2002). Sedimentary 

organic material consists of approximately 60 percent 

(by mass) organic carbon; thus, the mean organic 

carbon content of the aquifer materials is estimated to 

be extremely small. Quantitative analysis of similar 

sedimentary materials yielded organic carbon content 

of 0.01 to 0.15 percent (Schwarzenbach and Westall, 

1981; Curtis and others, 1986; MacIntyre and Stauffer, 

1988). Therefore, sorption was not expected to be a 

significant process.

If significant sorption had occurred, it would 

have been apparent in the (Br/Cl)THM values in the 

extracted water. Because the log Kow values of the 

THMs vary systematically from 1.97 for CHCl3 to 2.38 

for CHBr3 (Mackay and others, 1993), the more 

brominated THMs are sorbed preferentially to 

sedimentary materials. Applying equations 1–4 to each 

of the four THM species yields an estimate of the 

(Br/Cl)THM value in the injection water after 

equilibration with the aquifer materials (Cwat) that can 

be compared to the (Br/Cl)THM value in the injection 

water prior to interaction with the aquifer (Cinit). Cinit

is represented by the injection water samples from the 

storage experiments and Cwat by the water extracted at 

the beginning of the extraction period. Samples of 

injection water with the least admixed ground water are 

most likely to be extracted at the beginning of the 

extraction period. For an organic carbon content of 

0.05 percent, equations 1–4 predict that the (Br/Cl)THM

values for the extraction water at the beginning of the 

extraction period (that is, injection water equilibrated 

with aquifer materials) would be 30 percent lower than 

the (Br/Cl)THM values for the water from the storage 

experiment (that is, injection water that has not 

interacted with aquifer materials). However, the 

(Br/Cl)THM values for the injection water from the 

storage experiment and for the extracted water at the 

beginning of the extraction period are approximately 

equal (fig. 5I). This analysis suggests that sorption of 

THMs to aquifer materials did not occur to a 

significant, measurable extent during the third 

injection, storage, and recovery cycle.

Mixing

Once significant biodegradation and sorption 

effects have been ruled out, (Br/Cl)THM can be used as 

an indicator of the relative concentrations of bromide 

and DOC in the water in which the THMs formed. For 

a constant DOC concentration, (Br/Cl)THM increases 

as the initial dissolved bromide concentration increases 

(for example, Amy and others, 1985; Symons and 

others, 1993), and for a constant bromide 

concentration, (Br/Cl)THM should increase as the DOC 

concentration decreases. Therefore, (Br/Cl)THM should 

increase as the ratio of bromide to DOC concentrations 

increases. Using the amount of bromine incorporated 

into the THMs in the THMFP experiment (0.013 mg/L) 

as a minimum estimate of the initial concentration of 

bromide in the water, the estimated bromide to DOC 

concentration ratio in the ground water (greater than or 

equal to 0.065 milligram of bromide per milligram of 

DOC) is approximately 10 times greater than the ratio 

for the injection water (greater than or equal to 0.006 

milligram of bromide per milligram of DOC). Thus, 

THMs formed in mixtures of ground water and injected 

water would have higher (Br/Cl)THM values than 

THMs formed in pure injection water. Mixing between 

ground water and injection water after THM formation 

would produce a pattern of decreasing THM 

concentrations and constant (Br/Cl)THM values in the 
II. Formation and fate of trihalomethanes during the injection, storage, and recovery cycle 31

extracted water as extraction proceeded. Mixing 

between ground water and injection water before THM 

formation is complete would produce a pattern of 

decreasing THM concentrations and increasing 

(Br/Cl)THM values in the extracted water as extraction 

proceeded.

The observed decrease in THM concentrations in 

the extracted water during the extraction period 

(fig. 5A) is consistent with mixing between injection 

water containing THMs and THM-free ground water. 

The increase in (Br/Cl)THM values in the extracted 

water during the first phase of extraction (fig. 5I) 

further implies that some of the mixing happened 

before THM formation was complete. The relatively 

constant (Br/Cl)THM values in the extracted water 

during the second phase of extraction (fig. 5I) suggests 

that the mixing that produced these water samples 

occurred primarily after THM formation was complete.

Mass Balance of Chloride, Dissolved Organic Carbon, and 

Trihalomethanes

The preceeding discussion identified dilution 

from mixing as the dominant process controlling the 

progressive decrease in THM concentrations in the 

water extracted from well 4-32 during the third 

injection, storage, and recovery cycle. In this section, 

mass balance calculations are used to estimate the 

proportions of injected water and ground water in the 

extracted water and the percentage of THMs injected 

into the aquifer that was retrieved. 

The amount of injected water that was extracted 

was determined using three constituents present in the 

injection water: chloride, DOC, and THM. The mass 

balances for chloride and DOC were determined for 

only one point in the extraction cycle—at the end of the 

first phase of extraction from well 4-32 (October 24, 

1998). At this point, the total volume of water extracted 

from well 4-32 was 132 percent of the total volume of 

water injected (fig. 3B). 

Mass Balance Calculations

Mass balances were calculated using estimates 

for the concentrations of the constituents in the average 

injected water and average ground water (the injection 

water and ground water end-members, respectively) 

and for the measured concentrations of the constituents 

in the extracted water. The concentrations of chloride, 

DOC, and THMs in the injection water and ground 

water end-members are listed in table 4.

Table 4. Average concentrations of constituents in injected water and 

ground water used in mass-balance, tracer mixing model, and descriptive 

mixing model calculations

[DOC, dissolved organic carbon; THM, trihalomethane; SF6, sulfur 

hexafluoride. mg/L, milligram per liter; μg/L, microgram per liter; 

μmol/L, micromole per liter; pmol/L, picomole per liter]

1Based on average concentration in injection water samples 2Based on average concentration in selected samples extracted from 

well 7N/12W-27P2 (4-32) and piezometers 7N/12W-27P6 and 27P7 3Selected from range of concentrations measured in ground water 4Measured in ground water prior to the third cycle 5Concentration assumed to be zero

Constituent

Injected water 

concentration 

(Cinj)

Ground water 

concentration

(Cgw)

Chloride 7.5 mg/L1 21 mg/L3

DOC 1.76 mg/L1 .2 mg/L4

THMs 58 μg/L2 0 μg/L5

THMs .48 μmol/L2 0 μmol/L5

SF6 63.6 pmol/L1 0 pmol/L4
32 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

The mass-balance calculations entailed a 

comparison of the mass of the constituent injected with 

the mass of the constituent extracted. Minj is the 

difference between the mass of the constituent in the 

injection water and the mass of the constituent in the 

ground water (fig. 9; equation 5). Minj may be positive 

or negative depending on whether the concentration of 

the constituent in the injection water is greater than or 

less than the concentration in the ground water. 

(5)

where 

The volumes of water injected and extracted are 

expressed as V, the equivalent volume. The value of V is 

given by equation 6 for the injection period and by 

equation 7 for the extraction period. V increases from 

−1 to 0 during the injection period and then increases 

from 0 to a positive number during the extraction 

period. When V equals 1, the volume of water extracted 

equals the volume of water injected.

(6)

(7)

where, at the time for which V is being calculated,

Mext is the difference between the mass of the 

constituent in the extracted and mass of the constituent 

in the ground water (fig. 9; equation 8). Mext may be 

positive or negative depending on whether the 

concentrations of the constituent in the extracted water 

are greater than or less than the concentration in the 

ground water. 

(8)

where

The percentage recovery of the injected 

constituent is the ratio of Mext to Minj (equation 9). The 

percentage recovery may be less than or greater than 

100 percent. Values greater than 100 percent indicate 

that there was a source of the constituent in the aquifer. 

Values less than 100 percent indicate that the 

constituent was not completely recovered. Incomplete 

recovery may be caused by either destruction of the 

constituent in the aquifer or dilution of the injected 

water by mixing with ground water.

(9)

If there are no sources or losses of the constituent 

in the aquifer and there is no mixing between the 

injected water and ground water, then the variation in 

concentration of the constituent in the extracted water 

can be described by figure 9A or B. The injected water 

is completely recovered during the extraction period 

between V = 0 and V = 1. If there is no mixing of the 

injected water with ground water, the concentration of 

the constituent in the extracted water, Cext(V), is 

represented by two line segments: Cext(V) equals the 

concentration in the injected water from V = 0 to V = 1 

during the extraction period and then equals the 

concentration in the ground water for extraction 

beyond V = 1.

If there is mixing between the injected water and 

ground water, the variation in concentration of the 

constituent in the extracted water can be described by 

figure 9C or D. The concentration of the constituent in 

the extracted water varies as a function of V. For the 

mass-balance calculations presented in this report, 

Cext(V) was determined by using least squares 

regression to fit a polynomial expression to the data. 

The polynomial expression can be integrated and 

evaluated for any point V in the extraction period to 

calculate Mext at that point in the extraction period.

Vinj is the total volume of water injected,

Cinj is the average concentration of the 

constituent in the injected water, and

Cgw is the average concentration of the 

constituent in the ground water.

is the cumulative volume of water 

extracted, and

is the cumulative volume of water injected. 

Minj VinjCinj VinjCgw = –

V

Vinj*

Vinj

= ---------- – 1

V

Vext

Vinj

= --------

Vext

Vinj*

Cext(V) is the function fit to the extraction data.

Mext Cext( ) V dV

0

Vext

∫ VextCgw = –

recovery

Mext

Minj

= ---------- × 100
II. Formation and fate of trihalomethanes during the injection, storage, and recovery cycle 33

Figure 9. Method for calculating mass balance of a conservative constituent without (A,B) and with (C,D) mixing between injected water and ground 

water.

CONCENTRATION OF THE CONSTITUENT

Cgw

Cinj

Minj Mext Cext(V)

Cgw

Cinj

Minj Mext Cext(V)

-1.0 -0.5 0.0 0.5 1.0 1.5 2.0

V, EQUIVALENT VOLUME INJECTED OR EXTRACTED

Cgw

Cinj

Minj

Mext

Cext(V)

Cgw

Cinj

Minj

Mext

Cext(V)

CONCENTRATION OF THE CONSTITUENT

A

B

C

D

Cinj

Cgw

Cext(V)

V

Minj

Mext

mean concentration of

constituent in injected water

concentration of constituent in

ground water

concentration of constituent in

extracted water as a function

of V

volume of water injected or

extracted relative to total

volume of water injected

difference between masses of

constituent in the volume of

injected water and in an

equivalent volume of ground

water

difference between masses of

constituent in the volume of

extracted water and in an

equivalent volume of ground

water
34 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

Mass Balance Results

Chloride

We constructed a chloride mass-balance 

calculation for the first phase of extraction, V = 0 to 

V = 1.32 (June 30, 1998, to October 24, 1998), when 

the chloride concentrations in water extracted from 

well 4-32 rose smoothly from 7 to about 20 mg/L 

(fig. 10A). Minj was calculated using equation 5, the 

estimates of Cinj and Cgw given in table 4, and Vinj = 1. 

Cext(V) was determined by fitting the chloride 

concentrations in the extraction water for the period

V = 0 to V = 1.32 with a cubic function using the curvefitting function in SigmaPlot2001(fig. 10A). Mext was 

calculated using equation 8, the estimate of Cgw given 

in table 4, and Vext = 1.32. The calculation yielded a 

68 percent recovery of injected chloride. Because 

chloride is a conservative constituent, chloride 

recovery is a proxy for recovery of the injected water. 

Therefore, on the basis of chloride concentrations, only 

68 percent of the injected water was recovered by the 

time the volume of water extracted reached 132 percent 

of the volume of water injected. 

The error associated with the mass balance 

calculation is estimated by considering the 

uncertainties in the estimates of Cinj and Cgw. Cext(V) is 

assumed to have no error. Variation of Cinj between 

7 and 9 mg/L results in variation in the percentage 

recovery of chloride between 65 and 76 percent, 

respectively. Variation of Cgw between 20 and 23 mg/L 

results in variation in the percentage recovery of 

chloride between 63 and 76 percent, respectively. Cgw

cannot be less than 20 mg/L because the maximum 

concentration of chloride in the extraction water during 

the first phase of extraction was 20 mg/L 

(figs. 5F and 10A).

 It is noteworthy that after pumping from 

well 4-32 resumed on February 22, 1999, the chloride 

concentrations in the extracted water followed a trend 

substantially different from that of the first part of the 

recovery phase of the cycle. Chloride concentrations 

decreased from 15 to 11 mg/L between V = 1.32 and 

V = 2.50 (fig. 10A), suggesting that the injection water 

was mixing with different ground water end-members 

during phases 1 and 2. The ground water end-member 

during the second phase of extraction apparently 

contained approximately 11 mg/L of chloride. This 

apparent shift in the composition of the ground-water 

end-member may be explained by the heterogeneity of 

the ground water near well 4-32 before the third cycle, 

the movement of water in the aquifer caused by 

pumping at well 4-34 when well 4-32 was idle, and the 

increased drawdown at well 4-32 during the period 

when both wells were being pumped.

Dissolved Organic Carbon

DOC concentrations in the water extracted from 

well 4-32 during the first phase of extraction declined 

smoothly from 1.70 mg/L at V = 0 to 0.27 mg/L at 

V = 1.32. The data were fit with a cubic function 

(fig. 10B) using the curve-fitting algorithm in 

SigmaPlot2001 to give Cext(V). Minj and Mext were 

calculated using equations 5 and 8, respectively, the 

estimates of Cinj and Cgw given in table 4, Vinj = 1, and 

Vext = 1.32. These calculations yielded only a 

39 percent recovery of injected DOC at V = 1.32. 

Recovery of DOC was lower than that for chloride, 

indicating that DOC did not behave conservatively in 

the aquifer. Even if Cinj was as high as 2 mg/L (the 

maximum DOC concentration in the injection water 

samples), the calculated percentage recovery of DOC 

would only be 51 percent, which is still significantly 

less than the recovery of chloride. 

The low percentage recovery of DOC suggests 

that injected DOC degraded in the aquifer. Field and 

laboratory studies involving introduction of DOC from 

surface waters into aquifers composed of clastic 

sediments report substantial degradation of the DOC 

(Quanrud and others, 1996; Barber and others, 1997). 

DOC concentrations in samples from the piezometers 

support the conclusion that injected DOC degraded in 

the aquifer. The water samples collected from 

piezometers 27P6 and 27P7 on June 15, 1998, had 

concentrations of THMs (fig. 5A), chloride (fig. 5F), 

and other major ions (Fram and others, 2002) similar to 

concentrations in samples from well 4-32, but had 

significantly lower concentrations of DOC (fig. 5B). 

This suggests that DOC in the injected water degraded 

as the water travelled from well 4-32 to the 

piezometers. If this degradation occurred rapidly, 

before all the residual free chlorine in the injected 

water was exhausted, it may also partially explain why 

the estimated amount of THMs formed from the 

injected water after injection into the aquifer (58 μg/L) 

was less than the amount of THMs formed in the 

storage experiment (86 μg/L).
II. Formation and fate of trihalomethanes during the injection, storage, and recovery cycle 35

Figure 10. chloride (A), dissolved organic carbon (B) and trihalomethane (C) concentrations in ground-water collected before the first, second, and third 

cycles (April 1996 through April 1999), and in injection and extraction water collected from well 7N/12W-27P2 (well 4-32) during the third injection, storage, 

and recovery cycle (March 1998 through April 1999), Lancaster, Antelope Valley, California. 

V is the cumulative volume of water extracted (positive numbers) or injected (negative numbers) relative to the total volume of water injected at well 4-32. 

Equations of the best-fit curves to the concentration data in the first phase of extraction are shown on panels A and B and adjacent to panel C. Data are from 

Fram and others (2002), Metzger and others (2002) and table 3. CHLORIDE, IN MILLIGRAMS PER LITER

0

5

10

15

20

25

30

DISSOLVED ORGANIC CARBON,

IN MILLIGRAMS PER LITER

0.0

0.5

1.0

1.5

2.0

2.5

V, EQUIVALENT VOLUME FRACTION

-1.0 -0.5 0.0 0.5 1.0 1.5 2.0 2.5

TRIHALOMETHANES,

IN MICROMOLES PER LITER

0.0

0.1

0.2

0.3

0.4

0.5

0.6

EXPLANATION

Injection water

Extraction water

Ground water (well 4-32, well 4-34, piezometers)

A

B

C

Injection

Extraction

phase 1

Extraction

phase 2

Ground

water

Injection

Extraction

phase 1

Extraction

phase 2

Ground

water

Injection

Extraction

phase 1

Extraction

phase 2

Ground

water

43

36

y = 6.5+18x-11x2+3.4x3

y = 1.7-3.1x+2.4x2-0.62x3

a

b

c d

e

f

g

h

i j

line slope intercept

a -0.442 0.485

b -0.375 0.445

c -0.300 0.392

d -0.175 0.292

e -0.088 0.214

f -0.800 1.250

g -0.350 0.620

h -0.225 0.432

i -0.068 0.181

j -0.025 0.082

piez. 27P8, pre-cycle 3

well 4-32, pre-cycle 1

well 4-34, pre-cycles 1 and 2

well 4-32, pre-cycle 2

well 4-32, pre-cycle 3

well 4-34, pre-cycle 3

piez. 27P6 and 27P7, pre-cycle 3

Curve fit to extraction water data

Estimate of end-member composition
36 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

Trihalomethanes

The percentage of the THMs injected into the 

aquifer that were extracted from well 4-32 was 

calculated using a method similar to that used for the 

chloride and DOC mass-balance calculations. The 

THM concentrations in the extracted water could not 

be fit adequately using a single mathematical function 

(as were the data for chloride and DOC for V equaling 

0 through 1.32); thus, Cext(V) was determined by fitting 

the data to a series of 10 linear segments (fig. 10C). 

The 10 segments were fit to the high concentration side 

of the data to maximize the estimated mass of THMs 

extracted. THM concentrations are plotted in molar 

units rather than mass units to eliminate variation 

owing to the differences in mass of THMs having 

different (Br/Cl)THM values. Minj was calculated using 

equation 5, the estimates of Cinj and Cgw given in table 

4, and Vinj = 1. Values of Mext were calculated at values 

of Vext ranging from 0 to 2.50 using equation 8, the 

estimate of Cgw given in table 4, and the appropriate set 

of lines (a through j, fig. 10) for Cext(V). The total 

number of moles of THM extracted (Mext) and the 

percentage recovery of THMs (eq 9) are plotted as the 

solid curve in figure 11. By V = 1.32, 70 moles of 

THMs had been extracted and by V = 2.50, 84 moles of 

THMs had been extracted. 

The THM mass-balance calculation indicates 

that 66 percent of the injected THMs were recovered 

by V = 1.32. The recovery of THMs is similar to that of 

chloride, a conservative constituent, and higher than 

that of DOC, a nonconservative one. This suggests that 

THMs behaved conservatively in the aquifer (after 

accounting for the THMs formed in the aquifer after 

injection due to reaction of the residual chlorine in the 

injected water). 

Figure 11. Cumulative moles of trihalomethanes (THM) recovered during the extraction period of the third injection, storage, and recovery cycle (March 

1998 through April 1999) Lancaster, Antelope Valley, California.

Note different scales for moles of THMs extracted (left) and percentage of THMs extracted (right).

V, EQUIVALENT VOLUME EXTRACTED

0.0 0.5 1.0 1.5 2.0 2.5

MOLES OF TRIHALOMETHANE EXTRACTED

0

20

40

60

80

100

120

PERCENTAGE OF TRIHALOMETHANES EXTRACTED

0

20

40

60

80

100

THM injected assuming

no formation after

injection

THM injected assuming

formation after injection

THM extracted

from Well 4-32

Additional THM extracted

from Well 4-34
II. Formation and fate of trihalomethanes during the injection, storage, and recovery cycle 37

THM recovery from well 4-32 increased only to 

84 moles, or 80 percent of the amount injected, as V

increased to 2.50 (fig. 11). Extrapolating the trend of 

THM extraction as a function of the volume of water 

extracted suggests that extracting 100 percent of the 

THMs injected would require extraction of a nearly 

infinite volume of water. 

Implications for Water Flow in the Aquifer

The results of the mass-balance calculations and 

the water-quality and water-level monitoring at 

well 4-32 and the nested piezometers can be combined 

to form a conceptual model of the injection, storage, 

and recovery cycle. The conventional model of 

injection, storage, and recovery envisions the injected 

water forming a discrete dome or “balloon” around the 

injection well (for example, Singer and others, 1993; 

Pyne, 1995) (model A in fig. 12). Extraction from the 

well then deflates the balloon, and all of the injected 

water is recovered. In this model, ground water is only 

extracted from the well if the volume of water extracted 

exceeds the volume of water injected or the balloon of 

injected water is advected away from the well by the 

regional ground-water flow prior to extraction. 

Applying the balloon model (model A) to data from 

injection, storage, and recovery projects can 

profoundly affect interpretation of the data. For 

example, in the injection, storage, and recovery 

projects examined by Singer and others (1993), the 

known hydraulic gradients in the aquifers were too 

small to move the balloon of injected water away from 

the wells prior to extraction. Therefore, Singer and 

others (1993) concluded that the extracted water must 

have been injected water with no admixed ground 

water, and any observed decrease in THM 

concentrations in the extracted water must have been 

due to biodegradation.

The conventional conceptual model, the balloon 

model, does not explain the data collected during the 

aquifer injection, storage, and recovery tests for this 

study. Mass-balance results from the third cycle 

indicated that only 80 percent of the injected water had 

been recovered by the time the volume of water 

extracted reached 250 percent of the volume of water 

injected, implying that a large amount of mixing had 

occurred between injected water and ground water. 

Furthermore, the variations in (Br/Cl)THM values 

suggest that some of the mixing occurred during or 

shortly after injection while DOC and residual free 

chlorine in the injected water were still reacting to form 

THMs. The data collected for this study suggest a new 

conceptual model for injection, storage, and recovery 

that is illustrated by model B in figure 12.

Mixing between the injected water and ground 

water occurs primarily by advective transport and 

hydrodynamic dispersion. Hydrodynamic dispersion 

describes the spreading of a solute by mechanical 

dispersion and molecular diffusion (Bear, 1979). 

Mechanical dispersion is caused by velocity variations 

at the microscopic scale, and molecular diffusion is 

caused by the solute concentration gradient (Bear, 

1979). Molecular diffusion is on a much smaller scale 

than mechanical dispersion and its effect on mixing can 

be neglected. Dispersion occurs as individual parcels of 

fluid take different torturous paths through a porous 

medium. The greater the dispersion, the more the 

injected water will mix with the native ground water. 

Streetly (1998) used a numerical model of fluid flow to 

investigate the effect of dispersion on mixing between 

injected water and ground water. The numerical model 

consisted of an aquifer with no hydraulic gradients and 

no density contrasts between the ground water and 

injected water, and an equal volume of water was 

injected and extracted. Recovery of the injected water 

increased from 45 to 75 percent as the coefficient of 

dispersion was decreased from 40 to 2.5 meters 

(Streetly, 1998). 
38 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

Figure 12. A simple balloon conceptual model (A) and a more realistic conceptual model (B) for water flow in the aquifer during the third injection, 

storage, and recovery cycle (March 1998 through April 1999), Lancaster, Antelope Valley, California.

Model A Model B

Injection

period

Extraction

period

After

completion

of cycle

water level

injected water

ground

water

1

2

3
II. Formation and fate of trihalomethanes during the injection, storage, and recovery cycle 39

Advective transport also causes mixing of the 

injected water and ground water at the Lancaster site 

because the injection and extraction flow paths are not 

mirror images in a heterogeneous unconfined aquifer. 

The screened interval in the injection/extraction well 

used for this study (well 4-32) almost fully penetrated 

the upper and middle aquifers, extending continuously 

from 282 to 717 ft below land surface (fig. 2). Cores 

collected during the installation of the nested 

piezometers indicate that both of these aquifers consist 

of sand layers interbedded with layers of small gravel 

and layers of fine sand and silt (Fram and others, 2002). 

The simulation/optimization model, developed to 

evaluate the injection, storage, and recovery tests at 

Lancaster, required hydraulic conductivity values of 

17 and 1.7 ft per day for the upper and middle aquifers, 

respectively, to simulate the water levels observed 

during the tests (Phillips and others, in press). 

Therefore, a higher percentage of the injected water is 

expected to have entered the upper aquifer because of 

its significantly higher hydraulic conductivity. This is 

shown schematically in panel 1 of model B, figure 12.

During the extraction period, aquifer loss and 

well loss caused by turbulent flow through the well 

screen and inside the well to the pump intake caused a 

water-level drawdown in well 4-32 of greater than 70 ft 

(Phillips and others, in press). This drawdown reduced 

the effective aquifer thickness of the upper aquifer and 

thereby likely reduced the percentage of flow from the 

upper aquifer and likely increased the percentage of 

flow from the middle aquifer. In addition, the 

drawdown in the extraction well would effectively 

strand injected water in the upper part of the upper 

aquifer because it could no longer flow laterally into 

the extraction well, and the fine sand and silt layers 

present in the aquifer would retard its downward 

movement. The slow downward movement of this 

stranded injected water into the underlying producing 

aquifer zone could explain the continued presence of 

low concentrations of THMs in the extraction water, 

even after the volume of water extracted reached 250 

percent of the volume of water injected.

The change in the composition of water 

extracted from well 4-32 after the 4-month hiatus in 

extraction also provides insights into the nature of fluid 

flow in the aquifer. THM concentrations in the water 

extracted from well 4-32 were significantly higher at 

the start of the second phase of extraction than they had 

been at the end of the first phase (figs. 5A and 10C) and 

the chloride concentrations in the extraction water at 

the end of the second phase of extraction were lower 

than concentrations in the extracted water sampled at 

the end of the first phase of extraction (figs. 5F and 

10A). After the pump failed at well 4-32, there was no 

extraction from the site for about 2 months. This 

cessation of pumping allowed water-levels to rise 

(fig. 3C), and ground water from the upper aquifer that 

had relatively high THM and chloride concentrations 

(sampled by piezometer 27P8 [fig. 5A,F]) mixed with 

water in the underlying aquifer zones. When pumping 

resumed, this water—with its higher THM and chloride 

concentrations—was extracted quickly, resulting in a 

pulse of higher THM and chloride concentrations in the 

extracted water (fig. 5). The continued pumping from 

both wells 4-32 and 4-34 resulted in the largest waterlevel declines for piezometer 27P6 observed in the 

third cycle (fig. 3C). Pumping from both wells 

produced a large drawdown at well 4-32, which 

effectively eliminated the contribution of water high in 

THM and chloride concentrations from the upper part 

of the upper aquifer (this part of the upper aquifer is 

sampled by piezometer 27P8). Therefore, the chloride 

concentration in the extraction water at the end of the 

second phase of extraction reflects mixing between the 

remaining injected water and the native ground water 

that has a higher proportion of water from deeper zones 

that contain relatively low concentrations of chloride 

(about 4 mg/L in samples from piezometers 27P7 and 

27P6 collected prior to the third injection cycle; 

(fig. 5F). 
40 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

The results of this study indicate that the 

injection, storage, and recovery cycle at Lancaster 

cannot be described by a simple balloon model. We 

propose an alternative conceptual model that includes 

extensive mixing between injected water and ground 

water, and heterogeneities in the aquifer system. These 

heterogeneities, combined with significant drawdown 

in the extraction well, effectively strand injected water 

in the upper part of the aquifer system. This stranded 

injected water will be difficult to remove without 

modifying the extraction process. One possible 

modification would be reducing the pumping rate 

during the extraction period to reduce the drawdown in 

the extraction well. This would allow more of the 

injected water in the upper part of the aquifer system to 

directly enter the extraction well. Another would be 

installing additional extraction wells screened only in 

the upper aquifer. Without a modified extraction 

program, the results of this study indicate that the water 

quality of the aquifer system will be affected by the 

water quality of the injected water, even if 250 percent 

more water is recovered from the aquifer system than is 

injected into the aquifer system. Even if the extraction 

program is modified, the extensive mixing between 

injected water and ground water may make recovery of 

all injected water difficult.

Conclusions

The results presented in this section of the report 

support the following conclusions: 

1. The amount of continued THM formation in 

the injection water after injection into the aquifer was 

limited by the amount of residual chlorine present in 

the injection water at the time of injection. Reaction 

between DOC and residual chlorine to produce THMs 

continued for approximately four weeks after injection. 

Re-chlorinating injection water after extraction would 

result in additional THMs. The mean concentration of 

THMs formed in the injection water, including THMs 

formed before and after injection into the aquifer, was 

estimated to be 58 μg/L. Dechlorination of the 

injection water immediately before injection would 

reduce the total amount of THMs formed.

2. The changes in the concentrations of dissolved 

constituents in the extraction water during the 

extraction period were consistent with changes that 

would be caused by mixing the injection water with the 

ground water. The observed increasing (Br/Cl)THM

values concomitant with decreasing THM 

concentrations during the first phase of extraction (June 

30, 1998, through October 24, 1998) was inconsistent 

with predicted changes caused by biodegradation or 

sorption of THMs in the aquifer, but is consistent with 

predicted changes caused by mixing between THMfree ground water with a high bromide to DOC 

concentration ratio and injection water containing 

THMs and having a low bromide to DOC 

concentration ratio. At least some of the mixing must 

have occurred before all the residual chlorine in the 

injection water was consumed. 

3. Mass-balance calculations showed that 

67 percent of the chloride and THMs injected into the 

aquifer were recovered by the time that 132 percent of 

the volume of water injected had been extracted. 

Continued extraction of water to 250 percent of the 

volume injected only increased the percentage recovery 

of injected THMs to 80 percent. 

4. Dilution due to mixing of injection water and 

ground water was sufficient to explain the decrease of 

THM concentrations in the extracted water during the 

extraction period. Once we accounted for THM 

formation from reaction of the residual chlorine in the 

injected water, THMs behaved as conservative 

constituents in the aquifer.

5. The response of the ground-water flow system 

in the aquifer to injection, storage, and recovery cycles 

is extremely complex and does not follow a simple 

balloon model. Extensive mixing of injection water and 

ground water occurs and the water flow paths during 

injection and recovery are different; therefore, the 

injection water will be difficult to remove without 

modifying the extraction process. Without a modified 

extraction program, the results of this study indicate 

that the water quality of the aquifer system will be 

affected by the water quality of the injected water, even 

if 250 percent more water is recovered from the aquifer 

system than is injected into the aquifer system. Even if 

the extraction program is modified, the extensive 

mixing between injected water and ground water may 

make recovery of all injected water difficult
III. Modeling Dissolved Constituents, Trihalomethanes, and Sulfur Hexafluoride Tracer Concentrations in Extracted Water 41

III. MODELING DISSOLVED 

CONSTITUENTS, TRIHALOMETHANES, 

AND SULFUR HEXAFLUORIDE TRACER 

CONCENTRATIONS IN EXTRACTED WATER

By Brian A. Bergamaschi and Jordan F. 

Clark

The goal of the research presented in this section 

is to understand the decrease in trihalomethane (THM) 

concentrations in the extracted water: that is, were 

THM concentrations lower because of THM loss in the 

aquifer resulting from biodegradation or sorption, or 

were THM concentrations lower because of dilution by 

mixing with ground water? Results presented by 

Thomas and others (2000) suggest a potential for 

THMs to biodegrade in an oxic aquifer. Other studies 

report significant adsorption of THMs to soils (Walton 

and others, 1992) and humic coals and shales 

(Grathwohl, 1990). Miller and others (1993) and 

Thomas and others (2000) found that the observed 

decrease in THM concentrations in extracted water 

from injection, storage, and recovery cycles at the 

Las Vegas, Nevada site was primarily due to mixing 

between the injected water and ground water.

In this section, we further explore whether 

dilution of the injected water by ground water can 

explain the observed decrease in THM concentrations 

in the extraction water. Two independent models were 

used to estimate THM concentrations in extracted 

water; both assumed that mixing was the only process 

that reduced THM concentrations. The first modeling 

approach used a conservative tracer to predict THM 

concentrations in extracted water. The second 

modeling approach was to derive a simple descriptive 

mixing model to estimate physical processes in the 

aquifer. If the model-derived THM concentrations 

accurately represent the observed concentrations, then 

the assumptions inherent in the models are presumed to 

be correct, and mixing is presumed to cause the 

observed decline. If, on the other hand, model-derived 

concentrations diverge significantly from the observed 

values, then mixing cannot explain the observed 

decline, and the assumptions inherent in the models 

must be inaccurate. This approach relies on the fact that 

significant loss of THMs in the aquifer owing to 

adsorption to aquifer sediments or bacterial 

degradation would cause divergence from modelcalculated values.

Sulfur Hexafluoride as a Tracer

A tracer was necessary to model the THM 

concentrations because without a tracer there was no 

way to directly measure the relative amount of ground 

water and injected water in a given sample of extracted 

water. A tracer can be defined as a measurable 

substance, naturally present in a system or artificially 

introduced, that can be used to evaluate 

biogeochemical and hydrologic processes within that 

system. In this case, it is important that the tracer is 

conservative; tracer concentration should not be altered 

by sorption onto aquifer materials or by bacterial 

degradation.

Because tracer concentration is not dependent on 

total mass balance, it is insensitive to changes in 

pumping rate, a hiatus in pumping, or simultaneous 

extraction from nearby wells. Thus, it provides an 

accurate assessment of the contribution of ground 

water to each individual sample. For a tracer that is not 

present in the ground water, the fractional contribution 

of ground water to a sample of extracted water is 

simply one minus the ratio of the measured 

concentration of the tracer in a sample to the measured 

concentration in the injection water.

Some natural tracers have been used to 

characterize flow and mixing within aquifers; for 

example, chloride and bromide concentrations in water 

and the oxygen isotopic composition (δ18O) of the 

water commonly are used. However, most of these 

tracers were not deemed reliable for this experiment. 

Chloride concentrations varied substantially within the 

aquifer. Bromide reacted with the residual chlorine in 

the injection water and became incorporated into 

THMs and other disinfection by-products. δ18O values 

for injection water were not sufficiently different from 

those for the native ground water or the ground water 

influenced by previous cycles to permit its use to 

accurately assess mixing.
42 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

Sulfur hexafluoride (SF6) was introduced as a 

tracer into the injection flow stream (Fram and others, 

2002) because no reliable natural tracers existed for 

this system. Results of the experiment indicated that 

chloride could be used as a tracer. We chose SF6 as a 

tracer because it was important that the introduced 

tracer not alter the water-quality with respect to its use 

as drinking water. SF6 has extremely low toxicity, even 

at extraordinarily high levels and, therefore, is suitable 

for use in potable water systems. As an example of its 

low toxicity, animals breathing 80 percent SF6 and 

20 percent oxygen exhibited no symptoms of distress 

(Hathaway and others, 1991). The Federal standard 

allows 6 grams of SF6 per cubic meter of breathing air. 

The target concentration in this experiment was 100 

picomole per liter (pmol/L), or 15 parts per trillion, in 

the injection water. The actual concentration in the 

injection water was estimated to be approximately 

63 pmol/L.

From the standpoint of chemistry, SF6 is an 

excellent tracer. It is a gas at room temperature, boiling 

at 20°C. It is slightly soluble in water, but it is readily 

purged and may be measured at extremely low levels 

using a gas chromatograph fitted with an electron 

capture detector. The hexavalent, symmetric character 

of the molecule causes it to be chemically stable, 

resistant to biodegradation, and relatively unreactive 

with mineral surfaces. SF6 usually does not occur at 

measurable concentrations in nature. In the 

environmental sciences, it has been used extensively to 

characterize flow and gas exchange in surface-water 

systems (for example, Wannikhof and others, 1987; 

Clark and others, 1994). In engineering applications, it 

commonly is used as a leak indicator. More recently, it 

has been used to investigate infiltration of surface water 

into a ground-water basin (Gamlin and others, 2001).

The use of a tracer alone, however, does not 

provide a clear picture of the physical processes in the 

aquifer. To help understand these processes and to 

estimate the cumulative effect of repetitive cycles of 

injection, storage, and recovery, we applied a simple 

descriptive mixing model, using measured or estimated 

boundary conditions. In doing so, we were able to 

analytically estimate the physical dynamics of mixing 

within the aquifer. Comparing the concentrations 

predicted by the mixing model with chloride and SF6

tracer data indicated that reasonable estimates were 

obtained from the model. By using the model to 

simulate 10 successive cycles, we were able to asses 

potential long-term effects of annual subsurface 

injection, storage, and recovery cycles.

Experimental Methods

The procedures for collecting water samples 

from well 4-32 and the analytical methods used to 

measure THM, residual chlorine, and chloride 

concentrations and pH, temperature, and THM 

formation potential (THMFP) are described in detail by 

Fram and others (2002).

The procedures for adding SF6 to the injection 

water and collecting water samples for SF6 analysis, 

and the analytical method used to measure SF6

concentrations also are described in detail by Fram and 

others (2002) and are summarized briefly here. SF6

was added to the injection water flow stream by 

bubbling a calibrated gas mixture of 100 parts per 

million SF6 in nitrogen through a gas diffuser in the 

center of the flow stream. The gas flow rate was 

controlled at 70 milliliters per minute to achieve a 

target concentration in the injection water of 

100 pmol/L. Due to fluctuations in operational 

conditions, the actual concentration was less than the 

target concentration. Water samples were collected 

from a sampling port on well 4-32 and put directly into 

100-milliliter (mL) gas tight syringes that had secure 

luer port valves. Each water sample was equilibrated 

with 20 mL of ultrapure nitrogen gas that was added to 

the syringe, and then the headspace gas was drawn into 

a 20-mL evacuated container. The gas was analyzed by 

a gas chromatograph fitted with an electron capture 

detector (Wannikhof and others, 1987; Clark and 

others, 1994); the detection limit of the method was 

0.04 pmol/L.
III. Modeling Dissolved Constituents, Trihalomethanes, and Sulfur Hexafluoride Tracer Concentrations in Extracted Water 43

Tracer Results 

The goal of the tracer portion of this experiment 

was to provide an independent method of estimating 

the amount of injected water in a sample at any given 

time during extraction. We used two independent 

modeling approaches to derive the amount of extracted 

water and thus the THM concentration expected if 

mixing was the dominant reason for changes during 

extraction. For both models, we used the injected 

tracer, SF6, as well as a natural tracer, chloride. 

Chloride concentrations varied smoothly during the 

modeled extraction period, suggesting that chloride 

may be useful as a secondary tracer. Chloride is 

conservative in ground water systems and is not subject 

to sampling and analytical artifacts that are common to 

volatile constituents such as SF6 and THMs. To derive 

the concentration values expected during extraction, 

both modeling approaches used a mean injection water 

concentration for THMs, chloride, and SF6. 

Estimations of the mean concentrations of THMs and 

chloride in injection water have been presented 

previously (table 4), but a concentration was needed for 

SF6.

The target concentration for SF6 in the injection 

water was 100 pmol/L. However, measurements of SF6

concentrations during the injection period showed a 

large amount of variability, and the concentrations 

were almost universally below this value (fig. 13). 

Measured values for SF6 concentrations ranged from 

23 to 105 pmol/L (RSD = 43 percent); the median was 

40 pmol/L (fig. 13). In addition to the high variability 

between samples, replicate samples also had a much 

greater variability than expected. The average RSD for 

the 22 replicate samples analyzed was 19 percent. 

Analytical replicates typically are reproducible to 

within less than 2 percent of the measured value (Clark 

and others, 1994).

Figure 13. Sulfur hexafluoride (SF6) concentrations in injection and extraction water collected from well 7N/12W-27P2 (well 4-32) during the third 

injection, storage, and recovery cycle (March 1998 through April 1999), Lancaster, Antelope Valley, California. 

Data are from Fram and others (2002).

DATE

AM J J A SO ND J F MA

SULFUR HEXAFLUORIDE, IN PICOMOLES PER LITER

0

25

50

75

100

125

Injection Extraction

M

1998 1999

average

injection

water
44 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

The high replicate variability provided clues to 

the reasons for erratic values during injection and led 

us to discard many values in determining the mean 

concentration. Variations in SF6 concentration may be 

the result of (1) a large variability in pumping rate, 

(2) poor equilibration of SF6 during the transit between 

the diffuser and the sampling point, (3) bubble 

formation in the injection line, or (4) losses during 

sampling. Because the pumping rate was monitored 

continuously (fig. 3), it could be eliminated as a 

possible confounding factor. All remaining scenarios 

would result in measured SF6 concentrations lower 

than actual injected concentrations. Anecdotal 

observations suggest that all scenarios contributed to 

the observed high variability (Fram and others, 2002). 

We chose the mean of the upper 50 percent of the 

measurements to represent the injected SF6

concentrations because it provided a statistically robust 

estimate of the mean injection water concentration, but 

without skewing the mean by incorporating 

anomalously low values. The mean SF6 concentration 

in the injection water thus obtained was 63.6 pmol/L 

(RSD = 29 percent) (fig. 13).

The variations in injected SF6 and THM 

concentrations during the extraction period provide a 

good indicator of important processes occurring in the 

aquifer. Following completion of the injection period, 

well 4-32 remained idle for 2 weeks before the 

extraction period began (fig. 3A). Once extraction 

began, the THM and SF6 concentrations in extraction 

water samples (figs. 5A and 13) declined roughly 

exponentially during the first stage of extraction 

(June 30 to October 24, 1998), as may be expected 

from a process controlled by mixing (Berner, 1980). 

Samples collected 24 days after the onset of extraction, 

when 31 percent of the injection water volume had 

been extracted, contained SF6 concentrations of 

50 pmol/L (fig. 13), corresponding to 78 percent of the 

concentration in the injection water. Near the time that 

the extracted water volume equaled the injected water 

volume, samples collected contained SF6

concentrations of 14 pmol/L, corresponding to 

22 percent of the concentration in the injection water. 

SF6 concentrations declined until pump failure on well 

4-32 on October 24, 1998, caused a hiatus in 

extraction. A few days prior to pump failure, on 

October 21, 1998, extraction water samples contained 

11 pmol/L of SF6 (fig. 13), corresponding to 17 percent 

of the concentration in the injection water.

SF6 concentrations increased during the 

pumping hiatus at well 4-32, suggesting a higher 

fraction of injection water migrated into the highvelocity zone of the aquifer during this time (see the 

second section of this report for further discussion). 

The SF6 concentration again declined following 

resumption of extraction. Even 10 months after 

extraction began and after extracting 250 percent of the 

volume of water that was injected (as well as a similar 

volume from nearby well 4-34) (fig. 3), SF6

concentrations declined to only 9 pmol/L (fig. 13), 

corresponding to 14 percent of the concentration in the 

injection water.

Modeling

The initial hypothesis was that systematic 

decrease in the THM concentrations in the extraction 

water during the extraction period was not significantly 

different from the decrease expected from mixing 

between injection water and ground water in the 

aquifer. To test this hypothesis, the tracer model and the 

descriptive mixing model explicitly assume that mixing 

is the only process that reduced THM concentrations in 

the extraction water. The modeling was simplified by 

assuming that variability in many of the parameters did 

not strongly affect the results. In particular, we 

assumed that the concentrations of constituents in the 

injection water and the ground water were adequately 

represented by mean concentrations.

Tracer Mixing Model

If THMs behave conservatively and the only 

process affecting their concentration is mixing, the 

THM concentration in any sample of extracted water 

can be predicted using the tracer mixing model. A 

significant difference between the predicted THM 

concentrations in the extracted water and the measured 

concentrations would indicate that processes other than 

mixing must have affected the THM concentrations. 

The tracer mixing model only requires data for the 

concentrations of the tracer and of the constituents of 

interest in the ground water, the injected water, and the 

extracted water. No data for water volumes and no 

assumptions concerning the physical scenario in which 

mixing occurs are needed. The tracer is assumed to be 

perfectly conservative, and samples of extracted water 

are assumed to be simple mixtures of injected water 
III. Modeling Dissolved Constituents, Trihalomethanes, and Sulfur Hexafluoride Tracer Concentrations in Extracted Water 45

and ground water. The predicted concentration of a 

conservative constituent in any sample of extracted 

water, Ct, is calculated from the known concentrations 

of the tracer and the constituent in the ground water 

and injected water and the measured concentration of 

the tracer in the sample of extracted water (equations 

10 and 11):

(10)

where 

(11)

and where

The tracers SF6 and chloride were used in the 

tracer mixing model to predict concentrations of THMs 

in the extraction water samples. There was good 

correspondence between the THM concentrations 

predicted using SF6 or chloride as the tracer (fig. 14) 

and the concentrations measured in the extracted water 

during the first phase of extraction before the pump 

failed on October 24, 1998. These results indicate that 

THMs behaved as a conservative constituent; no 

systematic deviation was observed between the modelpredicted and the measured THM concentrations as 

would have been expected if significant biodegradation 

or sorption of THMs had occurred. Thus, the results of 

the tracer mixing model suggest that mixing between 

injected water and ground water was the primary 

mechanism controlling the concentration of THMs in 

the extracted water. 

The THM concentrations predicted using SF6 as 

the tracer also correspond well to the THM 

concentrations measured in the extracted water during 

the second phase of extraction, after resumption of 

pumping at well 4-32 on February 22, 1999 (fig. 14A). 

However, the THM concentrations predicted using 

chloride as the tracer do not match the measured THM 

concentrations during this period (fig. 14B). The failure 

of the chloride tracer during the second phase of the 

extraction period likely reflects a change in the 

composition of the ground water near well 4-32 caused 

by pumping from nearby well 4-34 (see the second 

section of this report for more discussion).

Another way of looking at the tracer data is to 

examine the relation between the SF6 tracer 

concentrations and the THM concentrations in the 

extraction water samples (fig. 15). Linear regression of 

these two constituents yielded a slope of 0.72 μg THM 

per pmol SF6 and an r2 value of 0.55, indicating that 

they are correlated —as would be expected if they had 

been diluted simultaneously by mixing with ground 

water containing no SF6 or THMs. The slope 

approximates the average expected concentration ratio 

of 0.91 THM:SF6 (fig. 15), the initial THM 

concentration in the injected water in the aquifer 

(58 mg/L) divided by the SF6 concentration in the 

injected water (63.6 pmol/L), supporting the 

conclusion that THM concentrations and SF6 tracer 

concentrations in extraction water samples were 

controlled by the same process. The same calculation 

was not done for chloride because chloride does not 

trace mixing over the entire extraction period.

Ct is the concentration of the constituent 

in the extraction water at any time t.

ft is the fraction of injected water in the 

water extracted at any time t,

C0 is the concentration of the constituent 

in the injection water,

Cgw is the concentration of the constituent 

in the ground water, and

Ct

tracer is the concentration of the tracer in the 

extraction water at any time t, 

C0

tracer is the concentration of the tracer in the 

injection water, 

Cgw

tracer is the concentration of the tracer in the 

ground water,

Ct ftC0 1 ft = + ( ) – Cgw

ft

Ct

tracer Cgwtracer –

C0

tracer Cgwtracer –

= -----------------------------------------
46 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

Figure 14. Measured and sulfur hexafluoride (SF6) tracer-derived (A) and chloride tracer-derived (B) trihalomethane concentrations in extraction water 

collected from well 7N/12W-27P2 (well 4-32) during the third injection, storage, and recovery cycle (March 1998 through April 1999), Lancaster, Antelope 

Valley, California. TRIHALOMETHANES, IN MICROGRAMS PER LITER

0

10

20

30

40

50

60

70

MEASURED

SF6 TRACER DERIVED

DATE DURING EXTRACTION PERIOD

JJ AS O N D J FMA M

0

10

20

30

40

50

60

70

MEASURED

CHLORIDE TRACER DERIVED

A

B

1998 1999
III. Modeling Dissolved Constituents, Trihalomethanes, and Sulfur Hexafluoride Tracer Concentrations in Extracted Water 47

Descriptive Mixing Model 

A simple model was used to evaluate if mixing 

of native ground water with the injection water could 

explain the observed THM concentrations in the 

extraction water. The model selected is a single-zone 

mixing model commonly used to assess the 

exponential decrease in concentration of conservative 

constituents associated with dilution and mixing (for 

example, Berner, 1980). The correspondence between 

the model and the physical environment may be 

conceptualized by assuming that water injected into the 

aquifer displaces the ground water surrounding the 

injection well and that upon extraction, the injected 

water mixes homogeneously with the displaced water 

within the mixing zone (fig. 16). Thus, the volume of 

the mixing zone (V) is defined in the model to equal the 

volume of injected water, and constituents dissolved in 

the water within this mixing zone are assumed to be 

homogeneously distributed by mixing. Thus, the 

concentration of dissolved constituents in extracted 

water at any time (t) during the extraction period is 

assumed to equal the concentration of the constituents 

in the mixing zone at that time (Ct). The concentrations 

of constituents in the ground water entering the mixing 

zone from the surrounding aquifer (Cgw) are assumed 

to equal the concentrations measured in samples from 

the well collected prior to the injection period and 

invariant with time. To maintain mass-balance, the 

volume of ground water entering the mixing zone from 

the surrounding aquifer per unit time is set equal to the 

volume of water extracted from the well (Q).

Figure 15. Trihalomethane and sulfur hexafluoride concentrations in extraction water collected from well 7N/12W-27P2 (well 4-32) during the third 

injection, storage, and recovery cycle (March 1998 through April 1999), Lancaster, Antelope Valley, California. 

The concentration ratio predicted from the average injected water composition and the linear regression of the concentration data are shown as solid lines.

0 10 20 30 40 50 60 70

TRIHALOMETHANES,

IN MICROGRAMS PER LITER

Extracted water samples

Average injected water

Predicted concentration

ratio,y= 0.91x

Linear regression

y = 0.72x + 4.2

r 2 = 0.55

SULFUR HEXAFLUORIDE,

IN PICOMOLES PER LITER

0

10

20

30

40

50

60
48 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

Note that this model is not intended to represent 

the actual physical mixing processes within the aquifer. 

Undoubtedly physical mixing occurs both during 

injection and extraction, and the mixing process is not 

simply the mixing of water from two sources. Evidence 

of inhomogeneous distribution of ground water and 

injection water surrounding the injection well is 

obtained from microgravity surveys, velocity log data, 

and water-quality data (Fram and others, 2002; 

Metzger and others, 2002; Phillips and others, in 

press). Nevertheless, this model is a useful heuristic 

tool to assess the effects of mixing during this 

injection, storage, and recovery cycle as well as other 

operational scenarios.

The change in mass (M) of any conservative 

constituent in the mixing zone over time (t) is given by 

the following equation: 

 

(12)

where 

M is the mass of a conservative constituent in 

the mixing zone,

t is any time.

Qgw is the volume of water that enters the mixing 

zone per unit time from ground water,

Cgw is the concentration of the constituent in the 

ground water and the concentration of the 

constituent in the mixing zone, 

Qext is the volume of water extracted from the 

mixing zone per unit time by pumping, and

Ct is the time dependent concentration of the 

constituent in the mixing zone (Ct is the 

same as the concentration in the water 

extracted by pumping).

dM

dt

--------- QgwCgw QextCt = –

Figure 16. Diagram of the descriptive mixing model for mixing between injection water and ground water during extraction.

Extraction from well, Q

Zone of displaced ground water

Mixing zone = volume

of injected water, V

Displaced ground water

entering mixing zone

during extraction

Instantaneous mixing

in mixing zone
III. Modeling Dissolved Constituents, Trihalomethanes, and Sulfur Hexafluoride Tracer Concentrations in Extracted Water 49

To obtain information about concentration, the 

mass (M) of a constituent in the mixing zone can be 

separated into the volume of the mixing zone (V), 

which does not vary over time, and the concentration of 

the constituent in the mixing zone (Ct), which does:

(13)

where 

Solving the equation for Ct, then integrating over 

the extraction time gives a continuous function of 

concentration during the extraction period: 

(14)

where 

The value of β may be obtained by solving the 

expression at the initial condition of the model:

(15)

where

For conservative constituents, Ct = 0 equals the 

concentration of the constituent in the injection water 

(Cinj). The single-zone mixing model was used to 

estimate the changes in concentration of SF6, chloride, 

and THMs due to mixing during the extraction period 

for the third injection, storage, and recovery cycle. 

Calculating Ct using equations 14 and 15 requires 

values for Cinj and Cgw. Values of Cinj and Cgw for SF6, 

THMs, and chloride are given in table 4.

The form of the solution (equation 14) shows 

that the model-derived change in concentration of a 

conservative constituent is an exponential function in 

time, but has a linear dependence on the concentration 

of the constituent in both injection and ground water. 

According to the initial assumptions, the exponential 

argument is a function of Q, the extraction rate, and V, 

the total volume of water injected. Model-derived 

concentrations of SF6, chloride and THMs for Q/V

values of 0.010 and 0.024 per day bracket the majority 

of the concentrations measured in extraction water 

samples (fig. 17A–C). The average pumping rate over 

the modeled period (June 30 to October 24, 1998) was 

670,000 gallons per day (fig. 3A), and the total volume 

of water injected was 58 million gallons (fig. 3B), 

which corresponds to a Q/V of 0.012 per day. Using a 

value of 0.012 per day for Q/V in the model yields 

computed concentrations that are greater than the 

measured SF6 and THM concentrations and are less 

than the measured chloride concentrations 

(fig. 17A–C). Since the initial concentrations of SF6, 

chloride, and THMs, and the pumping rate are known, 

the divergence between the model-derived 

concentrations and the observed concentrations may 

indicate that one of the initial assumptions was 

inaccurate, namely, the assumption that the mixing 

zone was equal to the total volume of water injected. A 

Q/V of 0.014 per day yielded the smallest difference 

between the measured SF6 and chloride concentrations 

and the model-derived concentrations. The difference 

was quantified using the sum of the squares of the 

differences between the measured and calculated 

values. A Q/V of 0.014 per day corresponds to a mixing 

zone volume of 48 million gallons, about 10 million 

gallons less than the volume actually injected. Recall 

that results from this study suggest that drawdown in 

the extraction well effectively isolated injected water in 

the shallow part of the aquifer system (see section 

“Implications for Water Flow in the Aquifer” in the 

second section of this report), preventing it from 

mixing. The modeled THM concentrations using a Q/V

of 0.014 per day reasonably matched the measured 

THM concentrations. 

Q is equal to Qext, which equals Qgw because the 

model is constrained to maintain mass 

balance of water in the mixing zone

β is the integration constant.

Ct = 0 is the concentration of the constituent in 

the extracted water at the beginning of 

the extraction period.

V

dCt

dt

-------- Q Cgw Ct = ( ) –

Ct Cgw βe–( ) Q V⁄ t = +

β Ct = 0 Cgw = –
50 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

Figure 17. Measured and model-derived sulfur hexafluoride concentrations (A), chloride concentrations (B), and trihalomethane concentrations (C) in 

extraction water collected from well 7N/12W-27P2 (well 4-32) during the third injection, storage, and recovery cycle (March 1998 through April 1999), 

Lancaster, Antelope Valley, California. TRIHALOMETHANES, IN MICROGRAMS PER LITER

0

10

20

30

40

50

60

CHLORIDE,

IN MILLIGRAMS PER LITER

4

8

12

16

20

24

SULFUR HEXAFLUORIDE,

IN PICOMOLES PER LITER

0

10

20

30

40

50

60

70

Q/V = 0.014

Q/V = 0.010

Q/V = 0.010

June July August September October

DATE DURING EXTRACTION PERIOD IN 1998

Model-derived

Model-derived

Measured

A

B

C

Q/V = 0.010

Q/V = 0.012

Q/V = 0.024

Q/V = 0.014

Q/V = 0.012

Q/V = 0.024

Q/V = 0.014

Q/V = 0.012

Q/V = 0.024

Ratio of extraction

rate to injection

Q/V
III. Modeling Dissolved Constituents, Trihalomethanes, and Sulfur Hexafluoride Tracer Concentrations in Extracted Water 51

Based on the good correspondence between the 

rate of change in the concentrations of THMs, SF6, and 

chloride and the rate predicted by the model using 

realistic estimates of the concentrations in the injected 

water and ground water, we concluded that the simple 

mixing of the injected water with ground water can 

explain the observed time-dependent variations in SF6, 

chloride, and THM concentrations in the extraction 

water during the third injection, storage, and recovery 

cycle. In addition, comparing model results with 

concentrations measured in extraction water samples 

suggests that not all of the water injected during a cycle 

is easily extracted during that cycle. Finally, the good 

correspondence between the measured concentrations 

and the model-derived values indicates that using a 

Q/V of 0.014 in the model reasonably reproduced 

concentration variations over this cycle and could be 

used to predict the effects of repetitive injection, 

storage, and recovery cycles. This value was used for 

all subsequent model calculations.

Potential Implications of Mixing for Repetitive 

Cycles of Injection, Storage, and Recovery

The concordant results from the two independent 

tracers as well as the THMs indicated that the 

simplified mixing model was sufficiently accurate as a 

descriptive tool to use it for estimating mixing 

processes within the aquifer during successive cycles 

of injection and extraction and during different 

scenarios of volumes of water injected and extracted. 

Using the model to test scenarios and forecast the 

results of successive cycles assumes that the simple 

descriptive modeling approach will continue to provide 

a reasonable estimate of the complex physical 

processes occurring in the aquifer and that advection 

does not affect the concentration of conservative 

constituents in the ground water by moving “new” 

ground water into the mixing zone.

The first scenario tested was ten annual 

repetitions of a cycle like the one at well 4-32 

described in this report—2 months of injection 

followed by 6 months of extraction. The ratio of the 

rate of extraction to the total volume of water injected 

(Q/V) was held at 0.014 per day during all ten cycles. 

At the onset of each cycle, the concentration of THMs 

(or of any conservative constituent) in the mixing zone 

was set to equal 100 percent of the concentration in the 

injection water, and the concentration in the zone of 

displaced ground water was set to equal the 

concentration in the water in the mixing zone at the end 

of the previous cycle (the residual concentration in the 

aquifer). The model was run for 10 annual cycles of 

injection and extraction to assess whether constituents 

from the injection water accumulate in the aquifer.

Constituents from the injection water 

accumulated in the mixing zone (fig. 18A) during the 

ten cycles. The rate of decline of the concentration in 

the mixing zone decreased in each subsequent cycle, 

resulting in an increase in the concentration in the 

mixing zone at the end of each subsequent cycle (the 

residual concentration). This residual concentration 

was 57 percent of the injected water concentration after 

10 years (fig. 18A). Note that this model scenario was 

based on pumping 2.5 times more water from the 

aquifer than was injected every year for ten years. So in 

essence, this scenario represents a best-case scenario 

for preventing build-up of injected conservative 

constituents in the aquifer. As can be demonstrated by 

modeling more realistic injection, storage, and 

recovery operations, if the volume of water extracted is 

closer to the volume injected, then the residual 

concentration after each cycle will be higher.

Subsequent tests used a more realistic 

operational scenario wherein the injection and 

extraction periods were each six months, and the 

volume of water injected was equal to the volume of 

water extracted. This cycle was repeated to assess the 

cumulative effect of ten annual cycles (fig. 18B). The 

residual concentrations in the water in the aquifer at the 

end of each cycle were much higher in this scenario 

than those in the original scenario, and the rate of 

increase of the residual concentrations during the first 

five annual cycles was also much greater (fig. 18A,B). 

The residual concentration in the water in the aquifer 

was indistinguishable from the concentration in the 

injection water after 10 annual cycles.
52 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

Figure 18. Predicted concentrations of a conservative injected constituents in an aquifer during 10 annual cycles of injection, storage, and recovery for two 

different injection period durations (A,B) and four different amounts of mixing between cycles (B,C,D,E). 

Inj (injection) and ext (extraction) are the durations of the injection and extraction periods in months, and mix is the percentage of mixing between cycles. REMAINING INJECTED CONSTITUENT, IN PERCENT

YEAR

1 2 3 4 5 6 7 8 9 10

A

B

C

D

E

0

25

50

75

100

0

25

50

75

100

0

25

50

75

100

0

25

50

75

100

0

25

50

75

100

inj=2

ext=6

mix=0

inj=6

ext=6

mix = 25

inj=6

ext=6

mix = 95

inj=6

ext=6

mix = 50

inj=6

ext=6

mix=0
III. Modeling Dissolved Constituents, Trihalomethanes, and Sulfur Hexafluoride Tracer Concentrations in Extracted Water 53

However, the model assumes no mixing between 

the model boundary and surrounding, constituent-free 

ground water, such as may occur in the aquifer. Mixing 

across the model boundary between the zone of 

displaced ground water and surrounding, constituentfree ground water would result in greater dispersion, 

but slower increase in residual concentration. To 

simulate possible outcomes of such mixing, the 

descriptive mixing model was revised to 

instantaneously lower the residual concentration at the 

end of each cycle by 25 percent, 50 percent, and 

95 percent and to use the modified value as the model 

input for the ground water concentration in the next 

cycle (fig. 18C,D,E).

Note that increasing the mixing of the residual 

ground water lowers the ultimate extent of 

accumulation but not the rate (fig. 18C,D,E). Also, 

when 95 percent of the residual THM is lost by 

dispersion between cycles, the eventual accumulation 

of THM in the mixed zone is nearly 40 percent of that 

in the original injected water. Regardless of the mixing 

of the displaced ground water, these results suggest that 

THM and other conservative constituents present in 

injection water will accumulate in the aquifer, and they 

will accumulate more rapidly as injection times are 

longer and extraction times are shorter, regardless of 

the extracted volume. Longer extraction times will 

result in greater removal.

Removal of this residual concentration at a later 

time may be difficult because the ratio of injected water 

to ground water in the extracted water declines 

asymptotically owing to mixing. Thus, removing the 

residual concentration would require pumping several 

times the injected-water volume. Reducing the 

concentration of a constituent in the water in the 

mixing zone in the aquifer to 10 percent of its 

concentration in the injected water would require 

removal of 2.3 times the volume of water injected, 

reducing it to 1 percent would require removing 

4.6 volumes, and reducing it to 0.1 percent would 

require removal of 9 volumes. Thus, based on the 

model results for two scenarios of 10 annual cycles of 

injection, storage, and recovery, we conclude that there 

are substantial potential long-term consequences to the 

injection of conservative constituents into the aquifer.

Conclusions

The results of these modeling efforts support the 

following conclusions:

1. The measured variation in THM 

concentrations during the extraction period of the third 

injection, storage, and recovery cycle may be explained 

fully by mixing of injected water with ground water, as 

shown by the tracer mixing and the descriptive mixing 

models. Extensive mixing with ground water appears 

to be the cause of the observed decline in THM 

concentrations during the extraction period.

2. THMs behaved conservatively in the aquifer. 

THM formation continues in the aquifer due to reaction 

between residual chlorine and DOC in the injection 

water at the time of injection. However, there was no 

evidence that biodegradation or sorption within the 

aquifer caused a decrease in THM concentrations, nor 

was there evidence for enhanced formation of THMs in 

the aquifer.

3. By using a simple descriptive mixing model 

calibrated to measured data, it was possible to forecast 

the effects of repeated injection, storage, and recovery 

cycles. Repeated cycles may increase the residual 

concentration of THMs (or any conservative 

constituent with a concentration greater in the injected 

water than in the ground water) in the aquifer. The 

extent of this increase was directly related to the ratio 

of the rate of extraction to the total volume injected 

(Q/V) and to amount of mixing across model 

boundaries. For realistic ratios of the volume of 

extracted to injected water, repeated cycles should 

yield residual concentrations in the aquifer 

approaching 100 percent of injection-water 

concentrations by the end of 10 annual cycles if 

residual concentrations are not lowered by further 

mixing.
54 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

IV. THE POTENTIAL FOR BIODEGRADATION 

OF TRIHALOMETHANES BY AQUIFER 

BACTERIA

By Kelly D. Goodwin

The purpose of this part of the study was to 

determine whether bacterial degradation of THMs can 

attenuate THM concentrations in the ground water 

after injecting surface water into the aquifer system. 

Three types of studies were performed: sediment 

microcosm experiments, water enrichment microcosm 

experiments, and measurements of bacterial densities 

in water samples. The sediment microcosms consisted 

of aquifer sediment and ground water, and the water 

enrichment microcosms consisted of ground water or 

extraction water amended with bacteria and particles 

concentrated from a larger volume of water. Live and 

sterilized vials of sediment microcosms and water 

enrichment microcosms were prepared, and many were 

amended with nutrients and vitamins. CHCl3 and 

CHBr3 were added to the vials and the amounts were 

monitored during an incubation period. Biodegradation 

of the CHCl3 or CHBr3 by bacteria in the aquifer 

sediment or in the water samples would be indicated in 

these experiments by a decrease in the amount of 

CHCl3 or CHBr3 detected in the live vial relative to the 

amount in the corresponding sterile vial. Bacterial 

densities were measured in water samples collected 

from the wells and the nested piezometers to determine 

if the population of bacteria in the aquifer was affected 

by the injection, storage, and recovery cycle. Detailed 

descriptions of the experimental and analytical 

methods used are given by Fram and others (2002).

Biodegradation of halogenated organic 

compounds has been observed in a variety of 

environments, including anaerobic ground water 

(McCarty and others, 1984), freshwater, estuarine 

water, seawater, and water from a hyper saline lake 

(Goodwin and others, 1998). Phylogenetically diverse 

bacteria having the ability to degrade halogenated 

organic compounds have been isolated. Biodegradation 

typically is carried out by one of three mechanisms: cometabolic dehalogenation, aerobic metabolism, or 

reductive dehalogenation. 

When dehalogenation occurs co-metabolically, 

halogenated organic compounds degrade without 

bacterial growth. The bacteria seem to receive no 

benefit; instead, compounds are dehalogenated by an 

enzyme that has broad substrate specificity (Wackett 

and others, 1992). Methane-, ammonia-, toluene-, and 

butane-oxidizing bacteria have been shown to degrade 

halogenated organic compounds co-metabolically 

(Henson and others, 1988; McClay and others, 1996; 

Hamamura and others, 1997; Moran and Hickey, 

1997). Degradation of CHCl3 and CHBr3 can be 

mediated by certain methane- (Bartnicki and Castro, 

1994) and toluene-oxidizing bacteria (McClay and 

others, 1996). Furthermore, acclimating soils to an air 

and natural gas mixture stimulated the biological 

oxidation of CHCl3 to carbon dioxide (Strand and 

Shippert, 1986).

Metabolism (degradation with growth) occurs 

when dehalogenation creates a compound that is used 

in the metabolic pathway of the organism, allowing the 

organism to use the halogenated organic compound as 

a source of carbon and energy (Leisinger and Bader, 

1993). Bacteria that can grow on methyl chloride 

(Vannelli and others, 1998), methyl bromide (Connell 

Hancock and others, 1998), dichloromethane 

(Doronina and Trotsenko, 1994), and dibromomethane 

(Doronina and Trotsenko, 1994; Goodwin and others, 

1998) have been isolated. Some of these bacteria are 

facultatively methylotrophic, meaning that they can 

grow by consuming organic compounds other than 

halogenated methanes. For example, a bacterium 

isolated from agricultural soil, strain ImB-1, can grow 

on methylamines and glucose as well as on 

halogenated methanes (Connell Hancock and others, 

1998). A bacterium isolated from a seawater 

enrichment culture (Goodwin and others, 1998), 

Leisingeria methylohalidivorans strain MB2, can grow 

on halogenated methanes, yeast extract, and casein 

(Schaefer and others, 2002). The ability to grow on a 

variety of substrates probably is an important survival 

strategy for bacteria that can metabolize halogenated 

organic compounds and live in nonpolluted 

environments because concentrations of halogenated 

organic compounds are very low in nonpolluted 

environments—typically at the picomole per liter 

(pmol/L) level (Manley and others, 1992; Moore and 

Tokarczyk, 1993). Bacteria do consume halogenated 
IV. The Potential for Biodegradation of Trihalomethanes by Aquifer Bacteria 55

organic compounds present at nanomole per liter 

(nmol/L) (Goodwin and others, 1998) and pmol/L 

concentrations (King and Saltzman, 1997; Hines and 

others, 1998). For example, L. methylohalidivorans and 

strain ImB-1 can consume 2.4 pmol/L (equivalent to 

approximately 12 parts per trillion) of methyl bromide, 

though both routinely are grown on 100 μmol/L 

(micromole per liter) of methyl bromide (Goodwin and 

others, 2001). 

Methanes with three or more halogen atoms 

seem to be more difficult to metabolize in aerobic 

environments than methanes with one or two halogen 

atoms. Aerobic metabolism of trihalomethanes has not 

been observed definitively, and bromoform (CHBr3) 

has been found to be resistant to biodegradation in 

several aerobic, aquatic environments (Goodwin and 

others, 1998). However, carbon tetrachloride was 

biodegraded in aerobic ground water and soils (Happell 

and Wallace, 1998). Studies of the fate of THMs in the 

aerobic aquifer system used for injection, storage, and 

recovery in Las Vegas, Nevada, have yielded 

conflicting results. Some investigators concluded that 

the decline in THM concentrations during water 

extraction was due to biodegradation (Singer and 

others, 1993; Pyne and others, 1996), whereas others 

concluded that the declines in THMs were due solely to 

dilution of the injected water by mixing with the 

ground water (Miller and others, 1993; Bernholtz and 

others, 1995). Most recently, Thomas and others (2000) 

concluded that both cases are true. In the early years of 

the Las Vegas project, declines were due solely to 

dilution, but in later years, biodegradation of 

brominated THM species seems to have contributed to 

the observed decline in THM concentrations during 

extraction. The decline in concentrations of brominated 

THM species was greater than expected for dilution 

alone, and the inorganic chemistry of the system did 

not change significantly with time, making changes in 

abiotic processes an unlikely explanation. However, the 

population of bacteria in the aquifer increased greatly 

with time as indicated by heterotrophic plate counts 

and bacterial growth on well screens. An acclimation 

period—the time period when no biological destruction 

of a chemical is detected—was observed in many 

environments for a variety of chemicals (Alexander, 

1994). In fact, absence of an acclimation period usually 

indicates chemical, not biological, mechanisms of 

destruction (McCarty and others, 1984). 

THMs are susceptible to reductive 

dehalogenation. This process typically is mediated by 

anaerobic bacteria (Bagley and Gossett, 1995; KohlerStaub and others, 1995), although aerobic reductive 

dehalogenation has been observed in the laboratory 

(Castro, 1993). Numerous field studies have shown 

evidence for degradation of halogenated organic 

compounds in anaerobic aquifers (Bouwer and others, 

1981, 1984; McCarty and others, 1984; Roberts, 1985) 

including aquifers used for injection, storage, and 

recovery projects (Singer and others, 1993). It is 

possible that aerobic aquifers contain microzones that 

have little or no oxygen. Happell and Wallace (1998) 

maintained that the carbon tetrachloride degradation 

observed in aerobic ground water and soils could have 

been due to such microzones. It also is possible that 

microzones of low oxygen could explain the results in 

the report by Thomas and others (2000). Although 

oxygen concentrations generally were in the range of 

4 to 6 mg/L in the Las Vegas aquifer, sulfate- and ironreducing bacteria sometimes were noted during plate 

counts, which indicated that anaerobic conditions 

existed near some of the wells used for injection and 

recovery.

Sediment Microcosm Experiments

Sediment microcosms were constructed using 

water collected from well 4-32 on March 4, 1998 

(before the injection period) and sediment from the 

core taken from the depth corresponding to that of the 

screened interval of piezometer 27P7 (Fram and others, 

2002). Enriched sediment microcosms were made by 

amending some of the sediment microcosms with 

nutrients and vitamins. The microcosms were placed in 

sealed vials and were spiked with known masses of 

CHCl3 or CHBr3 prior to incubation for 145 days 

under aerobic or anaerobic conditions. Anaerobic 

conditions were established by flushing the headspace 

of the vial with nitrogen to remove oxygen. Sterile 

controls were prepared by filtering the water through 

sterile 0.2-micron pore-size filters prior to filling some 

of the vials, or by autoclaving the microcosms. The 

mass of CHCl3 or CHBr3 in the headspace of each vial 

was measured several times during the incubation 

period using head-space gas chromatography and a gas 

chromatograph fitted with an electron-capture detector.
56 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

Aerobic sediment microcosms spiked with 

CHCl3 or CHBr3 showed no significant bacterial 

degradation of CHCl3 or CHBr3 after 145 days of 

incubation, as indicated by the overlap of the data for 

the live and sterile vials for each measurement time 

(fig. 19A,B). In addition, the slopes of the mass of 

CHCl3 or CHBr3 versus time were calculated for the 

live and the sterile control samples; the standard errors 

of the two slopes overlapped (not shown). The standard 

error of the slope is the error in the slope of the linear 

regression equation fit to the data (for example, Helsel 

and Hirsch, 1995). Overlap of the standard error of the 

slopes for the live and sterile control samples indicates 

that regression of the two data sets yields lines with 

statistically indistinguishable slopes—in other words, 

the change in the mass of CHCl3 or CHBr3 during the 

incubation period is the same for the live and sterile 

control samples. 

Aerobic enriched sediment microcosms were 

made by amending sediment microcosms with 

potassium dihydrogen phosphate (KH2PO4) 

(0.02 gram per liter), ammonium chloride (NH4Cl) 

(0.5 gram per liter), and vitamins including B12

(1 milliliter per liter) (Pfennig, 1978). Adding minerals 

and vitamins should have increased the likelihood of 

biodegradation compared with unamended sediment 

microcosms. Aerobic enriched sediment microcosms 

showed no significant bacterial degradation of CHCl3

or CHBr3 after 145 days of incubation, as indicated by 

the overlap of the data for the live and sterile control 

samples at most measurement times (fig. 20A,B), and 

the overlap of the standard errors for the slopes of the 

mass versus time data for the live and sterile control 

samples (not shown).

Biodegradation of halogenated organic 

compounds has been reported in anaerobic aquifers 

(Bouwer and others, 1981, 1984), including those used 

for injection, storage and recovery (Singer and others, 

1993). Therefore, a second set of enriched sediment 

microcosm samples was incubated under anaerobic 

conditions. Anaerobic conditions were established by 

flushing the headspace of the vial with nitrogen gas to 

remove oxygen.

The live and sterile control anaerobic, enriched 

sediment microcosms behaved differently than the 

aerobic microcosms during the incubation period. An 

equal amount of CHCl3 was present in live and sterile 

control anaerobic enriched sediment microcosms after 

11 days of incubation (fig. 21). However, the amount of 

CHCl3 in live and sterile control samples differed 

significantly at 53 days of incubation (the Student t

test, α = 0.05). However, between 53 and 145 days of 

incubation, the mass of CHCl3 in the live samples did 

not decline relative to the mass in the sterile control 

samples, as indicated by overlap in the standard error 

of the slopes for live and control samples for 

days 53–145. The live samples contained 26 percent 

less CHCl3 than did the autoclaved sterile control 

samples at 53 days and 28 percent less at 145 days of 

incubation (fig. 21).

Bacterial degradation of CHBr3 was observed in 

one of the two live anaerobic enriched sediment 

microcosms. In the microcosm that showed 

biodegradation, live B, the mass of CHBr3 was 

undetectable after 14 days of incubation (fig. 22). 

Additions of CHBr3 (respikes) at 14.4, 43.5, and 56.5 

days of incubation were also consumed by the bacteria 

in the sample. However, the mass of CHBr3 in the 

replicate live sample, live A, was not significantly 

different from the mass of CHBr3 in the sterile control 

sample during the entire incubation period. 

Inconsistent behavior in replicate vials has been 

observed in other experiments, and probably occurs 

because not all vials contain an adequate number of the 

bacteria that degrade the halogenated organic 

compounds to establish a viable culture (Goodwin 

1996).
IV. The Potential for Biodegradation of Trihalomethanes by Aquifer Bacteria 57

Figure 19. Mass of chloroform (CHCl3) (A) and bromoform (CHBr3) (B) in sediment microcosms containing sediment and ground water from Lancaster, 

Antelope Valley, California.

Data points represent the mean of two to four replicate analyses and the error bars represent plus or minus one standard deviation about the mean. Data 

are from Fram and others (2002). CHCl3, I

N MICROGRAMS

Sterile Control

Live

INCUBATION TIME, IN DAYS

0 25 50 75 100 125 150

CHBr3, I

N MICROGRAMS

Sterile Control

Live

0

0.5

1.0

1.5

2.0

0

2

4

6

8

A

B
58 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

Figure 20. Mass of chloroform (CHCl3) (A) and bromoform (CHBr3) (B) in aerobic, enriched sediment microcosms containing sediment and ground water 

from Lancaster, Antelope Valley, California. 

Data points represent the mean of two to four replicate analyses and the error bars represent plus or minus one standard deviation about the mean. Data 

are from Fram and others (2002). CHCl3, I

N MICROGRAMS

Sterile Control

Live

INCUBATION TIME, IN DAYS

0 25 50 75 100 125 150

CHBr3, I

N MICROGRAMS

Sterile Control

Live

A

B

4

6

8

10

12

14

10

15

20

25

30

35
IV. The Potential for Biodegradation of Trihalomethanes by Aquifer Bacteria 59

Figure 21. mass of chloroform (CHCl3) in anaerobic, enriched sediment microcosms containing sediment and ground water from Lancaster, Antelope 

Valley, California. 

Data points represent the mean of two to four replicate analyses and the error bars represent plus or minus one standard deviation about the mean. Data 

are from Fram and others (2002).

INCUBATION TIME, IN DAYS

0 25 50 75 100 125 150

CHCl3, I

N MICROGRAMS

0

2

4

6

Sterile Control

Live

Figure 22. Mass of bromoform (CHBr3) in anaerobic, enriched sediment microcosms containing sediment and ground water from Lancaster, Antelope 

Valley, California. 

Data points represent the mean of two to four replicate analyses and the error bars represent plus or minus one standard deviation about the mean. Data 

are from Fram and others (2002).

INCUBATION TIME, IN DAYS

0 20 40 60 80

CHBr3, I

N MICROGRAMS

0

20

40

60

80

Sterile Control

Live A

Live B
60 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

Water Enrichment Microcosm Experiments

The water enrichment microcosm experiments 

consisted of ground water or extraction water amended 

with KH2PO4, NH4Cl, and vitamins, plus the bacteria 

and particles concentrated from a larger volume of the 

same type of water. No sediment was added to these 

microcosms. Centrifugation or filtration was used to 

concentrate bacteria and particles for use in these 

microcosms. Extraction water was centrifuged at 

14,000 rotations per minute and then the lower third of 

the water was used for the microcosms. Bacteria and 

particles were also concentrated from ground water or 

extraction water by filtering 2 liters of water through a 

0.2-micrometer pore-size filter. The filters were then 

placed in serum vials containing 30 mL of ground 

water or extraction water. The vials were spiked with 

known masses of CHCl3 and CHBr3 (equivalent to 

approximately 195 μg/L and 380 μg/L, respectively, in 

the water), and incubated for up to 83 days. One set of 

sterile control microcosms consisted of autoclaved, 

enriched ground water or extraction water, and the 

second set consisted of extraction water sterilized by 

filtration through a sterile, 0.2-μm pore-size filter, plus 

or minus a sterile filter. All of the water enrichment 

microcosms were incubated under aerobic conditions.

No significant bacterial degradation of CHCl3

was observed in any of the water enrichment 

microcosms after 83 days of incubation, as indicated 

by the similar behavior of the autoclaved sterile control 

and the live samples (fig. 23A). Bacterial degradation 

of CHBr3 also was not observed in microcosms 

containing bacteria concentrated by centrifugation 

(fig. 23B). However, the mass of CHBr3 did decrease in 

microcosms containing bacteria concentrated by 

filtration. The slopes of the mass versus time data for 

both live samples were significantly different from the 

slope for the autoclaved sterile control sample (the 

Student t test, α = 0.05). The amount of CHBr3 loss in 

the two microcosms was remarkably consistent, even 

though they contained two different kinds of water. The 

average percentage loss in bottles concentrated by 

filtration relative to controls was 37 percent at 13 days, 

54 percent at 34 days, and 60 percent at 83 days. 

Although CHCl3 was also present in these vials, there 

was no analogous CHCl3 loss (fig. 23A). 

The second set of sterile control microcosms was 

used to determine if the loss of CHBr3 was an artifact 

caused by adsorption of CHBr3 onto the filter paper. 

Microcosms with and without filters behaved similarly, 

showing no significant difference in loss of either 

CHCl3 or CHBr3 after 30 days of incubation (fig. 24). 

This result suggests that the pattern of CHBr3 loss 

shown in figure 24B was not an artifact caused by 

CHBr3 adsorption, and thus, the loss may have been 

due to biodegradation. Landmeyer and others (2000) 

and Thomas and others (2000) observed 

biodegradation of CHBr3 in the Las Vegas, Nevada 

aquifer and suggested that it may have occurred in 

anaerobic microzones within the otherwise aerobic 

aquifer. The vials were not expected to contain such 

microzones, although the existence of such microzones 

in the pores of the filters could not be ruled out. 

Cometabolism of CHBr3 could not have occurred in 

the microcosms because no cometabolic substrates 

were added. However, aerobic biodegradation of 

CHBr3 has not been reported previously; therefore, 

these results need to be reproduced before confidence 

can be placed in this interesting outcome.

Bacterial Density

Bacterial density was monitored in water 

samples collected from the injection and extraction 

well (well 4-32) and from three of the nested 

piezometers (27P6–8) (fig. 2). Water samples were 

collected in sterile, 2-mL cryotubes, preserved with 

gluteraldehyde, and stored at −70°C until analysis. Cell 

numbers were determined by acridine orange direct 

count (AODC) (Hobbie and others, 1977), and sterile 

sodium citrate was added during filtration to remove 

background fluorescence (Harvey, 1987).

Ground water collected from well 4-32 on March 

6, 1998, (before the beginning of the third cycle 

injection period) contained a lower concentration of 

bacteria than did extraction water collected in August, 

September, and October 1998, although the cell counts 

were within 1 standard deviation of each other (fig. 25). 

Injection water sampled at well 4-32 in June contained 

almost no bacteria; in fact, the concentration was not 

significantly different from zero (fig. 25). This result 

was expected because the injection water was 

chlorinated.
IV. The Potential for Biodegradation of Trihalomethanes by Aquifer Bacteria 61

Figure 23. Mass of chloroform (CHCl3) (A) and bromoform (CHBr3) (B) in water enrichment microcosms stored for an incubation period and containing 

extraction water or ground water from Lancaster, Antelope Valley, California. 

Data points represent the mean of two to four replicate analyses and the error bars represent plus or minus one standard deviation about the mean. Data 

are from Fram and others (2002). CHCl3, I

N MICROGRAMS

3

4

5

6

7

INCUBATION TIME, IN DAYS

0 20 40 60 80 100

CHBr3, I

N MICROGRAMS

0

5

10

15

20

EXPLANATION

Sterile Control

Extraction water and centrifuged bacteria

Extraction water and filtered bacteria

Ground water and filtered bacteria

A

B
62 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

Figure 24. Mass of chloroform (CHCl3) (A) and bromoform (CHBr3) (B) in sterile control samples for water enrichment microcosms stored for an incubation 

period and containing sterile filters and extraction water from Lancaster, Antelope Valley, California. 

Data points represent the mean of two to four replicate analyses and the error bars represent plus or minus one standard deviation about the mean. Data 

are from Fram and others (2002). CHCl3, I

N MICROGRAMS

5

6

7

8

9

No filter

Filter

INCUBATION TIME, IN DAYS

0 5 10 15 20 25 30

CHBr3, I

N MICROGRAMS

8

10

12

14

16

No filter

Filter

A

B
IV. The Potential for Biodegradation of Trihalomethanes by Aquifer Bacteria 63

Figure 25. Bacterial densities in water samples collected from well 7N/12W-27P2 (well 4-32) during the third injection, storage, and recovery cycle 

(March 1998 through April 1999), Lancaster, Antelope Valley, California. 

Data points represent the mean of three replicate samples and the error bars represent plus or minus one standard deviation about the mean. Data are from 

Fram and others (2002).

SAMPLING DATE, IN 1998

FM AM J J A S O

BACTERIAL DENSITY, IN CELLS PER MILLILITER

Ground water

Injection water

Extraction water

0

50,000

100,000

150,000

200,000

250,000

300,000
64 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

Bacterial density in samples collected from 

piezometers at different depths (fig. 2) did not differ 

significantly except in October 1998 (fig. 26) when 

bacterial density was significantly higher in samples 

from the shallower piezometers, 27P7 and 27P8, than 

in samples from the deepest piezometer, 27P6. 

Bacterial densities were higher in October than in other 

months, and the bacteria were noticeably larger, as 

determined by visual inspection. In October, samples 

also were collected from wells 4-13, 4-33, and 4-42, 

which were outside the plume of injection water (see 

Metzger and others, 2002). Bacterial density in 

samples from the nested piezometers and well 4-32 

were significantly higher than in samples from 

wells 4-13, 4-33, and 4-42 (fig. 27), indicating that 

concentration of bacteria in the aquifer increased in 

response to the injection, storage, and recovery process 

at well 4-32. An increase in the concentration of 

bacteria in the aquifer near injection wells has been 

observed by other researchers; indeed, such an increase 

with time was cited as evidence for the development of 

CHBr3 biodegradation in the Las Vegas, Nevada 

injection, storage, and recovery project (Thomas and 

others, 2000).

Figure 26. Bacterial densities in water samples collected from nested piezometers 7N/12W-27P6, 27P7, and 27P8 during the third injection, storage, and 

recovery cycle (March 1998 through April 1999), Lancaster, Antelope Valley, California. 

Data points represent the mean of three replicate samples and the error bars represent plus or minus one standard deviation about the mean. Data are from 

Fram and others (2002).

SAMPLING DATE, IN 1998

AUG 4 SEP 3 OCT 7 NOV 4

BACTERIAL DENSITY, IN CELLS PER MILLILITER

7N/12W-27P8

7N/12W-27P7

7N/12W-27P6

0

100,000

200,000

300,000

400,000

500,000
IV. The Potential for Biodegradation of Trihalomethanes by Aquifer Bacteria 65

Conclusions

The experimental results support the following 

conclusions: 

1. No significant bacterial degradation of 

chloroform (CHCl3) or bromoform (CHBr3) was 

observed in aerobic sediment microcosms or in aerobic 

water enrichment microcosms. In anaerobic 

experiments, there was some initial loss of CHCl3 in 

live samples compared to sterile control samples, but 

that degradation was not sustained between days 53 

and 145. However, biodegradation of CHBr3 was 

observed in an anaerobic sediment microcosm. This 

result suggests that although the Lancaster aquifer is 

aerobic, bacteria capable of degrading CHBr3 under 

anaerobic conditions are present. If anaerobic 

microzones developed in the aquifer, CHBr3 could 

biodegrade. The development of such microzones and 

acclimation of the bacterial community probably 

accounts for the CHBr3 biodegradation seen in the 

Las Vegas, Nevada injection, storage, and recovery 

project; a similar situation could develop in the 

Lancaster aquifer. However, the water in fully 

anaerobic aquifers can acquire unpleasant odors and 

tastes and would represent a shift from the aquifer’s 

natural redox state; therefore, the level of oxygenation 

of the aquifer should be closely monitored in the 

extraction water with time. Because the dominant 

THM in the extraction water during the third cycle was 

CHCl3, biodegradation of CHBr3 would not have 

reduced THM concentrations significantly. It is 

possible that acclimation of the bacterial community 

could result in CHBr3 biodegradation in subsequent 

injection, storage, and recovery cycles, as has been 

Figure 27. Bacterial densities in water samples collected from wells 7N/12W-27P2 (well 4-32), 7N/12W-27J4 (well 4-13), 7N/12W-27H3 (well 4-33), and 

7N/12W-27J6 (well 4-42), and the nested piezometers 7N/12W-27P6, 27P7, and 27P8, during the third injection, storage, and recovery cycle (March 1998 

through April 1999), Lancaster, Antelope Valley, California. 

Data points represent the mean of three replicate samples and the error bars represent plus or minus one standard deviation about the mean. Data are from 

Fram and others (2002).

PIEZOMETER SAMPLES

27P8 27P7 27P6 4-32 4-13 4-33 4-42

BACTERIAL DENSITY, IN CELLS PER MILLILITER

WELL SAMPLES

October 7, 1998

0

100,000

200,000

300,000

400,000

500,000
66 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

observed in the Las Vegas project (Thomas and others, 

2002). However, it should be noted that CHCl3

biodegradation was not observed in the field nor in 

microcosms constructed using the Las Vegas aquifer 

(Landmeyer and others, 2000) or the Lancaster aquifer 

(this study) sediments. Therefore, even if CHBr3

biodegradation were to occur, CHCl3 is expected to 

persist. 

2. No biodegradation of CHCl3 was observed in 

the water enrichment microcosms containing bacteria 

concentrated by centrifugation or filtration, but a 

significant decline in CHBr3 was observed in 

microcosms containing bacteria concentrated by 

filtration, although the loss was not complete. 

However, the pattern of CHBr3 loss was remarkably 

similar for both waters, which raised suspicion that the 

result was due to some type of experimental artifact. 

Loss from leaks was ruled out because both CHCl3 and 

CHBr3 were present in the vials, and leaks would have 

affected both compounds. No specific adsorption of 

CHBr3 to the filters was observed in sterile control 

experiments. These results indicate that the loss of 

CHBr3 might have been due to some aerobic, 

biological mechanism, but this conclusion must viewed 

with caution until it can be reproduced.

3. Water samples from well 4-32 and nested 

piezometers 27P6–8 contained significantly higher 

concentrations of bacteria than samples from wells 

outside the influence of the injection well, indicating 

that the injection, storage, and recovery process 

increased bacterial numbers near the injection site. 

V. SUMMARY AND CONCLUSIONS

By Miranda S. Fram, Roger Fujii, Brian A. 

Bergamaschi, and Kelly D. Goodwin

This study addresses the formation and fate of 

trihalomethanes (THMs) in the aquifer system used for 

the injection, storage, and recovery demonstration 

project in Lancaster, Antelope Valley, California. 

Injection, storage, and recovery is a water-management 

method that involves injection of surface water into an 

aquifer system for storage during periods when surface 

water is plentiful and recovery of the stored water by 

extraction during periods when surface water is scarce. 

Surface water from the State Water Project’s California 

Aqueduct was treated before injection into the 

Lancaster aquifer. Disinfection (chlorination in this 

case) is necessary to prevent biofouling and 

introducing microbial contaminants into the aquifer. 

Natural organic matter, mostly dissolved organic 

carbon, present in surface water reacts with chlorine 

during treatment to form carcinogenic disinfection byproducts such as THMs. The concentration of THMs in 

drinking water is regulated by the U.S. Environmental 

Protection Agency. Therefore, when treated surface 

water is injected into an aquifer system, the possible 

persistence of THMs in the aquifer poses a potential 

problem for the long-term feasibility of injection, 

storage, and recovery because of the need to assure that 

ground-water quality is not significantly degraded. 

Thus, this study was designed to address the following 

questions:

 1. What controls the continued formation of 

THMs in the aquifer after injection?

2. What causes the continued presence of low 

concentrations of THMs in the extracted water after all 

the injection water presumably has been retrieved?

3. What causes the decrease in THM 

concentrations as extraction proceeds?

4. Are there natural attenuation mechanisms that 

can decrease the THM concentrations in the aquifer?

Our monitoring and experimental results lead us 

to conclude that the major factor controlling the 

continued formation of THMs in the aquifer after 

injection (question 1) was the concentration of residual 

chlorine in injected waters. The injection water 

contained averages of 27.5 μg/L THM and 0.79 mg/L 

residual free chlorine at the time of injection. An 

experiment consisting of storing samples of injection 

water for up to 16 weeks prior to analysis showed that 

THM formation caused by reaction between the 

dissolved organic carbon and residual chlorine in the 

injection water stopped after 4 weeks, when the 

residual chlorine was essentially consumed. The mean 

total concentration of THMs formed from the injection 

water after injection and storage in the aquifer was 

estimated as 58 μg/L, based on samples of extraction 

water and samples from the nested piezometers that 

were judged to represent pure injected water with no 

admixed ground water. Adding dechlorination 

treatment of the injected water immediately before 

injection would reduce the total amount of THMs 

formed from the injected water. The concentration of 

THMs formed in the injected water after injection and 

storage in the aquifer was an average of 89 μg/L less 

than the THM concentrations determined by THM 
V. Summary and Conclusions 67

formation potential experiments, indicating that rechlorinating the water after extraction would result in 

formation of additional THMs. 

Evidence clearly shows that dilution due to 

mixing of injected water with ground water was the 

major process affecting the concentration of THMs in 

water extracted during the third cycle. Mixing of 

injected water with ground water can explain the 

continued low levels of THMs in the extracted water 

after all of the injection water had presumably been 

extracted (question 2) and the decrease in THM 

concentrations as extraction proceeded (question 3). 

Mass balance calculations, results of the sulfur 

hexafluoride (SF6) tracer study, and inferences from a 

mathematical mixing model for the third cycle 

indicated that mixing of the injected water with the 

ground water occurred in the aquifer. Mass balance 

calculations showed that only about 67 percent of the 

chloride and THMs injected into the aquifer system 

were recovered by the time that 132 percent of the 

injected water volume had been extracted. Continued 

extraction to 250 percent of the injected water volume 

only increased the recovery to 80 percent of the 

injected THMs. The similar recoveries for chloride and 

THMs indicate that THMs behaved conservatively in 

the aquifer (after accounting for the THMs formed 

from the residual free chlorine in the injected water).

A conservative tracer, SF6, was added to the 

injection water throughout the injection period and 

then its concentration in the extraction water was 

monitored. SF6 concentrations were used to calculate 

the relative proportions of injected water and ground 

water in extraction water samples, and then these 

proportions were used to predict THM concentrations 

in the extraction water. Close agreement between 

predicted and measured THM concentrations in the 

extraction water indicated that THMs behaved 

conservatively (after accounting for the THMs formed 

from the residual free chlorine in the injected water) 

and that attenuation of THMs and SF6 was due to 

dilution of injected water with ground water that 

contained no THMs or SF6. Calculations using 

chloride as a tracer yielded similar results. 

A descriptive mixing model representing a 

simple scenario of mixing between injected water and 

ground water during the extraction process was derived 

and used to predict concentrations of conservative 

constituents in the extracted water. The concentrations 

of THMs, SF6, and chloride in the extracted water 

predicted by the model closely matched the measured 

concentrations. The low concentrations of THMs in 

water extracted at the end of the third cycle represent 

mixtures consisting of a small proportion of injected 

water and a large proportion of ground water. Thus, the 

results suggest that mixing of injected water and 

ground water can adequately explain extraction water 

composition without further consideration of other 

mechanisms, such as microbial degradation or 

sorption.

The descriptive mixing model was used to 

forecast the accumulation of conservative constituents 

present in the injection water after 10 annual cycles of 

injection, storage, and recovery. For realistic ratios of 

the volumes of injected to extracted water, the model 

predicts that the concentrations of injected constituents 

in the water remaining in the aquifer at the injection 

site after the 10th cycle would approach 100 percent of 

injection water concentrations. Removal of this 

residual concentration at a later time, using existing 

extraction methods, may be difficult because the ratio 

of injected water to ground water in the extracted water 

declines asymptotically owing to mixing. Thus, 

removing the residual concentration would require 

pumping several times the injected water volume. 

Reducing the concentration of a constituent in the 

water near the injection/extraction well to 10 percent of 

its concentration in the injected water would require 

removal of 2.3 times the volume of water injected, 

reducing it to 1 percent would require removing 

4.6 volumes, and reducing it to 0.1 percent would 

require removing 9 volumes. The results of this study 

indicate that the concentrations of THMs (or any 

conservative constituent present in the injection water 

at a concentration exceeding that in the native ground 

water) will increase in the aquifer surrounding the 

injection site.

Our results indicate that the natural attenuation 

mechanisms, biodegradation and sorption, did not 

significantly decrease the concentration of THMs in the 

aquifer (question 4). The potential for biodegradation 

of THMs by aquifer bacteria was assessed using two 

types of experiments: sediment microcosms prepared 

from aquifer sediment and ground water incubated 

with and without amendation with nutrients, incubated 

under aerobic and anaerobic conditions, and water 

enrichment microcosms prepared from ground water or 

extraction water amended with nutrients and with 

bacteria and particles concentrated from a larger 

volume of water and incubated under aerobic 

conditions. Results from these experiments showed no 
68 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

bacterial degradation of chloroform (CHCl3) or 

bromoform (CHBr3) under aerobic conditions, such as 

those in the aquifer in this study. Bacterial degradation 

of CHBr3 under anaerobic conditions was observed. 

However, because the Lancaster aquifer is aerobic and 

because CHBr3 comprises only a small proportion of 

the THMs, biodegradation is not considered an 

important attenuation mechanism for THMs in this 

aquifer.

Measurements of bacterial densities in water 

samples collected during the cycle showed that water 

samples collected from the injection well and the 

nearby nested piezometers during the extraction period 

contained significantly more bacteria than water 

samples collected from wells farther away. This result 

suggests that the injection process may increase 

bacterial population in the aquifer near the injection 

site.

Comparison of the (Br/Cl)THM values in the 

injection water from the storage experiments with the 

values in the extraction water sampled at the beginning 

of the extraction period suggests that sorption was not 

an important process for controlling THM 

concentrations in this aquifer. No evidence was 

observed for the preferential removal of brominated 

THM species as would have been expected had 

sorption occurred. In addition, the close 

correspondence of the change in THM concentrations 

during the extraction period to the changes in the 

concentrations of SF6 and chloride (species known not 

to be sorbed to the aquifer material) also suggests that 

sorption of THMs to aquifer materials was not an 

important process.

Finally, our results provided some insights into 

the ground-water flow system during the injection, 

storage, and recovery cycle. The data and modeling 

results lead us to conclude that the aquifer-system 

response to injection, storage, and recovery cycles was 

very complex. Extensive mixing between injected 

water and ground water occurred, and the water flow 

paths during injection and extraction seemed to be 

different. We proposed a conceptual model for water 

flow in the aquifer during the third injection, storage, 

and recovery cycle at the Lancaster site. Although the 

injection/extraction well was continuously screened 

through the upper and middle aquifers, injected water 

preferentially flowed into high hydraulic conductivity 

zones in the upper aquifer. Mechanical dispersion 

resulted in some mixing between injected water and 

native ground water during injection. During 

extraction, water-level drawdown around the 

injection/extraction well resulted in inefficient 

transport of water from the upper portions of the 

aquifer to the well, thus increasing the proportion of 

water extracted from deeper portions of the aquifer. As 

a result, injected water was stranded in upper portions 

of the aquifer. The water extracted from the 

injection/extraction well consisted of mixtures of 

injected water and native ground water. This hydraulic 

regime resulted in incomplete recovery of the injected 

water.

REFERENCES CITED

Alexander, M., 1994, Biodegradation and bioremediation: 

San Diego, Calif., Academic Press, 302 p.

Amy, G.L., Chadik, P.A., Chowdhury, Z.K., King, P.H., and 

Cooper, W.J., 1985, Factors affecting incorporation of 

bromide into brominated trihalomethanes during 

chlorination, in Jolley, R.L., Bull, R.J., Davis, W.P., 

Katz, S., Roberts, M.H. Jr., Jacobs, V.A., eds., Water 

chlorination: chemistry, environmental impact and 

health effects, v. 5: Boca Raton, Lewis Publishers, Inc., 

p. 907–922.

Babcock, D.B., and Singer, P.C., 1979, Chlorination and 

coagulation of humic and fulvic acids: Journal of the 

American Water Works Association, v. 71, no. 3, 

p. 149–152.

Bagley, D.M., and Gossett, J.M., 1995, Chloroform 

degradation in methanogenic methanol enrichment 

cultures and by Methanosarcina barkeri 227: Applied 

and Environmental Microbiology, v. 61, no. 9, 

p. 3195–3201. 

Barber, L.B., II, Brown, G.K., Kennedy, K.R., Leenheer, 

J.A., Noyes, T.I., Rostad, C.E., and Thorn, K.A., 1997, 

Organic constituents that persist during aquifer storage 

and recovery of reclaimed water in Los Angeles 

County, California, in Kendall, D.R., ed., Proceedings 

of AWRA Symposium, Conjunctive Use of Water 

Resources, Aquifer Storage and Recovery: Herndon, 

Virginia, American Water Resources Association, 

TPS-97-2, p. 261–272.

Bartnicki, E.W., and Castro, C.E., 1994, Biodehalogenation: 

rapid oxidative metabolism of mono- and 

polyhalomethanes by Methylosinus trichosporium OB3b: Environmental Toxicological Chemistry, v. 13, 

no. 2, p. 241–245. 

Bear, Jacob, 1979, Hydraulics of groundwater: New York, 

McGraw-Hill, 569 p.

Berner, R.A., 1980, Early diagenesis, a theoretical approach: 

Princeton, N.J., Princeton University Press, 237 p.
References Cited 69

Bernholtz, A., Brothers, K., and Katzer, T., 1995, Artificial 

ground-water recharge in Las Vegas Valley, Clark 

County, Nevada: Storing today, treating tomorrow?, in 

Johnson, A.I., and Pyne, R.D.G., eds., Artificial 

recharge of ground water, II, Proceedings of the Second 

International Symposium on Artificial Recharge of 

Ground Water: New York, American Society of Civil 

Engineers, p. 548–557.

Bloyd, R.M., Jr., 1967, Water resources of the Antelope 

Valley–East Kern Water Agency area, California: U.S. 

Geological Survey Open-File Report [67-20], 73 p.

Bouwer, E.J., McCarty, P.L., Bouwer, H., and Rice, R.C., 

1984, Organic contaminant behavior during rapid 

infiltration of secondary wastewater at the Phoenix 23rd 

Avenue Project: Water Research, v. 18, no. 4, 

p. 463–472.

Bouwer, E.J., Rittmann, B.E., and McCarty, P.L., 1981, 

Anaerobic degradation of halogenated 1- and 2-carbon 

organic compounds: Environmental Science and 

Technology, v. 15, no. 5, p. 596–599.

Brothers, K., and Katzer, T., 1990, Water banking through 

artificial recharge, Las Vegas Valley, Clark County, 

Nevada: Journal of Hydrology, v. 114, p. 77–103.

Campbell, B.G., Conlon, K.J., Mirecki, J.E., and Petkewich, 

M.D., 1997, Evaluation of aquifer storage recovery in 

the Santee Limestone/Black Mingo Aquifer near 

Charleston, South Carolina, 1993–95: U.S. Geological 

Survey Water-Resources Investigations Report 

96-4283, 89 p.

Carlson, C.S., Leighton, D.A., Phillips, S.P., and Metzger, 

L.F., 1998, Regional water table (1996) and water-table 

changes in the Antelope Valley ground-water basin, 

California: U.S. Geological Survey Water-Resources 

Investigations Report 98-4022, 2 sheets.

Castro, C.E., 1993, Biodehalogenation: The kinetics and 

rates of the microbial cleavage of carbon-halogen 

bonds: Environmental Toxicology and Chemistry, v. 12, 

no. 9, p. 1609–1618.

Chin, Y-P., Aiken, G., and O’Loughlin, E., 1994, Molecular 

weight, polydispersity, and spectroscopic properties of 

aquatic humic substances: Environmental Science and 

Technology, v. 28, no. 11, p. 1853–1858.

Chiou, C.T., Peters, L.J., and Freed, V.H., 1979, A physical 

concept for soil-water equilibria for nonionic organic 

compounds: Science, v. 206, no. 4420, p. 831–832.

Chiou, C.T., Porter, P.E., and Schmedding, D.W., 1983, 

Partition equilibria of nonionic organic compounds 

between soil organic matter and water: Environmental 

Science and Technology, v. 17, no. 4, p. 227–231.

Clark, J.F., Wannikhof, R., Schlosser, P., and Simpson, H.J., 

1994, Gas exchange in the tidal Hudson River using a 

dual tracer technique: Tellus, v. B46, p. 274–285.

Connell Hancock, T.L., Costello, A.M., Lidstrom, M.E., and 

Oremland, R.S., 1998, Strain ImB-1, a novel bacterium 

for the dissipation of methyl bromide in fumigated 

agricultural soils: Applied and Environmental 

Microbiology, v. 64, no. 8, p. 2899–2905. 

Curtis, G.P., Roberts, P.V., and Reinhard, M., 1986, A natural 

gradient experiment of solute transport in a sand aquifer 

4. Sorption of organic solutes and its influence on 

mobility: Water Resources Research, v. 22, no. 13, 

p. 2059–2067.

Dibblee, T.W., Jr., 1967, Areal geology of the western 

Mojave Desert, California: U.S. Geological Survey 

Professional Paper 522, 153 p., 4 pls. in pocket.

Doronina, N.V., and Trotsenko, Y.A., 1994, Methylophilus 

leisingerii sp. nov., a new species of restricted 

facultatively methylotrophic bacteria: Microbiology, 

v. 63, p. 298–302.

Dutcher, L.C., and Worts, G.F., 1963, Geology, hydrology, 

and water supply of Edwards Air Force Base, Kern 

County, California: U.S. Geological Survey Open-File 

Report [63-05], 225 p.

Fram, M.S., Berghouse, J.K., Bergamaschi, B.A., Fujii, R., 

and Clark, J.F., 2002, Water quality monitoring and 

studies of the formation and fate of trihalomethanes 

during the third injection, storage, and recovery test at 

Lancaster, Antelope Valley, California, March 1998 

through April 1999: U.S. Geological Survey Open-File 

Report 02-102, 48 p.

Fujii, R., Ranalli, A.J., Aiken, G.R., and Bergamaschi, B.A., 

1998, Dissolved organic carbon concentrations and 

compositions, and trihalomethane formation potentials 

in waters from agricultural peat soils, Sacramento–San 

Joaquin Delta, California: Implications for drinkingwater quality: U.S. Geological Survey Water Resources 

Investigations Report 98-4147, 75 p.

Gamlin, J.D., Clark, J.F., Woodside, G., and Herndon, R., 

2001, Large-scale tracing of ground water with sulfur 

hexafluoride: Journal of Environmental Engineering, 

v. 127, no. 2, p. 171–174.

Goodwin, K.D., 1996, Natural cycles of brominated 

methanes: Macroalgal production and marine microbial 

degradation of bromoform and dibromomethane: Ph.D. 

thesis, California Institute of Technology.

Goodwin, K.D., Schaefer, J.K., and Oremland, R.S., 1998, 

Bacterial oxidation of dibromomethane and methyl 

bromide in natural waters and enrichment cultures: 

Applied and Environmental Microbiology, v. 64, no. 12, 

p. 4629–4636.

Goodwin, K.D., Varner, R.K., Crill, P.M., and Oremland, 

R.S., 2001, Consumption of tropospheric levels of 

methyl bromide by C1 bacteria and comparison to 

saturation kinetics: Applied and Environmental 

Microbiology, v. 67, no. 12, p. 5437–5443. 
70 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

Grathwohl, P., 1990, Influence of organic matter from soils 

and sediments from various origins on the sorption of 

some chlorinated aliphatic hydrocarbons: Implications 

of KOC correlations: Environmental Science and 

Technology, v. 24, no. 11, p. 1687–1693.

Gschwend, P.M., and Wu, S., 1985, On the constancy of 

sediment-water partition coefficients of hydrophobic 

organic pollutants: Environmental Science and 

Technology, v. 19, no. 1, p. 90–96.

Hamamura, N., Page, C., Long, T., Semprini, L., and Arp, 

D.J., 1997, Chloroform cometabolism by butane-grown 

CF8, Pseudomonas butanovora, and mycobacterium 

vaccae JOB5 and methane-grown methylosinus 

trichosporium OB3b: Applied and Environmental 

Microbiology, v. 63, no. 9, p. 3607–3613. 

Happell, J.D., and Wallace, D.W.R., 1998, Removal of 

atmospheric CCl4 under bulk aerobic conditions in 

groundwater and soils: Environmental Science and 

Technology, v. 32, no. 9, p. 1244–1252, 1998.

Harvey, R.W., 1987, A fluorochrome-staining technique for 

counting bacteria in saline, organically enriched, 

alkaline lakes: Limnology and Oceanography, v. 32, 

p. 993–995.

Hathaway, G.J., Proctor, N.H., Hughes, J.P., and Fischman, 

M.L., eds., 1991, Proctor and Hughes' chemical hazards 

of the workplace: New York, Van Nostrand Reinhold, 

666 p.

Helsel, D.R., and Hirsch, R.M., 1995, Statistical methods in 

water resources: Amsterdam, The Netherlands, Elsevier 

Science Publishers, Studies in Environmental Science 

Series, v. 49, 529 p., software in pocket (3 1/2 inch 

diskette).

Henson, J.M., Yates, M.V., Cochran, J.W., and Shackleford, 

D.L., 1988, Microbial removal of halogenated 

methanes, ethanes, and ethylenes in an aerobic soil 

exposed to methane: FEMS Microbial Ecology, v. 53, 

p. 193–201. 

Hines, M.E., Crill, P.M., Varner, R.K., Talbot, R.W., Shorter, 

J.H., Kolb, C.E., and Harris, R.C., 1998, Rapid 

consumption of low concentrations of methyl bromide 

by soil bacteria: Applied and Environmental 

Microbiology, v. 64, no. 5, p. 1864–1870. 

Hobbie, J.E., Daley, R.J., and Jasper, S., 1977, Use of 

nucleopore filters for counting bacteria by fluorescence 

microscopy: Applied and Environmental Microbiology, 

v. 33, p. 1225–1228.

Howle, J.F., Phillips, S.P., Denlinger, R.P., and Metzger, L.F., 

in press, Determination of specific yield and changes in 

water-table surface using temporal microgravity 

surveys during freshwater injection tests, Lancaster, 

Antelope Valley, California, 1996–97: U.S. Geological 

Survey Water-Resources Investigations Report 03-

4019.

Ikehara, M.E., and Phillips, S.P., 1994, Determination of 

land subsidence related to ground-water-level declines 

using global positioning system and leveling surveys in 

Antelope Valley, Los Angeles and Kern Counties, 

California, 1992: U.S. Geological Survey WaterResources Investigations Report 94-4184, 101 p.

Karickhoff, S.W., Brown, D.S., and Scott, T.A., 1979, 

Sorption of hydrophobic pollutants on natural 

sediments: Water Research, v. 13, no. 3, p. 241–248.

King, D.B., and Saltzman, E.S., 1997, Removal of methyl 

bromide in coastal seawater: Chemical and biological 

rates: Journal of Geophysical Research, v. 102, no. C8, 

p. 18715–18721. 

Kohler-Staub, D., Frank, S., and Leisinger, T., 1995, 

Dichloromethane as the sole carbon source for 

Hyphomicrobium sp. Strain DM2 under denitrification 

conditions: Biodegradation, v. 6, p. 229–235. 

Landmeyer, J.E., Bradley, P.M., and Thomas, J.M., 2000, 

Biodegradation of disinfection by-products as a 

removal process during aquifer storage extraction: 

Journal of the American Water Resources Association, 

v. 36, no. 4, p. 861–868.

Larson, R.A., and Weber, E.J., 1994, Reaction mechanisms 

in environmental organic chemistry: Boca Raton, Lewis 

Publishers, 433 p.

Leighton, D.A., and Phillips, S.P., 2003, Simulation of 

ground-water flow and land subsidence in the Antelope 

Valley ground-water basin, California: U.S. Geological 

Survey Water-Resources Investigations Report 

03-4016., available only on the Internet at URL 

http://water.usgs.gov/pubs/wri/wrir034016/text.htm.

Leisinger, T., and Bader, R., 1993, Microbial dehalogenation 

of synthetic organohalogen compounds, hydrolytic 

dehalogenases: Chimia, v. 47, p. 116–121. 

Londquist, C.J., Rewis, D.L., Galloway, D.L., and 

McCaffrey, W.F., 1993, Hydrogeology and land 

subsidence, Edwards Air Force Base, Antelope Valley, 

California, January 1989–December 1991: U.S. 

Geological Survey Water-Resources Investigations 

Report 93-4114, 74 p.

Los Angeles County Department of Public Works, 2000, 

Lancaster subbasin aquifer storage and recovery (ASR) 

demonstration project final report: Los Angeles County 

Department of Public Works, Waterworks and Sewer 

Maintenance Division, Waterworks District 40, 

Antelope Valley, Water Quality and Engineering 

Section, Water Resources Unit, 40 p. plus appendices.

Mabey, D.R., 1960, Gravity survey of the western Mojave 

Desert, California: U.S. Geological Survey Professional 

Paper 316-D, p. 51–73.

MacIntyre, W.G., and Stauffer, T.B., 1988, Liquid 

chromatography applications to determination of 

sorption on aquifer materials: Chemosphere, v. 17, 

no. 11, p. 2161–2173.
References Cited 71

Mackay, D., Shiu, W.Y., and Ma, K.C., 1993, Illustrated 

handbook of physical-chemical properties and 

environmental fate for organic chemicals: v. III, Volatile 

organic chemicals: Boca Raton, Lewis Publishers, 

916 p.

Manley, S.L., Goodwin, K., and North, W.J., 1992, 

Laboratory production of bromoform, methylene 

bromide and methyl iodide by macroalgae and 

distribution in nearshore southern California waters: 

Limnology and Oceanography, v. 37, no. 8, p. 1652-

1659.

McCarty, P.L., Rittmann, B.E., and Bouwer, E.J., 1984, 

Microbiological processes affecting chemical 

transformations in groundwater, in Bitton, G., and 

Gerba, C.P., eds., Groundwater Pollution Microbiology, 

New York, John Wiley & Sons, p. 90–115.

McClay, K., Fox, B.G., and Steffan, R.J., 1996, Chloroform 

mineralization by toluene-oxidizing bacteria: Applied 

and Environmental Microbiology, v. 62, no. 8,

p. 2716–2722.

Metzger, L.F., Ikehara, M.E., and Howle, J.F., 2002, Verticaldeformation, water-level, microgravity, geodetic, waterchemistry, and flow-rate data collected during injection, 

storage, and recovery tests at Lancaster, Antelope 

Valley, California, September 1995 through September 

1998: U.S. Geological Survey Open-File Report 

01-414, 149 p.

Miller, C.J., Wilson, L.G., Amy, G.L., and Brothers, K., 

1993, Fate of organochlorine compounds during aquifer 

storage and recovery: The Las Vegas experience: 

Groundwater, v. 31, no. 3, p. 410–416.

Mirecki, J.E., Campbell, B.D., Conlon, K.J., and Petkewich, 

M.D., 1998, Solute changes during aquifer storage 

recovery testing in a limestone/clastic aquifer: Ground 

Water, v. 36, no. 3, p. 394–403.

Moore, R.M., and Tokarczyk, R., 1993, Volatile biogenic 

halocarbons in the northwest Atlantic: Global 

Biogeochemical Cycles, v. 7, p. 195–210. 

Moran, B.N. and Hickey, W.J., 1997, Trichloroethylene 

biodegradation by mesophilic and psychrophilic 

ammonia oxidizers and methanotrophs in groundwater 

microcosms: Applied and Environmental Microbiology, 

v. 63, no. 10, p. 3866–3871. 

Morris, J.C., 1978, The chemistry of aqueous chlorine in 

relation to water chlorination, in Jolley, R.L., ed., Water 

Chlorination: Environmental Impact and Health 

Effects, v. 1: Ann Arbor, Mich., Ann Arbor Science 

Publishers, Inc., p. 21–33.

Oliver, B.G., 1980, Effect of temperature, pH, and bromide 

concentration on the trihalomethane reaction of 

chlorine with aquatic humic material, in Jolley, R.L., 

Brungs, W.A., and Cumming, R.B., eds., Water 

Chlorination: Environmental Impact and Health 

Effects, v. 3, Ann Arbor, Mich., Ann Arbor Science 

Publishers, Inc., p. 141–149

Peng, D., and Dural, N.H., 1998, Multicomponent 

adsorption of chloroform, carbon tetrachloride, and 

1,1,1-trichloroethane on soils: Journal of Chemical 

Engineering Data, v. 43, p. 283–288.

Pfennig, N., 1978, Rhodocyclus purpureus gen. nov. and sp. 

nov., a ring-shaped, vitamin B12-requiring member of 

the family rhodospirillaceae: International Journal of 

Systematic Bacteriology, v. 28, p. 283–288. 

Phillips, S.P., Carlson, C.S., Metzger, L.F., Howle, J.F., 

Sneed, Michelle, Galloway, D.L., Ikehara, M.E., 

Hudnut, K.W., Behr, J.A.,King, N.E., and Lewis, A.S., 

in press, Analysis of pilot tests of subsurface injection, 

storage, and recovery of freshwater in Lancaster, 

Antelope Valley, California: U.S. Geological Survey 

Water-Resources Investigations Report 03-4061.

Pyne, R.D.G., 1995, Groundwater recharge and wells; a 

guide to aquifer storage and recovery: Boca Raton, 

Lewis Publishers, CRC Press, Inc., 376 p.

Pyne, R.D., Singer, P.C., and Miller, C.T., 1996, Aquifer 

storage and extraction of treated drinking water: 

Denver, American Water Works Association Research 

Foundation and American Water Works Association, 

176 p.

Quanrud, D.M., Arnold, R.G., Wilson, L.G., Gordon, H.J., 

Graham, D.W., and Amy, G.L., 1996, Fate of organics 

during column studies of soil aquifer treatment: Journal 

of Environmental Engineering, v. 122, p. 314–321.

Reckhow, D.A., Singer, P.C., and Malcolm, R.L., 1990, 

Chlorination of humic materials: By-product formation 

and chemical interpretations: Environmental Science 

and Technology, v. 24, no. 11, p. 1655–1664.

Roberts, P.V., 1985, Field observations of organic 

contaminant behavior in the Palo Alto Baylands, in

Asano, T., ed., Artificial Recharge of Groundwater: 

Boston, Butterworth Publishers, p. 647–679.

Roberts, P.V., Goltz, M.N., and Mackay, D.M., 1986, A 

natural gradient experiment on solute transport in a 

sand aquifer 3. Retardation estimates and mass balances 

for organic solutes: Water resources Research, v. 22, no. 

13, p. 2047–2058.

Roberts, P.V., Schreiner, J., and Hopkins, G.D., 1982, Field 

study of organic water quality changes during 

groundwater recharge in the Palo Alto baylands: Water 

Research, v. 16, p. 1025–1035.

Rook, J.J., 1977, Chlorination reactions of fulvic acids in 

natural waters: Environmental Science and Technology, 

v. 11, no. 5, p. 478–483.
72 Processes Affecting Trihalomethane Concentrations Associated with the Third Injection, Storage, and Recovery Test at Lancaster, Antelope Valley, CA

Rook, J.J., Grass, A.A., and Van der Heijden, J., 1978, 

Bromide oxidation and organic substitution in water 

treatment: Journal of Environmental Science and 

Health, v. 13, p. 91–116.

Schaefer, J.K., Goodwin, K.D., McDonald, I.R., Murrell, 

J.C., and Ormland, R.S., 2002, Leisingeria 

methylohalidivorans gen. nov., sp. nov., a marine 

methyltroph that grows on methyl bromide: 

International Journal of Systematic and Evolutionary 

Microbiology, v. 52, no. 3, p. 851–859.

Schwarzenbach, R.P., and Westall, J., 1981, Transport of 

nonpolar organic compounds from surface water to 

groundwater. Laboratory sorption studies: 

Environmental Science and Technology, v. 15, no. 11, 

p. 1360–1367.

Singer, P.C., Pyne, R.D.G., Mallikarjun AVS, Miller, C.T., 

and Mojonnier, C., 1993, Examining the impact of 

aquifer storage and recovery on DBPs: Journal of the 

American Water Works Association, v. 85, no. 11, 

p. 85–94.

Strand, S.E., and Shippert, L., 1986, Oxidation of 

chloroform in an aerobic soil exposed to natural gas: 

Applied and Environmental Microbiology, v. 52, 

p. 203–205. 

Streetly, M.J., 1998, The use of modeling to predict the 

behavior of ASR systems: in Peters, J.H., ed., Artificial 

recharge of groundwater: Proceedings of the Third 

International Symposium on Artificial Recharge of 

Groundwater, TISAR 98, Amsterdam, The Netherlands, 

21–25 September, 1998: A.A. Balkema, Rotterdam, p. 

263–267.

Stumm, W., and Morgan, J.J., 1996, Aquatic Chemistry (3d 

ed.): New York, John Wiley & Sons, Inc., 1022 p.

Symons, J.M., Krasner, S.W., Simms, L.A., and Sclimenti, 

M., 1993, Measurement of THM and precursor 

concentrations revisited; The effect of Bromide ion: 

Journal of the American Water Works Association, 

v. 85, no. 1, p. 51–62.

Thomas, J.M., McKay, W.A., Cole, E., Landmeyer, J.E., and 

Bradley, P.M., 2000, The fate of haloacetic acids and 

trihalomethanes in an aquifer storage and recovery 

program, Las Vegas, Nevada: Groundwater, v. 38, no. 4, 

p. 605–614.

Traina, S.J., Novak, J., and Smeck, N.E., 1990, An 

ultraviolet absorbance method of estimating the percent 

aromatic carbon content of humic acids: Journal of 

Environmental Quality, v., 19, no. 1, p. 151–153.

Vannelli, T., Studer, A., Ketesz, M., and Leisinger, T., 1998, 

Chloromethane metabolism by Methlobacterium sp. 

Strain CM4: Applied and Environmental Microbiology, 

v. 64, no. 5, p. 1933–1936. 

U.S. Environmental Protection Agency, 1998, Environmental 

Protection Agency: 40 CFR parts 9, 141, and 142: 

National Primary Drinking Water Regulations: 

Disinfectants and disinfection by-products, final rule: 

Federal Register, v. 63, no. 241, p. 69390–69476.

Wackett, L.P., Logan, M., Block, F.A., and Bao-li, C., 1992, 

A mechanistic perspective on bacterial metabolism of 

chlorinated methanes: Biodegradation, v. 3, p. 19–36. 

Walton, B.T., Hendricks, M.S., Anderson, T.A., Griest, W.H., 

Merriweather, R., Beauchamp, J.J., and Francis, C.W., 

1992, Soil sorption of volatile and semivolatile organic 

compounds in a mixture: Journal of Environmental 

Quality, v. 21, no. 4, p. 552–558.

Wannikhof, R., Mulholland, P.J., Broecker, W.S., and 

Hamilton, M., 1987, Gas exchange rates on Mono Lake 

and Crowley Lake, California: Journal of Geophysical 

Research, v. 92, p. 14567–14580.