In Re: Methyl Tertiary Butyl Ether ("MTBE") Products Liability Litigation

Filing 4625

MEMORANDUM OF LAW in Opposition re: (609 in 1:08-cv-00312-VSB-DCF) MOTION to Dismiss . . Document filed by New Jersey Department of Environmental Protection, The Commissioner of the New Jersey Department of Environmental Protection. (Attachments: # 1 Exhibit 1, # 2 Exhibit 2 part 1 of 4, # 3 Exhibit 2 part 2 of 4, # 4 Exhibit 2 part 3 of 4, # 5 Exhibit 2 part 4 of 4, # 6 Exhibit 3, # 7 Exhibit 4, # 8 Exhibit 5, # 9 Exhibit 6, # 10 Exhibit 7)Filed In Associated Cases: 1:00-cv-01898-VSB, 1:08-cv-00312-VSB-DCF.(Kaufmann, Leonard)

Figure 5.4. Example of changes in TBA and MTBE concentrations as attributed to
biodegradation in USEPA 2005 (see page 27 USEPA (2005); Site 00UT038, Well MW08). Review of site documents indicates that the observed changes in concentrations are
likely the result of DPE remediation (Eweis et al., 2007).
5.5.2 Isotopic fractionation as a proxy indicator of biodegradation
The last few years have seen considerable interest in the use of compound specific isotope analysis (CSIA)
as a tool to document biodegradation of MTBE in the subsurface (Kolhatkar et al., 2002; Kuder et al.,
2005; Wilson et al., 2005a, Zwank et al., 2005; USEPA 2005). Simply put, the methodology stems from the
principle that microorganisms preferentially biodegrade the chemical with lighter isotope (for example 12C,
which is more common), resulting in enrichment in the heavier isotopes (e.g., 13C, which is less common),
a process referred to as isotopic fractionation. For MTBE, CSIA was seen as a powerful tool, making it
possible to “unequivocally identify and measure anaerobic biodegradation” (USEPA 2005). In applying this
methodology it was assumed that other processes that can fractionate compounds in the subsurface, such as
volatilization, adsorption, diffusion, etc…have a negligible isotopic effect compared to biodegradation.
Recent research however, indicates otherwise: “Use of CSIA is no silver bullet and will be most useful and
cost effective when applied within the context of the hydrological, geological, geochemical and
microbiological parameters at the site. The factors that affect contaminant transport and degradation over
time as well as space must be identified and evaluated.” (USEPA, 2008).
Here, I discuss:

74

•

The limitations of CSIA application to field sites as an indicator of MTBE biodegradation, given
recent studies that show that other physical and chemical processes can in fact contribute to
isotopic fractionation.

•

The published reports on the use of CSIA to document anaerobic MTBE biodegradation,
especially with regards to Orange County, California, and the related MTBE “natural” attenuation
research performed by the USEPA (2005).

5.5.2.1 Limitations in CSIA application at field sites
A fundamental limitation in the application of CSIA to groundwater systems stems from both the known
and the unknown potential for other physical and chemical processes to produce similar isotopic
fractionation signals to those produced by biodegradation. The study of the effects of physical and chemical
processes on isotopic fractionation is an evolving and active field of research (Pernaton et al., 1996;
Prinzhofer and Pernaton, 1997; Poulson et al., 1997; Harrington et al., 1999; Zhang and Kroos, 2001;
Ellis et al., 2002; Schloemer and Kroos, 2004; Richter et al., 2006; LaBolle et al., 2006; Kuder et al., 2007,
Van Breukelen 2007; Bouchard et al., 2008; Kuder et al., 2009). In a recent publication, USEPA (2008)
noted: “Because CSIA is a new approach, there are no widely accepted standards for accuracy, precision
and sensitivity, and no established approaches to document accuracy, precision, sensitivity and
representativeness” (USEPA 2008).
For example, recently published results (LaBolle et al., 2006) indicate that diffusion in water can lead to
isotopic enrichment effects similar to those attributed to aerobic biodegradation. Figure 5.5 compares the
computed isotopic enrichment of 13C MTBE by diffusion in an alluvial aquifer with that measured during
MTBE biodegradation in laboratory microcosms showing how they may produce similar effects (LaBolle et
al., 2008).

Figure 5.5. Computed 13C MTBE isotopic enrichment resulting from transport in a thin alluvial aquifer
bounded by aquitards is compared with measured isotopic fractionation resulting from aerobic
biodegradation of MTBE (Hunkeler, 2001). Results show the potential for diffusion to confound efforts to
assess biodegradation using CSIA in the field.
75

Research in the late 1990’s demonstrated that volatilization has a significant effect on the isotopic
fractionation of VOCs. A seminal compendium of papers on isotopic fraction appeared in the journal
Organic Chemistry in 1999 (Vol. 30). The introductory note to those publications called attention to the
role of volatilization:
“A reasonable conclusion to draw from this set of papers is that while physical fractionation of
carbon and chlorine isotopes may be small under vapor/liquid equilibrium conditions, such
isotopic fractionation may be magnified at large degrees of evaporation. Such potential Rayleigh
reservoir effects must therefore be taken into account in natural or engineered systems.” (Abrajano
and Lollar, 1999).
However, that advice and its importance, especially with regards to remediation, were largely ignored volatilization was generally assumed to have negligible effects on isotopic fractionation, until recently.
Current research on the volatilization of MTBE indicates significant isotopic effects (Philp and Kuder,
2006; Kuder et al., 2007). Based on laboratory studies, Kuder et al., (2009) concluded: “Experimental
results on MTBE behavior in diffusive volatilization and dynamic vapor extraction show measurable
changes in the isotope ratios of the MTBE remaining in the aqueous or nonaqueous phase liquid (NAPL)
matrix”.
Even so, the simplified bench-top laboratory models used by Kuder et al., (2009) only touched on the
possible extent of this phenomenon at field sites, which span tens to hundreds of feet. Remediation
processes at field sites, such as dual phase extraction (DPE), air sparging (AS) and soil vapor extraction
(SVE) involve the four-phase transfer of MTBE between air, water, soil and free-phase gasoline, as well as
the transport of groundwater and gasoline vapors over tens to hundreds of feet, and the redistribution of
gasoline from a vapor to a free-phase in the vicinity of the extraction well. Isotopic fractionation will likely
occur in all of these mass transfer processes (Eweis et al., 2007).
Abiotic transformation of MTBE to TBA can also contribute to isotopic fractionation. For example, a
recent study (Elsner et al., 2007) reported significant isotopic fractionation of MTBE resulting from acid
hydrolysis (abiotic degradation). As discussed earlier, MTBE may also be reduced abiotically to TBA
under anaerobic (methanogenic) conditions (Bradley et al., 2006). Given that this is relatively new
research, the effect of abiotic MTBE reduction on isotopic fractionation, to my knowledge, is unknown at
this point in time. Nevertheless, studies on other contaminants indicate that isotopic fractionation due to
reductive processes can be significant. VanStone et al. (2004) for example, reported significant enrichment
in 13C in chlorinated ethylenes as a result of abiotic reductive dechlorination. Similarly, Bill et al. (2001)
reported that lighter isotopes reacted preferentially in abiotic reductive dechlorination of TCE resulting in
13

C enrichment. As a result, USEPA (2008) notes: “The relative importance of biodegradation versus

processes of abiotic degradation at the site must be considered. In the past few years, the principles of
Rayleigh controlled isotope fractionation of organic contaminants in ground water have been shown to
apply to abiotic degradation as well as biodegradation”

76

Further work is needed therefore, both in the laboratory and in the field, to evaluate the effects of
remediation and abiotic transformation processes on isotopic fractionation. To my knowledge, the effects of
SVE and DPE, as well as other remediation processes and hydrogeologic factors, on the isotopic
fractionation of MTBE have not been studied under field conditions.

5.5.2.2 The USEPA (2005) Report on “Monitored Natural Attenuation of MTBE”
The USEPA document “Monitored Natural Attenuation of MTBE as a Risk Management Option at
Leaking Underground Storage Tank Sites” (USEPA 2005) relied heavily on the use of isotopic
fractionation as an indicator of anaerobic MTBE biodegradation in the subsurface asserting that: “This
makes it possible for the first time to unequivocally identify and measure anaerobic biodegradation of
MTBE at field Scale” (USEPA 2005). The EPA document (USEPA 2005) relied on the work of J.T. Wilson
(USEPA) and others and evolved out of a need to interpret data from MTBE-contaminated sites where
relatively high concentration ratios of TBA to MTBE were observed in monitoring wells. Theoretically,
high TBA to MTBE could result from MTBE biodegradation to TBA.
In this section, I will discuss the research findings of Wilson et al. (2005a) and Kuder et al. (2005) and the
data that they collected from thirteen Orange County, California MTBE-contaminated sites, and which
formed the basis for USEPA (2005)22. I will also provide details from the recent analysis and review of this
research by Eweis et al. (2007). The research that led to the publication of Eweis et al. (2007) was
conducted with the cooperation of Dr. J.T. Wilson (USEPA) who provided my colleagues and I with details
regarding the analytical results from the 13 Orange County sites that he studied. He did not participate in
the actual analysis. We also gathered additional publicly available data from the California State Water
Resources Control Board Geotracker database, including quarterly reports that describe the investigation
and remediation activities performed at the 13 Orange county MTBE-contaminated sites discussed in
USEPA (2005) (see Table 4.3).
The Researchers. Site specific apparent indications of anaerobic MTBE biodegradation, based on isotopic
fractionation, have been collected and have been published during the last decade by three main authors
(J.T. Wilson, T. Kuder and R. Kolhatkar who have worked and published together), and their respective
colleagues (see Table 5.3 below)
The Study Sites. Much of the data comes from sites in Orange County, California, with one additional site
from Parsippany, New Jersey. A common characteristic of these sites is a relatively high concentration of
TBA (a co-contaminant of MTBE and a possible degradation product) to MTBE. Data was collected from
thirteen sites (each containing multiple monitoring wells) in Orange County during 2002 and 2003. The
sites discussed in Kuder et al. (2005) are a subset of the ones discussed in Wilson et al. (2005a). However,
22

USEPA 2005 was authored by J.T. Wilson, P.M. Kaiser and C. Adair all with the
USEPA Office in Ada, Oklamhoma. The document has sometimes been cited as Wilson
et al., 2005, which makes things confusing given that Wilson and colleagues had two
other publications that year,
77

while Wilson et al. (2005a) identified each site by the case number assigned to it by the Regional Water
Control Board, Kuder et al., (2005) kept the site name and address ambiguous, assigning each the acronym
OC (Orange County) followed by a number from 1 to 8 (Table 5.4). This discrepancy gives the mistaken
impression that the two papers (Kuder et al., 2005 and Wilson et al., 2005a) are unrelated and that there
have been many more than 13 sites studied using isotopic fractionation in Orange County. Similarly, data
from the site in Parsippany, New Jersey (USEPA 2005) was used in more than one paper, by the same
group of authors (Kolhakar et l., 2002; Wilson et al., 2005b, Kuder et al., 2005), again leading to the
impression that the data came from more than one site (Table 5.4).
Prior and ongoing remediation, and the differential transport properties of MTBE and TBA, were
ignored. While the USEPA sought to establish whether or not MTBE was “naturally” attenuating in
groundwater, the USEPA (2005) overlooked the prior and ongoing remediation at the study sites and its
potential effects on the isotopic fractionation of MTBE and potential impacts on the relative concentrations
of MTBE and TBA (as explained in sections 5.5.1 and 5.5.2 above). In fact, of 12 sites with evidence of
MTBE fractionation, 11 sites had either undergone remediation, or were in remediation at the time of
sampling (Table 5.4). Groundwater samples were collected from within, or from wells very nearby,
remediation wells. The USEPA also ignored the differential fate and transport properties of MTBE and
TBA, properties that naturally lead to their fractionation (and changing concentration ratios) in the
environment (see Section 5.5.1). These oversights allowed for the mistaken assumption that elevated
and/or rising TBA to MTBE ratios, and the isotopic fractionation of MTBE were necessarily the result of
natural processes, mainly biodegradation.
TBA was in the gasoline. To clarify, TBA is a common impurity in MTBE, accounting for up to 10% of
the MTBE added to gasoline (USEPA 2005; Eweis et al., 2007). As explained in Section 5.5.1, given the
differential fate and transport properties of MTBE and TBA, elevated TBA to MTBE ratios in groundwater
are to be expected under natural conditions because these chemicals partition differently between soil,
water and air and migrate at different speeds in groundwater. For example, once in groundwater TBA is
relatively nonvolatile compared to MTBE. Therefore, remediation processes that include volatilization
(evaporation), such as Dual Phase Extraction (DPE), Soil Vapor Extraction (SVE), and Air Sparging (AS),
will remove MTBE from groundwater, leaving most of the TBA behind. As discussed in Section 5.5.1,
Eweis et al. (2007) showed that volatilization will eventually lead to TBA concentrations that compare
favorably to those predicted from the dissolution of gasoline, and MTBE concentrations that are much
lower than predicted through dissolution, a signal that could understandably be misinterpreted as MTBE
biodegradation.
Study sites were not under natural conditions. In light of this background information, Eweis et al.
(2007) summarized the research of Wilson et al. (2005a) as follows:
“Wilson et al. [2005a] compared field-measured MTBE and TBA concentrations at contaminated
sites in Orange County, California, to those predicted from dissolution of MTBE-oxygenated
gasoline into water. Predicted concentrations were based on the assumption that gasoline
78

contained 11 vol % MTBE, and 0.22 vol % TBA (based on 2% TBA by weight in MTBE). Wilson
et al. [2005a] noted that measured TBA concentrations at these sites were in agreement with those
predicted from dissolution, whereas the measured MTBE concentrations were commonly one or
more orders of magnitude less than predicted from dissolution of MTBE oxygenated gasoline into
water. The researchers theorized that “some process or processes” are acting to lower MTBE and
increase TBA concentrations [Wilson et al., 2005a]. Using carbon isotope fractionation in a
focused study of 13 sites, with TBA/MTBE ratios greater than those expected from the dissolution
of fresh MTBE oxygenated gasoline, Wilson et al. [2005a] concluded that MTBE biodegradation
could explain the majority of TBA concentrations at a fraction of the sites (4 out of 13), but could
not explain the majority of TBA concentrations at a majority of the sites (9 out of 13).”
Wilson et al. (2005a) reported some evidence of biodegradation23 for 12 of the 13 sites, and no evidence for
MTBE biodegradation at one site (at site 99UT015 the two wells tested showed “no evidence of
biodegradation” Wilson et al., 2005a).
However, as discussed in Eweis et al. (2007), Wilson et al. (2005a) overlooked the prior and ongoing
remediation by volatilization conducted at sites (Table 5.4), as well as a concurrent rise in the water table
into the previously unsaturated zone where volatilization had taken place. The latter is important because,
as the water table rises, the contamination affected by volatilization in the unsaturated zone enters the
saturated groundwater where it can then be detected in the monitoring wells (See Section 5.5.1 and Eweis et
al., 2007). Our review of site files from the 13 sites found that (Table 5.4):
•

substantial remediation was performed for at least 11 of the 12 sites with reported evidence of
biodegradation, and

•

no significant remediation was performed at the one site where Wilson et al. (2005a) reported “no
evidence of biodegradation”.

In other words, the potential for biodegradation was not studied under natural conditions. Moreover, the
isotopic fractionation and elevated and/or rising TBA to MTBE ratios appeared to be associated with the
remediation. Not noted by Eweis et al. (2007) was that the remaining site where no reported remediation
was performed, and for which Wilson et al. reported evidence of biodegradation, appears to have been be
associated with a vapor leak.24
Rather than present findings from all 13 sites, in Eweis et al. (2007) we chose to present findings from two
sites that the USEPA (2005) chose as exemplary of MTBE “biodegradation”. The two figures presented in
Eweis et al. (2007) (one in the text and one in the online material of the Journal publication) contain the
data presented in USEPA (2005) Figure 5.3 “Transition from MTBE to TBA in monitoring wells at
gasoline spill sites in Orange County, California.” The Figure from Eweis et al. (2007) appears in Figure
5.4 presented in the previous section (5.5.1) of this report, discussed earlier in the context of TBA and
MTBE volatilization. With regards to the data in the figure, the USEPA (2005) writes:

23

Data for isotopic fractionation varied by monitoring well. For example, while one well
may have had a strong shift in the isotopic signal, a nearby well had an insignificant shift
and was therefore seen as not having exhibited signs of biodegradation.
24
The site general lacks detectable BTEX.
79

“The best indication for MTBE biodegradation that is available from conventional monitoring
parameters is an abrupt and persistent increase in the ratio of TBA to MTBE in ground water.”
However, as the figures from Eweis et al. (2007) show, the researchers of Wilson et al. (2005a) were
sampling sites and specific wells, with concurrent DPE remediation. The additional analysis in Eweis et al.
(2007) shows that the observed trends in TBA and MTBE concentrations in those wells are consistent with
that expected from the remediation. Accordingly, the most likely cause for the observed changes in TBA to
MTBE ratios was clearly the remediation, and not biodegradation25.
At the time of the Wilson et al. (2005a) study, the fractionation of stable carbon isotopes (in particular the
enrichment of heavy isotopes) detected through CSIA was thought to provide “an unequivocal indication of
MTBE biodegradation” (USEPA, 2005). It is noteworthy that similar statements have occurred in numerous
publications since the late 1990’s. Specifically, Kolhatkar et al. (2002), Kuder et al. (2005), Wilson et al.
(2005a) and USEPA (2005) thought that the isotopic fractionation effects from physical processes were
negligible. However, as explained in this report, more recent findings showing that sorption, diffusion and
volatilization can all significantly affect isotopic fractionation contradict those notions.
Eweis et al. (2007) called attention to the evolving nature of the research and what we do not yet know
about isotopic fractionation. For example, DPE and SVE involve much more than volatilization. Both
processes involve the four-phase transfer of MTBE between air, water, soil and free-phase gasoline, as well
as the transport of gasoline vapors over tens of feet, and potentially the redistribution of gasoline from a
vapor to a free-phase in the vicinity of the extraction well. Isotopic fractionation will occur in all of these
mass transfer processes.
As was pointed out in Eweis et al., (2007):
“To our knowledge, the extent to which MTBE isotopically fractionates from these, and other,
physical processes has not been assessed at field sites undergoing DPE and/or SVE remediation”.

25

13 Note that similar trends might be observed even if the remediation were not
concurrent, but rather occurred earlier in time, because of the regional water level rise
into the unsaturated zone above the water table affected by remediation
80

Table 5.3: Select papers on the anaerobic biodegradation of MTBE:
Citation and year
published
Kolhatkar et al.,
2002

Authors

Publication Title

Publishing entity

R. Kolhatkar, T.
Kuder, P. Philp, J.
Allen, and J.T.
Wilson

Environmental
Science and
Technology

Kuder et al., 2005

T. Kuder, J.T.
Wilson, P. Kaiser, R.
Kolhatkar, P. Philp,
and J. Allen
J.T. Wilson, R.
Kolhatkar, T. Kuder,
P. Philp, and S.
Daugherty
J.T. Wilson, C.
Adair, P.M. Kaiser,
and R. Kolhatkar
J.T. Wilson, P.M.
Kaiser, and C. Adair,

Use of compound-specific stable
carbon isotope analyses to
demonstrate anaerobic biodegradation
of MTBE in groundwater at a gasoline
release site
Enrichment of stable carbon and
hydrogen isotopes during anaerobic
biodegradation of MTBE: microcosm
and field evidence
Stable isotope analysis of MTBE to
evaluate the source of TBA in ground
water
Anaerobic biodegradation of MTBE at
a gasoline spill site

Ground Water
Monitoring and
Remediation
USEPA

Wilson et al.,
2005a

Wilson et al.,
2005b
USEPA 2005

USEPA 2007

J.T. Wilson and C
Adair

Monitored Natural Attenuation of
MTBE as a Risk Management Option
at Leaking Underground Storage Tank
Sites
Monitored natural attenuation of
tertiary butyl alcohol (TBA) in ground
water at gasoline spill sites

Environmental
Science and
Technology
Ground Water
Monitoring and
Remediation

USEPA

81

Table 5.4: Remediation performed at sites used in the natural-attenuation/isotopic-fractionation studies of Wilson et al. (2005a) and Kuder et al. (2005). Thirteen
of fourteen sites were in Orange County, California (Page 1 of 2).
Site Name
Mobil #18F1M

Address
5502 Orangethorpe, La
Palma, Orange County,
California
15001 Goldenwest,
Huntington Beach, Orange
County, California

Mobil #18G3W

Wilson et al.
(2005b) and
USEPA (2005)
96UT028

Kuder et al.
(2005)
OC1

00UT038

OC2

Mobil #18H7Q

13872 Red Hill Ave.,
Tustin, Orange County,
California

88UT138

OC3

Mobil #18-HL8

26051 La Paz, Mission
Viejo, Orange County,
California

87UT211

OC4

Arco #1888

16501 Golden West,
Huntington Beach, Orange
County, California

88UT198

OC5

18025 Magnolia, Fountain
Valley, Orange County,
California

85UT114

OC 6

b

16969 Brookhurst,
Fountain Valley, Orange
County, California

98UT010

OC 7

b

30011 Crown Valley,
Laguna Niguel, Orange
County, California

91UT086

OC 8

a

Arco #1905

Shell Oil

Shell Oil

Remediation prior to or concurrent
with 2002- 2003 Study
SVE
DPE (1999-2000) DPE (2001-2002)
SE
DPE (2001) DPE/GWPT system
installed (2001)
DPE (2002)
over-purging (2002).
LPH pump outs (1990-1991)
Over-purging (1998-2000)
DPE testing (1999)
Tank upgrades (2000)
150 ton SE (2000) SVE (2003).
Tank upgrades (1987)
SE (1987)
SVE (1991)
DPE (1999-2001)
Over-purging (1993)
SVE (1993, 1995)
SVE (2000-2003)
GWPT system (2003)
SE (1992)
SVE (1992-2000) Free Product
removal (hand bailing; 1997-1999) DPE
(2000-2003)
No Cleanup actions exist.

Manual Separate-phase hydrocarbons
(SPH) recovery (1986-1990)
SVE (1996)
Over-purging (1999)
SE and treat (2003)

Remediation
After 2003
Monthly over-purging (2004)
Bioreactor startup (2004)
GWPT (2004-2011).

Current Site
Open/Closure
Info
OPEN
Begin date 1996

GWPT system started (2004)
GWE (2008).

OPEN
Begin Date 2000

Over-purging (2006)
Oxygen emitters installed
(2007)
Over-purging (2010)
Chemical oxidation injection
(2010).
167 ton SE (2006) Fuel
system removal activities
(2006)
Installed soil vapor intrusion
barrier (2008).
Over-purging events (20062009)
11 iSOC units (2010-2011)

OPEN
Begin Date 1988

DPE (2004-2006) DPE which
consist of separate SVE and
GWE systems:
SVE (2007-2010)
GWE (2007-2011)
No Cleanup actions exist
Oxygen pulse sparging
proposed and six sparge wells
installed (2004)
None

OPEN
Begin Date 1985

OPEN
Begin Date 1987

OPEN
Begin Date 1988

OPEN
Begin Date 1997
Closed 8/2005
Begin Date 1991

CONTINUED ON FOLLOWING PAGE

82

Table 5.4 (CONTINUED): Remediation performed at sites used in the natural-attenuation/isotopic-fractionation studies of Wilson et al. (2005a) and Kuder et al.
(2005). Thirteen of fourteen sites were in Orange County, California (Page 2 of 2).
Site Name
b

Address

Wilson et al.
(2005b) and
USEPA (2005)
86UT062

Kuder et al.
(2005)

Remediation prior to or concurrent
with 2002- 2003 Study

-

SVE (1988-1990 and 1992)
GWPT (1989-1996)
DPE (SVE-GWE) (2001)
SE and tank removal (1986, 2000, and
2003)
SE and tank removal (1993)
AS-SVE system (1994-1995)
DPE system (2002)
GWPT (2003).
Tank removal (1986)
Free product removal (1987-1988)
GWE and Nutrient enhancement with
subsequent AS (1991-1994)
SE and GWE and Free Product removal
(1995-1996)
Periodic pump outs using vacuum
truck (2002)
SE (1990 & 1991) excavation pit
dewatering (1992)

Shell Oil

27101 Ortega Highway,
San Juan Capistrano,
Orange County, California

Farjami or
Stanton Mobil
Station

12493 Beach Blvd.,
Stanton, Orange County,
California

99UT032

-

Former Texaco
Service
c
Station

6011 Manchester Blvd.,
Buena Park, Orange
County, California

86UT175

-

Tosco-76
#7329

34306 Pacific Coast, Dana
Point, Orange County,
California

99UT015

-

Unocal #5385
or (76 Service
Station No.
255385)

34131 Doheny Park Rd.,
Dana Point (Also listed
under Capistrano Beach) ,
Orange County, California

89UT007

-

SE
Tank, and pipe replacement (1990)
SVE system test (1992)
SVE (1995-1996)
EBD (1995-1996)
ORC (1998)

Remediation
After 2003
None

Current Site
Open/Closure
Info
Closed 6/2006
Begin Date 1986

GWPT (2004-2008)
Oxygen infusion device
(2009-2010)
GWPT (2010-2011).
Proposed and denied oxygen
injection pilot test (2004)
tank removal (2006)
Over purging (2006).

OPEN
Begin Date 1997

Free Product removal
(LNAPL) (2008)
DPE (2008-2009)
132 ton SE (2010)
Onsite and offsite ozone
injection and DPE (2010)
monthly LPH removal events
(2011)
oxygen injection (2011).
AS ozone air sparging system
(2004-2009)

OPEN
Begin Date 1998

Closed 7/2007
Begin Date 1986

OPEN
Begin Date 1988

?
Parsippany, NJ
Parsippany
yes
UNKNOWN
UNKNOWN
UNKNOWN
Site used in Busch-Harris et al. (2008) Biotrap study; bSite used in Shell DeVaull et al. (2003) study; cSite considered in Shell DeVaull et al. (2003) study.
SE: Soil extraction/excavation
GWPT: Groundwater pump and treat system
SVE: Soil vapor extraction
iSOC unit: In Situ Submerged Oxygen Curtain
DPE: Dual phase extraction
EBD: Enhanced Biodegradation
AS: Air sparging
ORC: Oxygen Releasing Compound (Injected)
GWE: Groundwater extraction
a

83

5.5.2.3 The use of Bio-Traps to verify MTBE biodegradation:
More recently, 13C-labeled chemicals, such as MTBE and TBA, have been used to better measure the
occurrence of biodegradation in situ. In general terms, a bio-trap is made up of a perforated container,
packed with activated carbon, which is then allowed to adsorb the chemical in question. When placed in the
subsurface (usually for several weeks), the bio-trap acts as a microcosm, where bacteria can grow. If the
bacteria degrade the 13C-labeled chemical, it can incorporate it into its cell mass. CSIA can then be used to
analyze cellular material (such as phospholipid fatty acids, or amino acids) to draw a more direct link
between the baited chemical and microbial degradation. This method can be useful to show “degradation
potential rather than degradation rate” (Bombach 2010).
In a study by Busch-Harris et al., (2008), bio-traps were used at one of the original thirteen Orange County
sites studied by Wilson et al. (2005a), to examine MTBE and TBA biodegradation in situ. The site at 18025
Magnolia street (85UT114; see Table 4.4) had been undergoing remediation, first through SVE (since
1992) and then through DPE starting in 2000. The 13C-MTBE and 13C-TBA amended bio-traps were
deployed in 2005 and 2006 in wells MW-15 and MW16. Both wells (as well as several others on this site)
were used for DPE operation, and remediation in the two wells had to be turned off temporarily while the
bio-traps were being incubated in the subsurface. Because of “limited success due to low levels of total
biomass collected in the bio-traps” the first experiment with the 13C-MTBE amended bio-traps had to be
repeated using a larger setup, with “five to six times the number of Bio-Sep beads per bio-trap as those
used in the first deployment” (Busch-Harris et al., 2008). However, very little cellular mass could be
extracted form the bio-traps: “no enriched fatty acids were detected in bio-traps from well MW15”, and
only one fatty acid was detected at high enough concentration in well MW16. Yet, the authors (BuschHarris et al., 2008) concluded that the results demonstrated the “existence of degradation potential” in the
aquifer. However, a similar study with bio-traps (Fiorenza et al., 2009) concluded that 13C-phospholipid
fatty acids within the range reported by Busch-Harris et al. (2008) were not considered favorable for natural
attenuation through biodegradation.
Of course none of the foregoing details address the salient question: can natural attenuation control the
plumes? Some researchers apparently believe that the data from site 85UT114, including isotopic
fractionation and elevated TBA to MTBE ratios (Wilson et al., 2005a and Kuder et al., 2005; Busch-Harris
et al., 2008), stem from biodegradation. As such, site 85UT114 represents one of the more optimistic
examples of apparent biodegradation in Orange County, or the country for that matter. Despite this, the
plumes are not being controlled, even with substantial remediation efforts. The release was discovered in
1985; remediation, which started at this site in 1992, continues as of the last groundwater monitoring report
available on Geotracker (1st quarter 2011); and the State Geotracker system describes the “Groundwater
Contamination Plume Not Stable or Decreasing”.26

26

http://geotracker.swrcb.ca.gov/profile_report.asp?global_id=T0605900033
84

5.5.2.4 Shell Study in Orange County
In November of 2003, Shell Global Solutions submitted a report to Seth Daugherty of the Orange County
Health Care Agency addressing the presence of TBA at selected UST remediation sites in Orange County
(DeVaull et al., 2003). Shell scientists noted that, depending on the commercial process used to produce
MTBE, “TBA may be present in MtBE either as a co-product, an impurity, or as an unreacted initial
component, depending on the [MTBE] manufacturing process…Typical gasoline-grade MtBE contains 96
to 99% MTBE by weight” (DeVaull et al., 2003). As such, TBA was often present in the MTBE added to
gasoline in California. DeVaull et al., (2003) however, could not explain observed concentrations at UST
sites (high TBA, low MTBE and low BTEX) simply through physical processes such as “dissolution,
partitioning, mixing, and dilution”, and looked for “biological transformation” (of MTBE to TBA) as a
potential source for measured high TBA concentrations. Like USEPA (2005), DeVaull et al (2003) did not
consider other possible explanations for the elevated TBA to MTBE ratios, e.g., volatilization and
remediation (as explained by Eweis et al., 2007).
DeVaull et al. (2003) built laboratory microcosms with soil and groundwater samples from three sites in
Orange County (see Table 5.4). While six sites were initially considered, only three were sampled for the
laboratory biodegradation experiments; those sites were part of the original 13 sites chosen by Wilson et al.
(2005a). Biodegradation of MTBE and TBA was examined under different conditions, including aerobic,
cometabolic, anaerobic with methanol added (methanogenic) and anaerobic (the latter two being most
common at release sites, and therefore more relevant to natural attenuation).
While MTBE and TBA appeared to degrade in many of the microcosms, some of the results seemed to be
inconsistent. MTBE and TBA were degraded under aerobic conditions, but under co-metabolic conditions
some microcosms from the same site showed degradation while others did not. The authors noted that:
“the results from the co-metabolism tests are inconclusive” (DeVaull et al., 2003). Under anaerobic
conditions (with methanol added to create methanogenic conditions) MTBE was degraded in 8 out of 9
microcosms, while TBA was degraded in 1 out of 9. Finally, under anaerobic conditions (without the
addition of methanol), MTBE was degraded in 7 out of 9 microcosms, while TBA was degraded in another
7 out of 9. MTBE and TBA were degraded in the same microcosm in only 5 out of 9 microcosms.
The Shell scientists did not conduct any follow up work to enrich for the microbial cultures, and their work
was not published in any scientific journals. To my knowledge, the work by DeVaull et al. (2003) did not
undergo an impartial27 peer review process. This is important considering that some of the laboratory
techniques appear suspect or flawed. For example, all of the 18 anaerobic microcosms (designed to be
deprived of oxygen) had measurable concentrations of oxygen at the end of the incubation period. In fact, 3
out of 9 microcosms labeled as “anaerobic with methanol” and 6 out of 9 microcosms labeled “anaerobic”

27

The report (DeVaull et al., 2003) lists G.E. Spinnler, also a Shell Global Solutions
scientist, as the only reviewer.
85

had oxygen concentrations high enough to qualify as “aerobic,” as defined by DeVaull et al., (2003)28.
Furthermore, the results obtained by DeVaull et al., (2003), particularly with regards to TBA degradation
under anaerobic conditions, seem inconsistent with the work conducted by the USEPA. In microcosms
from six different contaminated sites in the U.S., TBA was degraded under anaerobic conditions in
microcosms from only one site (USEPA, 2007). USEPA (2007) noted:
“In summary, the microcosm experiments did not present compelling evidence that anaerobic
TBA biodegradation was widespread at gasoline spill sites.”
Furthermore, and based on the literature available at the time, USEPA (2007) noted:
“A close examination of the available information indicates that a default presumption that TBA is
readily degraded in anaerobic ground water is not justified”
It should also be noted that DeVaull et al., (2003) were careful not to suggest that the results of their work
be extrapolated to other sites in Orange County without further examination:
“Finally, we note that the selected sites exhibited low groundwater gradients and inferred low
groundwater seepage velocities. We would expect the effects of biological degradation to be more
significant and observable at these types of sites, than at sites with higher groundwater flow rates
(high permeability/high recharge sites). This is especially true if degradation in anaerobic
conditions is the significant transformation and removal mechanism” (DeVaull et al., 2003).
Interestingly, later publications from Shell Oil scientists seem to emphasize the difficulty in attenuating
MTBE and TBA through biodegradation by natural microbial cultures in the subsurface:
•

A patent publication by Shell Oil Company scientists for a technology that incorporates patented
aerobic microorganisms and granular activated carbon for above ground water treatment notes:
“The use of bacteria or naturally occurring microbes for biodegradation of a wide range
of organic contaminants is known. However, attempts to degrade MTBE and/or TBA
using bacterial cultures have generally met with little success” (Byers et al., 2004).

•

Also in a patent publication by Shell Oil Company scientists for a technology that involves
introducing patented aerobic microbial cultures and oxygen into the subsurface for the treatment
of “ethers and/or alcohol” the scientists noted:
“Now, having used oxygenated fuels for several years, it has become clear that these
cleaner-burning fuels pose great threats to groundwater resources. In particular, many
oxygenate chemicals are very soluble in water and are slow to degrade in the
environment; hence they tend to accumulate in water resources once released to the
environment” (Salanitro et al., 2005).
The scientists continued to state:
“It is also now known that when oxygenate chemicals including alkyl ethers, such as
MTBE and tertiary butyl alcohol [TBA], are found in the subsurface, then they are
resistant to biodegradation under natural conditions. This is the main reason for their
persistence and accumulation in soil and groundwater” (Salanitro et al., 2005).

As such, the USEPA (2007) document on the natural attenuation of TBA states:
“If a transport and fate model is used as part of the risk evaluation, there is no justification to
include biodegradation in the model. Including biodegradation in the model requires direct
28

DeVaull et al., (2003) noted: “aerobic criteria for dissolved oxygen > 1 mg/L”
86

evidence that bacteria in the contaminated aquifer are capable of degrading TBA, or in fact, have
degraded TBA” (USEPA 2007)
Finally, it is important to note that several of the sites originally considered for the study by DeVaull et al.
(2003) continue to undergo engineered remediation and were open as of the last monitoring period reported
on Geotracker in 2011.

5.6

Migration of MTBE to Supply Wells

Contamination entering groundwater can travel to water supply wells. Because contamination moves with
the groundwater, travel times from contaminant source areas to supply wells commonly range from years to
centuries, depending on hydrologic conditions and pumping rate. Because supply wells commonly pump a
mix of water from various recharge locations, contamination entering from any particular release location
will tend to mix with uncontaminated water from other locations, diluting detectable concentrations in the
well (Einarson and Mackay, 2001). This process is called wellbore dilution.
An MTBE plume may be diluted in a well due to wellbore mixing (Johnson et al., 2000). Because publicsupply wells typically have long screened intervals, wellbore dilution is generally appreciable.
Consequently, low-level MTBE contamination from non-point sources, such as atmospheric deposition, is
unlikely to result in detectable (with standard methods) levels of contamination in an operating supply well.
This is an important point because atmospheric sources of MTBE contamination have been shown to
produce surface water or shallow groundwater MTBE concentrations of about 1 µg/L: “maximum
concentrations of MTBE that might occur in shallow groundwater as a result of atmospheric deposition
would be approximately 1 µg/L” (NJDEP, 2001). Thus, the detection of MTBE concentrations in public
water supply wells generally implicates a contaminant source having a much higher concentration than that
of a nonpoint source. The mere detection of MTBE in a public supply well strongly implicates point
sources, such as LUFT sites, of contamination. These concepts are illustrated in Figure 5.6a and 5.6b.
Importantly, if the source is a LUFT and the core of the plume emanating from the LUFT is en route to the
well, then continued pumping of the contaminated water supply well will typically result in ever increasing
MTBE concentrations in drinking water produced by the well.

87

(a)

(b)

Figure 5.6. (a) Schematic representation of low-concentration, diffuse source of MTBE to
a public water supply well. Dilution with uncontaminated groundwater entering the well
will commonly result in the measured MTBE concentration at very low values, or below
detection. (b) Schematic representation of a high-concentration, point-source of MTBE,
which is likely to produce measurable or elevated concentrations of MTBE in public
water supply well water samples.

88

5.7

The Role of Time

“[Groundwater] velocities under typical hydraulic gradients can range from a few millimeters per year to a
meter per day” (Squillace et al., 1998). As such, travel times from recharge locations to discharge locations,
such as an individual supply well, commonly range from years to centuries (e.g., Fogg et al., 1998a;
Johnson et al., 2000). The time a ‘packet’ of water spends in the groundwater is referred to as its age.
Although groundwater ages are commonly expressed as average values, water pumped from a well is
derived from a mix of different recharge locations resulting in a mix of different ages (times of travel to the
well) in any given sample (e.g., Busenberg and Plummer, 1992; Fogg et al., 1998a; Weissmann et al.,
2002; Bethke and Johnson, 2002). Additionally, supply wells will commonly draw a significant fraction of
older water (Fogg et al., 1998a; Weissmann et al., 2002). Therefore, it would not be surprising, for
example, to find groundwater with an average age of 50 years resulting from a mix of waters that range
from tens to hundreds of years in age.

5.7.1 Volatile Organic Compounds
The USGS studied the occurrence of 86 volatile organic compounds (VOCs) in groundwater to assess the
vulnerability of particular supply wells in the Los Angeles, California basin (Shelton et al., 2001). The
basin is recharged primarily from engineered facilities. Of the 86 VOCs, the study found MTBE to have the
most localized distribution (primarily near recharge facilities). The authors concluded that: “this nearly
exclusive occurrence of MTBE close to the area of recharge is consistent with the relatively recent
introduction of MTBE as a fuel additive.” In contrast, chloroform contamination introduced in the early
1900’s was detected more widely and at greater distances from the recharge facilities than all other VOCs.
Similarly, a vulnerability assessment of California wells in the Livermore-Amador and Niles Cone
groundwater basins found “MTBE is a good indicator of very recently recharged water.”
Because MTBE is generally associated with “recently recharged [young] water,” not enough time has
elapsed for many MTBE plumes to spread to drinking water wells. Travel times from contaminant source
locations to wells can range from days to a century or more, with times on the order of decades quite
common in many basins (Fogg, et al., 1998a; Weissmann et al., 2002). The recent introduction of MTBE
relative to other persistent groundwater contaminants can be seen in Table 5.3 below.

89

Table 5.3. Historical Use of Selected Anthropogenic Groundwater Contaminants in
California.
Compound

Principal Uses and Source(s) to Groundwater

Widespread Use in US

MCL
(µg/l)e

Nitrate
TCE
PCE

Agricultural fertilizer, feed lots

~1950 – present

45,000

d

5.0

d

Industrial solvent

5.0

a,b,c

0.2

~1930 – present

Industrial solvent

~1930 – present
a

DBCP

Agricultural soil fumigant /nematocide
(banned in 1977)

~1960 – 1977

MTBE

Gasoline LUFTs

1992-2006

5.0f

a: http://www.epa.gov/ttn/atw/hlthef/dibromo-.html, Accessed March 24, 2004
b: Raski, D. J., I. J. Thomason, J J. Chitambar, and H. Ferris, September, 2003, A History of
Nematology in California, available at
http://plpnemweb.ucdavis.edu/nemaplex/History%20of%20Nematology%20in%20California.pdf
c: Burow et al. (1999)
d: http://www.concentric.net/~Rnk0228/solhist.html
e: http://www.epa.gov/safewater/hfacts.html - Synthetic
f: Secondary taste and odor MCL for California.

5.7.2 Vulnerability to Contamination
The vulnerability of a water supply well to contamination depends upon the travel time from contaminant
source areas to the well screen or open hole in bedrock, which in turn, depends in part upon hydrologic
conditions, well placement, well construction, and pumping rate. As a general rule, shallow groundwater is
more vulnerable to contamination than deep groundwater, due to its proximity to the land surface, and
therefore contaminant source areas. Private wells are generally more susceptible to contamination than
public wells, as they tend to be shallower and of poorer construction with regards to minimizing
contamination (Happel et al., 1998). Drinking water wells in New Jersey are highly vulnerable to
contamination as evidenced by several of the sites at issue in this case.
The advent of widespread use of MTBE in gasoline is relatively recent, and therefore many MTBE plumes
are still young. By comparison, many other contaminants that are less soluble (Figure 5.1) and less mobile
than MTBE, for example solvents such as perchloroethylene (PCE) and trichloroethylene (TCE), have been
in widespread use for many decades (Shelton et al., 2001; Table 5.3). Contaminants like these, released into
groundwater in decades past, are commonly more widespread than MTBE in older, deeper groundwater.
Their observed fate alerts us to the likely long-term fate of MTBE.

90

6
Remediation and Natural Attenuation of MTBE in
Groundwater
6.1

Remediation

Persistent contamination can render groundwater unusable for decades to centuries, absent costly and timeconsuming remediation or expensive treatment. “Once a groundwater supply is polluted, it is [often]
difficult and expensive to clean up” (Helpern et al., 2001). This is largely due to the same groundwater
flow and transport processes that cause significant variations in groundwater age within the same aquifer
system. That is, once an aquifer is exposed to a dissolved contaminant like MTBE, much of the
contaminant migrates into regions of extremely slow groundwater flow (e.g., silts and clays). From there,
the contaminants can migrate, or “bleed,” out very slowly from the slow-flow regions over an extended
period of time (LaBolle and Fogg, 2001; Zhang et al., 2007). A phenomena, often called “tailing,”
describes a situation in which the rate of bleed-out is slow enough to allow the contaminant to persist at
problematic concentrations for decades to centuries. This “tailing” is the primary reason that
decontamination of aquifers can take decades to centuries longer than contamination of aquifers (LaBolle
and Fogg, 2001; Zhang et al., 2007).

6.2

Natural Attenuation

Natural attenuation is the term used to describe contaminant removal from the environment (typically the
subsurface) through natural processes, without engineered intervention. These natural processes can be
physical, chemical, or biological. Where applicable, the use of monitored natural attenuation as a
remediation strategy can reduce the cost of site cleanup by millions of dollars.
The increased reliance on natural attenuation for cleanup of contaminated groundwater in the 1990’s,
prompted the release of “Natural Attenuation for Groundwater Remediation” by the National Research
Council (NRC, 2000). The objective of the study, which was conducted by 15 experts in the fields of
science, technology, and policy of natural attenuation, was to provide guidance on natural attenuation and
when and how it should be applied to remediate contaminated soil and groundwater (Rittmann, 2003).
Natural attenuation was deemed acceptable when the processes involved resulted in the destruction or
immobilization of the contaminant (in this context, processes such as dilution, dispersion, and partial
breakdown of a contaminant to toxic byproducts, would be considered unacceptable). The Environmental
Engineering Committee (EEC) of the USEPA Science Advisory Board (SAB) (USEPA, 2001) reviewed the
findings of NRC (2000). The committee (SAB) endorsed the findings of NRC 2000, stating:
“During the course of this review, the Subcommittee [Natural Attenuation Research
Subcommittee] agreed with and endorsed the overall strategy that NRC (2000) provides for
deciding whether or not MNA protects human health and the environment. According to NRC
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(2000), MNA should be selected only when a cause-and-effect relationship is established
between loss of contaminant and destruction or immobilization processes.” (USEPA 2001,
emphasis added).
The report (USEPA, 2001) further stated:
“the Subcommittee considers MNA to be a remediation approach based on understanding and
quantitatively documenting naturally occurring processes at a contaminated site that protect
humans and ecological receptors from unacceptable risks of exposure to hazardous contaminants.
The Subcommittee stresses that, MNA is a “knowledge-based” remedy because, instead of
imposing active controls, as in engineered remedies, scientific and engineering knowledge is used
to understand and document naturally occurring processes.” (emphasis as published).
Given that MTBE is not strongly adsorbed to soil matter, NRC (2000) listed biotransformation, or
biodegradation, as the “dominant attenuation process” for MTBE. USEPA 2001 explains:
“The NRC (2000) report de facto restricted the acceptable (or eligible) natural attenuation
processes to biodegradation and strong immobilization. This is a more restrictive definition of that
used by EPA…and reflects NRC’s finding that, while other processes occur and can contribute to
natural attenuation, the processes that destroy or strongly immobilize contaminants are more
reliable in reducing risks.” (emphasis as published).
In evaluating MTBE and the potential for natural attenuation as a remediation strategy, NRC (2000)
referred to MTBE as “generally resistant to biodegradation”. NRC (2000) also rated the likelihood that
natural attenuation will succeed as a remediation strategy for MTBE as low. In contrast, the BTEX group
was given a rating of high.
In light of more recently published literature on MTBE biodegradation, Rittmann (2003), chairman of the
committee responsible for NRC (2000), conducted a reevaluation of the NRC (2000) guidance on MTBE
natural attenuation. In this update, Rittmann (2003) concluded:
“Based on laboratory and field experience so far, the most promising type of MTBE
biodegradation is aerobic and involves initial monooxygenase reactions. The situation requires
that the supply of dissolved oxygen be significant and that MTBE degraders expressing the
monooxygenase enzymes be present in significant numbers. Having both requirements present is
not a high probability circumstance for natural attenuation; therefore, the original NRC (2000)
judgment that the likelihood of success is low for natural attenuation of MTBE seems to remain
valid.”
Intrinsic bioremediation is described by the National Research Council (NRC, 1994) as “the innate
capabilities of naturally occurring microbes to degrade contaminants without taking any engineering steps
to enhance the process.” Evidence collected in the scientific literature thus far suggests that MTBE lacks
the “intrinsic bioremediation” ability that naturally regulates BTEX releases. Three examples from oil
industry (Shell Global Solutions) studies are given below:
(1) Salanitro et al. (1999) state:
“MTBE, the common gasoline oxygenate that has been detected in groundwater, is inherently
difficult to biodegrade in aquifer soils because of its ether bond and tertiary carbon features.
92

Microbes grow slowly and produce low cell yields when cultured on MTBE. Characterization of
plumes and aquifer material from many sites indicate that the transport of ether may extend
beyond BTEX plumes. The low concentration of indigenous degraders, presence of low dissolved
oxygen and the poor growth of microbes on MTBE indicate that natural bioattenuation will be
difficult to demonstrate in groundwater plumes.”
(2) Similarly, Spinnler et al. (2001b) have concluded:
“So far, the evidence indicating intrinsic bioremediation of MTBE is inconclusive. In fact, the
extreme length of some MTBE plumes implies intrinsic bioremediation does not occur at rates
capable of controlling plume migration. … Microcosm experiments done in our laboratory from
soil collected at MTBE contaminated sites indicated MTBE-degrading microorganisms were
present at few sites. The presence of natural degraders does not, by itself, insure biodegradation
will be successful. Degraders need to be present in sufficient quantities and respond to
stimulation.”
(3) Spinnler et al. (2001a) also stated:
“MTBE degraders have not been observed to be as ubiquitous in the environment as BTEX
degraders. Even when present, MTBE degraders may not have sufficient population density to
degrade high concentrations of MTBE. Laboratory data indicates MTBE is an extremely poor
growth substrate for the organisms, so rapid population growth may not occur in contaminated
aquifers.”
These observations on the potential for MTBE attenuation through biodegradation under natural conditions
have not changed much over the years: For example, based on a review of the literature on MTBE and
TBA biodegradation in the subsurface, Schmidt et al. (2004) concluded:
“In general, compared with hydrocarbons present in gasoline, fuel oxygenates biodegrade much
slower, if at all. The presence of MTBE and related compounds in groundwater therefore
frequently limits the use of in situ biodegradation as remediation option at gasoline-contaminated
sites. Though degradation of MTBE and TBA in field studies has been reported under oxic
conditions, there is hardly any evidence of substantial degradation in the absence of oxygen.”
Schmidt et al. (2004) also added: “Reliable in situ degradation rates for MTBE and TBA under various
geochemical conditions are not yet available”.
Similarly, and in reference to assessing degradation rates for field conditions, the USEPA document on the
natural attenuation of MTBE (USEPA, 2005) cautions:
“Most studies in support of MNA do not provide a solid estimate of the rate of degradation of the
contaminant in the ground water along the flow path. Without a solid estimate of the rate of
degradation in the ground water, a conservative evaluation of the risk to a receptor is restricted to
the assumption that the contaminant does not degrade at all.”
Further, USEPA-OGWDW (2008) noted:
“MTBE has several properties that increase its persistence and mobility in the environment once
released” and “Once dissolved in ground water, it generally resists degradation and moves at
nearly the same velocity as ground water”.
At the same time, Youngster et al, (2008) stated:
93

“The prevalence of anoxic conditions in gasoline-polluted groundwater means that anaerobic
MTBE degradation will likely be a necessary process of remediation. MTBE degradation is a rare
process and the responsible organisms are sparsely distributed, even in contaminated sediments.”
Natural Attenuation and Transport, MTBE compared to BTEX: The fact that MTBE is poorly
attenuated in the unsaturated zone (the zone between the land surface and the water table) may cause it to
be the only contaminant detected in groundwater after a gasoline spill, as was observed in one case in
Maine (Garrett et al., 1986). Similar observations were noted during a field study in Wisconsin (McLinn
and Rehm, 1997) designed to evaluate the fate of residual hydrocarbons in the vadose zone. During 2 years
of monitoring, MTBE was the only gasoline constituent detected in the groundwater. In the same time
period, pore gas samples showed elevated carbon dioxide concentrations and low oxygen concentrations,
suggesting that the petroleum hydrocarbons were being degraded in the unsaturated zone above the water
table, before making it to the groundwater.
Similarly, because of its transport properties and poor natural attenuation, MTBE plumes tend to travel
farther and deeper than BTEX plumes (Table 6.1). In a survey of state experiences with MTBE and other
oxygenate contamination at LUST sites (NEIWPCC, 2003) 32 states indicated that MTBE plumes tend to
be longer than typical BTEX plumes, yet only 11 states require three-dimensional characterization of the
plumes. Less than half of the states said that they are taking extra steps to ensure that oxygenates do not
migrate beyond standard monitoring parameters.
Table 6.1: Summary of field observations comparing MTBE plumes to those of BTEX
Author

Sites Studied

Findings

Dernbach, 2000

Tahoe Basin

MTBE plumes spread at least 3 to 6 times the

distance of the largest BTEX plumes.
Salanitro et al., 2000

Port Hueneme, California

MTBE plume traveled more than 1,500m down

gradient from the spill source, compared to just 50m
for the BTEX components.
Thuma et al., 2001

Deer Park, New York

MTBE plume extended to about 8000 ft, compared

to approximately 700 ft for BTEX.
Wilson et al., 2002

Vandenberg Air Force Base

MTBE plume traveled 17 to 34 times the distance

in Central California

that the BTEX plume did (520 meters for MTBE
compared to 15-30 meters for BTEX).

Martienssen et al.,

2006

Leuna, East Germany

MTBE plume about 1700 m long and several

hundred meter wide. Benzene (longest of the BTEX)
plume less than 300 m long.

94

These observations are consistent with the scientific understanding of MTBE and BTEX transport in
groundwater (Johnson et al., 2000). For example, in the Tahoe basin, Dernbach (2000) noted that the
MTBE plumes spread at least 3 to 6 times the distance of the largest BTEX plumes. In comparison, the
MTBE plume at Vandenberg Air Force Base in Central California traveled 17 to 34 times the distance that
the BTEX plume did (Wilson et al., 2002). At the Port Hueneme site, California, the MTBE plume traveled
more than 1,500 meters (4,920ft) down gradient from the spill source, compared to just 50 meters (164ft)
for the BTEX components (Salanitro et al., 2000). In Leuna, East Germany, a MTBE plume extends about
1700 m (5,577ft) in length and several hundred meters wide. The corresponding benzene plume (longest of
the BTEX compounds) is less than 300 m (984ft) long. In comparison, BTEX plumes in California rarely
exceed 250 feet (76 meters) (Rice et al., 1995). In 1995, MTBE in the then nearly 7,000 ft long plume in
East Patchogue, N.Y. had already traveled approximately 1,000 ft farther than BTEX (Weaver et al., 1996).
In fact, because of its persistence and mobility in the subsurface, MTBE has been relied upon as a
“conservative tracer” to study the fate of BTEX constituents in the subsurface (Robbins and Gilbert, 2000;
Rifai et al., 1995), and to conduct forensic investigation of gasoline spills, including the estimation of the
timing or source of a gasoline spill (Davidson and Creek, 2000).
MTBE, TBA and Natural Attenuation: Another potential problem with MTBE biodegradation is the
formation of byproducts, particularly TBA, a chemical that is more toxic than MTBE (Keller et al., 2000).
TBA has been shown to accumulate, before being degraded at a rate slower than that of MTBE (Salanitro
et al., 1994; Bradley et al., 2001a, Wilson et al., 2002). In other instances, TBA persists or degrades very
slowly (Mormile et al., 1994; Hardison et al., 1997; Hyman et al., 1998; Vainberg et al., 2006; Skinner et
el., 2008). However, although TBA is sometimes observed to be a byproduct of MTBE biodegradation, its
mere presence should not be considered proof of natural attenuation or biodegradation of MTBE.
Moreover, as discussed previously, elevation in the TBA/MTBE ratio is not proof of natural attenuation or
biodegradation of MTBE either.

6.2.1 Site characterization and natural attenuation
Many MTBE contamination sites are poorly characterized. One problem stems from the fact that
monitoring networks were often developed with BTEX contamination in mind. However, MTBE and
BTEX plumes often require different levels of monitoring. The American Petroleum Institute (API, 2000)
explained:
“If a BTEX release at a given site predates an ether oxygenate [MTBE] release, then the BTEX
plume may be as long (or even longer) than the oxygenate plume. However, the older BTEX
plume is more likely to have been affected by changing flow directions or fluctuating water levels,
while the younger oxygenate plume is more likely to have experienced fewer variations in
groundwater conditions. Given these factors, the older BTEX plume may be wider and thicker,
and the younger oxygenate [MTBE] plume may be narrower and thinner. Thus, an ether oxygenate
[MTBE] plume may be more difficult to detect and delineate, even when its overall length is
similar to that of the associated BTEX plume.”
This characteristic of MTBE plumes has been noted by other authors as well:
95

“MTBE plumes are often narrow compared to BTEX plumes. Unless monitoring well networks
are spaced fairly closely across the width of the plume, it is possible to miss the MTBE plume
entirely as it slips through between monitoring points. During site characterization, closely spaced
direct-push sampling points at multiple depths placed at right angles to the long axis of the plume
can be used as a tool for locating a narrow plume and facilitating proper placement of permanent
monitoring wells.” (Ellis, 2000)

In addition, Fiorenza and Rifai (2003) noted that:

“One of the challenges in assessing natural attenuation at MTBEcontaminated sites has been due to the limitations of the site
characterization. For example, the monitoring well network at a site may
not be extensive enough, and it might lack depth-discrete sampling
intervals that would allow for the characterizations of MTBE and cocontaminants.”
Often, MTBE plumes may be so large that they have migrated offsite. Existing monitoring networks would
be inadequate to detect them as they spread. In fact, Mace et al. (1997) in their survey of more than 600
gasoline contamination sites concluded that:

“almost 40 percent of the [600 plus] sites [surveyed] have dissolved

hydrocarbon plumes that reportedly extend offsite” (Mace et al., 1997)
Note that the survey was conducted in the 1990s and already MTBE plumes had migrated offsite.
Similarly, in a study of 127 sites by Rifai and Rixey (2004) (also discussed in Rifai et al., 2003):

“The MTBE and benzene plumes at many of the sites had traveled off-site

and the existing monitoring networks did not capture the full extent of the
plume. During the course of the research, it was determined that the
monitoring well network at 91 sites [out of 127 examined] was not
sufficient for delineating the dissolved phase hydrocarbon plumes; in
almost every case this was true for both MTBE and benzene plumes.”
Clearly monitoring was inadequate.
Quite often, monitoring of MTBE plumes does not account for the vertical migration of contamination.
Very often, monitoring is limited to shallow depths and does not account for vertical contaminant migration
deeper into the aquifer. For example, deficiencies in plume monitoring were highlighted in a study of a
gasoline release in East Patchogue, New York, that resulted in a 5000 feet long benzene plume and a
similarly sized MTBE plume (due to its more recent release), where researchers concluded that:

96

“If the service station LUST site had been characterized in a traditional manner, without vertical
delineation, the estimated plume length and detachment of the MTBE plume would have been completely
misinterpreted; the benzene plume at the site would have been judged to be 3,500 feet shorter than actual
and the MTBE pulse would have been mischaracterized as below the State of New York’s level of concern.
[detected MTBE concentrations were as high as 9,600 ppb]” (Delaware Department of Natural Resources
and Environmental Control, August, 2000).

For all these reasons, Shih et al. (2004) stated:
“Clearly, plume length as defined is two-dimensional. The lack of depth-specific data and
other site-specific knowledge across the population of LUFT sites investigated in this
paper preclude evaluation of plume transport in the vertical direction. In areas of
significant recharge, this can bias the measurements toward shorter plumes, since a
typical monitoring well screened across the water table may fail to detect the leading
edge of the plume as it is deflected downward in response to the infiltration of recharge
from above. Further, fluctuating flow directions as well as errors in their determination
can result in monitoring well network configurations that create additional biases in
plume length measurement.”
Similarly, USEPA (2005):

“Conventional monitoring wells can provide an incomplete picture of the true distribution
of MTBE in ground water. If the screen of a monitoring well is long compared to the
thickness of the plume of contamination, it can sample the plume of contaminated ground
water and cleaner ground water above or below the plume, giving a false impression of
natural attenuation from one well to another. Long plumes of MTBE may dive below the
screens of monitoring wells altogether. Any evaluation of natural attenuation between
monitoring wells should consider the screened intervals of the wells, the depth interval
contaminated with gasoline (if that information is available), and the lithological features
sampled by the wells.”
In addition, MTBE plumes can migrate laterally in perched groundwater above the water table. In such cases, a
plume could migrate offsite before it even reaches the water table. Wells drilled to saturated groundwater on site
could miss the plume entirely.

6.2.2 Lessons Learned from Other Recalcitrant Groundwater Contaminants
The fact that MTBE has been shown to biodegrade in some studies does not justify reliance on natural attenuation as
a contaminant regulation mechanism in the environment. For perspective, contaminants such as DBCP, nitrate, PCE,
TCE, and TCP also have been known to biodegrade under a variety of conditions; yet, as discussed previously,
concentrations of these compounds have increased in deeper groundwater on a decadal time scale (Fogg et al.,
1998a; Dubrovsky et al., 1998). Some 30 years since it was banned from use and some 45 – 50 years after its first

widespread use in California, DBCP29, known to slowly biodegrade, is still one of the most commonly detected
contaminants in PWS wells at concentrations above its MCL (CDHS database, March, 2006). The mere existence of
conditions under which a compound might biodegrade neither guarantees, nor implies, that the contamination will
generally be eliminated or reduced significantly through biodegradation.

7

Use of MTBE and other Oxygenates in Gasoline

The following information on the use of MTBE and other common oxygenates [ethyl tertiary butyl ether (ETBE),
tertiary amyl methyl ether (TAME) and ethanol] as gasoline fuel additives in the United States comes from
information available through the U.S. Energy Information Administration (EIA)30:

Oxygenates were first required in 39 areas of the country with the 1992 wintertime oxygenated
gasoline program, arising from the Clean Air Act and designed to reduce emissions of carbon
monoxide. This program required oxygen at a minimum level of 2.7 percent by weight (which by
volume is equivalent to 15.0 percent MTBE or 7.4 percent ethanol). California obtained a waiver
from this minimum 2.7 percent oxygen level because of concerns it would increase automotive
emissions of oxides of nitrogen (NOx). California’s requirement was amended to require between
1.8 and 2.2 percent by weight oxygen. The reformulated gasoline program, aimed at reducing
emissions of ozone-forming VOCs, nitrogen oxides, and other air pollutants, was implemented
year round in certain areas of the United States, and during the summer months in other parts,
beginning in December 1994. The program required a minimum of 2.1 percent oxygen by weight
(which by volume is equivalent to approximately 11.7 percent MTBE or 5.8 percent ethanol).
MTBE usage as an octane booster grew in the early 1980's in response to the phasing out of lead
from gasoline, and an increase in demand for gasoline. The petroleum industry responded to the
oxygenated and reformulated gasoline programs by amending gasoline with various oxygenates,
principally MTBE and ethanol (Table 7.1). The use of MTBE increased significantly from about
83,000 in 1990 to about 161,000 barrels per day in 1994. MTBE use in gasoline increased to an
estimated 256,000 barrels per day by 1997, following the reformulated gasoline program in
December 1994 (Table 7.1).
In 2005, the New Jersey Legislature passed legislation that would ban the sale of gasoline containing more than
0.5% MTBE in the State, effective on January 1, 2009.31

29

Reported half lives for conditions similar to in-situ groundwater range from 6.1 to 141 years (Burow et al., 1999).
Also see US EPA Technical Fact Sheet on Dibromochloropropane, DBCP, available at:
http://www.epa.gov/OGWDW/dwh/t-soc/dbcp.html
30
http://www.eia.doe.gov/emeu/steo/pub/special/mtbe.html#Who%20gets%20gasoline%20with%20oxygenates last
accessed 2/15/07
31

New Jersey Department of Environmental Protection http://www.state.nj.us/dep/dsr/mtbe/mtbereport.htm last accessed Oct. 26, 2012.

Table 7.1: Estimated Oxygenate Use in 1997 by Petroleum Administration Defense District (PADD)
(expressed in thousand barrels per calendar day).

Oxygenated
OxygenatedReformulated

PADD 2

PADD 3

MTBE

109.80

4.20

25.81

-

97.23

237.03

ETBE

0.00

0.00

0.00

-

0.34

0.34

TAME

8.66

0.02

2.97

-

2.94

14.58

Ethanol

0.74

22.54

0.00

-

0.95

24.22

MTBE

Reformulated

PADD 1

PADD 4

PADD 5

Total

-

0.00

0.00

0.34

0.46

0.80

ETBE

-

0.00

0.00

0.00

0.00

0.00

TAME

-

0.00

0.00

1.08

0.00

1.08

Ethanol

-

8.96

1.72

2.48

5.14

18.29

MTBE

17.83

-

-

-

0.10

17.93

ETBE

0.00

-

-

-

0.00

0.00

TAME

0.00

-

-

-

0.00

0.00

Ethanol

1.54

-

-

-

1.22

2.76

Data From Energy Information Administration
URL: ftp://ftp.eia.doe.gov/pub/forecasting/steo/special/rpt/rfg4.xls
PADD locations are illustrated in Figure 7.1.

Figure 7.1. Petroleum administration defense districts (PADDs) (From Energy Information
Administration).

8

MTBE in Groundwater Resources

Due to its widespread use, MTBE is now “… ubiquitous and can be detected at very low concentrations nearly
anywhere on the Earth’s surface (Science and Technology Review, May 2003).” Widespread detections of MTBE in
groundwater resources were seen across the nation (Report by the Blue Ribbon Panel on Oxygenates in Gasoline,
1999). In the Executive Summary and Recommendations section of the Report by the Blue Ribbon Panel on
Oxygenates in Gasoline (July 1999), the authors state:

“At the same time, the use of MTBE in the program [the Federal Reformulated Gasoline
Program] has resulted in growing detections of MTBE in drinking water, with between 5% and
10% of drinking water supplies in high oxygenate use areas showing at least detectable amounts
of MTBE. The great majority of these detections to date have been well below levels of public
health concern, with approximately one percent rising to levels above 20 ppb [parts per billion
which is the same as µg/l]. Detections at lower levels have, however, raised consumer taste and
odor concerns that have caused water suppliers to stop using some water supplies and to incur
costs of treatment and remediation. The contaminated wells include private wells that are less
well protected than public drinking water supplies and not monitored for chemical
contamination.”

8.1 The Onset of MTBE Contamination in New Jersey
One of the first published reports of MTBE contaminating a community water supply well comes from the
Rockaway Township in north-central New Jersey. The town of 20,000 people relied primarily on three wells drilled
into unconsolidated glacial deposits at depths of 100 to 200 ft (McKinnon and Dyksen, 1984). In November 1979,
TCE was discovered in two of the wells, which resulted in those wells being taken out of service. The remaining
well, the largest of the three (pumping at 1,000 gpm) “was not affected and remained in service” (McKinnon and
Dyksen, 1984). However, in October 1980, the fuel oxygenates, diisopropyl ether (DIPE) and MTBE “were found
at sufficiently high levels to cause severe taste and odor problems in the water…As a result, numerous taste and
odor complaints were received from the township’s residents” (McKinnon and Dyksen, 1984).
To treat the groundwater, the township decided to use granular activated carbon (GAC). Based on the TCE levels
detected in the wells initially, the GAC beds were projected to last six to eight months before it was necessary
regenerate the beds. However, “after three months of operation, the concentration of DIPE in the contactor effluent
was 14 µg/L, MTBE was 23 µg/L and TCE was not detected. Therefore, it was necessary to replace the carbon
every two months at a cost of about $32,000. By the end of 1981, the carbon was being replaced at a rate of about
once every four to six weeks” (McKinnon and Dyksen, 1984). As a result, an aeration system was added to the
treatment train, to help strip the dissolved contaminants form the water before it was passed through the GAC beds.
This helped cut the cost of treatment significantly as it added to the lifetime of the GAC beds. However, this case
highlighted the difficulty in treating MTBE contaminated drinking water.

8.2. USGS studies
The United States Geologic Survey (USGS) has conducted numerous studies on water quality across the nation.
Several of those studies addressed contamination in the Eastern United States, and in New Jersey in particular. In
many of those studies MTBE was found to be among the most frequently detected volatile organic compounds
(VOCs) as explained below.

•

“Occurrence of nitrate, pesticides, and volatile organic compounds in the KirkwoodCohansey aquifer system, southern New Jersey” (Stackelberg et al., 1997). This was
one of the earliest reports to document the widespread detections of MTBE in New Jersey
groundwater. As part of the National Water-Quality Assessment (NAWQA) program,
USGS scientists sampled a network of 72 shallow monitoring wells in the KirkwoodCohansey aquifer system in the Glassboro area of southern New Jersey. The study area
was chosen partly because of its heavy reliance on groundwater, and its rapidly
increasing population. The “wells were randomly distributed among agricultural, urban,
and undeveloped areas to provide data representative of chemical conditions of
groundwater underlying each of these land-use settings.” The wells were sampled in 1996
for a range of nutrients (e.g. nitrate), pesticides and VOCs. Results indicated that MTBE
was the second (after chloroform) most frequently detected VOC (at a detection level of
0.1 µg/L) with a detection frequency of 44.4%. Though most VOC detections were low
concentrations, MTBE also had the highest maximum concentration detected at 43.8
µg/L. It is also important to note that benzene (the most mobile and most toxic of the
gasoline BTEX group) was not among the top 17 VOCs detected in the sampling event,
despite a long history of use as part of gasoline.

•

“A Review of Literature for Methyl tert-Butyl Ether (MTBE) in Sources of
Drinking Water in the United States” (Delzer and Ivahnenko, 2003b). In a summary of
the available literature on MTBE detections in water supplies across the country, USGS
scientists summarized findings reported from 13 different states including New Jersey. In
one study, domestic supply wells were sampled (on a random basis) from four different
regions within the Highlands, Piedmont, and Coastal Plain Provinces of New Jersey.
With a reporting limit of 0.1µg/L, MTBE was detected in all four regions with an overall
detection frequency of 35.6 percent (37 out of 104 wells sampled). The concentrations

were relatively low ranging from 0.1 to a maximum of 30.2 µg/L (Delzer and Ivahnenko,
2003).
•

“Quality of Water from Crystalline Rock Aquifers in New England, New Jersey,
and New York, 1995-2007” (Flanagan et al., 2012). This study assessed data collected
as part of NAWQA. The data included sampling of 26 domestic wells in a small area
with crystalline bedrock in northern New Jersey. The area is part of the “Long IslandNew York-New Jersey Coastal Drainages (LINJ) well network.” The wells in New Jersey
were sampled in 1997. Out of the 26 wells sampled in the area, 6 wells (23%) had MTBE
detections. Five wells (19%) had MTBE detections at concentrations between 1 and 13
µg/L, and one well had a detection above 13 µg/L. Flanagan et al (2012) noted:
“MtBE was detected more often in water samples with apparent ages of less than
25 years than in water samples with apparent ages greater than 25 years. This
finding is consistent with the time period of high MtBE use in areas in the United
States where reformulated gasoline was mandated.”

Other regional Studies
As expressed above, the United States Geological Survey (USGS) has conducted several studies on the occurrence
of MTBE in water sources across the nation. In these studies the data are collected from a wide array of sources and
cover public and private supply wells, rural, urban, deep, and shallow, and from throughout the United States. In
many places, these studies have found MTBE to be either the most commonly detected, or the second most
commonly detected volatile organic compound (VOC). The widespread detection of MTBE in water sources in the
United States is clearly expressed in several USGS reports
•

“Occurrence and Distribution of Methyl tert-Butyl Ether and Other Volatile Organic Compounds in
Drinking Water in the Northeast and Mid-Atlantic Regions of the United States, 1993-98” (Grady and
Casey, 2001). In this report, USGS scientists examined the presence of VOCs in data supplied by “2,110
randomly selected community water systems (CWSs) in 12 Northeast and Mid-Atlantic States,” including
New Jersey, between 1993 and 1998. The study found MTBE to be the most frequently detected VOC in
drinking water after the trihalomethane compounds (THMs), compounds that are “potentially formed
during the process of disinfecting drinking water with chlorine”.
The study noted that the State with the highest MTBE detection frequency in drinking water was New
Jersey, with a detection frequency of 21.5 percent. While VOC data spanned 6 years, the MTBE data
collected for New Jersey was limited to two years (1997-1998). Grady and Casey (2001) noted: “With
detections reported by 21.5 percent of randomly selected CWSs with data for MTBE, New Jersey reported
the greatest frequency of MTBE in drinking water at concentrations at or above 1.0 µg/L, and Virginia (1.3

percent) reported MTBE least frequently.”
This USGS study also found that: “The frequency of MTBE detections in drinking water is significantly
related to high-MTBE-use patterns. Detections are five times more likely in areas where MTBE is or has
been used in gasoline at greater than 5 percent by volume as part of the oxygenated or reformulated
(OXY/RFG) fuels program.” In this study, it was estimated that: “Thirty-four percent of the population
served by randomly selected CWSs (community water systems) with data for MTBE may have been
exposed to the gasoline additive in their drinking water.”
•

USGS Fact Sheet entitled: “MTBE and other volatile organic compounds- New findings and
implications on the quality of source waters used for drinking-water supplies” (Zogorski et al., 2001).
In this report, it was concluded that: “MTBE was detected in about 5 percent of ground-water samples
collected by NAWQA [National Water Quality Assessment] across the Nation.” …“Specifically, MTBE
was detected in about 4 percent of community systems serving less than 10,000 people, and in nearly 15
percent of systems serving greater than 50,000 people.”

•

“Occurrence and Status of Volatile Organic Compounds in Ground Water from Rural, Untreated,
Self-Supplied Domestic Wells in the United States, 1986-99” (Moran et al., 2002). This study examined
data collected from 1,926 rural domestic wells from 39 states between 1986-99. Of 55 possible VOCs,
MTBE again was found to be the second most frequently detected (following trichloromethane), with a
detection frequency of 2.2% (compared to 0.3% for benzene).

•

“A National Survey of Methyl tert-Butyl Ether and Other Volatile Organic Compounds in DrinkingWater Sources: Results of the Random Survey” (Grady, 2003). In the random survey, water samples
collected from 954 CWSs from all 50 states, Native American Lands, and Puerto Rico, were tested for the
presence of 66 VOCs, including MTBE and 3 other gasoline oxygenates. The study was conducted between
May 1999 and October 2000. The results showed MTBE detections in 8.7% of the samples (compared to
0.3% for benzene), with MTBE being the second most detected VOC after chloroform, (a THM).

•

“Occurrence and Temporal Variability of Methyl tert-Butyl Ether (MTBE) and Other Volatile
Organic Compounds in Select Sources of Drinking Water: Results of the Focused Survey” (Delzer
and Ivahnenko, 2003). In the focused survey, water samples collected from 134 CWSs that were suspected
or known to contain MTBE, were analyzed for the presence of the 66 VOCs, as above. The study was
conducted between September 1999 and April 2001. The results showed MTBE to be the most frequently
detected VOC, with a detection frequency of 55.5% (compared to 7.6% for benzene).

•

“Water Resources Data New York Water Year 2002” (Spinello et al, 2003). A USGS report on the
water resources data for Long Island included sampling of 50 wells in the four aquifers (Glacial, Jamaico,
Magothy, and Lloyd) in Kings and Queens counties, with analysis “for 275 organic and inorganic
constituents to assess the ground water as a potable supply” (Spinello et al., 2003). Results showed MTBE
to be the “most frequently detected contaminant” at 76% detection frequency.

•

“Volatile organic compounds in the Nation’s Groundwater and drinking- Water Supply wells:
Circular 1292” (Zogorski et al., 2006). In this latest USGS study, MTBE was again found to be the second
most frequently detected VOC (the first being chloroform) in samples collected from domestic (private)
and public drinking water supply wells.

•

“Occurrence of Volatile Organic Compounds in Aquifers of the United States”
(Carter et al., 2008). A study by USGS summarizing the results of analysis for volatile
organic compounds in ambient groundwater in aquifers across the United States. The
study covered analysis conducted between 1985 and 2002, from 3,498 wells in 98
different aquifer studies. The source of the data included “(1) NAWQA sampling of
ground water in aquifer studies during 1993-2002, and (2) data from prior USGS studies
and from other Federal, State, and local agencies that sampled ground water in aquifer
studies during 1985-1997” (Carter et al., 2008). Each study was based on random
sampling, with no specific land-use type being targeted for sampling. Samples were
analyzed for 55 volatile organic compounds at assessment levels of 0.02 and 0.2 µg/L.
The results of the analysis showed MTBE was the third most frequently detected VOC (at
assessment level of 0.2 µg/L) after chloroform and PCE. The study also (Carter et al.,
2008) concluded that New Jersey had one of the highest detection frequencies of VOCs
in wells (57-60%) emphasizing the vulnerability of the aquifer system to contamination.
Carter et al (2008) noted:
“The fate of VOCs in ground water is largely dependent on their persistence
under the conditions present in the aquifer. VOCs that are persistent in water are
more likely to be detected in ground water because they can travel greater
distances from their source before degradation and dilution occur.”

Collectively these studies emphasize the vulnerability of drinking water resources to contamination by MTBE,
particularly in light of its relatively short history of production and use. Consider for example, that chloroform was
first used as an anesthetic in 1847, and first used for chlorination of a water supply system in the USA in 1908,
while MTBE was first approved for use as an octane booster by the USEPA in 1979 (Zogorski et al., 2006). Given
the relatively short history of use of MTBE in the United States, such widespread detections of MTBE in water
sources nationwide is alarming.
In testimony before the U. S. House of Representatives (November 1, 2001), Robert Hirsch, the Associate Director
for Water for the USGS stated that:

“Through our interpretations of this large data set we have also determined that low-levels of

MTBE are detected in about 1 out of 5 wells in MTBE high-use areas. Although we do not expect
to see great change in these results over time, we recognize that there may be a delay in the
detection of MTBE in some wells-particularly those that are deeper and may be farther from the
source of contamination. MTBE is the second most frequently detected volatile organic
compound (VOC). Chloroform, a drinking-water disinfection by-product and a commercial
solvent, is the most frequently detected VOC.”
In other words, MTBE, a compound whose widespread use only began relatively recently, was already detected in
20% of wells in high-use areas and was the second most frequently detected VOC in 2001. Further, this
contamination was expected to migrate deeper over time, eventually impacting deep water supply wells. These facts
suggest that we now are only beginning to see the initial developments of potential groundwater quality impacts due
to widespread MTBE contamination. I must reiterate that our past experience with other groundwater contaminants
demonstrates that the full impacts from compounds like MTBE evolve over a decadal time scale, not over a mere
few months or years, in groundwater basins like those upon which California and other states depend for drinking
water supply.

9

Ethanol

Alcohols were first added to gasoline as octane boosters in the 1930s, and ethanol blended gasoline was sold in the
Midwest in the 1930s and 1940s (Moyer, 2003). The use of ethanol as an alternative, or renewable, energy source
was stimulated in the mid 1970s, in response to the oil embargoes of 1973 and 1979 (CRS, 2003). More recently,
with the introduction of the Clean Air Act Amendments of 1990, ethanol was also used as a fuel oxygenate, though
mostly in the Midwestern states (Table 7.1). In 1998 it was the “second most common” fuel oxygenate present in
about 15% of oxygenated gasoline (USEPA 2004). With the ban or phase-out of MTBE in many States, the
production of ethanol, a replacement oxygenate, has risen sharply, doubling in a matter of few years (Zhang et al.,
2006). Yet, despite a long history of use, surprisingly little field data is available on environmental impacts of
ethanol-blended gasoline.
As a gasoline oxygenate, less ethanol by volume (compared to MTBE) is needed to achieve the oxygen
requirements necessary to meet the Amendments to the Clean Air Act. For Oxygenated Fuel (2.7% by weight
oxygen) 7.3% ethanol by volume would be added to gasoline, compared to concentrations of up to 15% by volume
MTBE (USEPA 2008). For Reformulated Gasoline (2% by weight oxygen) 5.4% ethanol by volume would be
added to gasoline, compared to 11% by volume MTBE (USEPA 2008).
Some recent research has suggested that at some sites, the use of ethanol might increase groundwater contamination
relative to the use of conventional gasoline. However, and as will be discussed next, studies have concluded that the
impacts from ethanol to groundwater are expected to be “less problematic” (NEIWPCC and NESCAUM, 2001) and
“more manageable” (Powers et al., 2001b) compared to MTBE, and that ethanol is “an acceptable substitute” to
MTBE (Dakhel et al., 2003).

9.1

Properties of Ethanol in Soil and Groundwater

Ethanol is miscible in water, and therefore tends to partition preferentially into water from gasoline. From its pure
phase in contact with water, ethanol can reach aqueous concentrations that are many times those of the maximum
aqueous concentration attainable by MTBE. Once in water, ethanol is considerably less volatile than MTBE.
Laboratory and field studies indicate that ethanol is easily, and preferentially (relative to BTEX), biodegraded by
indigenous microorganisms in groundwater under different conditions (Corseuil and Alvarez, 1996; Alverez et al.,
2001; Powers et al. 2001b; Deeb et al., 2002; Ruiz-Aguilar et al., 2002; Mackay et al., 2006; Zhang et al. 2006).
Because of its ability to biodegrade easily, ethanol itself “is unlikely to be a significant groundwater pollutant”
(Powers et al., 2001b). For example, only 5 states require sampling and analysis for ethanol in groundwater at
gasoline release sites, compared to 31 states that require sampling and analysis for MTBE (NEIWPCC, 2003). States
that use mostly ethanol as oxygenate, such as Minnesota and North Dakota, consider the risks to be low because
ethanol degrades easily as indicated by a survey of the 50 states experience with oxygenates at LUST sites,
conducted by the New England Interstate Water Pollution Control Commission (NEIWPCC, 2003). For example the
responses from both Minnesota and North Dakota indicated that the oxygenate used in those states (ethanol) is “not
perceived to be a problem”. The Minnesota response indicated “the human risks associated with ethanol are very
low. Also, ethanol degrades very quickly.” North Dakota’s response was similar: “ethanol will degrade naturally if
released into the environment.”

9.2

Cosolvency and Mobility Effects of Ethanol on BTEX

At concentrations exceeding those required for RFG, ethanol in gasoline can have concentration-dependent cosolvency and mobility effects. Co-solvency effects are thought to arise from a reduction in the polarity of water,
effecting the partitioning of BTEX compounds between NAPL, water, soil, and vapor phases and resulting in higher
equilibrium, aqueous concentrations of BTEX (Heermann and Powers, 1998; Da Silva and Alverez, 2002).
Research indicates that significant cosolvency effects are not likely to arise from a spill of ethanol oxygenated
gasoline (up to 10% by volume ethanol (Da Silva and Alverez, 2002; Heermann and Powers 1998; Powers and
Heermann, 1999). However, higher concentrations, e.g., 50% by volume ethanol in gasoline, can lead to the free
migration of benzene, toluene, and o-xylene (BToX) within the local groundwater (Da Silva and Alverez, 2002).
This raises concerns regarding neat ethanol spills, e.g., from bulk storage facilities, in the presence of other
petroleum hydrocarbons (Powers and McDowell, 2001), like the one that occurred at the Pacific Northwest Terminal
site spill described later.
Ethanol in gasoline can also affect the mobility of BTEX compounds through a reduction in the retardation of BTEX
and the surface tension of water (Heermann and Powers, 1998; Da Silva and Alverez, 2002; Powers and McDowell,
2001). These effects are seen as having potentially positive and negative impacts on groundwater quality. On the one
hand, increases in NAPL mobility in the unsaturated zone may affect leaching of contaminants to groundwater. On
the other hand, increased mobility would likely improve the efficiency of some remediation efforts (Powers and
McDowell, 2001). Even in the presence of cosolvency and mobility effects, both ethanol and BTEX compounds are
expected to attenuate through biodegradation as explained in the following section.

9.3

Inhibitory effects of Ethanol on the Biodegradation of BTEX Compounds

Microbial degradation appears to be the dominant natural attenuation mechanisms for ethanol and BTEX in
groundwater (Alverez et al., 2001). To carry out the biodegradation of ethanol and BTEX, indigenous
microorganisms rely on other constituents referred to as electron acceptors. Because ethanol is more easily
biodegraded than BTEX, the preferential biodegradation of ethanol can interfere with the biodegradation of BTEX
compounds by depleting electron acceptors used for BTEX biodegradation (Corseuil et al., 1996). Still, effects of
the presence of ethanol on BTEX degradation can vary. For example, laboratory experiments have found that
ethanol can actually enhance toluene degradation in the presence of excess electron acceptors (Alverez et al., 2001).
Similarly, ethanol, at relatively low concentrations, can enhance benzene degradation under denitrifying conditions
(Wu et al., 2009).
As discussed in the following sections, field and modeling studies suggest that for a large enough spill that results in
significant ethanol and BTEX groundwater plumes, the effects caused by the preferential biodegradation of ethanol
in groundwater can either increase the length of, or prolong the duration of, BTEX plumes (Alverez et al., 2001;
Buscheck et al., 2001; Mackay et al., 2006). On the other hand, natural attenuation (biodegradation) in the
unsaturated zone, or locally in groundwater, may virtually prevent the spread of ethanol and BTEX contamination
from small liquid or vapor phase releases of ethanol-oxygenated gasoline (10% by volume) (Dakhel et al., 2003).
Furthermore, because ethanol and BTEX are relatively easy to biodegrade, engineered remediation (introduction of
oxygen, nutrients) can be very effective at attenuating contamination and preventing its spread to sensitive receptors
(Heaston et al., 2010).
The following field and modeling studies illustrate some of the effects of ethanol addition to gasoline that are
mentioned above.

9.4

Field Studies

Small Volume Releases (Dakhel et al., 2003)
To mimic a small-volume gasoline spill, a controlled experiment involving the release of an artificial gasoline
containing MTBE and ethanol (at 5% by weight of each) was performed in a lysimeter having a 2.3 m thick vadose
zone of alluvial sand overlying a gravel aquifer. The experiment consisted of three sequential phases. In the first
phase of 41 days, no recharge water was applied to the surface. During this phase, MTBE and volatile gasoline
hydrocarbon vapors migrated by diffusion to, and accumulated in, groundwater. Ethanol, on the other hand,
naturally attenuated in the vadose zone and did not accumulate in groundwater. During the second phase of 30 days,
5mm/day recharge was applied to the surface over a period of 30 days to simulate infiltration due to a rain event.
During this phase, MTBE, ethanol, and soluble gasoline constituents (e.g., BTEX) were flushed to the groundwater.
During the third phase of 120 days, no recharge was applied. Once in groundwater, ethanol, benzene and other
detectable petroleum hydrocarbons (with the exception of isooctane) were rapidly biodegraded. More than 99% of
the ethanol disappeared. In contrast, there was no evidence of MTBE biodegradation. MTBE reached peak
concentrations of more than 250,000 µg/L in groundwater and persisted at concentrations greater than 125,000 µg/L

for the duration of the experiment lasting about 6 months. Higher MTBE concentrations would be expected to result
from a spill of oxygenated gasoline with 11% by volume MTBE. The study’s authors concluded that: “groundwater
quality will be affected considerably less at small volume release sites when MTBE-free gasoline or gasohol
[ethanol blended gasoline] are spilled” and that “MTBE-free gasoline would be less harmful for groundwater
resources and that ethanol is an acceptable substitute.”
These results have important implications for the small volume vapor- and liquid-phase releases that beset USTs and
UST systems, including upgraded 1998 compliant ones. Ethanol in small volume releases at a gasoline station is
likely to bioattenuate significantly in the vadose zone and, in the event that it reaches the water table, locally within
groundwater. In contrast, MTBE is expected to migrate to, and persist in, groundwater.

Neat Ethanol Spill Site at the Pacific Northwest Terminal Refinery (Buscheck and
O’reilly, 200132; Powers and McDowell, 2001)
In March 1999, a 19,000-gallon (about 72,000 liters) release of neat ethanol from an above ground storage tank at
the Pacific Northwest Terminal Refinery produced an ethanol plume that ultimately extended more than 250 ft (76
m) down-gradient. While the ethanol plume appeared to degrade, the spill affected NAPL behavior and aqueous
phase benzene concentrations of the preexisting hydrocarbon contamination on site. Benzene concentrations
increased by as much as a factor of 15 and the LNAPL thickness in many of the wells increased. Both observations
may have been due to cosolvency effects and the inhibition of biodegradation due to the spill of (pure) neat ethanol.

Transport of Ethanol in a Sand and Gravel Aquifer (Zhang et al., 2006).
This study monitored plumes originating from an injected pulse containing approximately 220 mg/L ethanol and 16
mg/L bromide in a sand and gravel aquifer for 2.5 months. Observations suggest that ethanol migration was
attenuated only through biodegradation. Ethanol rapidly degraded with a computed half-life of 2.2 days.

Site 60, Vandenberg Air Force Base, CA (Mackay et al., 2006)
Two experiments were performed side-by-side, simultaneously, over a period of 9 months in an aquifer (with sulfate
reducing conditions) underlying a former fuel station to evaluate the effects of ethanol on the biodegradation of
BToX (benzene, toluene and oxylene). On one side, BToX was continuously injected; on the other side, BToX with
ethanol was continuously injected. The BToX plumes on both sides migrated approximately the same distance and
then retracted. The BToX plumes on the side with no ethanol retracted at a significantly greater rate, and further,
than the BToX plumes on the side with ethanol. The researchers hypothesized (1) that the plumes reached the same
length initially due to an acclimation period before biodegradation began, and (2) that the slower retraction of BToX
on the ethanol side was due to the preferential biodegradation of ethanol and associated depletion of dissolved
sulfate, leading to methanogenic/ acetogenic conditions less effective for biodegradation of BToX. These results
provide field evidence for predicted increases in BTEX plume lengths as a result of ethanol addition to gasoline.

32

http://www-erd.llnl.gov/ethanol/proceed/terminl.pdf Buscheck and O’reilly, 2001 last accessed 2/27/07

LUFT sites in Iowa and Kansas (Ruiz-Aguilar et al., 2003).
This study compared the lengths of plumes originating from ethanol-oxygenated gasoline (10% ethanol by volume)
and conventional gasoline LUFT sites in Kansas and Iowa, respectively. The authors concluded that plumes for
benzene in ethanol-containing gasoline were significantly longer than plumes for benzene in conventional gasoline.
The same relationship was not found for toluene plumes. Median benzene plume lengths were found to be 36%
longer for ethanol-containing versus conventional gasoline sites. These data are consistent with the expected
inhibitory effects of biodegradation indicated by the modeling studies discussed below. Even so, caution is
warranted when comparing plumes lengths between different geologic settings and from sites that are not well
monitored.

9.5

Modeling Studies

Researchers have developed mathematical models aimed at studying the effects of ethanol in gasoline (10% by
volume) on plume behavior and the size of BTEX groundwater plumes (Dooher, 1999; McNab et al., 1999; Molson
and Barker, 2000; Powers et al., 2001a; McNab, 2001; Molson et al., 2002; Deeb et al., 2002). Overall, these
models predict increases in BTEX plume lengths due to the addition of ethanol, a result consistent with some of the
field studies discussed above. Maximum benzene plume lengths originating from conventional gasoline spills at
LUFT sites are typically several hundred feet in length (Fogg et al., 1998b; Happel et al., 1998). Results from these
models predict that ethanol may increase benzene plume lengths from less than 10% to as much as 150% depending
on the modeling approach and specified subsurface conditions. The approaches used in many of these studies are
generally considered to be conservative in that they rely on a number of simplifying assumptions that, if incorrect,
would tend to overestimate the effects of ethanol on BTEX plume lengths (Powers et al., 2001a).
In addition, in all of these modeling studies it is assumed that ethanol and BTEX compounds freely migrate as a
homogenous mixture from the unsaturated-zone to the water table. However experiments suggest that in the
unsaturated zone, ethanol migration is preferentially attenuated relative to BTEX (Powers and McDowell, 2001).
This phenomenon could increase the time available for the biodegradation of ethanol in the unsaturated zone and
decrease the rate at which ethanol leaches to the saturated groundwater. In this case, less ethanol would reach
groundwater leading to potentially smaller effects on the lengthening of BTEX plumes than predicted by the
modeling studies discussed above. Accordingly, Powers and McDowell (2001) state:
“Attempts to mathematically model the overall fate of ethanol and BTEX in the subsurface must

incorporate the mechanisms seen in these experiments in order to more accurately depict the
mechanisms that affect the initial distribution of ethanol and gasoline in the unsaturated zone. The
results presented here suggest that the retention of ethanol in the unsaturated zone may be
significant.”

9.6

Summary

As discussed previously, the most important factor contributing to the differential fate of MTBE and ethanol in the

subsurface appears to be biodegradability. The poor biodegradability of MTBE allows for long persistent plumes,
exposing larger volumes of water to contamination. A detailed review of the health and environmental impacts of
the addition of ethanol to gasoline in the Northeast (NEIWPCC and NESCAUM, 2001), jointly prepared by the New
England Interstate Water Pollution Control Commission (NEIWPCC) and the Northeast States for Coordinated Air
Use Management (NESCAUM, 2001) concluded with recommendations for removal of the oxygen mandate for
reformulated gasoline. However, in the absence of a removal, the report concluded that: “from a water quality
protection standpoint, the potential adverse impacts associated with ethanol are clearly less problematic than for
MtBE-blended gasoline.” (NEIWPCC and NESCAUM, 2001). Scientists at Lawrence Livermore National
Laboratory (Powers et al., 2001b) reached similar conclusions: “given the choice between MTBE and ethanol, and
with our current level of understanding, it appears that the groundwater resource impacts associated with the use of
ethanol will be less significant and more manageable than those associated with MTBE”.
Evidence from the state of Minnesota, where the use of ethanol in gasoline at 10% (E10) was mandated in 1997
(Toso, 2009), bear out the above predictions. In a recent article on the use of ethanol in gasoline in Minnesota, Toso
(2009) from the Minnesota Pollution Control Agency (MPCA) noted that: “after an informal review, the potential
differences between an E10 and an ethanol-free gasoline release were deemed minor and covered under existing
investigation and cleanup policy, which is based on groundwater plume delineation. No ethanol specific issues have
arisen from E10 releases since then.” (Toso, 2009). Furthermore, it should be noted that although ethanol has been in
use in gasoline since the 1930s, to my knowledge, there is no evidence to date that it has posed a problem as a
contaminant in groundwater drinking supplies.
Based on all of the above, and given the dispersed nature of fuel contamination, it is clear that ethanol (at 5.4%
requirement in reformulated gasoline) would have been a much better choice as an oxygenate in gasoline, from a
groundwater protection standpoint.

10

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11 Glossary of Terms
Acclimation period: The time elapsed between the introduction of contaminant(s) and observation of microbial
activity (biodegradation).
Activated Carbon: A highly adsorbent form of carbon used to remove odors and toxic substances from liquid or
gaseous emissions.

Activated Sludge: sludge particles (flocs) produced by the growth of large groups of organisms
(mostly bacteria, but also fungi and protozoa) during the secondary treatment process of
wastewater. This process involves agitation of the liquid and aeration, to maintain high oxygen
concentration to maximize microbial activity.
Alluvial aquifer: A geologic formation of alluvium (materials such as clay, silt, sand, and gravel deposited by
streams or rivers) that can yield significant quantities of water to wells or springs.
Bioaugmentation: the addition of bacteria or other microorganisms that can biodegrade a specific organic
compound.
Biobarrier: a biologically active zone installed in the subsurface, in a direction perpendicular to that of the normal
flow of contaminated water, to allow for the biodegradation of the contaminant.
Biodegradation: biologically induced structural transformation of a chemical compound.
Bioremediation: the use of microorganisms in removing contaminants from water, soil or air.
BTEX: benzene, toluene, ethyl benzene and xylene (m-, o-, and p- xylene). These volatile organic compounds are of
particular concern in gasoline, because of their water solubility and toxicity.
BToX: benzene, toluene, and o-xylene.
CDHS: California Department of Health Services
CDWR: California Department of Water Resources
Cosolvency: An effect wherein the presence of more than one organic contaminant causes a change in the
partitioning behavior of one or more of these contaminants, as compared to the condition where they are present
alone.
CWS: community water supply (Johnson et al., 2000); community water system (Water-Resources Investigations
Report WRIR00-4228, 2001)
Diffusion: the movement of a chemical resulting the random motion of molecules.
Dilution: reduction of contaminant concentration by mixing with less-contaminated, or uncontaminated, water.
DLR: detection limit for reporting purposes (DLR) of 3 µg/L. The DLR is the level at which DHS is confident
about the quantity being reported. Results at or above the DLR are required to be reported to DHS; some
laboratories may report results at lower concentrations. As defined in the CDHS web page.
(http://www.dhs.cahwnet.gov/ps/ddwem/chemicals/MTBE/mtbeindex.htm)
Enzyme: a protein produced by a living organism, which functions as a biochemical catalyst, facilitating a specific
chemical reaction.

Fractured medium: A rock or clay formation which is mostly solid but contains cracks that can
transmit or store water.
GAC: granular activated carbon. A highly adsorbent form of carbon (possessing a high porosity
and high surface activity) typically used to remove organic contaminants from water.
Half-life: the time required for the initial concentration of a compound to be reduced by half, often used in reference
to radioactive decay, and sometimes applied to biodegradation and other reactions.
Humic substances: material derived from the degradation of plants algae and other organisms, which accounts for
much of the naturally occurring organic carbon33.
Inoculum: material that serves as a source of microbial cells, such as activated sludge, soil, or sediment
Intrinsic biodegradation: biodegradation of contaminants by indigenous microorganisms.
LUFT: leaking underground fuel tank
LUST: leaking underground storage tank
Lysimeter: a device used for measuring the percolation of water through soil, and for determining the soluble
constituents removed through drainage.
MCL: maximum contaminant level, refers to “enforceable regulatory standards under the Safe Drinking Water Act
and must be met by all public drinking water systems to which they apply” (from:
33

Scott et al., 1998

http://www.dhs.ca.gov/ps/ddwem/chemicals/MCL/mclindex.htm). In New York, the MCL for MTBE is 10 µg/liter.
In some States, a primary MCL is set on health-based standards, while a different secondary MCL is set on aesthetic
standards such as taste and odor. In California, the secondary MCL for MTBE is 5 µg/liter, while the primary MCL
for MTBE is 13 µg/liter.
Microcosm: a laboratory-scale model used to study the interactive relationships of microbial communities and the
ability to biodegrade contaminants, among other things.
Mineralization: the complete biodegradation of an organic compound into carbon dioxide, water, inorganic ions
and molecules, and possibly cellular material.
The MTBE Blue Ribbon Panel: “was created by a Charter from the Clean Air Act Advisory Committee (CAAAC)
to provide independent advice and counsel to EPA on policy issues associated with the use of MTBE and other
oxygenates in gasoline” (http://www.epa.gov/oms/consumer/fuels/oxypanel/blueribb.htm).
Monooxygenase: an enzyme that catalyzes the addition of one atom of molecular oxygen to a molecule to be
transformed.
PADD: petroleum administration defense district.
ppb: parts per billion US; parts per 1,000 million; microgram per liter, µg/l.
PWS: public water supply.
Recalcitrant: resistant to microbial degradation.
RFG: reformulated gasoline
SWRCB: State Water Resources Control Board
Trihalomethanes (THM): Disinfectant by-products formed when chlorine reacts with organic compounds in water.
These include trichloromethane (chloroform) dibromochloromethane, bromodichloromethane, and tribromomethane
(bromoform).
USEPA: United States Environmental Protection Agency.
USGS: United States Geologic Survey, defined on the web site (http://www.usgs.gov/) as: “Federal source for

science about the Earth, its natural and living resources, natural hazards, and the environment.”
UST: underground storage tank.
VOC: volatile organic chemical.

Appendix A: C.V. of Dr. Graham E. Fogg

Curriculum Vitae
Graham E. Fogg
Hydrology Program, Department of Land, Air, and Water Resources
Hydrologic Sciences Graduate Group
One Shields Avenue
University of California, Davis 95616
[phone] FAX: [phone]
E-mail: [e-mail]
ACADEMIC BACKGROUND
B.S. Hydrology, University of New Hampshire, 1975
M.S. Hydrology, University of Arizona, 1978
Ph.D. Geology, The University of Texas at Austin, 1986
HONORS
Phi Kappa Phi Honor Society
Geological Society of America Birdsall-Dreiss Distinguished Lecturer, 2002
Fellow, Geological Society of America
O.E. Meinzer Award, 2011
AREAS OF EXPERTISE
A. Contaminant transport in complex porous media; Geologic/geostatistical characterization of
subsurface heterogeneity
B. Numerical modeling of ground-water systems
C. Regional system hydrogeology: ground-water hydraulics, sustainability of groundwater quality,
groundwater and surface water interaction
PROFESSIONAL EXPERIENCE
A. Present Position: Professor of Hydrogeology, Hydrology Program, Department of Land, Air, and
Water Resources, University of California, Davis (July 1, 1996 - present). Vice Chair for
Hydrology 1998-2001.
B. Associate Professor of Hydrogeology, Hydrology Program, Department of Land, Air, and Water
Resources, University of California, Davis (March 15, 1989 - June 30, 1996).
C. Research Associate, Bureau of Economic Geology, The University of Texas at Austin. (Sept.
1985 - Jan. 1989).
Research multiphase flow hydrodynamics of hydrocarbon reservoirs in West Texas and stochastic
methods of characterizing reservoir interconnectedness. Research regional hydrogeologic
processes as related to potential impacts of lignite mining on water quality and quantity. Develop
improved methods of describing aquifer heterogeneity using geologic information.
D. Research Scientist Associate, Bureau of Economic Geology, The University of Texas at Austin
(January 1978 - August 1985).
Conduct research on hydrogeologic suitability of salt domes in the East Texas Basin for storage

of high-level nuclear wastes. Assemble and interpret hydrogeologic data bases; design and
supervise hydrologic test-drilling programs; construct detailed two and three dimensional groundwater flow models for analyzing near-dome flow systems and potential radionuclide pathways to
the biosphere; research new methods of modeling mass transport and of describing aquifer
heterogeneity.
PROFESSIONAL SOCIETIES
American Geophysical Union, Water Resources Division
Association of Ground Water Scientists and Engineers
International Association of Hydrogeologists
Geological Society of America
Groundwater Resources Association of California
SELECTED PROFESSIONAL ACTIVITIES (last 10 years)
Manuscript reviewer, Hydrogeology Journal, Ground Water journal, Water Resources Research,
Mathematical Geology, Advances in Water Resources, Transport in Porous Media, among others.
Proposal reviewer, National Science Foundation
Member, Peer Review College of the Danish Council for Strategic Research
Member, Committee to evaluate research performance of GEUS (Geological Survey of Denmark
and Greenland), Sept.-Nov., 2007
Co-Chair of Subsurface Characterization Panel, Dept. of Energy, Basic Research Needs Program in
Subsurface Science (to determine basic research needs for the future, concerning subsurface CO2
sequestration and nuclear waste isolation), 2007-08.
SELECTED UNIVERSITY ACTIVITIES
College of Agricultural and Environmental Sciences Academic Planning Committee, 2009.
Member, Search Committee for Director of John Muir Institute for the Environment
Chair, Search Committee for Associate/Full Cooperative Extension Specialist and Associate Full
Professor in Water Resources (Robert Hagan Endowed Chair), 2008 – 2009.
Member, delegation to University of Concepcion, Chile, to develop collaborative water programs,
March – present 2009.
Vice Chair for Hydrology, Department of Land, Air and Water Resources, 1998 – 2001.
Chair, Hydrologic Sciences Graduate Group, 1993 – 1998, 2006 – present.
Graduate Advisor, Hydrologic Sciences Graduate Group, 1991 – 1993, 1998 – 2000, 2001–present.
Graduate Advisor, Water Science Graduate Group, 1989 – 1991.

COURSES TAUGHT
The University of Texas at Austin Geology Dept.: Groundwater Modeling, 1987.
UC Davis: WSC 298, Contaminant Transport Processes, 1989 (graduate).
WSC 149B, HYD 145B, Groundwater Hydrology, 1990, 1991, 1992, 1993, 1994, 1996, 1997
(undergraduate).
HYD 10, Water and Power, 2001, 2003, 2004, 2006, 2007, 2008, 2009.
HYD 146, Hydrogeology and Contaminant Transport, 1998, 1999, 2000, 2001, 2003, 2004,

2007, 2008, 2009.
HYD 269, Numerical Modeling of Groundwater Systems, 1990, 1991, 1992, 1993, 1994,
1996, 1997, 1998, 1999, 2000, 2001, 2004, 2006, 2008 (graduate).
HYD 273, Introduction to Geostatistics, 1990, 1992, 1994, 1996, 1998, 2000, 2002, 2004,
2006, 2008 (graduate).

SELECTED SHORT COURSES AND WORKSHOPS DEVELOPED AND TAUGHT
“Modeling of Heterogeneity and Transport in Geologic Media: Transition Probability, Markov
Chains and Random-Walk Particle Methods (TProGS & RWHET)” June 22-27, 2009. University of
Copenhagen, Denmark.
“Hydrofacies Modeling with TProGS, a Geostatistical Approach based on Transition Probability and
Markov Chains.” February 6-9, 2006. U.S. Geological Survey, Sacramento State University,
Sacramento, CA.
“Hydrofacies Modeling with TProGS, a Geostatistical Approach based on Transition Probability and
Markov Chains.” October 20-24, 2003. Geocenter Copenhagen, Denmark.
"Principles of Groundwater Flow and Transport Modeling." 3-day short course.
Sponsor/Coordinator: University Extension, University of California, Davis. Taught one to two
times per year since 1998.
"The Groundwater Workshop." 1-day workshop. Sponsor/Coordinator: The Water Awareness
Committee of Monterey County, Inc. March 19, 1998. Salinas, CA.
"Applied Groundwater Hydrology: Principles, Measurements, and Interpretation." 2-day short
course. Sponsor/Coordinator: Cal/EPA Department of Toxic Substances Control, CA Groundwater
Resources Association and the University of California, Davis. May 19-20, 1998. Berkeley, CA.
"Introduction to Applied Groundwater Hydrology: Principles, Measurements, and Interpretation." 3day short course. Sponsor/Coordinator: University of California Davis. November 6-7, 1996;
Fresno County, CA; November 1997, Sacramento, CA.
SELECTED EXPERT TESTIMONY (public service and litigation)
•

Expert witness and testimony in Imperial Irrigation District vs. Coachella Valley Water District,
1987-1988.

•

Testimony to California Assembly Committee dealing with registration of groundwater
professionals. 1991.

•

Testimony to California Joint Legislative Audit Committee. “Is Groundwater Quality Falling
Through the Legislative Cracks?” February 9, 1999.

•

Testimony at CalEPA Public Hearing on MTBE in California, Diamond Bar, CA. Impacts of
MTBE on California Groundwater. February 19, 1999.

•

Testimony at CalEPA Public Hearing on MTBE in California, Sacramento, CA. Impacts of MTBE
on California Groundwater. February 23, 1999.

•

Testimony to EPA Blue Ribbon Panel on MTBE Issues. Impacts of MTBE on California
Groundwater. March 25, 1999.

•

Expert witness and testimony in South Lake Tahoe Public Utility District vs. ARCO et al., 20002001.

•

Expert witness and testimony in Citizens for a Better Environment vs. Union Oil Corp.et al., 20002001.

•

Expert witness and testimony to International Tribunal in the Arbitration Under Chapter 11 of the
North American Free Trade Agreement (NAFTA) Between Methanex Corporation (Canada) and
The United States of America, World Bank, Washington, D.C., June 11, 2004.

•

Expert witness and testimony to arbitration panel in City of Santa Monica vs. Shell Oil Co. et al.,
2005.

•

Expert witness and testimony in Fruitridge Vista Water Company vs. ARCO et al., 2007.

•

Expert witness and testimony in County of Suffolk and Suffolk County Water Authority vs.
Amerada Hess Corp., et al., and Unitied Water New York vs. Amerada Hess Corp, et al., 2008.

•

Expert witness and testimony in City of New York v. Amerada Hess Corp., et al. 04 Civ. 3417,
2009.

INVITED TALKS AND LECTURES (2000-present)
Dichotomy of Scale Concepts in Geology and Hydrology, Aquifer Sedimentology Conference (SEPM
special conference), Sante Fe, Sept. 29, 2000.
Heterogeneity, the Pumping Test and the Plume: An Evolution (or Devolution?) of Hydrogeologic
Concepts, Michigan State University, Oct. 19, 2000.
Geologic Heterogeneity and Predictive Uncertainty in Hydrogeologic Models, AGU, Dec. 15, 2000.

2002 Geological Society of America Birdsall-Dreiss Distinguished Lectures
Talk A: Plume Behavior in Heterogeneous Geologic Systems: Natural Attenuation,
Remediation, and the Role of Diffusion
Talk B: Groundwater Vulnerability and the Meaning of Groundwater Age Dates
Talk C: A Geologic Approach to Simulation of Subsurface Hydrology
Total institutions visited: 45 (37 Universities; 2 USGS offices; 1 USGS workshop; 2 state surveys; 1 GS
of Canada office; 1 national lab; 1 consulting firm; 1 state professional society)
Total lectures delivered: 56
All lectures were by invitation.

Date of Lecture
1. January 14
2. January 15
3. January 16
4. January 17
5. January 18
6. January 28
7. January 29
8. February 1
9. February 5
10. February 6
11. February 12
12. February 19
13. February 21
14. February 25
15. February 27
16. March 1
17. March 4
18. March 4
19. March 12
20. March 13
21. March 14
22. April 2
23. April 3
24. April 9
25. April 10
26. April 11
27. April 23
28. April 24
29. April 24
30.
31.
32.
33.
34.
35.
36.
37.
38.
39.
40.
41.
42.
43.
44.
45.

April 30-May 1
May 2
May 3
September 13
September 16
September 17
September 19
September 23
October 4
October 7
October 10
October 16
October 17
October 18
October 29
November 1113

Location
Indiana University (Bloomington)
Wright State University (Dayton)
University of Kentucky, Lexington
Ohio State University (Columbus)
University of Akron
University of Mississippi (Oxford)
University of Memphis
University of Tennessee, Knoxville
Clemson University
Georgia State University (Atlanta)
Fresno State University
Boise State University
University of British Columbia
Louisiana State University (Baton Rouge)
Texas A&M University
University of Texas at Austin
Bureau of Economic Geology
Intera, Inc.
Penn State University
MIT
University of New Hampshire
Iowa State University
University of Wisconsin
University of Waterloo
Syracuse University
Canadian Geological Survey (Ottawa)
Northern Arizona University
USGS mtg. Tucson
University of Arizona
Ty Ferre
New Mexico Tech
Los Alamos National Laboratory
University of New Mexico (Albuquerque)
University of Nevada Reno
Johns Hopkins University
Delaware Geological Survey, U. of Delaware
University of Virginia
USGS – Reston
University of Hong Kong, China
University of Chang’an, Xian, China
Chinese University of Geosciences, Beijing, China
Stanford University
USGS – Menlo Park
University of California, Santa Cruz
GSA Denver.
University of Hawaii

Talk
B, C
C, A
B
C
B
C
B
A, B
A
A
B
C
B
B
A
A
B
C
B
B
B, A
A, B
A, B
B
C
C
B
C
A
C, B
A
C, B
B
B
C
A, B
B
B
B
B
A
C
B
B
A, B, C

Influence of Paleosols on Fluid Flow and Transport: Perspective on Alluvial Complexity and
Hydrogeology, Presentation before the DOE Nuclear Waste Technical Review Panel, Las Vegas, NV,
Jan. 28, 2003.
Review of IGSM, and Perspective on Future Incarnations, Annual Meeting of CA Bay-Delta Modeling
Forum, Asilomar, CA, Feb. 2003.
Hydrostratigraphy and Contaminant Transport Beneath Rancho Cordova and Vicinity….
Plus Discussion of Cosumnes Phenomena, American River Conference, CSU Sacramento, Nov.
2003.
Perspective on Heterogeneity, Groundwater Monitoring, and Natural Attenuation, Geological Society of
America Annual Meeting, Nov. 2003, Seattle, WA.
Perched Aquifer Hydrology: An Unexplored Frontier?, Geological Society of America Annual Meeting,
Nov. 2003, Seattle, WA.
Scientific Perspective on Groundwater, California Water Forum (Jay Lund), April 14, 2004.
Analysis of Human Exposure to Perchlorate Through Coupled Modeling of Groundwater and the Surface
Distribution System, US EPA Superfund Conference, Oct. 19, 2004, Sacramento, CA.
Fish Passage, Groundwater and Base Flow - The Case of the Cosumnes River. CALFED Science
Conference, Oct. 2004, Sacramento, CA.
The Stochastic Facies Model of Heterogeneity: A Necessary Tool or Unnecessary Tedium?, Geological
Society of America Annual Meeting, Nov. 2004, Denver, CO.
The Role of Perched Aquifers in Hydrological Connectivity and Biogeochemical Processes in Vernal
Pool Landscapes, Great Central Valley, California. Presentation to Sacramento Parks and Recreation
Board, Nov. 2004.
Contaminant Transport in Groundwater, With Special Focus on MTBE, HYD 198 Lecture (Peter Hernes’
class), Feb. 24, 2005.
Contaminant Transport in Groundwater, With Special Focus on MTBE, HYD 198 Lecture (Peter Hernes’
class), Mar. 3, 2005.
Keynote: Washington Hydrogeology Symposium, April 14, 2005, Tacoma, WA: Groundwater
Vulnerability and the Meaning of Age Dates.
Recent Perspectives on Groundwater Age: Estimation, Modeling, and Water Quality Sustainability, 25th
Biennial Conf. & CA Groundwater Resources Association Meeting, Oct. 2005, Sacramento, CA.
The Hydrofacies Approach and Why ln K σ2 <5-10 is Unlikely, American Geophysical Union Fall
Meeting, Dec. 2004.
Groundwater Quality Sustainability, Creeping Normalcy, and a Research Agenda, Geological Society of
America Annual Meeting, Oct. 17, 2005, Salt Lake City, UT.

Contaminant Transport in Groundwater, With Special Focus on MTBE, SAS 8, Jan. 2006.
Keynote: Computational Methods in Water Resources Conference, June 19, 2006, Copenhagen,
Denmark: If Heterogeneity Is So Important, Why Do We Still Ignore It?
Practical Approach For Dealing With Heterogeneity and Why It Is Essential for Modeling Transport,
California Groundwater Resources Association Conference on High-Resolution Site Characterization
and Monitoring, November 15, 2006.
Contaminant Hydrogeology: Rejuvenate the Research Agenda, Geological Society of America Annual
Meeting, Oct. 23,2006, Philadelphia, PA.
Keynote: If Heterogeneity Is So Important, Why Do We Still Ignore It?, DOE Environmental

Remediation Science Program, Oct. 23, 2006, Oak Ridge TN.
Contaminant Transport in Groundwater, With Special Focus on MTBE, SAS 8 Guest Lecture, Jan. 2007.
What’s in a Plume Model and Why Should We Care?, Geological Society of America Annual Meeting,
Oct. 29, 2007, Denver, CO.
Motivation of Synthesis, With an Example on Groundwater Quality Sustainability, American
Geophysical Union Fall Meeting, Dec. 12, 2007, San Francisco, CA.
The Graduate Group Model: An Innovative Structure That Is Still Maturing, Annual Meeting, ASCE
Environmental and Water Resources Institute, May 15, 2008, Honolulu, Hawaii.
Climate Change, Water Resources and Deep CO2 Sequestration, UCD Summer Science Institute, July 9,
2008, Davis, CA.
Groundwater Contaminant Transport and the Utility of a Subsurface Observatory, Consortium of
Universities for the Advancement of Hydrologic Sciences (CUAHSI) Biennial Colloquium, July 14,
2008, Boulder, CO.
Climate Change, Water Resources and Deep CO2 Sequestration, Future Farmers of America, August 26,
2008, Davis, CA.
Keynote: New Paradigms for Groundwater and Surface Water Interaction, HydroPredict 2008,
International Interdisciplinary Conference on Predictions for Hydrology, Ecology, and Water
Resources Management: Using Data and Models to Benefit Society, Prague, Czech Republick, Sept.
16, 2008.
Climate Change and Subsurface Storage, Conference Honoring Shlomo Neuman’s 70th Birthday, Oct. 25,
2008, Tucson, AZ.
Subsurface Storage and Recovery: Perspective on Climate Change and Sustainability of Groundwater
Quantity and Quality in CA, Joint Monthly Meeting of American Water Resources Association and
California Groundwater Resources Association, Nov. 8, 2008, Sacramento, CA.
Subsurface Storage and Recovery: Perspective on Climate Change and Sustainability of Groundwater
Quantity and Quality in CA, University of Concepcion, Chile, March 18, 2009.

Subsurface Storage and Recovery: Perspective on Climate Change and Sustainability of Groundwater
Quantity and Quality in CA, Special webinar presentation to CA state agency Geosymposium, April
28, 2009.
Subsurface Storage and Recovery: Perspective on Climate Change and Sustainability of Groundwater
Quantity and Quality in CA, Bay Area Chapter of California Groundwater Resources Association,
May. 20, 2009, Berkeley, CA.
Keynote: The Disconnect Between Monitoring and Modeling, NOVCARE Conference, Leipzig,
Germany, May 13, 2009.

PUBLICATIONS
Articles and Books
Fleckenstein, J.H. and G.E. Fogg. 2008. Efficient upscaling of hydraulic conductivity in
heterogeneous alluvial aqufers, Hydrogeology Journal, DOI 10.1007/s10040-008-0312-3.
Niswonger, R.G. and G.E. Fogg. 2008. Influence of perched groundwater on baseflow, Water
Resources Research, 44, W03405, doi:10.1029/2007WR006160.
Niswonger, R.G., D.E. Prudic, G.E. Fogg, D.A. Stonestrom and E.M. Buckland. 2008. Method
for estimating spatially variable seepage loss and hydraulic conductivity in intermittent
and ephemeral streams, Water Resources Research, 44, 10.1029/2007WR006626.
Rains, M.C. G.E. Fogg, T. Harter, R.A. Dahlgren, and R.J. Williamson. 2008. Geological
control of physical and chemical hydrology in vernal pool wetlands, Central Valley,
California, Wetlands Journal, 28(2), p. 347-362
LaBolle, E.M., Leaist, D.G., J. Gravner, G.E. Fogg and J.W. Eweis. 2008. Diffusive
fractionation of isotopes in groundwater, Water Resources Research, 44, W07405,
doi:10.1029/2006WR005264.
Eweiss, J.B., E.M. LaBolle, D. Benson and G.E. Fogg. 2007. The role of volatilization in changing TBA
and MTBE concentrations at MTBE contaminated sites, Environmental Science and Technology,
41(19): 6822-6827.
DePaulo, D.J., F.M. Orr Jr., S.M. Benson, M. Celia, A. Felmy, K.L. Nagy, G.E. Fogg, R. Snieder, J.
Davis, K. Pruess, J. Friedmann, M. Peters, N. B. Woodward, P. Dobson, K. Talamini, M. Saarni.
2007. Basic research needs for geosciences: Facilitating 21st century energy systems. Office of Basic
Energy Sci., U.S. Dept. of Energy, 287 p.
Lee, S.-Y., S.F. Carle and G.E. Fogg. 2007. Geologic heterogeneity and a comparison of two
geostatistical models: sequential Gaussian and transition probability-based geostatistical simulation,
Advances in Water Resources, 30, doi:10.1016/j.advwatres.2007.03.005.
LaBolle, E.M., G.E. Fogg and E.B. Eweis. 2006. Diffusive fractionation of 3H and 3He in groundwater
and its impact on groundwater age estimates, Water Resources Research, 42, W07202,
doi:10.1029/2005WR004756.
Fogg, G.E. and E.M. LaBolle. 2006. Motivation of synthesis, with an example on groundwater quality
sustainability, Water Resources Research (special forum on synthesis in the hydrologic sciences),
42, W03S05, doi:10.1029/2005WR004372.
Fleckenstein, J.H., R.G. Niswonger and G.E. Fogg. 2006. River-aquifer interactions, geologic
heterogeneity, and low-flow management, Ground Water, doi: 10.1111/j.1745-6584.2006.00190.x,
1-16.
Rains, M.C., G.E. Fogg, T. Harter, R.A. Dahlgren, and R.J. Williamson, 2006, The role of perched
aquifers in hydrological connectivity and biogeochemical processes in vernal pool landscapes,
Central Valley, California. Hydrological Processes, 20(5), p. 1157-1175.

Knudby, C., J. Carrera, J.D. Bumgardner and G.E. Fogg, 2006, An empirical method for the evaluation of
effective conductivity of binary media. Advances in Water Resources, 29, p. 590-604.
Purkey, D.R., W.W. Wallender, N. Islam, G.E. Fogg and B. Sivakumar. 2006. Describing near surface,
transient flow processes in unconfined aquifers below irrigated lands: A deforming finite element
model for heterogeneous aquifers. Journal of Hydrology, 330, 10.1016/j.jhydrol.2006.04.006.
Purkey, D.R., W.W. Wallender, G.E. Fogg and B. Sivakumar, 2004, Describing near surface, transient
flow processes in unconfined aquifers below irrigated lands: Model application in the western San
Joaquin Valley. Journal of Irrigation and Drainage Engineering, Nov./Dec. p. 451-459.
Weissmann, G.S., Y. Zhang, Fogg, G.E., and Mount, J.F., 2004, Influence of incised valley fill deposits
on hydrogeology of a glacially-influenced, stream-dominated alluvial fan. SEPM special publication
on Hydrogeophysics and Hydrostratigraphy (edited by John Bridge and David Hyndman), p. 15-28.
Fleckenstein, J.H., Anderson, M, Fogg, G.E. and Mount, J., 2004, Managing Surface watergroundwater to restore fall flows in the Cosumnes River. Journal of Water Resources
Management and Planning, July/August, 301-310 (received Best Paper Award).

Hopmans, J.W. and G.E. Fogg. 2003. Soil water flow under saturated conditions. Encyclopedia of Water
Science, Marcel Dekker, New York, 871-874.
Zhang, Y. and G.E. Fogg. 2003. Simulation of multi-scale heterogeneity of porous media and parameter
sensitivity analysis. Science in China (Series E), 46(5), 459-474.
Lu, S., F.J. Molz, G.E. Fogg and J.W. Castle. 2002. Combining stochastic facies and fractal models for
representing natural heterogeneity. Hydrogeology Journal, 10: 475-482.
Knudby, C., J. Carrera, and G.E. Fogg. 2002. An empirical method for the evaluation of the equivalent
conductivity of low-permeable matrices with high-permeable inclusions. IAHR/IAHS International
Groundwater Symposium, March 25-28, Berkeley, CA.

Weissmann, G.S., Z. Yong, .E.M. LaBolle, and G.E. Fogg. 2002. Dispersion of groundwater age
in an alluvial aquifer system, Water Resources Research. 38(10):16.1-16.8.

Weissmann, G.S., J.F. Mount, and G.E. Fogg. 2001. Glacially-driven cycles in accumulation and
sequence stratigraphy of a stream-dominated alluvial fan, San Joaquin Valley, California. Journal of
Sedimentary Research Section A-Sedimentary Petrology & Processes. 72(2): 241-250.
Fleckenstein, J., E. Suzuki and G. Fogg. 2001. Options for conjunctive water management to restore fall
flows in the Cosumnes River basin, California. In: M.A. Mariño and S.P. Simonovic (ed.),
Integrated Water Resources Management, IAHS Publication no. 272, IAHS Press, Wallingford,
U.K., (ISBN 1-901502-76-7).

LaBolle, E.M. and G.E. Fogg. 2001. A Review of California’s Integrated Groundwater and SurfaceWater Model (IGSM). In: Seo, H.S., E. Poeter, C. Zheng, and O. Poeter (ed.), MODFLOW 2001
and Other Modeling Odysseys – Conference Proceedings, September 11-14, 2001, 349-355.
LaBolle, E.M. and G.E. Fogg, 2001, Role of molecular diffusion in contaminant migration and recovery
in an alluvial aquifer system, Transport in Porous Media (Special Issue on Modeling Dispersion),
42, p. 155-179.
Fogg, G. E., S.F. Carle and C. Green, 2000, A connected-network paradigm for the alluvial aquifer
system, in press, in Zhang, D. and Winter, C.L. (eds.), Theory, Modeling and Field Investigation in
Hydrogeology: a Special Volume in Honor of Shlomo P. Neuman’s 60th Birthday, Geological
Society of America Special Paper 348, p. 25-42.
LaBolle, .E.M., J. Quastel, G.E. Fogg and J. Gravner. 2000. Diffusion processes in composite porous
media and their numerical integration by random walks: Generalized stochastic differential
equations with discontinuous coefficients. Water Resources Research 36(3): 651-662.
LaBolle, E.M. and G.E. Fogg. 2000. Reply to comment on "Diffusion theory for transport in porous
media: Transition probability densities of diffusion processes corresponding to advection-dispersion
equations" by Philippe Ackerer. Water Resources Research 36(): .
Scow, K.M., G.E. Fogg, D.E. Hinton and M.L. Johnson (ed.), 2000, Integrated Assessment of
Ecosystem Health, Lewis Publishers, Boca Raton, 358 p.
Ruud, N.C., T. Harter, G. Weissmann, G.E. Fogg, 2000, Conditional geostatistical modeling of an
alluvial aquifer system characterized by poor-quality shallow groundwater, in Proceedings
of the ModelCARE '99 conference; Calibration and reliability in groundwater modelling;
coping with uncertainty Stauffer, F. et al. (ed.), IAHS-AISH Publication, Vol. 265, p. 360368.
Burow, K.R., S.Y. Panshin, N.M. Dubrovsky, D. Van Brocklin, and G. E. Fogg, 1999, Evaluation
of processes affecting 1,2-dibromo-3-chloropropane (DBCP) concentrations in ground
water in the eastern San Joaquin Valley, California: analysis of chemical data and groundwater flow and transport simulations, U.S. Geological Survey Water-Resources
Investigations Report 99-4059, 57 p.
Weissmann, G.S. and G.E. Fogg, 1999, Multi-scale alluvial fan heterogeneity modeled with
transition probability geostatistics in a sequence stratigraphic framework, J. of Hydrology
226, p. 48-65.
Weissmann, G.S., S.F. Carle and G.E. Fogg, 1999, Three-dimensional hydrofacies modeling
based on soil surveys and transition probability geostatistics. Water Resources Research.
35(6). 1761-1770.

Fogg, G. E., E.M. LaBolle, and G.S. Weissmann. 1999, Groundwater vulnerability assessment:
Hydrogeologic perspective and example from Salinas Valley, California. In Corwin, D.L.,
K. Loague and T.R. Ellsworth, Application of GIS, Remote Sensing, Geostatistical and
Solute Transport Modeling. AGU Geophysical Monograph.
Fogg, G.E., D.E. Rolston, D.L. Decker, D.T. Louie, and M.E. Grismer, 1998, Spatial variation in nitrogen
isotope values beneath nitrate contamination sources. Ground Water, 6(3), p. 418-426.
Carle, S.F., E.M. LaBolle, G.S. Weissmann, D. Van Brocklin and G.E. Fogg, 1998, Geostatistical
simulation of hydrostratigraphic architecture: a transition probability/Markov approach, in Concepts
in Hydrogeology and Environmental Geology No. 1, SEPM (Society for Sedimentary Geology)
Special Publication, p. 147-170.
Fogg, G.E., C.D. Noyes and S.F. Carle, 1998, Geologically-based model of heterogeneous hydraulic
conductivity in an alluvial setting, Hydrologeology Journal, 6(1), p. 131-143.
LaBolle, E.M., J. Quastel and G.E. Fogg, 1998, Diffusion theory for transport in porous media: Transition
probability densities of diffusion processes corresponding to advection-dispersion equations, Water
Resources Research. 34(7), p. 1685-1693.
Carle, S.F. and G.E. Fogg, 1997, Modeling spatial variability with one- and multi-dimensional
continuous-lag Markov chains, Mathematical Geology, 29(5), p. 891-918.
LaBolle, E.M., G.E. Fogg and A.F.B. Tompson, 1996, Random-walk simulation of transport in
heterogeneous porous media: Local mass-conservation problem and implementation methods. Water
Resources Research. v. 32 no. 3. p. 583-393.
Carle, S.F. and G.E. Fogg, 1996, Transition probability-based indicator geostatistics. Mathematical
Geology. v. 28 no. 4 p. 453-476.
Rolston, D.E., G.E. Fogg, D.L. Decker, D.T. Louie and M.E. Grismer, 1996, Nitrogen isotope
ratios: a tool for identifying nitrate contamination sources. California Agriculture. v. 50 no.
2. p. 32-36.
Hoyle, B.L., K.M. Scow, G.E. Fogg and J.L. Darby, 1995, Effect of Carbon:nitrogen ratio on kinetics of
phenol biodegradation by Acinetobacter johnsonii in saturated sand. Biodegradation. v. 6. p. 283293.
Chen, J., J.W. Hopmans, and G.E. Fogg, 1995, Sampling design for soil moisture measurements in large
field trials. Soil Science, p. 155-161.
Tarboton, K.C., W.W. Wallender, G.E. Fogg, K. Belitz, 1995, Kriging of regional hydrogeologic
properties in the western San Joaquin Valley. Hydrogeology Journal, v. 3, no. 1, p. 5-23.
Fogg, G.E. and D.R. Nielsen, 1994, Modeling contaminant transport in the vadose zone: Perspective on
state of the art, in Wilson, G., L. Everett, and S. Cullen, Vadose Zone Characterization and
Monitoring: Principles, Methods, and Case Studies. Lewis Publishers, Chelsea, Michigan.
Ritzi, R. W., D. F. Jayne, A. J. Zahradnik, A. A. Field, and G. E. Fogg. 1994. Geostatistical modeling of
heterogeneity in glaciofluvial, buried-valley aquifers. Ground Water, 32(4), p. 666-674.

Loaiciga, H. A., Randall J. Charbeneau, Lorne G. Everett, Graham E. Fogg, Benjamin F. Hobbs, Shahrok
Rouhani. 1992. Review of groundwater quality monitoring network design. Journal of Hydraulic
Engineering, Vol. 118 No. 1, p. 11-37.
Fogg, G. E., F. J. Lucia, and R. K Senger, 1991, Stochastic simulation of interwell-scale
heterogeneity for improved prediction of sweep efficiency in a carbonate reservoir, in Lake,
L. W., H. B. Carroll, Jr., and T. C. Wesson, Reservoir Characterization II, Academic Press,
Inc., San Diego, pages 355-381.
Senger, R.K., F.J. Lucia, C. Kerans, M.A. Ferris, and G.E. Fogg, 1991, Dominant control on
reservoir-flow behavior in carbonate reservoirs as determined from outcrop studies, in,
Linville, B., T.E. Burchfield, and T.C. Wesson, Reservoir Characterization III, Penwell Books,
Tulsa, p. 107-150.
Fogg, G. E., 1990, Architecture and interconnectedness of geologic media: role of the lowpermeability facies in flow and transport. in Neuman, S.P. and Neretnieks, I., Hydrogeology
of Low Permeability Environments, from a special symposium at the 28the International
Geological Congress, Washington, D. C., July 9-19, 1989, Verlag Heinz Heise, p. 19-40.
Senger, R. K. and G. E. Fogg. 1990. Stream functions and equivalent freshwater heads for
modeling regional flow of variable-density groundwater, 1. Review of theory and
verification. Water Resources Research 26(9):2089-2096.
Senger, R. K. and G. E. Fogg. 1990. Stream functions and equivalent freshwater heads for
modeling regional flow of variable-density groundwater, 2. Applications and implications
for modeling strategy. Water Resources Research 26(9):2097-2106.
Kittridge, M. G., L. W. Lake, F. J. Lucia, and G. E. Fogg. 1990. Outcrop/subsurface comparisons
of heterogeneity in the San Andres Formation. Formation Evaluation (published by Society
of Petroleum Engineers), September. p. 233-240.
Lucia, F. J. and G. E. Fogg. 1990. Geologic/stochastic mapping of heterogeneity in a carbonate
reservoir. Journal of Petroleum Technology (published by Society of Petroleum Engineers),
October, p. 1298-1303.
Fogg, G. E., 1989, Emergence of geologic and stochastic approaches for characterization of
heterogeneous aquifers: Proceedings of the Conference “New Field Techniques for
Quantifying the Physical and Chemical Properties of Heterogeneous Aquifers”, Dallas,
Texas, March 20-23, p. 1-17.
Senger, R. K. and G. E. Fogg, 1987, Regional underpressuring in deep brine aquifers, Palo Duro
Basin, Texas, 1. effects of hydrostratigraphy and topography: Water Resources Research, v.
23, no. 8, p. 1481-1493.
Senger, R. K., C. W. Kreitler, and G. E. Fogg, 1987, Regional under pressuring in deep brine
aquifers, Palo Duro Basin, Texas, 2. the effect of Cenozoic basin development: Water
Resources Research, v. 23, no. 8 p. 1494-1504.

Fogg, G. E., 1986, Groundwater flow and sand-body interconnectedness in a thick, multipleaquifer system: Water Resources Research, v. 22, no. 5, p. 679-694.
Fogg, G. E. and R. J. Charbeneau, 1985, Impacts of lignite mining on groundwater in Texas, in
Smerdon, E. T. and W. R. Jordan, eds., Issues in groundwater management: Center for
Research in Water Resources, Water Resources Symposium No. 12, p. 361-377.
Fogg, G. E. and R. K. Senger, 1985, Automatic generation of flow nets with conventional
ground-water modeling algorithms: Ground Water, v. 23, no. 3, p. 336-344.
Fogg, G. E., W. R. Kaiser, M. L. Ambrose, and G. L. Macpherson, 1983, Regional aquifer
characterization for deep-basin lignite mining, Sabine Uplift area, northeast Texas: Sixth
International Coal Utilization Exhibition & Conference, v. 6, Lignite developments and
utilization, p. 22-49. Reprinted, 1984 in Kube, W. R, E. A. Sondreal, and C. D. Rao, eds.,
12th Biennial Lignite Symposium: Technology and Utilization of Low-Rank Coals
Proceedings: U. S. Department of Energy, Office of Fossil Energy, V. 1, 364-389.
Fogg, G. E. and C. W. Kreitler, 1981, Ground-water hydrology around salt domes in the East
Texas Basin: a practical approach to the contaminant transport problem: Association of
Engineering Geologists Bulletin, v. 18, no. 4, p. 387-411.
Neuman, S. P., G. E. Fogg and E. A. Jacobson, 1980, A statistical approach to the inverse
problem of aquifer hydrology: 2. case study: Water Resources Research, v. 16, no. 1, p. 3358.
Published Reports and Bulletins (refereed)
Green, C.T., E.M. LaBolle, and G.E. Fogg. Random walk particle tracking for simulating reactive
transport in heterogeneous aquifers: effects of concentration averaging. IAHS International
Groundwater Symposium. March 2002. Berkeley, CA.
Fogg, G. E., W. R. Kaiser, and M. L. Ambrose. 1991. The Wilcox Group and Carrizo Sand (Paleogene) in
the Sabine Uplift Area, Texas: Ground-Water Hydraulics and Hydrochemistry. The University of
Texas at Austin, Bureau of Economic Geology, Special Publication, 70 p., 19 plates.
Fogg, G. E. and F. J. Lucia. 1990. Reservoir modeling of restricted platform carbonates:
Geologic/geostatistical characterization of interwell-scale reservoir heterogeneity, Dune
Field, Crane County, Texas. The University of Texas at Austin, Bureau of Economic
Geology Report of Investigations No. 190. 66 pages.
Fogg, G. E., 1989, Stochastic analysis of aquifer interconnectedness: Wilcox Group, Trawick
area, East Texas: The University of Texas at Austin, Bureau of Economic Geology, Report of
Investigations No. 189, 68 p.
Bebout, D. G., Lucia, F. J., Hocott, C. R., Fogg, G. E., and Vander Stoep, G. W., 1987,
Characterization of the Grayburg Reservoir, University Lands Dune Field, Crane County,

Texas: The University of Texas at Austin, Bureau of Economic Geology Report of
Investigations No. 168, 98 p.
Senger, R. K., G. E. Fogg, and C. W. Kreitler, 1987, Effects of hydrostratigraphy and basin
development on hydrodynamics of the Palo Duro Basin, Texas: The University of Texas at
Austin, Bureau of Economic Geology Report of Investigations No. 165, 48 p.
Fogg, G. E. and D. A. Prouty, 1986, Ground-water hydraulics of the Wilcox-Carrizo aquifer
system, in Kaiser, W. R. and others, Geology and ground-water hydrology of deep-basin
lignite in the Wilcox Group of East Texas: The University of Texas at Austin, Bureau of
Economic Geology Special Publication. p. 60-70.
Fogg, G. E., 1986, Hydrogeologic drilling and testing program, Sabine Uplift area, in Kaiser, W.
R., and others, Geology and ground-water hydrology of deep-basin lignite in the Wilcox
Group of East Texas: The University of Texas at Austin, Bureau of Economic Geology
Special Publication, p. 77-84.
Fogg, G. E. and P. E. Blanchard, 1986, Empirical relations between Wilcox ground-water quality
and electric log resistivity, Sabine Uplift area, in Kaiser, W. R., and others, Geology and
ground-water hydrology of deep-basin lignite in the Wilcox Group of East Texas: The
University of Texas at Austin, Bureau of Economic Geology Special Publication, p. 115-118.
Fogg, G. E., and W. R. Kaiser, 1986, Regional hydrogeologic considerations for deep-basin
lignite development in Texas, in Kaiser, W. R., and others, Geology and ground-water
hydrology of deep-basin lignite in the Wilcox Group of East Texas: The University of Texas
at Austin, Bureau of Economic Geology Special Publication, p. 57-59.
Kaiser, W. R., M. L. Ambrose, W. B. Ayers, Jr., P. E. Blanchard, G. F. Collins, G. E. Fogg, D. L.
Gower, C. L. Ho, C. S. Holland, M. L. W. Jackson, C. M. Jones, Lewis, A. H., G. L.
Macpherson, C. A. Mahan, A. H. Mullin, D. A. Prouty, S. J. Tewalt, and S. W. Tweedy, 1986,
Geology and ground-water hydrology of deep-basin lignite in the Wilcox Group of East
Texas: The University of Texas at Austin, Bureau of Economic Geology Special Publication,
182 p.
Fogg, G. E., W. R. Kaiser, M. L. Ambrose, and G. L. Macpherson, 1983, Regional aquifer
characterization for deep-basin lignite mining, Sabine Uplift area, northeast Texas: The
University of Texas at Austin, Bureau of Economic Geology Geological Circular 83-3, 30 p.
Fogg, G. E., S. J. Seni, and C. W. Kreitler, 1983, Three-dimensional ground-water modeling in
depositional systems, Wilcox Group, Oakwood salt dome area, East Texas: The University
of Texas at Austin, Bureau of Economic Geology Report of Investigations No. 133, 55 p.
Senger, R. K. and G. E. Fogg, 1983, Regional modeling of ground-water flow in the Palo Duro
Basin, Texas Panhandle, p. 109-115, in Gustavson, T. C. and others, Geology and
geohydrology of the Palo Duro Basin, Texas Panhandle, a report on the progress of nuclear
waste isolation feasibility studies (1982): The University of Texas at Austin, Bureau of
Economic Geology Geological Circular 83-4, 156 p.

Fogg, G. E. and C. W. Kreitler, 1982, Ground-water hydraulics and hydrochemical facies in
Eocene aquifers of the East Texas Basin: The University of Texas at Austin, Bureau of
Economic Geology Report of Investigations No. 127, 75 p.
Simpkins, W. W. and G. E. Fogg, 1982, Preliminary modeling of ground-water flow near salt
dissolution zones, Texas Panhandle, P. 130-137, in Gustavson, T. C., and others, Geology
and geohydrology of the Palo Duro Basin, Texas Panhandle, a report on the progress of
nuclear waste isolation feasibility studies (1981): The University of Texas at Austin, Bureau
of Economic Geology Geological Circular 82-7, 212 p.
Fogg, G. E. 1981, Aquifer modeling of the Oakwood salt dome area, p. 139-149, in Kreitler, C.
W., and others, Geology and geohydrology of the East Texas Basin, a report on the progress
of nuclear waste isolation feasibility studies (1980): The University of Texas at Austin,
Bureau of Economic Geology Geological Circular 81-7, 207 p.
Fogg, G. E., 1981, Aquifer testing and monitoring around Oakwood salt dome, East Texas, p.
126-135, in Kreitler, C. W., and others, Geology and geohydrology of the East Texas Basin,
a report on the progress of nuclear waste isolation feasibility studies (1980): The University
of Texas at Austin, Bureau of Economic Geology Geological Circular 81-7, 207 p.
Fogg, G. E., 1981, Fluid-pressure versus depth relationships in the Wilcox-Carrizo aquifer
system, East Texas, p. 115-122, in Kreitler, C. W., and others, Geology and geohydrology of
the East Texas Basin, a report on the progress of nuclear waste isolation feasibility studies
(1980): The University of Texas at Austin, Bureau of Economic Geology Geological
Circular 81-7, 207 p.
Fogg, G. E. 1981, Total dissolved solids in the Eocene Wilcox-Carrizo aquifer around Keechi salt
dome, East Texas, p. 123-125, in Kreitler, C. W., and others, Geology and geohydrology of
the East Texas Basin, a report on the progress of nuclear waste isolation feasibility studies
(1980): The University of Texas at Austin, Bureau of Economic Geology Geological
Circular 81-7, 207 p.
Fogg, G. E., S. J. Seni, and S. P. Dutton, 1981, Core analysis of hydrologic parameters, Eocene
Wilcox aquifer, East Texas, p. 136-138, in Kreitler, C. W., and others, Geology and
geohydrology of the East Texas Basin, a report on the progress of nuclear waste isolation
feasibility studies (1980): The University of Texas at Austin, Bureau of Economic Geology
Geological Circular 81-7, 207 p.
Kreitler, C. W., S. P. Dutton, and G. E. Fogg, 1981, Geochemistry of ground water and cap rock
from Gyp Hill salt dome, South Texas, p. 167-171, in Kreitler, C. W., and others, Geology
and geohydrology of the East Texas Basin, a report on the progress of nuclear waste
isolation feasibility studies (1980): The University of Texas at Austin, Bureau of Economic
Geology Geological Circular 81-7, 207 p.
Fogg, G. E., 1980, Aquifer modeling at Oakwood Dome, p. 35-40, in Kreitler, C. W., and others,
Geology and geohydrology of the East Texas Basin, a report on the progress of nuclear

waste isolation feasibility studies (1979): The University of Texas at Austin, Bureau of
Economic Geology Geological Circular 80-12, 112 p.
Fogg, G. E., 1980, Drilling and hydrogeologic monitoring at Oakwood salt dome, p. 33-34, in
Kreitler, C. W., and others, Geology and geohydrology of the East Texas Basin, a report on
the progress of nuclear waste isolation feasibility studies (1979): The University of Texas at
Austin, Bureau of Economic Geology Geological Circular 80-12, 112 p.
Fogg, G. E., 1980, Regional aquifer hydraulics, East Texas Basin, p. 55-67, in Kreitler, C. W., and
others, Geology and geohydrology of the East Texas Basin, a report on the progress of
nuclear waste isolation feasibility studies (1979): The University of Texas at Austin, Bureau
of Economic Geology Geological Circular 80-12, 112 p.
Fogg, G. E., 1980, Salinity of formation waters, p. 68-72, in Kreitler, C. W., and others, Geology
and geohydrology of the East Texas Basin, a report on the progress of nuclear waste
isolation feasibility studies (1979): The University of Texas at Austin, Bureau of Economic
Geology Geological Circular 80-12, 112 p.
Fogg, G. E., and C. W. Kreitler, 1980, Impacts of salt-brining on Palestine salt dome, p. 46-54, in
Kreitler, C. W., and others, Geology and geohydrology of the East Texas Basin, a report on
the progress of nuclear waste isolation feasibility studies (1979): The University of Texas at
Austin, Bureau of Economic Geology Geological Circular 80-12, 112 p.
Fogg, G. E., C. W. Kreitler, and S. P. Dutton, 1980, Hydrologic stability of Oakwood Dome, p.
30-32, in Kreitler, C. W., and others, Geology and geohydrology of the East Texas Basin, a
report on the progress of nuclear waste isolation feasibility studies (1979: The University of
Texas at Austin, Bureau of Economic Geology Geological Circular 80-12, 112 p.
Kreitler, C. W. and G. E. Fogg, 1980, Geochemistry of ground water in the Wilcox aquifer, p. 7378, in Kreitler, C. W., and others, Geology and geohydrology of the East Texas Basin, a
report on the progress of nuclear waste isolation feasibility studies (1980): The University
of Texas at Austin, Bureau of Economic Geology Geological Circular 80-12, 112 p.
Published Reports and Bulletins (unrefereed)
Weissmann, G.S., LaBolle, E.M., and Fogg, G.E., 2000, Modeling environmental tracer-based
groundwater ages in heterogeneous aquifers, in Bentley, L.R., Sykes, J.F, Brebbia, C.A.,
Gray, W.G., and Pinder, G.F., eds., Computational Methods in Water Resources, Volume 2:
Computational Methods, Surface Water Systems and Hydrology, Proceedings of the XIII
International Conference on Computational Methods in Water Resources, Calgary, Alberta,
Canada, Jun 25-29, 2000, A.A.Balkema, Rotterdam, p. 805-811.
Green, C.T. and G.E. Fogg. 1998. Hydrogeologic factors in wetland function at subalpine pope marsh,
Lake Tahoe. Proceedings of the 5th National Watershed Conference, Reno, Nevada.
Carle, S.F. and G.E. Fogg. 1998. TProGS: A transition probability/Markov approach to geostatistical
modeling and simulation. User’s manual and software. 84 p.

Fogg, G.E., M.E. Meays, J. Trask, C. Green, E.M. Labolle and D.E. Rolston. 1998. Impacts of MTBE
on California groundwater. Contract report submitted to the Governor and Legislature of California
pursuant to Senate Bill 521.
Carle, S.F. and G.E. Fogg, Probability cokriging simulation, in proceedings of 1994 International
Association for Mathematical Geology Annual Conference, Quebec, Canada, Oct. 3-5, p. 62-67.
Fogg, G. E. and A. G. Journel. 1992. Quantifying geological continuity. In: American Association of
Petroleum Geologists Development Geology Reference Manual, Section IV, Statistical Methods.
Mohr, T. K. G., M. L. Davisson, R. E. Criss, and G. E. Fogg, 1992, Small Scale Application of Stable
Isotopes 18O and Deuterium to Delineate Migration Pathways at a Class III Landfill Site, in
Proceedings of the Sixth National Outdoor Action Conference on Aquifer Restoration, Ground-Water
Monitoring, and Geophysical Methods, National Ground-Water Association, May 11-13, 1992.
Senger, R. K., F. J. Lucia, C. Kerans, M. Ferris, and G. E. Fogg. 1992. Geostatistical/geological
permeability characterization of carbonate ramp deposits in San Andres Outcrop, Algerita
Escarpment, New Mexico. Society of Petroleum Engineers paper no. 23967, p. 287 -301.
Fogg, G. E., Lucia, F. J., and Senger, R. K., 1989, Stochastic simulation of interwell-scale
heterogeneity for improved prediction of sweep efficiency in a carbonate reservoir:
Proceedings of the Conference “Second International Reservoir Characterization Technical
Conference”, Dallas Texas, June 25-28.
Fogg, G. E. 1989. Emergence of geologic and stochastic approaches for characterization of
heterogeneous aquifers. Paper published in Proceedings of New Field Techniques for Quantifying
the Physical and Chemical Properties of Heterogeneous Aquifers, Dallas, Texas, March 20-23,
1989. p. 1-17.
Fogg, G. E., E. S. Simpson, and S. P. Neuman, 1979, Aquifer modeling by numerical methods
applied to an Arizona groundwater basin: University of Arizona, Department of
Hydrology and Water Resources Technical Report, Reports on Natural Resource Systems
No. 32.
Rolston, D.E., G.E. Fogg, D.L. Decker, and D.T. Louie, 1994, Nitrogen isotope ratios of natural
and anthropogenic nitrate in the subsurface. Presented at Water Down Under Conference,
Adelaide, Australia, 21-25 November.